Environmental Pollution 189 (2014) 161e168
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Environmental Pollution journal homepage: www.elsevier.com/locate/envpol
High tolerance of subalpine grassland to long-term ozone exposure is independent of N input and climatic drivers Matthias Volk*, Veronika Wolff, Seraina Bassin, Christof Ammann, Jürg Fuhrer Agroscope, Air Pollution/Climate Group, 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 25 November 2013 Received in revised form 25 February 2014 Accepted 28 February 2014
In a seven-year study, we tested effects of increased N and O3 deposition and climatic conditions on biomass of subalpine grassland. Ozone risk was assessed as exposure (AOT40) and as stomatal flux (POD0,1). We hypothesized that productivity is higher under N- and lower under O3 deposition, with interactions with climatic conditions. Aboveground biomass was best correlated with growing-degree days for May (GDDMay). Nitrogen deposition increased biomass up to 60% in the highest treatment, and 30% in the lowest addition. Also belowground biomass showed a positive N-response. Ozone enrichment had no effect on biomass, and no interaction between O3 and N was observed. Growth response to N deposition was not correlated to GDDMay or precipitation, but indicated a cumulative effect over time. Productivity of subalpine grassland is tolerant to increasing ozone exposure, independent of N input and climatic drivers. N deposition rates at current critical loads, strongly increase the grassland yield. Ó 2014 Elsevier Ltd. All rights reserved.
Keywords: Air pollution Nitrogen Critical load/level Phytotoxic ozone dose POD Productivity
1. Introduction The grassland habitat type s.l. covers some 40% of the terrestrial surface (White et al., 2000), and in subalpine regions of central Europe, highly diverse, semi-natural grasslands represent a major land cover type. Apart from their conservation value, these grasslands act as an important source of feed for livestock and are essential for the esthetic value of the landscape. They deliver ecosystem services like soil erosion and water quality protection, as well as sink potential for the greenhouse gas CO2. Understanding the sensitivity of these systems to changes in environmental conditions is crucial for evaluating future hazards and designing suitable management options and policies, to reduce the risk of losing their important ecological and agronomic functions. Changes in climate and increasing air pollution are among the most critical long-term risks to be considered. Air pollution by tropospheric ozone (O3) and reactive nitrogen (Nr) compounds, both resulting from industrial, traffic and agricultural emissions, is widespread, with potentially harmful effects across the northern hemisphere (Fuhrer et al., 2003; Ainsworth et al., 2012). Far from sources of O3 precursors, remote sites in the mid-latitudes are exposed to annual mean background
* Corresponding author. E-mail address: [email protected] (M. Volk). http://dx.doi.org/10.1016/j.envpol.2014.02.032 0269-7491/Ó 2014 Elsevier Ltd. All rights reserved.
concentrations of 20e45 ppb O3, representing approximately a doubling during the last century (Marenco et al., 1994; Vingarzan, 2004). Annual mean O3 concentrations are particularly high at sites above approximately 1000 m a.s.l., where diurnal variations in concentration are small due to reduced nocturnal O3 degradation and deposition (Chevalier et al., 2007). Least conservative scenarios expect mean annual background concentrations of O3 to possibly reach 80e90 ppb by 2100 (Vingarzan, 2004; Derwent et al., 2007), due to changes in emissions and climatic conditions (cf. Fuhrer, 2009). Present atmospheric N deposition rates in remote mountain areas are around 5 kg N ha1 yr1, but reach up to 60 kg N ha1 yr1 in regions with intensive livestock farming (Rihm and Kurz, 2001), thus exceeding critical loads defined to protect natural vegetation (Achermann and Bobbink, 2003). Due to long-range transport and increasing emission, N deposition over vast areas is predicted to almost double globally by 2050 (Galloway et al., 2004). Given these trends, it is evident that ecosystems in remote areas are e and will continue to be e at risk from air pollution. In individual plant species, elevated O3 has been found to cause visible injury (Bungener et al., 1999a; Bermejo et al., 2003), premature leaf senescence and damage to the photosynthetic system (Pell et al., 1997), guard cell damage (Mills et al., 2009; Hayes et al., 2012) and reduced growth (Hayes et al., 2006; Gimeno et al., 2004; Wyness et al., 2011; Ainsworth et al., 2012; Bungener et al., 1999b). It reduces plant productivity by entering leaves through the
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stomata, generating other reactive oxygen species and thus causing oxidative stress (cf. Ainsworth et al., 2012). However, translating experimental results from individual species to field conditions is difficult, and understanding community level responses requires experiments under natural conditions. Only limited experimental data exist, that describe effects of long-term O3 exposure on productivity and species composition in semi-natural grasslands. For instance, Volk et al. (2006) observed a significant decline in productivity over a 5-year period of exposure to elevated O3. Later Stampfli and Fuhrer (2010) found that the observed effect was confounded by plot management changes. From the few available field studies it was suggested that growth reduction due to O3 may be limited to productive habitats with high N availability (Bassin et al., 2007a). Typically, semi-natural ecosystems respond to high N loads with increased productivity and altered species composition (LeBauer and Treseder, 2008; Bobbink et al., 2010; Bassin et al., 2013). N-induced changes in the dominance of species may act as an additional factor changing the community sensitivity to additional stresses such as O3 or climate extremes (Bobbink et al., 1998). But it remains unclear how O3 and N availability interact with each other in semi-natural grassland communities. N-addition may increase plant sensitivity to O3 via stimulation of gas exchange leading to higher levels of canopy O3 uptake and higher growth rates. Alternatively, sensitivity may decrease through better oxidation damage repair, or via mitigation of sink limitation (Andersen, 2003). Finally, effects of air pollutants may depend on prevailing weather conditions, which strongly vary between and within seasons, and meteorological effects on productivity act not only immediately, but also delayed for several growing seasons via complex carry over effects (Arnone et al., 2008). Thus, only longerterm experiments may help to disentangle the complex interactions between different atmospheric factors. We conducted a free-air fumigation experiment for seven years to investigate the effects of combined O3- and N-deposition on subalpine grassland. During the initial three years of the experiment increased N-input caused a strong growth response, but there was no significant response to elevated O3 (Bassin et al., 2007b). It was then suggested that the lack of a response to O3 could be due to the short duration of the experiment, high intrinsic detoxification capacity under conditions prevailing at high altitudes, or high resistance to diffusive O3 uptake. In order to realistically estimate O3 risks, the UNECE Convention on Long-Range Transport (LRTAP) suggested to use an approach based on the ‘Phytoxic Ozone Dose’ (POD), which reflects the cumulative amount of O3 absorbed by plants (Mills et al., 2011), rather than using approaches based on indices (e.g. AOT40) describing the O3 concentration in the surrounding air (Mills et al., 2010). The report presented here contains for the first time the complete dataset and absolute numbers on seven years of aboveground yield (plus final harvest root mass), together with climate data for this full factorial, O3 N-deposition experiment. Also, the estimate of O3 exposure index POD based on canopy flux instead of leaf level data is unique. The objective is to test the multiple pollutant climate interaction effects on productivity. This is feasible only with this new dataset. We hypothesized that during the 7-year period, (1) climatic factors temperature and precipitation correlate positively with productivity, (2) the effect of N-addition and elevated O3 increases with favorable weather conditions and increased productivity, and (3) higher O3 sensitivity (i.e. reduced productivity) coincides with higher growth rates at high N loads, yielding a statistically significant interaction between the two factors.
2. Materials and methods 2.1. Experimental site The experiment was set up in the Central Alps at the Alp Flix grassland plateau (Sur, Canton Grisons, Switzerland, 2000 m a.s.l., 46 320 N, 9 390 E). The local natural tree line is at w2200 m a.s.l, c. 200 m above the anthropogenic forest line. Winters are cold with permanent snow cover from November until April. The vegetation is a Geo-Montani-Nardetum, a pasture community covering large areas in the subalpine zone of the European Alps and Pyrenees (EUNIS classification 35.1, http://eunis.eea. europa.eu/habitats-factsheet.jsp?idHabitat¼10122). Dry matter (DM) yield is approximately 120 g m2 a1. One-half of the cover consists of grasses (primarily Nardus stricta and Festuca rubra) and sedges (e.g. Carex sempervirens). The six most frequent forb species and a single legume species (Trifolium alpinum) each contribute approximately 2e3% to the total cover. The remaining 40% consists of more than 70 forb-, and a few grass- and legume-species (further details in Bassin et al., 2007b and Bassin et al., 2013). The soil is a slightly acidic leptic cambisol (WRB, 1998; pH 4.8e6) developed on Serpentinite bedrock. Soil depth is approximately 20 cm.
2.2. Meteorological data Meteorological data (air temperature and relative humidity, wind speed and direction, solar radiation, soil temperature) were recorded continuously (2 m height). Soil moisture was recorded manually in campaigns in 108 of 180 monoliths (cf. Volk et al., 2011). Continuous estimates of soil water content (SWC) were obtained by solving the balance equation of a single soil bucket of 20 cm depth (equivalent to depth of monoliths) and using a volumetric total pore space, field capacity and permanent wilting point at 50, 45 and 5%, respectively. Details of the balance equation can be found in Rodriguez-Iturbe et al. (1999). The soil bucket was assumed to receive input from infiltration equal to the sum of precipitation and melt water from snow cover, with the available pore space as upper limit. Snow accumulation and melting were computed following Kokkonen et al. (2006). Based on the unstressed evapotranspiration, estimated according to Allen et al. (1998) assuming an average crop coefficient of 1.2, evapotranspiration rates were computed as in Calanca (2007), assuming a critical volumetric soil water content of 30%. Water losses are evapotranspiration and runoff, the latter being evaluated as a residual. The SWC calculation was checked against measurements in 2006 and 2010 using TDR (Trime FM2, Imko, Ettlingen, Germany).
2.3. Experimental plots and treatments For about 60 years prior to the experiment, the grassland was grazed during 3e4 weeks (cattle) at w1.3 livestock units ha1 yr1. No extra manure or fertilizer was applied. Experimental plots were established as described before (Bassin et al., 2007b). Briefly, plots consisted of 180 intact turf monoliths (L W H ¼ 30 40 22 cm) that were excavated in the fall of 2003, placed in drained plastic boxes and randomly assigned to treatment combinations. Groups of 20 were placed in shallow pits, flush with the surrounding surface in the center of nine fumigation rings. To minimize confounding effects of microclimatic differences between rings, monoliths were rearranged annually between the fumigation rings, but kept within their respective O3 treatment. Both the O3 and N treatments were applied during the snow-free period from approximately May to October. Background N deposition at the study site e calculated from concentration measurements in air, rainwater and snow e amounted to 1 with progressing length of the experiment. Despite the strong effect of temperature on yield, there was no interaction of GDDMay or other climatic parameters with N-treatment. Instead, the relative growth response (YieldN54/Yieldcontrol) increased consistently over time, indicating a strong cumulative effect of N on biomass production (sign. N time interaction). 3.5. Belowground biomass and shoot/root ratio N deposition averaged over O3 treatments increased belowground biomass significantly by þ19% (N14) and þ22% (N54) (Table 5). There was no significant O3 main-effect (Table 6). In the O3 control treatment N14 caused a 28% and N54 caused a 34% increase in biomass, respectively. This N effect was similar in O3þþ, and much smaller in O3þ, nevertheless there was no N O3 interaction. In the absence of extra N (N4), we found increased belowground biomass in O3þ (þ8%) and O3þþ (þ17%). But averaged over all N treatments, belowground biomass was slightly reduced (7%) in O3þ, and increased in O3þþ (þ13%) compared to the control. Table 6 The shoot/root ratio (S/R) averaged over O3-treatments increased with N-treatment from 0.22 (0.013; N4control) to 0.23 (0.013, N14) and 0.29 (0.009, N54). On the other hand, the O3treatment (averaged over N-treatments) had no consistent effect. In O3þ, S/R increased markedly in N14 and N54, relative to N4. In O3þþ, changes in S/R were small in all N treatments, thus indicating that the belowground response was comparable to that of aboveground biomass. Over all, no significant N O3 interaction was found (not shown). 4. Discussion During the initial three years of this experiment, elevated O3 counteracted the positive effect of N on canopy greenness, but had no effect on plant growth (Bassin et al., 2007b). It was suggested that reduced life span could negatively affect growth, as found in experiments where O3 damage in leaves was followed by a decrease in productivity (Hayes et al., 2007), but only in the longer run. This expectation was supported by the slight decrease in canopy photosynthesis in O3þþ (Volk et al., 2011). However, the results after 7 years of treatment confirm the lack of a response of harvested biomass to elevated O3, both under ambient and increased N loads. This is in agreement with results from other field experiments that indicated high O3 tolerance of mature semi-natural grasslands (Thwaites et al., 2006; Stampfli and Fuhrer, 2010). On the plant community level, functional redundancy, high genetic variability, or complex species interactions could be reasons for the high tolerance. Under this scenario, strong detrimental effects on O3-sensitive functional groups (FGs), genotypes or species could be compensated as more O3-tolerant plants would gain competitive advantage and immediately fill developing gaps, as found at this site for Nardus stricta under high O3 treatment (Bassin et al., 2013). Also, on the organismic level, the O3 detoxification capacity may be better developed at higher altitudes and prevent oxidative stress at the subalpine site studied here, similar
M. Volk et al. / Environmental Pollution 189 (2014) 161e168
165
Fig. 1. [A] Time series of the soil water content (SWC, %), the temperature (Temp, C), and the sum of the rainfall (rainsum, mm). [B] O3 concentrations at canopy height throughout the exposure periods: bars show the interquartile range (25e75%) of 5 day O3 concentrations, dark squares show the 95%. [C] Time series of modeled Gs{O3} (in mmol m2 PLA s1) during vegetation period: bars show the interquartile range of 5 day daytime Gs, white squares show the median.
to findings by Polle and Rennenberg (1992) for Norway spruce at 1700 m a.s.l. With high intrinsic O3 tolerance, the added O3 exposure used here may have been too small to cause an effect that could be detected with the available power of the experiment. According to Mills et al. (2010), the critical level for this type of grassland is set to an AOT40 of 5 ppm h for 6 months, which is roughly half of the control treatment value in the present study, and much less than the AOT40s in O3þ and O3þþ treatments. But instead of using AOT40, Mills et al. (2011) recommend to assess O3 risks for vegetation with an O3 flux-based approach, determined by stomatal conductance Gs and O3 concentrations at the top of the canopy. Low rates of O3 uptake in this type of grassland community were suggested as a possible reason for the apparent lack of a response to O3. Therefore, we estimated the canopy’s stomatal O3
250 600
GDD May
200
GDD May
100
GDD May
GDD May
150
400
GDD May
GDD May
GDD May
200 50 0
2004
2005
2006
2007
2008
2009
2010
Precipitation sum (mm, May- July)
800
-2
Aboveground dry matter yield (g m ) Growing degree days (°C>0 in May)
300
0
Year Fig. 2. Aboveground yield (means 1 SE) of control plots in Alp Flix subalpine grassland (vertical bars, left y-axis). Growing degree-days in May (GDDMay text, left yaxis) and sum of precipitation MayeJuly (open circles, right y-axis) indicate the close correlation of climate parameters with plant productivity.
uptake using a model to calculate POD according to the agreed UNECE methodology (Mills et al., 2010). The model was parameterized based on meteorological data and measured H2O fluxes, collected to give best results for a Gmax estimate and its dependencies on meteorological parameters (temperature, radiation, water vapor deficit, and soil water availability). Despite 2010 was cool on average, our measurement campaign also covered a wide range of perfect Gmax conditions. On the other hand, we do not see a reason for any systematic under- or overestimation of the derived values from less than perfect parameterization under non-Gmax conditions. Due to limited measurement resources, we assumed that stomatal behavior in elevated O3 treatments would not differ from the control. Changes in stomatal functioning due to increased O3 exposure have been observed though, but the effect is not universal (see Ainsworth et al., 2012; Wagg et al., 2013). Earlier, we reported a weak decrease in stomatal conductance in some of the species present in this grassland type (Bassin et al., 2009). But because there was no O3 effect on the growth of dominant species during the first three years, it was assumed that a possible canopy stomatal response to elevated O3 would be negligible. The POD0 of 48.7 mmol m2 for seasonal exposure periods in the control treatment is similar to values determined for other grassland systems. For instance, Nussbaum and Fuhrer (2000) reported a value of about 50 mmol m2 in open-top chambers and 35 mmol m2 in ambient air. Using a model with parameters for productive grassland, Nussbaum et al. (2003) estimated values averaging 45 mmol m2 and Keller et al. (2007) between 70 and 80 mmol m2. Hence, the values obtained at this sub-alpine site can be considered typical for long-term exposure of perennial species. In O3þ and O3þþ, the phytotoxic ozone dose reached values found to be detrimental to tree growth (Mills et al., 2010). Thus, based on our data, the high O3 tolerance of the investigated vegetation cannot be explained by limited stomatal O3 uptake. But finding the same (high) stomatal flux associated with lower productivity is
M. Volk et al. / Environmental Pollution 189 (2014) 161e168
Block O3 N Time pH PrecipApreOct GDDMay NN O3 N N Time O3 Time Time Time N Time Time N PrecipApreOct N GDDMay
L-ratio
p-value
10.6 0.6 76.2 297.7 11.6 48.1 426.5 2.7 2.1 11.8 5.4 198.3 0.9 1.3 0.1
0.005 0.732 0.000 0.000 0.001 0.000 0.000 0.104 0.342 0.001 0.068 0.000 0.351 0.250 0.805
indeed an indicator for a non-substrate-limited photosynthesis. Probably growth of alpine plants is often rather sink limited instead, helping to explain their tolerance towards damage to the assimilation system. Given the comparatively constant ambient O3 concentrations at the experiment site (Fig. 1B), we had expected POD values to reflect the stomatal response to the varying meteorological conditions of the different years. Indeed POD0 in O3 control were similarly low in the two coolest and wettest years (2008 and 2010). But interestingly, the largest contrast in POD values between years (highest 2006 and lowest 2009) was observed in the two warmest and driest years of the study period. This may be explained by counteracting
Growth response (N4control/N9-54)
-2
Aboveground yield (DM g m )
350
N4 control N9 N14 N29 N54
250 200 150 100 50 0 1.7 1.6 1.5 1.4
-2
300
[B] N54 N29 N14 N9
1.3 1.2 1.1
O control O+ O ++
[B]
O+ O ++
250 200 150 100 50 0 1.5 1.4 1.3 1.2 1.1 1.0 0.9 0.8 0.7 0.6
2004
2005
2006
2007
2008
2009
2010
Year Fig. 4. [A] Annual aboveground yield (dry matter, DM) of harvested phytomass in Alp Flix subalpine grassland (>2 cm height), grouped by O3 exposure treatment (means 1 SE). O3 control ¼ No O3 added, O3þ ¼ ambient conc. 1.4, O3þþ ¼ ambient conc. 1.7 [B] Relative aboveground growth response as the ratio of Yield O3 control/Yield O3þ/þþ treatment.
Table 5 Mean (SE) belowground biomass (g dry matter m2)a.
1.0 0.9 0.8
[A]
effects of high temperature and low soil moisture on the stomatal conductance Gs{O3}. In 2009 the limiting effect of SWC (below 35%) was clearly stronger and longer lasting than in 2006. It was hypothesized that, with increasing N input, the sensitivity to O3 would change, as observed in previous grassland studies (e.g. Sanz et al., 2011; Jones et al., 2010; Wyness et al., 2011). But the results from this grassland community reveal no trend towards higher or lower O3 sensitivity of the community with increasing N input. The main effect of N was due to a shift in species composition, with sedges representing the biomass fraction with the strongest N response (Bassin et al., 2013). In contrast, in other component species, probably due to a strong P-co-limitation (Blanke et al., 2012), the N-induced growth increase was too small to cause physiological or morphological adaptations that could influence O3 sensitivity. While grasses are known to be O3 tolerant, O3 sensitivity in sedges is not well known. Assuming that sedges are equally insensitive to O3, the lack of a community biomass response to O3 could be related to the success of these insensitive FGs in this
[A]
300
350 Aboveground yield (DM g m )
Table 4 Repeated measures analysis of 2004e2010 aboveground biomass yield with effects of block, ozone (O3), nitrogen (N), time and selected interactions. Initial soil pH (2003), precipitation (PrecipApreOct) and growing degree-days in May (GDDMay) are continuous covariables.
Growth response (O3control/O3+/++)
166
O3 control
2004
2005
2006
2007 Year
2008
2009
2010
Fig. 3. [A] Annual aboveground yield (dry matter, DM) of harvested phytomass in Alp Flix subalpine grassland (>2 cm height), grouped by N-deposition treatment (means 1 SE). N4control ¼ no extra N deposition, N9 ¼ þ5 kg-, N14 ¼ þ10 kg-, N29 ¼ þ25 kg- and N54 ¼ þ50 kg ha1 a1 extra N deposition. [B] Relative aboveground growth response as the ratio of Yield N4control/Yield N9e54 treatment.
N4 N14 N54 Mean
545.0 699.7 730.0 658.2
48.19 47.54 25.58 30.00
O3 þ 589.5 636.7 601.4 609.2
O3þþ
12.26 25.07 36.05 15.08
637.1 769.5 833.1 746.6
O3 mean
65.14 75.15 134.52 56.00
590.5 702.0 721.5 590.5
27.23 31.78 49.94 27.23
O3 control ¼ No O3 added, O3þ ¼ ambient conc. 1.4, O3þþ ¼ ambient conc. 1.7. N4 ¼ no extra N deposition; N14 ¼ þ10 kg- and N54 ¼ þ50 kg ha1 a1 extra N deposition. a Measurement at 0e10 cm soil depth after 7 years of treatment (2010, end of experiment).
M. Volk et al. / Environmental Pollution 189 (2014) 161e168
5. Conclusions
Table 6 ANOVA for belowground biomassa.
Block O3 N pH O3 N
167
Df
denDf
L-ratio
p-value
2 2 1 1 2
4 4 95 95 95
4.94 4.26 9.07 4.09 0.86
0.085 0.119 0.003 0.043 0.65
a Measurement at 0e10 cm soil depth after 7 years of treatment (2010, end of experiment).
grassland. Possible effects on the more sensitive herb- or legume FGs, may be too small to be reflected in a biomass response at the community level, as observed by Wedlich et al. (2012) in upland grassland in the UK. Also, the potentially more sensitive FGs may to some extent be protected from O3 inside the canopy (Jäggi et al., 2006). But their low stature, barely exceeding the cutting height (2 cm), would rather exaggerate than obscure a small response. Despite the absence of an O3 effect on aboveground biomass, we tested whether belowground biomass would be more responsive. Earlier, belowground biomass was found to be reduced more relative to aboveground biomass under O3 enrichment (Jones et al., 2010; Wagg et al., 2012), thus leading to higher shoot/root ratio. This could be due to inhibited assimilate transport from sources to sinks, or to the preferential allocation of assimilates to new leaves in order to compensate for O3 damage and optimize light capturing (cf. Fuhrer and Booker, 2003). An increase in shoot/root ratio in response to O3 in the presence of extra N input was observed in O3þ, but not in O3þþ. For this grassland, Blanke et al. (2012) found the N-effect on shoot/root ratio to differ between species, but on the whole plant community level extra N increased the shoot/root ratio. In line with the optimal partitioning theory, suggesting that plants optimize the over all growth rate by adjusting their biomass allocation to capture the most limiting resource (Hilbert, 1990), O3 and N should act in the same direction, as observed in O3þ, but the absence of this interactive effect in O3þþ remains unexplained. The lowest yield (2006) occurred under low precipitation and low GDDMay, but the highest July temperatures (13.6 C compared to 7-yr median 9.7 C). The secondary role of precipitation must be seen on the background of an ecosystem that is not usually waterlimited during most of the main growing period. The melting snowpack in spring supplies the necessary soil water content during the most sensitive phase of plant development. Apparently, the basis for a high annual yield is determined mainly by high GDD values early in the year, favoring fast and early plant development. Favorable weather conditions should increase O3 sensitivity, but antagonistic effects of seasonal weather conditions make it difficult to identify interactions with O3 and N. However, POD values varied little between years, and the most productive year even yielded a slightly larger biomass harvest under O3þþ, which could hypothetically be related to compensatory growth, as found by Bielenberg et al. (2001) under conditions of O3-induced accelerated leaf senescence in trees with low N supply. Importantly, all responses in the N9 treatment up to year five (2008) were below 6%. But in the last year (2010) the N9 deposition level triggered a c. 31% increase in productivity, equivalent to 50% of the N54 effect. This late but strong response challenges present critical loads for Ndeposition in subalpine grassland (i.e. 5e10 kg ha1 yr1; Achermann and Bobbink, 2003; Bobbink and Hettelingh, 2011) and underlines the importance of carry-over effects. It implies first the availability of stored nutrients as resources for growth, and second the presence of large storage organs as sinks for newly acquired assimilates (Farrar and Jones, 2000). It suggests that such carry over effects in the perennial species of the subalpine grassland determine the responsiveness of the system to N enrichment.
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Ozone effects in a drier climate: implications for stomatal fluxes of reduced stomatal sensitivity to soil drying in a typical grassland species. Glob. Change Biol. 18 (3), 948e959. Hicks, B.B., Baldocchi, D.D., Meyers, T.P., Hosker, R.P., Matt, D.R., 1987. A preliminary multiple resistance routine for deriving dry deposition velocities from measured quantities. Water Air Soil Pollut. 36 (3e4), 311e330. Hilbert, D.W., 1990. Optimization of plant root: shoot ratios and internal nitrogen concentration. Ann. Bot. 66, 91e99. Jäggi, M., Ammann, C., Neftel, A., Fuhrer, J., 2006. Environmental control of profiles of ozone concentration in a grassland canopy. Atmos. Environ. 40, 5496e5507. Jones, M.L.M., Hodges, G., Mills, G., 2010. N mediates aboveground effects of ozone, but not below-ground effects in a rhizomatous sedge. Environ. Pollut. 158, 559e 565. Keller, F., Bassin, S., Ammann, C., Fuhrer, J., 2007. High-resolution modelling of AOT40 and stomatal ozone uptake in wheat and grassland: a comparison between 2000 and the hot summer of 2003 in Switzerland. Environ. Pollut. 146 (3), 671e677. Kokkonen, T., Koivusalo, H., Jakeman, T., Norton, J., 2006. Construction of a degreeday snow model in the light of the ten iterative steps in model development. In: Voinov, A., Jakeman, A.J., Rizzoli, A.E. (Eds.), Proceedings of the iEMSs Third Biennial Meeting: “Summit on Environmental Modelling and Software”. International Environmental Modelling and Software Society, Burlington, USA. July 2006. CD ROM. Internet: http://www.iemss.org/iemss2006/sessions/all.html. LeBauer, D.S., Treseder, K.K., 2008. Nitrogen limitation of net primary productivity in terrestrial ecosystems is globally distributed. Ecology 89 (2), 371e379. Marenco, A., Gouget, H., Nedelec, P., Pages, J.P., Karcher, F.,1994. Evidence of a long-term increase in tropospheric ozone from Pic du Midi data series e consequences e positive radiative forcing. J. Geophys. Res. Atmos. 99 (D8), 16617e16632. Mills, G., Hayes, F., Wilkinson, S., Davis, W.J., 2009. Chronic exposure to increasing background ozone impairs stomatal functioning in grassland species. Glob. Change Biol. 15 (6), 1522e1533.
Mills, G., et al., 2010. Chapter 3: Mapping critical levels for vegetation. In: Spranger, T., Lorenz, U., Gregor, H.-D. (Eds.), Manual on Methodologies and Criteria for Modelling and Mapping Critical Loads and Levels and Air Pollution Effects, Risks and Trends. Umweltbundesamt, Berlin. Available at: http://www. rivm.nl/media/documenten/cce/manual/ Ch3revisedsummer2010updatedJune2011.pdf. Mills, G., Pleijel, H., Braun, S., Büker, P., Bermejo, V., Calvo, E., Danielsson, H., Emberson, L., González Fernández, I., Grünhage, L., Harmens, H., Hayes, F., Karlsson, P.-E., Simpson, D., 2011. New stomatal flux-based critical levels for ozone effects on vegetation. Atmos. Environ. 45, 5064e5068. Nussbaum, S., Fuhrer, J., 2000. Difference in ozone uptake in grassland species between open-top chambers and ambient air. Environ. Pollut. 109, 463e471. Nussbaum, S., Remund, J., Rihm, B., Mieglitz, K., Gurtz, J., Fuhrer, J., 2003. Highresolution spatial analysis of stomatal ozone uptake in arable crops and pastures. Environ. Int. 29, 385e392. Pape, L., Ammann, C., Nyfeler-Brunner, A., Spirig, C., Hens, K., Meixner, F.X., 2009. An automated dynamic chamber system for surface exchange measurement of non-reactive and reactive trace gases of grassland ecosystems. Biogeosciences 6, 405e429. Pell, E.J., Schlagnhaufer, C.D., Arteca, R.N., 1997. Ozone-induced oxidative stress: mechanisms of action and reaction. Physiol. Plant. 100 (2), 264e273. Pinheiro, J.C., Bates, D.M., 1996. Unconstrained parameterizations for variancee covariance matrices. Stat. Comput. 6, 289e296. Polle, A., Rennenberg, H., 1992. Field studies on Norway spruce trees at high altitudes: II. Defence systems against oxidative stress in needles. New Phytol. 121, 635e642. R Development Core Team, 2011. R: a Language and Environment for Statistical Computing. R Foundation for Statistical Computing, Vienna. Rihm, B., Kurz, D., 2001. Deposition and critical loads of nitrogen in Switzerland. Water Air Soil Pollut. 130 (1e4), 1223e1228. Rodriguez-Iturbe, I., Porporato, A., Ridolfi, L., Isham, V., Cox, D.R., 1999. Probabilistic modelling of water balance at a point: the role of climate, soil and vegetation. Proc. Royal Soc. Lond. A 455, 3789e3805. Sanz, J., Bermejo, V., Muntifering, R., González-Fernández, I., Gimeno, B.S., Elvira, S., Alonso, R., 2011. Plant phenology, growth and nutritive quality of Briza maxima: responses induced by enhanced ozone atmospheric levels and nitrogen enrichment. Environ. Pollut. 159, 423e430. Stampfli, A., Fuhrer, J., 2010. Spatial heterogeneity confounded ozone-exposure experiment in semi-natural grassland. Oecologia 162 (2), 515e522. Thwaites, R.H., Ashmore, M.R., Morton, A.J., Pakeman, R.J., 2006. The effects of tropospheric ozone on the species dynamics of calcareous grassland. Environ. Pollut. 144 (2), 500e509. Vingarzan, R., 2004. A review of surface ozone background levels and trends. Atmos. Environ. 38, 3431e3442. Volk, M., Geissmann, M., Blatter, A., Contat, F., Fuhrer, J., 2003. Design and performance of a free-air exposure system to study long-term effects of ozone on grasslands. Atmos. Environ. 37 (9e10), 1341e1350. Volk, M., Bungener, P., Contat, F., Montani, M., Fuhrer, J., 2006. Grassland yield declined by a quarter in 5 years of free-air ozone fumigation. Glob. Change Biol. 12, 74e83. Volk, M., Obrist, D., Novak, K., Giger, R., Bassin, S., Fuhrer, J., 2011. Subalpine grassland carbon dioxide fluxes indicate substantial carbon losses under increased nitrogen deposition, but not at elevated ozone concentration. Glob. Change Biol. 17, 366e376. Wagg, S., Mills, G., Hayes, F., Wilkinson, S., Cooper, D., Davies, W.J., 2012. Reduced soil water availability did not protect two competing grassland species from the negative effects of increasing background ozone. Environ. Pollut. 165, 91e99. Wagg, S., Mills, G., Hayes, F., Wilkinson, S., William, D.J., 2013. Stomata are less responsive to environmental stimuli in high background ozone in Dactylis glomerata and Ranunculus acris. Environ. Pollut. 175, 82e91. Wedlich, K.V., Rintoul, N., Peacock, S., Cape, J.N., Coyle, M., Toet, S., Barnes, J., Ashmore, M., 2012. Effects of ozone on species composition in an upland grassland. Oecologia 168, 1137e1146. White, R., Murray, S., Rohweder, M., 2000. Pilot Analysis of Global Ecosystems (PAGE): Grassland Ecosystems. World Resources Institute, Washington D.C. WRB World reference base for soil resources, 1998. World Soil Resources Reports 84. Food and Agriculture Organization of the United Nations, Rome. Wyness, K., Mills, G., Jones, L., Barnes, J., Jones, D., 2011. Enhanced nitrogen deposition exacerbates the negative effect of increasing background ozone in Dactylis glomerata, but not Ranunculus acris. Environ. Pollut. 159 (10), 2493e2499.
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Environmental Pollution journal homepage: www.elsevier.com/locate/envpol
High tolerance of subalpine grassland to long-term ozone exposure is independent of N input and climatic drivers Matthias Volk*, Veronika Wolff, Seraina Bassin, Christof Ammann, Jürg Fuhrer Agroscope, Air Pollution/Climate Group, 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 25 November 2013 Received in revised form 25 February 2014 Accepted 28 February 2014
In a seven-year study, we tested effects of increased N and O3 deposition and climatic conditions on biomass of subalpine grassland. Ozone risk was assessed as exposure (AOT40) and as stomatal flux (POD0,1). We hypothesized that productivity is higher under N- and lower under O3 deposition, with interactions with climatic conditions. Aboveground biomass was best correlated with growing-degree days for May (GDDMay). Nitrogen deposition increased biomass up to 60% in the highest treatment, and 30% in the lowest addition. Also belowground biomass showed a positive N-response. Ozone enrichment had no effect on biomass, and no interaction between O3 and N was observed. Growth response to N deposition was not correlated to GDDMay or precipitation, but indicated a cumulative effect over time. Productivity of subalpine grassland is tolerant to increasing ozone exposure, independent of N input and climatic drivers. N deposition rates at current critical loads, strongly increase the grassland yield. Ó 2014 Elsevier Ltd. All rights reserved.
Keywords: Air pollution Nitrogen Critical load/level Phytotoxic ozone dose POD Productivity
1. Introduction The grassland habitat type s.l. covers some 40% of the terrestrial surface (White et al., 2000), and in subalpine regions of central Europe, highly diverse, semi-natural grasslands represent a major land cover type. Apart from their conservation value, these grasslands act as an important source of feed for livestock and are essential for the esthetic value of the landscape. They deliver ecosystem services like soil erosion and water quality protection, as well as sink potential for the greenhouse gas CO2. Understanding the sensitivity of these systems to changes in environmental conditions is crucial for evaluating future hazards and designing suitable management options and policies, to reduce the risk of losing their important ecological and agronomic functions. Changes in climate and increasing air pollution are among the most critical long-term risks to be considered. Air pollution by tropospheric ozone (O3) and reactive nitrogen (Nr) compounds, both resulting from industrial, traffic and agricultural emissions, is widespread, with potentially harmful effects across the northern hemisphere (Fuhrer et al., 2003; Ainsworth et al., 2012). Far from sources of O3 precursors, remote sites in the mid-latitudes are exposed to annual mean background
* Corresponding author. E-mail address: [email protected] (M. Volk). http://dx.doi.org/10.1016/j.envpol.2014.02.032 0269-7491/Ó 2014 Elsevier Ltd. All rights reserved.
concentrations of 20e45 ppb O3, representing approximately a doubling during the last century (Marenco et al., 1994; Vingarzan, 2004). Annual mean O3 concentrations are particularly high at sites above approximately 1000 m a.s.l., where diurnal variations in concentration are small due to reduced nocturnal O3 degradation and deposition (Chevalier et al., 2007). Least conservative scenarios expect mean annual background concentrations of O3 to possibly reach 80e90 ppb by 2100 (Vingarzan, 2004; Derwent et al., 2007), due to changes in emissions and climatic conditions (cf. Fuhrer, 2009). Present atmospheric N deposition rates in remote mountain areas are around 5 kg N ha1 yr1, but reach up to 60 kg N ha1 yr1 in regions with intensive livestock farming (Rihm and Kurz, 2001), thus exceeding critical loads defined to protect natural vegetation (Achermann and Bobbink, 2003). Due to long-range transport and increasing emission, N deposition over vast areas is predicted to almost double globally by 2050 (Galloway et al., 2004). Given these trends, it is evident that ecosystems in remote areas are e and will continue to be e at risk from air pollution. In individual plant species, elevated O3 has been found to cause visible injury (Bungener et al., 1999a; Bermejo et al., 2003), premature leaf senescence and damage to the photosynthetic system (Pell et al., 1997), guard cell damage (Mills et al., 2009; Hayes et al., 2012) and reduced growth (Hayes et al., 2006; Gimeno et al., 2004; Wyness et al., 2011; Ainsworth et al., 2012; Bungener et al., 1999b). It reduces plant productivity by entering leaves through the
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stomata, generating other reactive oxygen species and thus causing oxidative stress (cf. Ainsworth et al., 2012). However, translating experimental results from individual species to field conditions is difficult, and understanding community level responses requires experiments under natural conditions. Only limited experimental data exist, that describe effects of long-term O3 exposure on productivity and species composition in semi-natural grasslands. For instance, Volk et al. (2006) observed a significant decline in productivity over a 5-year period of exposure to elevated O3. Later Stampfli and Fuhrer (2010) found that the observed effect was confounded by plot management changes. From the few available field studies it was suggested that growth reduction due to O3 may be limited to productive habitats with high N availability (Bassin et al., 2007a). Typically, semi-natural ecosystems respond to high N loads with increased productivity and altered species composition (LeBauer and Treseder, 2008; Bobbink et al., 2010; Bassin et al., 2013). N-induced changes in the dominance of species may act as an additional factor changing the community sensitivity to additional stresses such as O3 or climate extremes (Bobbink et al., 1998). But it remains unclear how O3 and N availability interact with each other in semi-natural grassland communities. N-addition may increase plant sensitivity to O3 via stimulation of gas exchange leading to higher levels of canopy O3 uptake and higher growth rates. Alternatively, sensitivity may decrease through better oxidation damage repair, or via mitigation of sink limitation (Andersen, 2003). Finally, effects of air pollutants may depend on prevailing weather conditions, which strongly vary between and within seasons, and meteorological effects on productivity act not only immediately, but also delayed for several growing seasons via complex carry over effects (Arnone et al., 2008). Thus, only longerterm experiments may help to disentangle the complex interactions between different atmospheric factors. We conducted a free-air fumigation experiment for seven years to investigate the effects of combined O3- and N-deposition on subalpine grassland. During the initial three years of the experiment increased N-input caused a strong growth response, but there was no significant response to elevated O3 (Bassin et al., 2007b). It was then suggested that the lack of a response to O3 could be due to the short duration of the experiment, high intrinsic detoxification capacity under conditions prevailing at high altitudes, or high resistance to diffusive O3 uptake. In order to realistically estimate O3 risks, the UNECE Convention on Long-Range Transport (LRTAP) suggested to use an approach based on the ‘Phytoxic Ozone Dose’ (POD), which reflects the cumulative amount of O3 absorbed by plants (Mills et al., 2011), rather than using approaches based on indices (e.g. AOT40) describing the O3 concentration in the surrounding air (Mills et al., 2010). The report presented here contains for the first time the complete dataset and absolute numbers on seven years of aboveground yield (plus final harvest root mass), together with climate data for this full factorial, O3 N-deposition experiment. Also, the estimate of O3 exposure index POD based on canopy flux instead of leaf level data is unique. The objective is to test the multiple pollutant climate interaction effects on productivity. This is feasible only with this new dataset. We hypothesized that during the 7-year period, (1) climatic factors temperature and precipitation correlate positively with productivity, (2) the effect of N-addition and elevated O3 increases with favorable weather conditions and increased productivity, and (3) higher O3 sensitivity (i.e. reduced productivity) coincides with higher growth rates at high N loads, yielding a statistically significant interaction between the two factors.
2. Materials and methods 2.1. Experimental site The experiment was set up in the Central Alps at the Alp Flix grassland plateau (Sur, Canton Grisons, Switzerland, 2000 m a.s.l., 46 320 N, 9 390 E). The local natural tree line is at w2200 m a.s.l, c. 200 m above the anthropogenic forest line. Winters are cold with permanent snow cover from November until April. The vegetation is a Geo-Montani-Nardetum, a pasture community covering large areas in the subalpine zone of the European Alps and Pyrenees (EUNIS classification 35.1, http://eunis.eea. europa.eu/habitats-factsheet.jsp?idHabitat¼10122). Dry matter (DM) yield is approximately 120 g m2 a1. One-half of the cover consists of grasses (primarily Nardus stricta and Festuca rubra) and sedges (e.g. Carex sempervirens). The six most frequent forb species and a single legume species (Trifolium alpinum) each contribute approximately 2e3% to the total cover. The remaining 40% consists of more than 70 forb-, and a few grass- and legume-species (further details in Bassin et al., 2007b and Bassin et al., 2013). The soil is a slightly acidic leptic cambisol (WRB, 1998; pH 4.8e6) developed on Serpentinite bedrock. Soil depth is approximately 20 cm.
2.2. Meteorological data Meteorological data (air temperature and relative humidity, wind speed and direction, solar radiation, soil temperature) were recorded continuously (2 m height). Soil moisture was recorded manually in campaigns in 108 of 180 monoliths (cf. Volk et al., 2011). Continuous estimates of soil water content (SWC) were obtained by solving the balance equation of a single soil bucket of 20 cm depth (equivalent to depth of monoliths) and using a volumetric total pore space, field capacity and permanent wilting point at 50, 45 and 5%, respectively. Details of the balance equation can be found in Rodriguez-Iturbe et al. (1999). The soil bucket was assumed to receive input from infiltration equal to the sum of precipitation and melt water from snow cover, with the available pore space as upper limit. Snow accumulation and melting were computed following Kokkonen et al. (2006). Based on the unstressed evapotranspiration, estimated according to Allen et al. (1998) assuming an average crop coefficient of 1.2, evapotranspiration rates were computed as in Calanca (2007), assuming a critical volumetric soil water content of 30%. Water losses are evapotranspiration and runoff, the latter being evaluated as a residual. The SWC calculation was checked against measurements in 2006 and 2010 using TDR (Trime FM2, Imko, Ettlingen, Germany).
2.3. Experimental plots and treatments For about 60 years prior to the experiment, the grassland was grazed during 3e4 weeks (cattle) at w1.3 livestock units ha1 yr1. No extra manure or fertilizer was applied. Experimental plots were established as described before (Bassin et al., 2007b). Briefly, plots consisted of 180 intact turf monoliths (L W H ¼ 30 40 22 cm) that were excavated in the fall of 2003, placed in drained plastic boxes and randomly assigned to treatment combinations. Groups of 20 were placed in shallow pits, flush with the surrounding surface in the center of nine fumigation rings. To minimize confounding effects of microclimatic differences between rings, monoliths were rearranged annually between the fumigation rings, but kept within their respective O3 treatment. Both the O3 and N treatments were applied during the snow-free period from approximately May to October. Background N deposition at the study site e calculated from concentration measurements in air, rainwater and snow e amounted to 1 with progressing length of the experiment. Despite the strong effect of temperature on yield, there was no interaction of GDDMay or other climatic parameters with N-treatment. Instead, the relative growth response (YieldN54/Yieldcontrol) increased consistently over time, indicating a strong cumulative effect of N on biomass production (sign. N time interaction). 3.5. Belowground biomass and shoot/root ratio N deposition averaged over O3 treatments increased belowground biomass significantly by þ19% (N14) and þ22% (N54) (Table 5). There was no significant O3 main-effect (Table 6). In the O3 control treatment N14 caused a 28% and N54 caused a 34% increase in biomass, respectively. This N effect was similar in O3þþ, and much smaller in O3þ, nevertheless there was no N O3 interaction. In the absence of extra N (N4), we found increased belowground biomass in O3þ (þ8%) and O3þþ (þ17%). But averaged over all N treatments, belowground biomass was slightly reduced (7%) in O3þ, and increased in O3þþ (þ13%) compared to the control. Table 6 The shoot/root ratio (S/R) averaged over O3-treatments increased with N-treatment from 0.22 (0.013; N4control) to 0.23 (0.013, N14) and 0.29 (0.009, N54). On the other hand, the O3treatment (averaged over N-treatments) had no consistent effect. In O3þ, S/R increased markedly in N14 and N54, relative to N4. In O3þþ, changes in S/R were small in all N treatments, thus indicating that the belowground response was comparable to that of aboveground biomass. Over all, no significant N O3 interaction was found (not shown). 4. Discussion During the initial three years of this experiment, elevated O3 counteracted the positive effect of N on canopy greenness, but had no effect on plant growth (Bassin et al., 2007b). It was suggested that reduced life span could negatively affect growth, as found in experiments where O3 damage in leaves was followed by a decrease in productivity (Hayes et al., 2007), but only in the longer run. This expectation was supported by the slight decrease in canopy photosynthesis in O3þþ (Volk et al., 2011). However, the results after 7 years of treatment confirm the lack of a response of harvested biomass to elevated O3, both under ambient and increased N loads. This is in agreement with results from other field experiments that indicated high O3 tolerance of mature semi-natural grasslands (Thwaites et al., 2006; Stampfli and Fuhrer, 2010). On the plant community level, functional redundancy, high genetic variability, or complex species interactions could be reasons for the high tolerance. Under this scenario, strong detrimental effects on O3-sensitive functional groups (FGs), genotypes or species could be compensated as more O3-tolerant plants would gain competitive advantage and immediately fill developing gaps, as found at this site for Nardus stricta under high O3 treatment (Bassin et al., 2013). Also, on the organismic level, the O3 detoxification capacity may be better developed at higher altitudes and prevent oxidative stress at the subalpine site studied here, similar
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Fig. 1. [A] Time series of the soil water content (SWC, %), the temperature (Temp, C), and the sum of the rainfall (rainsum, mm). [B] O3 concentrations at canopy height throughout the exposure periods: bars show the interquartile range (25e75%) of 5 day O3 concentrations, dark squares show the 95%. [C] Time series of modeled Gs{O3} (in mmol m2 PLA s1) during vegetation period: bars show the interquartile range of 5 day daytime Gs, white squares show the median.
to findings by Polle and Rennenberg (1992) for Norway spruce at 1700 m a.s.l. With high intrinsic O3 tolerance, the added O3 exposure used here may have been too small to cause an effect that could be detected with the available power of the experiment. According to Mills et al. (2010), the critical level for this type of grassland is set to an AOT40 of 5 ppm h for 6 months, which is roughly half of the control treatment value in the present study, and much less than the AOT40s in O3þ and O3þþ treatments. But instead of using AOT40, Mills et al. (2011) recommend to assess O3 risks for vegetation with an O3 flux-based approach, determined by stomatal conductance Gs and O3 concentrations at the top of the canopy. Low rates of O3 uptake in this type of grassland community were suggested as a possible reason for the apparent lack of a response to O3. Therefore, we estimated the canopy’s stomatal O3
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uptake using a model to calculate POD according to the agreed UNECE methodology (Mills et al., 2010). The model was parameterized based on meteorological data and measured H2O fluxes, collected to give best results for a Gmax estimate and its dependencies on meteorological parameters (temperature, radiation, water vapor deficit, and soil water availability). Despite 2010 was cool on average, our measurement campaign also covered a wide range of perfect Gmax conditions. On the other hand, we do not see a reason for any systematic under- or overestimation of the derived values from less than perfect parameterization under non-Gmax conditions. Due to limited measurement resources, we assumed that stomatal behavior in elevated O3 treatments would not differ from the control. Changes in stomatal functioning due to increased O3 exposure have been observed though, but the effect is not universal (see Ainsworth et al., 2012; Wagg et al., 2013). Earlier, we reported a weak decrease in stomatal conductance in some of the species present in this grassland type (Bassin et al., 2009). But because there was no O3 effect on the growth of dominant species during the first three years, it was assumed that a possible canopy stomatal response to elevated O3 would be negligible. The POD0 of 48.7 mmol m2 for seasonal exposure periods in the control treatment is similar to values determined for other grassland systems. For instance, Nussbaum and Fuhrer (2000) reported a value of about 50 mmol m2 in open-top chambers and 35 mmol m2 in ambient air. Using a model with parameters for productive grassland, Nussbaum et al. (2003) estimated values averaging 45 mmol m2 and Keller et al. (2007) between 70 and 80 mmol m2. Hence, the values obtained at this sub-alpine site can be considered typical for long-term exposure of perennial species. In O3þ and O3þþ, the phytotoxic ozone dose reached values found to be detrimental to tree growth (Mills et al., 2010). Thus, based on our data, the high O3 tolerance of the investigated vegetation cannot be explained by limited stomatal O3 uptake. But finding the same (high) stomatal flux associated with lower productivity is
M. Volk et al. / Environmental Pollution 189 (2014) 161e168
Block O3 N Time pH PrecipApreOct GDDMay NN O3 N N Time O3 Time Time Time N Time Time N PrecipApreOct N GDDMay
L-ratio
p-value
10.6 0.6 76.2 297.7 11.6 48.1 426.5 2.7 2.1 11.8 5.4 198.3 0.9 1.3 0.1
0.005 0.732 0.000 0.000 0.001 0.000 0.000 0.104 0.342 0.001 0.068 0.000 0.351 0.250 0.805
indeed an indicator for a non-substrate-limited photosynthesis. Probably growth of alpine plants is often rather sink limited instead, helping to explain their tolerance towards damage to the assimilation system. Given the comparatively constant ambient O3 concentrations at the experiment site (Fig. 1B), we had expected POD values to reflect the stomatal response to the varying meteorological conditions of the different years. Indeed POD0 in O3 control were similarly low in the two coolest and wettest years (2008 and 2010). But interestingly, the largest contrast in POD values between years (highest 2006 and lowest 2009) was observed in the two warmest and driest years of the study period. This may be explained by counteracting
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2009
2010
Year Fig. 4. [A] Annual aboveground yield (dry matter, DM) of harvested phytomass in Alp Flix subalpine grassland (>2 cm height), grouped by O3 exposure treatment (means 1 SE). O3 control ¼ No O3 added, O3þ ¼ ambient conc. 1.4, O3þþ ¼ ambient conc. 1.7 [B] Relative aboveground growth response as the ratio of Yield O3 control/Yield O3þ/þþ treatment.
Table 5 Mean (SE) belowground biomass (g dry matter m2)a.
1.0 0.9 0.8
[A]
effects of high temperature and low soil moisture on the stomatal conductance Gs{O3}. In 2009 the limiting effect of SWC (below 35%) was clearly stronger and longer lasting than in 2006. It was hypothesized that, with increasing N input, the sensitivity to O3 would change, as observed in previous grassland studies (e.g. Sanz et al., 2011; Jones et al., 2010; Wyness et al., 2011). But the results from this grassland community reveal no trend towards higher or lower O3 sensitivity of the community with increasing N input. The main effect of N was due to a shift in species composition, with sedges representing the biomass fraction with the strongest N response (Bassin et al., 2013). In contrast, in other component species, probably due to a strong P-co-limitation (Blanke et al., 2012), the N-induced growth increase was too small to cause physiological or morphological adaptations that could influence O3 sensitivity. While grasses are known to be O3 tolerant, O3 sensitivity in sedges is not well known. Assuming that sedges are equally insensitive to O3, the lack of a community biomass response to O3 could be related to the success of these insensitive FGs in this
[A]
300
350 Aboveground yield (DM g m )
Table 4 Repeated measures analysis of 2004e2010 aboveground biomass yield with effects of block, ozone (O3), nitrogen (N), time and selected interactions. Initial soil pH (2003), precipitation (PrecipApreOct) and growing degree-days in May (GDDMay) are continuous covariables.
Growth response (O3control/O3+/++)
166
O3 control
2004
2005
2006
2007 Year
2008
2009
2010
Fig. 3. [A] Annual aboveground yield (dry matter, DM) of harvested phytomass in Alp Flix subalpine grassland (>2 cm height), grouped by N-deposition treatment (means 1 SE). N4control ¼ no extra N deposition, N9 ¼ þ5 kg-, N14 ¼ þ10 kg-, N29 ¼ þ25 kg- and N54 ¼ þ50 kg ha1 a1 extra N deposition. [B] Relative aboveground growth response as the ratio of Yield N4control/Yield N9e54 treatment.
N4 N14 N54 Mean
545.0 699.7 730.0 658.2
48.19 47.54 25.58 30.00
O3 þ 589.5 636.7 601.4 609.2
O3þþ
12.26 25.07 36.05 15.08
637.1 769.5 833.1 746.6
O3 mean
65.14 75.15 134.52 56.00
590.5 702.0 721.5 590.5
27.23 31.78 49.94 27.23
O3 control ¼ No O3 added, O3þ ¼ ambient conc. 1.4, O3þþ ¼ ambient conc. 1.7. N4 ¼ no extra N deposition; N14 ¼ þ10 kg- and N54 ¼ þ50 kg ha1 a1 extra N deposition. a Measurement at 0e10 cm soil depth after 7 years of treatment (2010, end of experiment).
M. Volk et al. / Environmental Pollution 189 (2014) 161e168
5. Conclusions
Table 6 ANOVA for belowground biomassa.
Block O3 N pH O3 N
167
Df
denDf
L-ratio
p-value
2 2 1 1 2
4 4 95 95 95
4.94 4.26 9.07 4.09 0.86
0.085 0.119 0.003 0.043 0.65
a Measurement at 0e10 cm soil depth after 7 years of treatment (2010, end of experiment).
grassland. Possible effects on the more sensitive herb- or legume FGs, may be too small to be reflected in a biomass response at the community level, as observed by Wedlich et al. (2012) in upland grassland in the UK. Also, the potentially more sensitive FGs may to some extent be protected from O3 inside the canopy (Jäggi et al., 2006). But their low stature, barely exceeding the cutting height (2 cm), would rather exaggerate than obscure a small response. Despite the absence of an O3 effect on aboveground biomass, we tested whether belowground biomass would be more responsive. Earlier, belowground biomass was found to be reduced more relative to aboveground biomass under O3 enrichment (Jones et al., 2010; Wagg et al., 2012), thus leading to higher shoot/root ratio. This could be due to inhibited assimilate transport from sources to sinks, or to the preferential allocation of assimilates to new leaves in order to compensate for O3 damage and optimize light capturing (cf. Fuhrer and Booker, 2003). An increase in shoot/root ratio in response to O3 in the presence of extra N input was observed in O3þ, but not in O3þþ. For this grassland, Blanke et al. (2012) found the N-effect on shoot/root ratio to differ between species, but on the whole plant community level extra N increased the shoot/root ratio. In line with the optimal partitioning theory, suggesting that plants optimize the over all growth rate by adjusting their biomass allocation to capture the most limiting resource (Hilbert, 1990), O3 and N should act in the same direction, as observed in O3þ, but the absence of this interactive effect in O3þþ remains unexplained. The lowest yield (2006) occurred under low precipitation and low GDDMay, but the highest July temperatures (13.6 C compared to 7-yr median 9.7 C). The secondary role of precipitation must be seen on the background of an ecosystem that is not usually waterlimited during most of the main growing period. The melting snowpack in spring supplies the necessary soil water content during the most sensitive phase of plant development. Apparently, the basis for a high annual yield is determined mainly by high GDD values early in the year, favoring fast and early plant development. Favorable weather conditions should increase O3 sensitivity, but antagonistic effects of seasonal weather conditions make it difficult to identify interactions with O3 and N. However, POD values varied little between years, and the most productive year even yielded a slightly larger biomass harvest under O3þþ, which could hypothetically be related to compensatory growth, as found by Bielenberg et al. (2001) under conditions of O3-induced accelerated leaf senescence in trees with low N supply. Importantly, all responses in the N9 treatment up to year five (2008) were below 6%. But in the last year (2010) the N9 deposition level triggered a c. 31% increase in productivity, equivalent to 50% of the N54 effect. This late but strong response challenges present critical loads for Ndeposition in subalpine grassland (i.e. 5e10 kg ha1 yr1; Achermann and Bobbink, 2003; Bobbink and Hettelingh, 2011) and underlines the importance of carry-over effects. It implies first the availability of stored nutrients as resources for growth, and second the presence of large storage organs as sinks for newly acquired assimilates (Farrar and Jones, 2000). It suggests that such carry over effects in the perennial species of the subalpine grassland determine the responsiveness of the system to N enrichment.
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