Dario A Fornara
email: d.fornara@ulster.ac.uk

Sibylle Steinbeiss

Niall Mcnamara

Gerd Gleixner

Simon Oakley

Paul Poulton

Andrew J Macdonald

Richard D Bardgett

Niall ; Mcnamara

Gerd ; Ceh Gleixner

Paul ; Poulton

N P Mcnamara

P R Poulton

Increases in soil organic carbon sequestration can reduce the global warming potential of long-term liming to permanent grassland

Keywords: agro-ecosystems, climate change mitigation, legumes, nitrogen fertilizer, soil density fractionation, soil microbial community agro-ecosystems, climate change mitigation, legumes, nitrogen fertilizer, Park Grass Experiment, soil density fractionation, soil microbial community

ou non, émanant des établissements d'enseignement et de recherche français ou étrangers, des laboratoires publics ou privés.

The application of calcium-and magnesium-rich materials to soil, known as liming, has long been a foundation of many agroecosystems worldwide because of its role in counteracting soil acidity. Although liming contributes to increased rates of respiration from soil thereby potentially reducing soils ability to act as a CO2 sink, the long-term effects of liming on soil organic carbon (Corg) sequestration are largely unknown. Here, using data spanning 129 years of the Park Grass Experiment at Rothamsted (UK), we show net Corg sequestration measured in the 0-23 cm layer at different time intervals since 1876 was 2 to 20 times greater in limed than in unlimed soils. The main cause of this large Corg accrual was greater biological activity in limed soils, which despite increasing soil respiration rates, led to plant C inputs being processed and incorporated into resistant soil organo-mineral pools. Limed organomineral soils showed: (1) greater Corg content for similar plant productivity levels (i.e. hay yields); (2) higher 14C incorporation after 1950s atomic-bomb-testing; and (3) lower C:N ratios than unlimed organo-mineral soils, which also indicate higher microbial processing of plant C. Our results show that greater Corg sequestration in limed soils strongly reduced the global warming potential (GWP) of long-term liming to permanent grassland suggesting the net contribution of agricultural liming to global warming could be lower than previously estimated. Our study demonstrates that liming might prove to be an effective mitigation strategy, especially because liming applications can be associated with a reduced use of nitrogen fertilizer which is a key cause for increased greenhouse gas emissions from agro-ecosystems.

Abstract

The application of calcium-and magnesium-rich materials to soil, known as liming, has long been a foundation of many agro-ecosystems worldwide because of its role in counteracting soil acidity. Although liming contributes to increased rates of respiration from soil thereby potentially reducing soils ability to act as a CO 2 sink, the long-term effects of liming on soil organic carbon (C org ) sequestration are largely unknown. Here, using data spanning 129 years of the Park Grass Experiment at Rothamsted (UK), we show net C org sequestration measured in the 0-23 cm layer at different time intervals since 1876 was 2 to 20 times greater in limed than in unlimed soils. The main cause of this large C org accrual was greater biological activity in limed soils, which despite increasing soil respiration rates, led to plant C inputs being processed and incorporated into resistant soil organo-mineral pools. Limed organo-mineral soils showed: (1) greater C org content for similar plant productivity levels (i.e. hay yields); (2) higher 14 C incorporation after 1950s atomic-bomb-testing; and (3) lower C:N ratios than unlimed organo-mineral soils, which also indicate higher microbial processing of plant C. Our results show that greater C org sequestration in limed soils strongly reduced the global warming potential (GWP) of long-term liming to permanent grassland suggesting the net contribution of agricultural liming to global warming could be lower than previously estimated. Our study demonstrates that liming might prove to be an effective mitigation strategy, especially because liming applications can be associated with a reduced use of nitrogen fertilizer which is a key cause for increased greenhouse gas emissions from agro-ecosystems.

Introduction

Global warming mitigation requires an effective strategy portfolio which includes reduction in CO 2 emissions, but also increased CO 2 sequestration in long-lived pools, such as world soils (Lal, 2004;Six et al., 2004). Agricultural soils in particular have the potential to act as C sinks following changes in land-use from arable cropping to longterm grasslands, or after the introduction of no-tillage systems, and changes to lime and fertilizer inputs (Lal, 2004;Six et al., 2004;West & McBride, 2005;[START_REF] Hamilton | Evidence for carbon sequestration by agricultural liming[END_REF][START_REF] Biasi | Direct experimental evidence for the contribution of lime to CO 2 release from managed peat soil[END_REF]. The application of calcium-and magnesium-rich materials to soil (liming), has long been a common agricultural practice worldwide, because it counteracts soil acidity and helps sustain and improve plant growth (Edmeades & Ridley, 1998;Norton & Chang, 2003). Lime applications are estimated to be ~ 30 Tg y -1 in the US alone [START_REF] Hamilton | Evidence for carbon sequestration by agricultural liming[END_REF] and globally are expected to increase threefold in the next 50 years due to the expansion of intensive agricultural practices (Tilman et al., 2001).

Emissions of CO 2 from lime are currently calculated with the mass fraction of C (12%) in either limestone (CaCO 3 ) or dolomite (MgCa(CO 3 ) 2 ) assuming an emission factor of 100%, whereby all C in lime is released into the atmosphere (IPCC, 1997[START_REF] Ipcc ; Egglestone | Turnover of nitrogen-15labeled fertilizer in old grassland[END_REF]. Liming is expected to increase soil C losses, mainly because of lime-induced increases in soil biological activity and therefore rates of soil CO 2 respiration (Zelles et al., 1987;[START_REF] Badalucco | Effect of liming on some chemical, biochemical, and microbiological properties of acid soils under spruce (Picea abies L.)[END_REF]Brady & Weil 2002;[START_REF] Biasi | Direct experimental evidence for the contribution of lime to CO 2 release from managed peat soil[END_REF]. Recent studies, however, suggest at least two potential pathways through which lime could enhance CO 2 sink capacity of agricultural soils. First, lime dissolution could increase carbonic acid (HCO 3 ¯) concentrations in soil water solutions which could sequester 25-50% of lime C [START_REF] Hamilton | Evidence for carbon sequestration by agricultural liming[END_REF]. Second, long-term liming and fertilization treatments may contribute to redistribute C org from labile to more humified soil pools which are resistant to decomposition (Manna et al., 2007). This may occur through positive calcium (Ca)-induced effects on soil aggregate stability [START_REF] Grant | Residual effects of additions of calcium compounds on soil structure and strength[END_REF] and on the physical protection of soil C org [START_REF] Clough | Physical and chemical protection of soil organic carbon in three agricultural soils with different contents of calcium carbonate[END_REF], which ultimately contribute to enhanced soil organic matter stabilization (Morris et al., 2007).

Despite the evidence of positive lime-induced effects on the soil C org pool, no studies have yet quantified the effects of long-term liming on net changes in C org sequestration in humified soils. Such data is essential to estimate the net contribution of liming to soil C dynamics and to quantify the global warming potential (GWP) of liming in agro-ecosystems. In particular, it is important to know how different agricultural practices contribute to net greenhouse gas emissions if we are to fully understand their impact at the global scale (Robertson et al., 2000;Smith et al., 2008).

Here we examine long-term data collected from the Park Grass Continuous Hay Experiment at Rothamsted (UK). Park Grass was established in 1856 (Lawes & Gilbert, 1859) and is the oldest ungrazed permanent grassland experiment the world. The data was used to: (1) determine whether long-term liming has influenced net changes in soil

Methods

Experimental design and data

The Park Grass Experiment is located at Rothamsted, Harpenden, Hertfordshire, in the south-east of England (51 0 48'N and 0 0 22'W). It was established in 1856 on 2.8 ha of land that had been permanent pasture for at least 100 years (Lawes & Gilbert 1900;Richardson, 1938;Thurston et al., 1976). Four plots with contrasting fertilizer treatments (see Table 3 in Supporting Information 1), whose effects on soil properties and plant community composition had been documented since the beginning of the experiment (Richardson, 1938;Thurston et al., 1976;Sohi et al., 2001;[START_REF] Crawley | Determinants of species richness in the Park Grass Experiment[END_REF][START_REF] Hopkins | Soil organic carbon contents in long-term experimental grassland plots in the UK 22 (Palace Leas and Park Grass) have not changed consistently in recent decades[END_REF], were selected for this study. The whole experiment received a small amount of lime in the 1880s; 4.0 tCaCO 3 ha -1 on the plots chosen for this study. A test of liming started on half-plots in 1903 (1920 on plot 14) with 4.0 tCaCO 3 ha -1 applied every four years. A new liming scheme was introduced in 1965; depending on the fertilizer treatment different amounts of lime were applied with the aim of maintaining soil pH at 7 in one of four quarter-plots within each treatment, whilst one sub-plot remained unlimed (see Table 3 in Supporting Information 1). By 2005 soil pH values of sub-plots given the largest lime dressings within plots 7, 14, 9 and 3 were 7.0, 7.0, 7.2 and 7.3 respectively, whereas pH values of the corresponding unlimed sub-plots were: 4.9, 6.1, 3.6 and 5.3.

The lack of treatment replication in the Park Grass Experiment (see also Data analysis) was overcome to some extent by using soil samples (0-23 cm) taken from the same treatments over an extended period of time. In particular, measurements of soil carbon were done on archived soil samples collected in 1870, 1876, 1959, 1971, 1976, 1984, 1991[START_REF] Drinkwater | Legume-based cropping systems have reduced carbon and nitrogen losses[END_REF][START_REF] Brady | The nature and properties of soils[END_REF][START_REF] Crawley | Determinants of species richness in the Park Grass Experiment[END_REF]. Soils for 1971and 1976 were available only for limed and 3 in Supporting Information 1). The unlimed sub-plot given (NH 4 ) 2 SO 4 (plot 9) has developed through time a thick mat of undecomposed vegetation which was kept separate when soil sampling (i.e. we sampled within c. 3-23 cm of mineral soil). Soil C values for 1932 were taken from literature (Richardson, 1938) and were measured in soil samples collected to a depth of 20 cm. Because some soils were ground (e.g. in 1959) they could not be used for density fraction analyses. We used soil weights measured in 1876 and 1959 to estimate tonnes of organic C per hectare which did not change substantially between 1959 and 2005 [START_REF] Hopkins | Soil organic carbon contents in long-term experimental grassland plots in the UK 22 (Palace Leas and Park Grass) have not changed consistently in recent decades[END_REF].

Finally, we used long-term data of plant aboveground productivity (i.e. hay yields) which are available from the online Electronic Rothamsted Archive (ERA: http://www.rothamsted.bbsrc.ac.uk/eRA/).

Soil organic matter fractionation

In 1870 and 1876, soils were sampled with an open-ended metal box (30 x 30 x 23cm deep); one hole was sampled in 1870 and three in 1876. In later years, 15 to 20 soil cores were taken with a 2 or 3cm diameter auger from each sub-plot and bulked together.

Archived soils from different years (Table 3 in Supporting Information 1) were sieved to remove roots, dried at 40 °C for 5 days and stored in glass vials. We separated soil organic matter fractions by floatation and sedimentation using a sodium iodide (NaI) solution at a density of 1.80 g cm -3 following a detailed protocol (Sohi et al., 2001; see http://www.rothamsted.bbsrc.ac.uk/aen/CarbonCycling/pdf/Fractionation_manual.pdf).

This method distinguishes between: (a) a free-organo-mineral fraction (FR-SOM) at density < 1.80 g cm -3 which represents discrete free organic particles located between which is considered more recalcitrant than the FR-SOM and IA-SOM fractions (Sohi et al., 2001). Both soil bulk samples and soil fractions were analyzed for total C and N by combustion and gas chromatography (COSTECH Analytical ECS 4010 instrument). To determine organic C concentration, bulk soil subsamples were burned for 16 hours at 550 °C in a muffle furnace. Organic carbon was then calculated as the difference between total soil C and inorganic C measured in the residue ashes after furnace burning (we repeated the same procedure using organo-mineral soil fractions). Total C in soils sampled in 2002 was determined by measuring CO 2 produced following high temperature oxidation using a LECO CNS-2000 elemental analyser.

Soil respiration analyses

Sixteen soil samples (6 cm diameter 10 cm depth) were randomly collected in each of the 8 sub-plots in 2008 stored in cooler box for 6 hours and then in a fridge at 4°C for one day. Four soil samples from each sub-plot were then allocated to each of four controlled environment rooms operated at 4, 10, 15 and 20°C. Soil CO 2 flux measurements were made using a static chamber approach where each core was sealed in a gas tight incubation chamber (1.8 L) with a rubber septa in the lid and 8 ml gas samples were taken at 0, 60, 120 and 240 minutes after chamber closure. Samples were stored in Exetainer gas vials (Labco Ltd, UK) and CO 2 concentrations were analysed using a Perkin Elmer Autosystem XL Gas Chromatograph fitted with a flame ionisation detector ppm CO 2 (BOC, UK). CO 2 fluxes rates per gram dry weight of soil were calculated using a standard approach [START_REF] Holland | Soil CO 2 , N 2 O and CH 4 exchange[END_REF] after applying linear regression to the CO 2 concentrations versus time data for each replicate.

The Q 10 values were then calculated from an exponential model fitted to the CO 2 flux data for the four incubation temperatures. The Q 10 is the proportional increase in respiration as the temperature increases by 10 o C where the Q 10 = (R2/R1) 10/(T2-T1) with R1 and R2 being the respiration rates at two given temperatures (T1, T2) separated by 10 o C.

Microbial analyses

Phospholipid fatty acid (PLFA) analysis was used to gain insight into the effects of fertilization and liming treatments on microbial community composition. Four fresh soil samples were collected in June 2008 from each subplot 0-10 cm soil depth, stored in a cooler box for 16 hours, and freeze-dried. Microbial lipids were extracted from 2 g of each freeze-dried soil with a solvent system that included methanol, chloroform, and a phosphate buffer. Total extracted lipids collected in the organic phase were fractionated into neutral, glyco-, and polar lipids with chloroform, acetone, and methanol using silicic acid chromatography. We used i15:0, a15:0, i16:0, i17:0 and a17:0 PFLAs as indicative of Gram + bacteria whereas 16:1ω7, 16:1ω9, 16:1ω5 and 18:1ω7 PFLAs were indicative of Gram -bacteria [START_REF] Frostegård | Shifts in the structure of soil microbial communities in limed forests as revealed by phospholipid fatty acid analysis[END_REF]Pawlett et al., 2000;[START_REF] Kramer | Variable use of plant-and soil-derived carbon by microorganisms in agricultural soils[END_REF]. was reduced to graphite at 600°C with an excess of hydrogen gas. The resulting graphite was analyzed by 14 C AMS. We dated the organo-mineral soil fraction only because this is a large and relatively stable soil carbon pool, and is likely to give reliable information on the potential accumulation of 14 C after nuclear bombs testing in the 1950s.

Global Warming Potential (GWP) of agricultural liming

We estimated the global warming potential (GWP) for each fertilization and liming treatments applied to the Park Grass plots between 1984 and 2005 based on contributions of individual gases (CO 2 , CH 4 and N 2 O-N; see Supporting Information 2). GWP gives an estimate of the cumulative radiative forcing of individual greenhouse gases relative to some reference gas, usually CO 2 , over a specific time span, here 21 years (see also Robertson et al., 2000). We estimated the net CO 2 fluxes taking into account: (1) net soil

Data analysis

The Park Grass Experiment was designed before modern statistical ideas about replication and randomization had been developed [START_REF] Crawley | Determinants of species richness in the Park Grass Experiment[END_REF]. However, the size of the experimental plots (>100 m 2 ) and the number of soil samples collected within each plot (15 to 20), together with the repeated samplings over time, go some way to compensate for the lack of replication, particularly because this grassland was reasonably uniform before the experiment began (Lawes & Gilbert, 1859). Our main goal was to

Results

Soil C org sequestration

We found that the C org sequestration to 23 cm soil depth since 1876, and averaged across the four different fertilization treatments, was significantly greater in limed than in unlimed soils (MANOVA, F 1,6 = 10.8, P = 0.015; Fig. 1a). In particular, the net C org increase in limed soils since 1876 was 2 times and 20 times greater in 1932 and 1991 than in unlimed soils, respectively. Also, the total C org content of limed soils, averaged for four recent years (i.e., 1991, 1998, 2002 and 2005), was 16.2 ± 2.2 t ha -1 greater than in unlimed soils. Soils that have not received lime (except for the small amount applied in the 1880s) show no significant changes in C org sequestration between 1876 and 2005 (P = 0.34, Fig. 1a). We found the increased C org content of limed soils was mainly due to increases in C org in the organo-mineral soil fraction (F 1,5 = 29.9, P = 0.002; Fig. 1b).

Liming however did not affect the C org content of more labile soil fractions (Table 1; Table 4 in Supporting Information 1). Increased soil C org sequestration in the limed plots occurred while air temperatures at Rothamsted also increased > 1.0 °C between 1876 and 2005 (Fig. 1b).

Soil C org and aboveground plant productivity

We tested whether soil C org accrual could result from greater plant productivity in limed grasslands (i.e. increased C inputs to limed soils). Aboveground plant biomass (averaged for 25 years; see Fig. 4c in Supporting Information 1) was significantly higher (P < 0.0001) in limed (6.1 ± 0.4 t ha -1 ) than unlimed plots (4.9 ± 0.3 t ha -1 ), and was 12 significantly lower (P < 0.0001) in unfertilized (3.2 ± 0.1 t ha -1 ) than fertilized plots (6.3 ± 0.2 t ha -1 ; see Fig. 4d in Supporting Information 1). However, we could not detect any significant relationship between plant aboveground productivity and total soil C org in our plots (F 1,22 = 0.21, P = 0.65), or between plant productivity and net changes in soil C org between 1984 and 2005 (F 1,7 = 2.08, P = 0.2). Instead, we found that for similar hay yields across years, the organo-mineral fraction of limed soils had ~137% more C org than the same fraction of unlimed soils (see Fig. 5 in Supporting Information 1). We also found that in both unlimed and limed plots fertilized with PKNaMg (but not N) the rates of soil C org accrual since 1984 increased significantly (P < 0.04) with the proportion of legume biomass measured in a previous study [START_REF] Crawley | Determinants of species richness in the Park Grass Experiment[END_REF].

To identify potential belowground mechanisms responsible for increased C org in limed organo-mineral soils, we performed a linear regression which included soil pH as predictor variable. This analysis (R 2 = 0.6, F 1,40 = 19.3, P < 0.0001) shows that the C org of the organo-mineral soil pool strongly increased with soil pH (Fig. 2a). We did not find any significant fertilization treatment effect on total soil C org of the organo-mineral pool (ANOVA, F 3,40 = 0.77, P > 0.51). Also we could not find any significant fertilization effect on soil C org of the organo-mineral pool across years (MANOVA, F 3,4 = 0.12, P = 0.94 ). However, we found that rates of soil C org change in the organo-mineral pool between 1984 and 2005 were 5.4 times greater in limed than unlimed plots under the (NH 4 ) 2 SO 4 treatment (see blue bars in the small bars plot framed in Fig. 2a), and 9.2 times greater in unfertilized limed soils (red bars).

Changes in soil microbial community composition and soil respiration rates

We tested whether potential liming and fertilization effects on soil C org content were associated with changes in soil biology. We did not find any difference in microbial biomass between limed and unlimed soils (F 1,7 = 0.0008, P = 0.97 ), but we found that soil pH exerted a strong negative effect on the ratio of Gram positive to Gram negative bacteria (G+:G-; P < 0.0001; Fig. 2b). Lower G+:G-ratios were associated with a higher temperature sensitivity of soil respiration (Q 10 ) (Estimate = -0.78, t-ratio = -11.3, P < 0.0001), with greater soil respiration rates, and with higher microbial metabolic quotients (a measure of the microbial respiration per unit biomass; see Table 5 in Supporting Information 1). Liming also had positive significant effects on soil respiration rates at room temperatures of soil incubation of 20°C (F 1,30 = 5.9, P = 0.02), but not at incubation temperatures of 4°C, 10°C or 15°C. Finally, low G+:G-ratios in limed plots were associated with greater soil C org of the organo mineral fraction (Fig. 6 in Supporting Information 1).

Radiocarbon measurements and C:N ratios of the organo-mineral fraction

We found that 14 C incorporation (expressed as percent of modern C (pMC)) into organomineral fractions after atomic bomb-testing in the 1950s was significantly higher in limed soils (R 2 = 0.48, F 1,22 = 19.1, P = 0.0003; Fig. 3a). We also found that the C:N ratio of organo-mineral limed soils was negatively associated with pMC (R 2 = 0.64, F 1,22 = 37.5, P < 0.0001; Fig. 3b) suggesting a faster incorporation of C into the organo-mineral fraction of limed soils. In support of this idea (that more C has been processed and 14 incorporated into organo-mineral soils) we found that limed organo-mineral soils had lower C:N ratios that those of unlimed soils (R 2 = 0.40, F 1,22 = 14.2, P = 0.001).

GWP associated with liming and fertilization treatments

We found that limed plots all exhibited a net GWP which was lower than that of unlimed plots (Table 2). The negative GWP of limed plots means higher mitigation potential of greenhouse gases by permanent grasslands that receive regular liming applications. The lower GWP of limed grasslands was mainly determined by their high soil C org sequestration ability (Table 2) which offset the emissions of CO 2 from the production, transport and application of lime and fertilizers. Across the unlimed plots the PKNaMg and PKNaMg + (NH 4 ) 2 SO 4 treatments showed a mitigation potential which was, however, lower than the mitigation potential of the co respective fertilization treatments under lime applications (Table 2). Finally, unlimed plots receiving no fertilization (Nil, in Table 2) or receiving PKNaMg + NaNO 3 showed a higher GWP which was mainly due to loss of soil CO 2 (positive values in Table 2) during the 21 years period.

Discussion

Overall our results indicate that regular lime applications to permanent grassland not only increased soil pH over time (Fig. 1a), but also increased and maintained the C sink capacity of the soil to a greater extent than in unlimed soils. Soils that had never received lime, except for a small amount in the 1880s, show no significant change in C org sequestration between 1876 and 2005 (see Fig. 1a), which is consistent with a long-term study at Park Grass that did not find any significant change in the C org content of unlimed [START_REF] Hopkins | Soil organic carbon contents in long-term experimental grassland plots in the UK 22 (Palace Leas and Park Grass) have not changed consistently in recent decades[END_REF]. The positive liming-effect on total soil C org content was observed as early as 1932 (Richardson, 1938) and 1959(Thurston et al., 1976) in the same Park Grass plots, however it is the long-term changes in total soil C org content revealed in our study that shows the magnitude of the liming effect. In the work reported here, the C org sequestration measured in the 0-23 cm layer at different time intervals since 1876 was 2 to 20 times greater in limed than in unlimed soils. Also total C org content of limed soils, averaged for 1991, 1998, 2002 and 2005, was 16.2 ± 2.2 t ha -1 greater than in unlimed soils. It is not clear what caused the low C org content of limed soils measured in 1984 (Fig. 1a), although this could be partly related to the temporary absence of previously regular lime applications in two experimental plots since 1972.

The positive liming effect on net changes in soil C org sequestration was observed across different fertilization treatments, and the magnitude of such effect increased at increasing soil pH differences between limed and unlimed plots under the same fertilization treatment. For example, the unlimed grassland fertilized with (NH 4 ) 2 SO 4 (soil pH = 3.6) has developed a thick 'mat' of undecomposed dead plant material which represents a large organic C pool (i.e. ~ 24 t C ha -1 as measured in 1959). However, even after adding the C org content of this 'mat' to the C org content of the underlying mineral soil (averaged for 1991, 1998, 2002 and 2005), the overall C org content (85 ± 1.3 t C ha -1 ) was still significantly lower than the C org content of limed soils fertilized with (NH 4 ) 2 SO 4 (103 ± 5. We found that the increased C org content of limed soils was mainly due to increases in C org in the organo-mineral soil fraction (Fig. 1b), which is considered to be a stableresistant soil fraction (Sollins et al., 1996;Sohi et al., 2001). Increased soil C org sequestration occurred even though air temperatures at Rothamsted also increased (by > 1.0 °C between 1876 and 2005; Fig. 1b); increases in temperature are known to enhance soil C losses due to enhanced decomposition and soil respiration (Zelles et al., 1987;[START_REF] Badalucco | Effect of liming on some chemical, biochemical, and microbiological properties of acid soils under spruce (Picea abies L.)[END_REF][START_REF] Biasi | Direct experimental evidence for the contribution of lime to CO 2 release from managed peat soil[END_REF][START_REF] Dorrepaal | Carbon respiration from subsurface peat accelerated by climate warming in the subarctic[END_REF], and were also detected in this study. Although increased aboveground productivity would be expected to play an important role in enhancing soil C org accrual in limed grasslands through greater C inputs to soil, we did not find evidence to support this in our study. Instead, we found that the organo-mineral fraction of limed soils had ~137% more C org than the same fraction of unlimed soils despite having similar hay yields across years. We also found that rates of soil C org accrual since 1984 increased significantly (P < 0.04) with the proportion of legume biomass in both unlimed and limed plots fertilized with PKNaMg (but not N; see Thurston et al., 1976;[START_REF] Crawley | Determinants of species richness in the Park Grass Experiment[END_REF], which suggests that legumes might play a role in soil C accumulation, as previously suggested in both grassland and cropping systems [START_REF] Drinkwater | Legume-based cropping systems have reduced carbon and nitrogen losses[END_REF][START_REF] Fornara | Plant functional composition influences rates of soil carbon and nitrogen accumulation[END_REF][START_REF] De Deyn | Vegetation composition promotes carbon and nitrogen storage in model grassland communities of contrasting soil fertility[END_REF].

Accumulation of C org in limed grasslands could also depend on increased root C inputs to these soils, which is consistent with results of other studies at Park Grass (Jenkinson et al., 2004) which showed that total root mass measured to 23 cm depth in limed soils receiving (NH 4 ) 2 SO 4 was > 2 times greater than that of unlimed soils receiving NaNO 3 (9.1 vs 3.3 t ha -1 respectively). However, unlimed unfertilized soils also showed high root [START_REF] Jensen | Simulating trends in soil organic carbon in long-term experiments using the soilplant-atmosphere model DAISY[END_REF], suggesting other processes (e.g. root C turnover rates, root exudation etc.) might play a significant role in C org sequestration in these grassland soils.

Changes in soil microbial community and soil respiration rates

Liming had strong effects on the microbial community composition, resulting in lower G+:G-ratios in limed compared to unlimed plots. Low G+:G-ratios were in turn associated with a higher temperature sensitivity of soil respiration, with greater soil respiration rates and with higher microbial metabolic quotients. This suggests less efficient microbial utilization of C in limed compared to unlimed soils, whereby microbes respire more C per unit of degradable C (see Wardle & Ghani, 1995). These patterns show similarities to soils that have received inputs of easily degradable C or lime [START_REF] Badalucco | Effect of liming on some chemical, biochemical, and microbiological properties of acid soils under spruce (Picea abies L.)[END_REF]Thirukkumaran et al., 2000), and support the notion that liming increases both the abundance of Gvs. G+ bacteria [START_REF] Frostegård | Shifts in the structure of soil microbial communities in limed forests as revealed by phospholipid fatty acid analysis[END_REF]Pawlett et al., 2000) and soil microbial activity (Zelles et al., 1987;[START_REF] Badalucco | Effect of liming on some chemical, biochemical, and microbiological properties of acid soils under spruce (Picea abies L.)[END_REF], resulting in higher C losses through soil respiration. If greater soil biological activity is responsible for higher soil respiration rates in our study, albeit measured using soils from a single sampling, it is possible that increased biological activity will also contribute to enhanced incorporation of plant C into more humified soil pools. Indeed, low G+:G-ratios in limed plots were associated with greater soil C org of the organo mineral fraction. A potential mechanism responsible for this may involve the role of G-bacteria which are often rootassociated, are largely responsible for the turnover and processing of labile plant material [START_REF] Kramer | Variable use of plant-and soil-derived carbon by microorganisms in agricultural soils[END_REF], and can increase their utilization of root exudates in limed (Rangel-Castro et al., 2005). Also there is evidence that C-based compounds from the mass of Gram -bacteria after cell death can significantly contribute to SOM formation and stabilization [START_REF] Kindler | Fate of bacterial biomass derived fatty acids in soil and their contribution to soil organic matter[END_REF].

The idea that more plant C is being processed in limed soils is supported by radiocarbon analyses, which showed higher 14 C incorporation into organo-mineral fractions of limed than unlimed soils after atomic bomb-testing in the 1950s. As a result of their higher C org turnover, the C:N ratio of organo-mineral limed soils was also lower than that of unlimed soils and was negatively associated with pMC (see Fig. 3b). Finally, despite limed soils receiving, on average, 2.8 ± 0.3 t C ha -1 (from CaCO 3 ) since 1990, we estimated that only 2.7% of this C input remained by 2005. Therefore, although this C input to limed soils contributed to the increase in inorganic C content, the bulk of the C accrual in limed soils was via organic C inputs.

Global warming potential of long-term liming to permanent grassland

The increased soil C org sequestration in limed soils was a key cause of their lower GWP when compared with the GWP of unlimed soils. This means the net contribution of agricultural liming to global warming could be lower than previous estimates suggest (IPCC 1997[START_REF] Ipcc ; Egglestone | Turnover of nitrogen-15labeled fertilizer in old grassland[END_REF]. Regular liming applications might prove to be a potential mitigation strategy to global warming if the magnitude of soil C org accrual we observed in our study is confirmed across a variety of grassland systems (see Soussana et al., 2007). But controls on the use of N fertilizers are also likely to be necessary (Robertson et al., 2000;Smith et al., 2008). Our results indicate that further reductions in the GWP of permanent grassland systems may be achieved by combining the use of lime and PKNaMg fertilizers 2) to increase the abundance of legumes [START_REF] Crawley | Determinants of species richness in the Park Grass Experiment[END_REF], which may provide the N necessary to sustain productivity and improve soil C sequestration.

Conclusions

Overall, our results show that regular lime applications to long-term grassland enhanced soil organic C sequestration, primarily by increasing the C org content of a resistant soil organo-mineral pool. Thus, the net contribution of agricultural liming to global warming could be lower than previously estimated (IPCC 1997[START_REF] Ipcc ; Egglestone | Turnover of nitrogen-15labeled fertilizer in old grassland[END_REF] at least for permanent grasslands. More importantly our study highlights two aspects that, if proved general, could affect current predictions on long-term soil C changes in managed grasslands. First, our results show that C sequestration in limed soils was enhanced by elevated soil C turnover rates, and that this was associated with increases in soil pH and changes in soil biological activity. Second, large variations in the organo-mineral C org content of limed soils over time (see Fig. 1b) raise questions about the long-term stability of these organomineral pools. Further studies are required in both of these areas if we are to improve our mechanistic understanding of the key processes controlling C dynamics in grassland soils. MANOVA was performed on C org values for 1876, 1932, 1959, 1984, 1991, 1998, 2002 and 2005. Archived soils were also available for limed plots in 1971 and for unlimed , 1984, 1991, 1998 and 2005); also mean increase in annual temperatures (solid line) since 1876 at Rothamsted. 1971,1976,1984,1991,1998

C

  org sequestration across experimental plots which have received four different long-term fertilization treatments (see Methods); (2) examine aboveground-belowground interactions that could influence long-term lime-induced effects on soil C org sequestration; (3) estimate the global warming potential (GWP) of agricultural liming on the basis of its impact on changes in soil C org sequestration and CO 2 fluxes associated with liming and fertilization treatments on the Park Grass Experiment.

  ; (b) an intra-aggregate organo-mineral fraction (IA-SOM) also at density < 1.80 g cm -3 which represents discrete organic particles located within stable soil aggregates; and (c) a residual heavy organo-mineral fraction at densities > 1.80 g cm -3

  Results were calibrated against a certified gas standard comprising 541

  Radiocarbon measurements were performed at the Max-Planck-Institute for Biogeochemistry in Jena, Germany. Samples were weighed into tin capsules and combusted by an elemental analyzer. The evolved CO 2 was transferred into a glass tube cooled with liquid nitrogen containing the iron catalyst necessary for reduction. The CO 2

  CO 2 changes between 1984 and 2005; (2) liming contribution to GWP through increased soil CO 2 emissions using IPCC factors(IPCC, 1997[START_REF] Ipcc ; Egglestone | Turnover of nitrogen-15labeled fertilizer in old grassland[END_REF]; (3) liming contributions to GWP due to the production and transport of lime; (4) GWP associated with the production and transport of fertilizers; (5) GWP associated with the emissions of N 2 O and CH 4 from soil; and (6) GWP associated with machinery emissions during field operations (i.e. liming, fertilization and hay cutting activities; see Supporting Information 2).

  address whether and why net changes in soil C org sequestration might occur across years depending on different liming and fertilization treatments rather than to compare the total soil C org content of different experimental plots. To determine potential liming and fertilization effects on soil C org sequestration across years we used repeated measures MANOVA. The distribution of soil C and N values for each year met the preconditions of a normal distribution (α = 0.01, p > 0.55 for multiple Shapiro-Wilk tests). Linear and multiple regression analyses were performed to address potential relationships between different soil and plant variables. To quantify how soil C org content of the organo-mineral fraction changed in limed and unlimed plots between 1984 and 2005 under different fertilization treatments, we calculated the annual rate of change in C org (r). To do this, we used the C org content of each sub-plot in 1984 and in 2005, with r being: r = Log e [(C org 1984 /C org 2005 ]/t, where t, time, is 21 years. This gives, for each sub-plot, the rate of change in the C org content of the organo-mineral fraction with units of yr -1 .

  3 t C ha -1 ; soil pH = 7.2). The question remains, however, about what mechanisms underlie such C org accrual in the limed soils of the Park Grass experiment.

  content of the organo-mineral soil fraction

  2004) Soil carbon sequestration impacts on global climate change and food security. Science, 304, 1623-1626. Lawes JB, Gilbert JH (1859) Report of experiments with different manures on permanent meadow land. Clowes, London. Lawes JB, Gilbert JH (1900) Agricultural, botanical and chemical results of experiments on mixed herbage of permanent grassland, conducted for many years in succession on the same land. Philosophical Transactions of the Royal Society Series B -Biological Sciences, 192, 139-210. Manna MC, Swarup A, Wanjari RH, Mishra B, Shahi, DK (2007) Long-term fertilization, manure and liming effects on soil organic matter and crop yields. Soil and Tillage Research, 94, 397-409. Morris SJ, Bohm S, Haile-Mariam S, Paul EA (2007) Evaluation of carbon accrual in afforested agricultural soils. Global Change Biology, 13, 1145-1156. Norton LD, Chang XC (1998) Liming to improve chemical and physical properties of soil. In: Handbook of Soil Conditioners: Substances that enhance the physical properties of soil (eds Wallace A, Terry RE), pp. 309-332. Marcel Dekker, New York. Pawlett M, Hopkins DW, Moffett BF, Harris JA (2009) The effect of earthworms and liming on soil microbial communities. Biology Fertility Soils, 45, 361-369. Rangel-Castro JI, Killham K, Ostle N et al. (2005) Stable isotope probing analysis of the influence of liming on root exudate utilization by soil microorganisms. Environmental Microbiology, 7, 828-838.

  1938) The nitrogen cycle in grassland soils with special reference to the Rothamsted Park Grass Experiment. Journal of Agricultural Science, 1, 73-121. Robertson GP, Paul EA, Harwood RR (2000) Greenhouse gases in intensive agriculture: Contributions of individual gases to the radiative forcing of the atmosphere. Science, 289, 1922-1925. Six J, Ogle SM, Breidt FJ, Conant RT, Mosier AR, Paustian K (2004). The potential to mitigate global warming with no-tillage management is only realized when practiced in the long term. Global Change Biology, 10, 155-160. Sohi SP, Mahieu N, Arah JRM, Powlson DS, Madari B, Gaunt JL (2001) A procedure to isolating soil organic matter fractions suitable for modeling. Soil Science Society America Journal, 65, 1121-1128. Sollins P, Homann P, Caldwell, BA (1996) Stabilization and destabilization of soil organic matter: mechanisms and controls. Geoderma, 74, 65-105. Soussana JF, Allard V, Pilegaard K et al. (2007) Full accounting of the greenhouse gas (CO 2 , N 2 O, CH 4 ) budget of nine European grassland sites. Agriculture Ecosystems and Environment, 121, 121-134. Smith P, Martino D, Cai Z et al. (2008) Greenhouse gas mitigation in agriculture. Philosophical Transactions of the Royal Society B, 363, 789-813. Thirukkumaran CM, Parkinson D (2000) Microbial respiration, biomass, metabolic quotient and litter decomposition in a lodgepole pine forest floor amended with nitrogen and phosphorous fertilizers. Soil Biology and Biochemistry, 32, 59-66. Thurston JM, Williams, ED, Johnston AE (1976) Modern developments in an experiment on permanent grassland started in 1856: Effects of fertilisers and lime on crop and soil analyses. Annales Agronomiques, 27, 1043-1082. Tilman D, Fargione J, Wolff B et al. (2001) Forecasting agriculturally driven global environment change. Science, 292, 281-284. Wardle DA, Ghani A (1995) A critique of the microbial metabolic quotient (qCO 2 ) as a bioindicator of disturbance and ecosystem development. Soil Biology and Biochemistry, 27, 1601-1610. West TO, McBride AC (2005) The Contribution of agricultural lime to carbon dioxide emissions in the United States: dissolution, transport, and net emissions. Agriculture Ecosystem and Environment, 108, 145-154. Zelles L, Stepper K, Zsolnay A (1987) The effect of lime on microbial activity in spruce (Picea abies L.) forests. Biology and Fertility of Soils, 9, 78-82.

Table 1 .

 1 Results of density fractionation analyses performed on soil samples collected from the Park Grass plots. Total C and N are expressed both as percentage and as unit weight per kilogram of soil (mean ± S.E.) and averaged for three recent years 1991, 1998 and 2005. FR-SOM = free-aggregate soil fraction; IA-SOM = intra-aggregate soil fraction. "Recovered" indicates the percentage of total C and N recuperated at the end of the density fractionation normalized to the initial mass of soil fractionated. C (%) N (%) C:N C (g C kg -1 soil) N (g N kg -1 soil)

Fig. 1 .

 1 Fig. 1. (a) Liming effects on bulk soil C org (t C ha -1 ) in the Park Grass Experiment.

  plots in 1976. The short vertical bars near the x axis indicate main lime applications (in 1976 lime was applied to only two plots). Solid lines indicate soil pH values for both limed and unlimed plots. (b) Liming effects on C org of the organo-mineral fraction for different years (bar histograms; MANOVA was performed on C org values for 1876

Fig. 2 .

 2 Fig. 2. (a)The relationship between soil C org of the organo-mineral fraction and soil pH

  and 2005. The small bars plot framed in the left-upper corner shows the rate of soil C org change for 21 years [Log (soil C 2005/soil C 1984)]/21 of the organo-mineral fraction (filled bars = limed plots; empty bars = unlimed plots within each fertilization treatment). (b) The relationship between Gram+:Gram-bacteria ratio on soil pH as both measured in 2008 (Estimate = -0.19, t-ratio = -10.8, P < 0.0001).

Fig. 3 .

 3 Fig. 3. (a)The relationship between the percent of modern C (pMC) in the organo-

  Fig. 1aFig.1b

Table 2 .

 2 Relative GWPs for different fertilization and liming treatments based on measurements and estimates of multiple CO 2 sinks and sources (see Supplementary material_2 for details). GWP was estimated for a21-year period between 1984 and 2005. Negative values indicate a global warming mitigation potential. Units are CO 2 equivalents (g CO 2 m -2 y -1 ).

	Bulk soil	Limed	3.89 ± 0.1	0.31 ± 0.01	12.6	39.2 ± 1.19	3.16 ± 0.11
		Unlimed	3 ± 0.08	0.24 ± 0.008	12.5	30.1 ± 0.88	2.39 ± 0.08
	FR-SOM	Limed	7.7 ± 0.84	0.42 ± 0.03	18.3	1.70 ± 0.28	0.09 ± 0.02
		Unlimed	7.5 ± 0.89	0.36 ± 0.01	20.8	1.73 ± 0.24	0.08 ± 0.01
	IA-SOM	Limed	11.4 ± 0.66	0.47 ± 0.02	24.3	0.59 ± 0.007	0.02 ± 0.002
		Unlimed	12.1 ± 0.59	0.45 ± 0.03	26.8	0.61 ± 0.006	0.02 ± 0.002
	Organo-mineral
		Limed	3.37 ± 0.09	0.29 ± 0.007	11.6	31.2 ± 0.89	2.67 ± 0.06
		Unlimed	2.5 ± 0.06	0.20 ± 0.005	12.5	23.4 ± 0.62	1.91 ± 0.05
	All fractions Limed				33.5 ± 1.20	2.78 ± 0.02
		Unlimed				25.7 ± 1.02	2.01 ± 0.02
	Recovered (%)
		Limed				85.4 ± 5.2	87.9 ± 6.1
		Unlimed				85.5 ± 5.9	84.1 ± 5.5

Global Change Biology

Acknowledgments: This research was supported by a Marie Curie Outgoing Fellowship issued to DAF within the Work Programme 2004, "Structuring the European Research Area" (2002-2006), and based at Lancaster University. We thank Helen Quirk and Gerlinde De Deyn who provided support during the laboratory work at Lancaster 20 University, and David Tilman for helpful comments on early drafts of the manuscript. We also thank NERC QUEST Thematic Programme for financial support to NPM. Finally, Rothamsted Research is an institute of the Biotechnology and Biological Sciences Research Council and receives support from the Lawes Agricultural Trust.


Leon Jl Van Den Berg

Philippine Vergeer

Tim Cg Rich

Simon M Smart

Dan Guest

Mike R Ashmore

Leon Jl Van Den Berg
email: l.vandenberg@science.ru.nl

Direct and indirect effects of nitrogen deposition on species composition change in calcareous grasslands

Keywords: Nitrogen deposition, species diversity, soil acidity, Festuco Brometalia grasslands, management Nitrogen deposition, species diversity, soil acidity, FestucoBrometalia grasslands

come   L'archive ouverte pluridisciplinaire

recently, analysis of data from a national survey covering all plant communities in the UK 10 corroborated these findings for both heathlands and acid grasslands, but failed to find a 11 significant relationship between N deposition and species richness in calcareous grasslands, 12 even when interacting environmental factors such as rainfall and altitude were included in the 13 models (Maskell et al. 2010). However, significantly more species characteristic of fertile 14 soil conditions, and a higher grass/forb ratio, was found in calcareous grassland plots with 15 higher Ndeposition. These findings suggest that the responses of calcareous grasslands to 16 increased N deposition differ from those of other semi-natural ecosystems, such as acid 17 grasslands and heathlands. This may be due to several factors, including (a) the importance of In addition, the limitations of the studies undertaken to date need to be considered. deposition and soil conditions, we analysed the data using both mixed effect models and 1 structural equation modelling (SEM). The SEM models allow explanatory variables to be 2 placed in 'series' rather than 'parallel', using sequential arrangements that reflect 3 hypothetical causal chains linking the effects. Such models provide insights that are not ). The survey method was specifically designed to be repeatable with a 5 minimum of observer bias.All the selected sites were nature reserves with active management 6 to maintain species diversity and characteristic species; 56 sites were selected, with 1-6 plots 7 on each site, resulting in a total of 128 plots. Sites were chosen to cover a wide geographical 8 range (Fig. 1), a large climatic gradient, a wide range in deposition of air pollutantsand all 9 types of calcareous grasslands known in the UK, using the National Vegetation Classification 12 Metal detectors were used for relocation of the plots. In each plot, 36 quadrats were 13 positioned in exactly the same pattern as in 1990-1993 and in each quadrat the 14 presence/absence of all vascular plants was recorded. One soil sample, composed of 5 subsamples, was taken from the top 10cmof each plot using an auger (3cm diameter). Soil 16 samples were stored at 4 °C until further analysis. Sward height and soil depth wererecorded Extractions were performed on fresh soil and corrected for moisture content after drying for This resulted in a dataset for only 52 species. Grass to herb ratio was calculated as the ratio 19 between the numbers of grass and herb species at the plot level. 11

For the 9 plots where detailed stocking was not provided we estimated the grazing intensity 12 from the sward height, based on the highly significant correlation (r 2 =0.58)between sward 13 height and livestock units obtained from the other plots. Linear regression between the total number of vascular plant species and Ndeposition did not 4 reveal a significant relationship in the 1990-1993 dataset (Fig. 2a), butNdeposition did show 5 a slight positive relationshipwith the grass to herb ratio (Fig. 2b).Almost no scarce and rare 6 species were observed at plots with Ndeposition above 25 kg N ha -1 y -1 (Fig. 2c); the average 7 number of scarce and rare species was 1.6 in the 77plots with deposition below 25 kg N ha -1 y - 8 1 compared to 0.3 in the 29 plots with deposition above 25 kg N ha -1 y -1 (ANOVA, df = 1, 9 F=13.62, p<0.001).

10 11

Minimum winter temperature and maximum summer temperature were highly 12 correlated;only maximum summer temperature was included in the model, as this was the 13 strongerpredictor according to the PCA.For similar reasons, organic content of the soil, Olsen 14 P and slope (which were all strongly correlated with pH) and mean annual rainfall (which was strongly correlated with maximum summer temperature) were excluded from the 16 analysis.Grazing pressure entered the analysis, but wasnot a significant explanatory variable 17 for any of the dependant parameters (Table 1a).

18

Maximum summer temperature and soil depth were the most important variables explaining 20 variation in species composition and speciesrichness in the 1990-1993 dataset ( 1b).In contrast to the spatial models, soil depth entered the modelsfor change in 1 number of oligotrophic, acidophilic and acidophobic species, and was never significant.

2 3
Soil pH showed a significant positive relationship with the change in the species richness, i.e.

4

there tended to be an increase in plots with a higher soilpH. In addition, soil pH was The models showed a significant negative effect of Ndeposition on changes in H' and 12 evenness (Table 1b). However, Ndeposition was not significantly related to change in the the model was a good fit. In addition, the CFI of 1.00, the NFI of 0.94, the TLI of 1.14and the 1 IFI of 1.04 that we calculated all indicated a good model fit.

3

As expected from the linear models (Table 1b), the SEM model showed that Ndeposition is a 4 significant explanatory parameter for the changes in H'of the plots (Fig. 4a, Table 2a). A 2a). Grazing pressure had an independent 14 direct negative effect on change in H'. 2b). The change in species richness was positively related to soil pH, as was herb species number, 16 indicating that plots of higher buffering capacity tended to gain species and plots with slightly Table 1:

2
Effect sizes of the predictor variables (expressed as the coefficient ± s.d.), for the subset of predictors that were included in the model with the

  the observed effects. For change in Shannon diversity, direct effects of N deposition were the dominant mechanism and there was an independent effect of change in grazing intensity. In contrast, for change in herb species number, indirect effects on soil acidity, linked to both N and S deposition, were more important than direct effects of N deposition.Atmospheric deposition of nitrogen (N) and sulphur (S) compounds increased dramatically in3 Europe in the second half of the 20th century, mainly as a result of increased fossil fuel 4 combustion and intensive agriculture (Galloway, 1995). Despite a strong decline in S 5 emissions and deposition in most parts of Europe since the 1990's (Vestrenget al. 2007), N 6 deposition has not declined to the same extent. Both N and S deposition has continued to 7 increase in other parts of the world, especially in parts of India and China (Galloway & 8 Cowling, 2002; Denteneret al. 2006).

N

  deposition can result in severe eutrophication via the direct input of nutrients but also 11 acidification via indirect processes such as nitrification (Galloway, 1995; Bobbink et al. 12 1998). N deposition has been identified as one of the most important global drivers of 13 biodiversity change in terrestrial ecosystems (Vitouseket al. 1997; Salaet al. 2000, Phoenix et 14 al. 2006), and has been linked to changes in plant species composition in a wide range of 15 terrestrial biomes from the Arctic to the tropics (e.g. Bobbink et al. 2010).Some of the most 16 affected terrestrial habitats are nutrient-poor semi-natural habitats such as heathlands and 17 grasslands. Calcareous grasslands are floristically amongst the most species rich communities 18 of North-Western Europe, containing many rare and endangered plant species (e.g. Willems, 19 1990). Their biodiversity is threatened by changes in management (e.g. Tansley, 1922; 20 Willems, 1978), habitat fragmentation (e.g. Fisher &Stöcklin, 1997) and disturbance (e.g. 21 Smart et al. 2003a), among other factors.

22 23 Experimental7

 23 studies in Dutch calcareous grasslandshave shown that increased N deposition 24 results in the loss of characteristic plant species and an increase of highly competitive grass 25 Brachypodiumpinnatum (e.g.Willems 1990; Bobbink et al. 1998). In the UK, 1 changes in species composition, a reduction in flowering, and a reduction in the forb/grass 2 ratio have been shown in a long-term N-addition experiment (Carroll et al. 2003), but 3 encroachment by B. pinnatum or other highly competitive grasses such as Bromus erectus 4 was not observed, possibly due to co-limitation by phosphorus (Morecroft et al. 1994; Wilson 5 et al. 1995; Carroll et al. 2003; Phoenix et al. 2003). 6 Stevens et al. (2004) found a significant negative spatial relationship between N deposition 8 and species richness in a national survey of one acidic grassland community in the UK. More

  9

  18 co-limitation by phosphorus, (b) active management and conservation to protect characteristic plant species and (c) the occurrence on shallow well buffered soils that are less 20 sensitive to acidification caused by N and S deposition.

23

  Experimental studies measure change over a relatively short period of time with N deposition 24 as the only variable factor. Field surveys which explore the spatial associations between N richness, or other response variables, do not allow an assessment of 1 whether these are due to long-term historical deposition or recent changes in deposition, such 2 as those simulated in experimental studies.Although long-term vegetation change in 3 calcareous grasslands in the UK since 1950 has been described by Bennie et al. (2006), their 4 analysis was not designed to assess the effect of N deposition, and there are no field surveys 5 relating recent changes in species composition or species numbers to atmospheric N 6 deposition, using plots that can be relocated with confidence.A further limitation to the 7 interpretation of field survey data is the multi-collinearity between predictor variables such as 8 N deposition and climate variables, and earlier studies have highlighted the limitations that 9 multiple regression or covariance analysis have in predicting the relative effects of the 10 explanatory variables on dependent variables such as species richness (e.g. Stevens et al. 11 2004, Maskell et al. 2010).12 13 We investigated the direct, indirect and interacting effects of atmospheric N deposition, other 14 confounding environmental variables, and site management on species richness and species 15 diversity in calcareous grasslands in the UK.We based our study on a network of 128 16 permanently marked plots in areas of calcareous grassland spread acrossthe UK that were 17 specifically established in the early 1990s, with the aim of monitoring long-term impacts of 18 climate change and air pollution, by John Rodwell and colleagues(Rich et al. 1993). Between 19 2006 and 2009, 48 plots were revisited and additional soil measurements were made. This 20 allowed us to assess the effect of N deposition and other factors on both spatial variation and 21 recent temporal changes in species richness and other relevant variables. 22 In order to test if changes in species richness and composition are significantly affected by 24 atmospheric Ndeposition in a matrix of effects of factors such as management, climate, S

4

  possible with standard covariance analyses into the relative importance of different direct and 5 indirect effects of N deposition. is a significant explanatory driverof spatial variation in species richness, 10 species diversity and the frequency of individual speciesin calcareous grasslands. 11 12 2) Ndeposition is a significant explanatory driverof change over the last 15 years in species 13 richness, species diversity and the frequency of individual species in calcareous grasslands.

2(

  changes in species richness, species diversity and the frequency of individual 16 species in calcareous grasslands can be explained by a combination of the effects of 17 management, direct effects of N deposition, and indirect effects of S and N deposition via soil Calcareous grassland survey: 1990-1993 survey 3 A large national survey of calcareous grasslands was carried out between 1990 and 1993 4 Rich et al. 1993

10(

  NVC)(Rodwell, 1998). On most sites, plots were chosen on both south and northfacing 11 slopes to account for effects of aspect. Each plot covered an area of 144 m 2 (typically 12 measuring 12×12 m or more rarely 6×24 m) and was established in a floristically 13 homogeneous stand of calcareous grassland. Individual plots were permanently marked at 14 unique places with buried copper coils. Detailed field notes, describing the exact location of 15 the plots, photographs, sketch maps and measurements from fixed features were also recorded, to ensure successful relocation of each plot in future surveys.The presence/absence 17 of all vascular plants was recorded in 36 50×50 cm quadrats positionedwithin each permanent 18 plot.Soil pH of the top 10 cm was determined in the field by mixing soil into a slurry with 19 distilled water (1:1vol:vol), equilibrating for 5 minutes, and measuring using a calibrated pH 20 meter. Soil type, soil depth, slope, aspect and sward height were recorded for each plot. For 21 each site, the size, its protection status and detailed information on local management 22 (grazing intensity) were noted.Data from the original reports could not be recovered for 7 23 plots (4 sites), resulting in available data for 121 plots.

  24

2

  In 2006, 35 out of the 56 sites were selected for re-survey (48 plots), covering a wide range of 3 Ndeposition.Sites were selected to reduce the effects of spatial autocorrelation in atmospheric 4 N deposition (e.g. by selecting high and low Ndeposition sites at similar latitudes). Sites and 5 plots withNVC communities CG1, CG2, CG10, CG11 and CG12 were targeted where 6 possible to reduce variation due to initial vegetation composition.Onlysixof the plots selected 7 for re-survey were identified as supporting different NVC communities (CG3, CG8, CG9), 8 mainly due to the presence of grasses such as Bromus erectus (CG3) or Sesleriaalbicans 9 (CG8&9).In broad terms, plots in the lowlandsof Wales and England focussed on stands of 10 CG1 and CG2, while those in the uplands of Wales and Scotland focussed on stands of 11 CG10, CG11 or CG12.

17

  at each plot and detailed information on the local management (grazing intensity) for the past 18 15 years was obtained from local managers at the plot level.

21

  Soil samples from the 2006-2009 survey were thoroughly mixed prior to extraction.

  22

  datafor each site wereobtained from the Centre of Ecology and 23 Hydrology (CEH) for two periods: 1993-1995 and 2006-2008 for the surveys of 1990-1993 24 2009 respectively. The Ndeposition ranged from 7.3 to 40.7 kg N ha -1 y -1 while 1 Sdeposition ranged from 8.0 to 27.6 kg S ha -1 y -1 .Climate data for each site wereobtained 2 from the 5km gridded climate database held by the UK Meteorological Office 3 (www.metoffice.co.uk). 10 year averaged data for the periods 1981-1991 and 1997-2006 4 respectively were used for the 1990-1993 and 2006-2009 surveys, including mean minimum 5winter temperature, mean maximum summer temperature, mean annual temperature and 6 mean annual rainfall. The aspect of each plot was converted to a 0-1 scale using 7 sin(a/(360/π)), where a = the aspect in degrees. This resulted in a gradient from North =0 to 8 South =1, with East and West both at 0.5. Where information on stocking was available from 9 local site managers, grazing intensity for each plot was calculated in terms of livestock units 10 (LU) per hectare per year using the conversion factors provided in Attwood &Heavey (1964).

16

  Principal componentanalysis (PCA) in CANOCO(TerBraak, 1988) was used to determine 17 the variables explaining most of the variation between the plots. For each survey, we selected 18 the strongest, non-correlated, variables for further analysis.

19 20(

 19 Possible combinations of predictor variables were explored to assess the effects of 21 Ndeposition, soil conditions, climate data and grazing pressure on species richness and 22 species composition. Data from the 1990-1993 survey were used in the spatial analysis, while 23 data from both surveys were used in the temporal analysis. All models were explicitly tested 24 in the response variable and residuals by inspection of semi-1 variograms. In the presence of spatial autocorrelation, generalized linear mixed-effect modes 2 (GLMM) were used to control for this spatial autocorrelation. In these models, a correlation 3 structure was added to correct for spatial autocorrelation. Correlation structures such as 4 corExp, and corSpher were used with the "form=~Easting+Northing" argument in the 5 correlation option to calculate the Euclidean distances (using Pythagoras theorem) between 6 sites with coordinates given by Easting and Northing. The models were ranked according to 7 the Akaike Information Criterion (AIC) and only the models and corresponding correlation 8 structure with the lowest AIC, representing the 'best fitting models', were selected. When no 9 spatial autocorrelation was observed, multiple linear regression models were used. Again, 10 models were ranked according to AIC and only the models with the lowest AIC were 11 selected. To test for multicollinearity, variance inflation factors (VIF) were computed for 12 each predictor variable in the model. Predictor variables that were highly correlated (VIF>10: 13 Gujarati, 1995) were not analysed in the same model; in those situations only the predictor 14 variable that explained most of the variance was selected. 15 16 For the change in Shannon diversity H', the ecological significance of the effect was 17 quantified by dividing the data into two groups, one with N deposition below the lowest, and 18 one with N deposition above the highest, critical load of N deposition for this vegetation type 19 (<15 and >25 kg N ha -1 y -1 respectively, Bobbink et al., 2003). The Cohen d-value was then 20 calculated according to: , where are the means of the data for sites 21 with high and low N deposition respectively and the standard deviations in these 22 groups.The calculated value of d was checked against Cohen's classification (Cohen, 231988).We also compared the predicted change in H' from our models at sites with N kg ha -1 y -1 with the value of H' in a reference dataset representing an 1 highly undesirable endpoint that could be realised given continued high N deposition when 2 grazing and/or cutting is maintained. We selected botanical data (n=327) from within the 3 randomly located plots of the Countryside Survey(CS) of Great Britain which independent 4 land cover mapping showed could be referenced to the Neutral Grassland Broad Habitat 5 Jackson 2000). The change in H' from 1990-93 to 2006-09 in the high N deposition plots 6 was then expressed as a percentage of the difference between the 1990-93 value and that of 7 the reference dataset, an approach similar to that used by Smart et al. (2003b).The analyses 8 were done in the 'R' (version 2.9.0) statistical and programming environment (R 9 Development Core Team, 2008). 10 11 Structural equation models (SEM) were used to further examine the relationships between N 12 deposition, soil conditions, climate data and grazing pressure on species richness and species 13 composition. The software package AMOS 17 was used to design the SEM models and 14 calculate path coefficients, squared multiple correlations, direct and indirect effects and 15 model fit. A significant χ 2 -statistic indicates that the SEM model provides a poor fit. χ 2 -16 statistics and four non-χ 2 -distributed fitting functions (Comparative Fit Index, CFI; Normed Fit Index, NFI; Tucker-Lewis Index, TLI; Incremental Fit Index, IFI) were compared in order 18 to test the goodness-of-fit of the models. In the SEM models, soil nutrient concentrations were represented by "mineral N" (NH 4

20a

  decrease at high Ndeposition (Fig.3).Ndeposition was not significantly related to the 21 change in any of the other dependent variables.22Instrong contrast to the spatial analysis for the 1990-93 data, maximum summer temperature 24 and aspect did not enter the best fit model for the change in any of the dependent variables 25

5

  associated positively withchange in the number of herbspecies and negatively with change in 6 grass to herb ratios. The mineral N content was also weakly positively associated with change 7 in the number of herb species. However, the changes in the number of acidophilic species, the 8 number of acidophobic species, and the cover weighted Ellenberg R index were not related 9 significantly to soil pH or any other variable. 10 11

1319

  number of eutrophic or oligotrophic species, or change in Ellenberg N index. The change in 14 grazing was significantly negatively related to the change in frequency weighted EllenbergN 15 index, while the change in the number of oligotrophic species showed a positive relationship 16 to the mineral N content of the soil. Structural equation models were used to further investigate the direct and indirect effects of 20 Ndeposition for two variables for which change between 1990-1993 and 2006-2009 was not 21 spatially auto-correlated: H' and number of herb species.Using maximum likelihood 22 estimation, we observed a χ 2 of 4.488 for the model for H'(df=7, p=0.722) for a goodness-of-23 fit test of the null hypothesis that the covariance matrix implied by the (expected) model 24 reproduces the observed covariance matrix. Failure to reject the null hypothesis indicates that 25

5

  direct negative standardised path coefficient of -0.302 (Fig.4a) on the change in H'was 6 observed. Standardised path coefficients represent the standardised change in a response 7 variable per standardised unit change in a predictor variable, i.e.when N deposition 8 increasedby one standard deviation, change in H' decreasedby 0.302 standard deviation. In 9 terms of ecological significance, this can be regarded as a moderate effect (Chin 1998). In 10 addition, there was a positive relationship between soil pH and increase in H'. Mineral N 11 content exerted an indirect effect on the changes via soil pH, as well as a direct effect. N 12 deposition was positively related to mineral N content. S deposition had no direct effect but a 13 negative indirect effect via soil pH (Fig. 4a, Table

  for herb species showed that Ndeposition only weakly explains the 17 changebetween 1990-1993 and 2006-2009 (Fig. 4b, Table2b) and that this was an indirect 18 effect via soil pH. A much stronger negative effect of Sdeposition on the change in herb 19 species was observed (Table2b),also indirectly via soil pH. Grazing pressure did not have an 20 effect and was eliminated from this model. Mineral N content exerted an indirect effect via 21 soil pH (Fig. 4b, Table

8

  Someof the 52 individual species that met our selection criteria showed very distinct changes 1 in frequency between the two surveys. We divided the plots into three groups based on the 2 2006-8 Ndeposition data,and calculated the mean change in frequency of individual species 3 for plots in each deposition class.Species which declined or increased in frequency over all 4 plots by more than 15% are listed in Table3, with the change in frequency split between the 5 three deposition categories.For the three species with an increase in frequency at the highest 6 Ndeposition, the overall relationship with Ndeposition was not significant. However, the five 7 species which declined in frequency in the highest class of N deposition (Galiumsterneri, Linum catharticum, Briza media, Gentianella amarella and Campanula rotundifolia) all 9 showed a significant correlation between the decline in frequency and Ndeposition. the temporal change between 1990-93 and 2006-9. Here we consider 5 the results from the spatial analysis and the analysis of temporal change in turn, placing our 6 findings into the wider context of the role of Ndeposition and other factors in influencing the 7 species richness, composition and diversity of calcareous grasslands.

17(

  lower pH tended to lose species. Despite the fact that calcareous soils are highly buffered, 18 experimental studies have shown that calcareous grasslands also leach base cations when N 19 deposition is increased, as a result of soil exchange processes and plant uptake of nitrogen 20 Horswillet al. 2008).An increase of soil pH due to the reduction in acid deposition over the 21 last 15 years (Emmett et al. 2010) may have contributed to increased species richness, 22 especially in those plots which were least affected by base cation depletion.

23 249

 23 Importance of direct and indirect effects of N deposition 25The use of structural equation models allowed us to identify and compare the comparative 1 strengths of directeffects of N deposition andindirect effects of both N and S deposition, 2 through soil pH, using hypothesised causal relationships that are consistent with experimental 3 studies. The responses of vegetation to changing N deposition are complex and involve 4 several mechanisms (Bobbink et al. 2010), and therefore statistical methods, such as 5 structural equation modelling, that incorporate this complexity and aim to disentangle 6 different direct and indirect effects through potential causal structures provide powerful new 7 tools to complement more conventional regression analysis.8The results show very different models for change in H'and for change in herb species 10 numbers. Changes in H' were dominated by a direct negative effect of N deposition; although 11 indirect negative effects of S and N depositionvia soil pH were also significant, they had a 12 much smaller contribution. The negative direct effect of N deposition on change in H' was 13 also found for changes in evenness in the mixed effect models, suggesting that eutrophication 14 rather than acidification is the dominant mechanism. In contrast, the direct effect of N 15 deposition on change in herb species numbers was not significant, and was small compared to 16 indirect effects via soil pH, indicating that acidification was the dominant mechanism.

17 18 7 A

 177 Soil pH, which had a significant positive relationship with both changes in H' and changes in 19 herb species numbers, was itself negatively influenced by S deposition and mineral N 20 content. The strong acidifying nature of S in deposition compared to N was revealed by the 21 stronger negative effects that were exerted by S deposition on soil pH. Calcareous grasslands 22 have been shown to retain most of the deposited N in the system with little leaching to deeper 23 soil layers or groundwater, thereby increasing mineral N content of the top soils (Phoenix et 24 al. 2008). The increased mineral N in the soils in turn can have a significant negative effect 25 soil pH (e.g. Horswillet al. 2008). This is consistent with the positive relationships 1 between N deposition and mineral N content and the significant negative relationships 2 between the mineral N content and soil pH in our study. These findings, using a different 3 method of analysis, are consistent with earlier studies of acid grasslands in which negative 4 effects of soil acidity on herb species numbers were recorded (e.g. Stevens et al. 2006; Duprè 5 et al. 2010, Stevens et al. 2010). 6 recent survey of changes in a local UK flora over the last 50 years showed the positive 8 effects of the designation of conservation sites on maintaining biodiversity, especially in 9 calcareous grasslands (Walker et al. 2009). In our structural equation model, change in 10 grazing also significantly improved the model fit, with a decrease in grazing pressure being 11 directly associated with a decrease in H'.A decrease in grazing pressure was also associated 12 with an increase in Ellenberg N score. These results may reflect an effect of poor 13 management on some sites over recent years, since the spatial analysis suggested that grazing 14 management in 1990-93 was optimal overall and hence did not affect spatial variation in 15 species richness or diversity. adverse effects of N deposition for calcareous grasslands, based on field 19 experiments, have been set as a range of 15-25 kg N ha -1 y -1 (Bobbinket al. 2003; Bobbink et 20 al, 2010), with the range indicating the variation in sensitivity.Our results provide the first 21 support for this critical load range from field data, as, in the plots where the critical load 22 range is exceeded, there is evidence of a decrease in species diversity and evenness, a decline 23 in the frequency of characteristic species, and a lower number of rare and scarce species.

24

  These effects are likely to be due to both direct effects of N deposition and indirect effects 25 soil pH.Since all sites had active management to maintain a high diversity 1 and characteristic species, our results imply that even focussed management on nature 2 conservation objectives cannot prevent adverse effects of high rates of N deposition. Given 3 the high and increasing rates of N deposition in many parts of the world, N deposition is a 4 wide-scale threat to the conservation of species-rich grassland communities, and their 5

3F

  lowest AIC for each response variable, for (a) the survey in 1990-93 and (b) the change between the surveys in 2006-9 and 1990-93. Predictors 4 significant at p<0.1, p<0.05, p<0.01 and p<0.001 are indicated by +, *, ** and ***, respectively. Analyses were based on a maximum of 106 5 calcareous grassland surveys in 1993 and a maximum of 42 comparisons between 2009 and 1993. All analyses were based on vascular plants 6 only. The number of acidophilic species was log(x+1) transformed.7

2322

  Figure legends

22

  Fig. 4a

23 R

 23 Johnson D, Grime JP, Booth RE (2008) Sustaining ecosystem services in 20 ancient limestone grassland: importance of major component plants and community 21 composition. Journal of Ecology, 96, 894-902 22 Development Core Team (2008) R: a language and environment for statistical computing. 24 Vienna, Austria: R Foundation for Statistical Computing, http://www.R-project.org. Cooper EA, Rodwell JS, Malloch AJC (1993) Effects of climate change and air 2 pollution on British calcicolous ecosystems.Final report to UK Department of the 3 Environment.

4 5

 4 Rodwell JS (ed) (1998). British Plant Communities, Vol. 3, Grasslands and Montane 6 Communities. Cambridge University Press, Cambridge.

7 8

 7 Sala OE, Chapin FS, Armesto JJ et al. (2000) Biodiversity-Global biodiversity scenarios for 9 the year 2100.Science, 287, 1770-1774.10 11 Smart SM, Clarke RT, van de Poll HM, Robertson EJ, Shield ER, Bunce RGH, Maskell LC 12 (2003a) National-scale vegetation change across Britain; an analysis of sample-based 13 surveillance data from the Countryside Surveys of 1990 and 1998. Journal of Environmental 14 Management, 67, 239-254.15 16 Smart SM, Robinson JC, Shield EJ, van de Poll HM (2003b) Locating eutrophication effects 17 across British vegetation between 1990 and 1998. Global Change Biology9, 1763-1774.

  1997) New Flora of the British Isles, second edition. Cambridge University Press, Dise NB, Gowing DJG, Mountford JO (2006) Loss of forb diversity in relation to 23 nitrogen deposition in the UK: regional trends and potential controls. Global Change Biology, Dise NB, Mountford JO, Gowing DJ (2004) Impact of nitrogen deposition on the 2 species-richness of grasslands. Science, 303, 1876-1879.

3 4

 3 Stevens CJ, Thompson K, Grime JP, Long CJ, Gowing DJG (2010) Contribution of 5 acidification and eutrophication to declines in species richness of calcifuge grasslands along a 6 gradient of atmospheric nitrogen deposition.Functional Ecology, 24, 478-484.

7 8

 7 Tansley AG (1922) Studies of the vegetation of the English chalk. II Early stages of 9 redevelopment of woody vegetation on chalk grassland. Journal of Ecology, 10, 168-177.

  1988) CANOCO, an extension of Decorana to analyze species-environment 12 relationships.Vegetatio, 75, 159-160.

  Figure2

  Table 1a).Ndeposition did not have a significant effect on speciesrichness or on the number of 1 eutrophic or oligotrophic species. The frequency weighted average Ellenberg N value was 2 also not significantly related toN deposition, but was best explained by the maximum summer

	3
	4	temperature.The only variable for which Ndeposition entered the final model was the grass to
	5	herb ratio, for which a positive effect, albeit at the p<0.1 level,was recorded.A significant
	6	negative effect of soil pH was found on the number of acidophilic species, in a model that
	7	also included a negative effect of maximum summer temperature. The frequency weighted
	8 9 10	Ellenberg R values increased with soil pH, in a model that also included a positive effect of F maximum summer temperature. o r
		R Analysis of change between surveys
		e v
		i e
		w
		O n l
		y
	21
	22	Spatial variation in species richness, the number of eutrophic species, the number of herb
		species, and H'could be explained solely by soil depth and maximum summer temperature,
	24
	25

23

with positive relationships with both variables in all cases.

11

The average soil pH in the 1990-1993 dataset was 7.41. The average soil pH in the more 12 recent survey wasmeasured more accurately and was 7.14. The significant difference (paired 13 t-test: t=2.52, df =35, p=0.016)was most likely due to different measurement methods and 14 cannot be interpreted in terms of environmental change. 15 16 A significant, but slight, decrease in speciesrichness, independent of Ndeposition, was 17 observed between 1990-1993 (average = 41) and 2006-2009 (average=39) (paired t-test t=-18 2.274, p=0.028, df =41).The change in both H'and E H over this period was significantly 19 negatively related to Ndeposition,with a general increase in H' and E H at low Ndeposition and

  Ritchie, 1998; Klimeket al. 2007). The absence of a management effect may result from the In addition, there was a significant decline in the frequency of five individual species when N 1 deposition exceeded 25 kg ha -1 y -1 .All five individual species (Linum catharticum,Briza

	1	prevent the dominance of eutrophic species. Grazing at moderate stocking rates maintains amongst the strongest parameters explaining spatial variation in vegetation composition of
	2	Atlantic European calcareous grasslands.In addition, Bennie et al. (2006) found that
	3 3 3	calcareous grassland plots on deeper soils were more likely to be invaded by (often
	4 4 4	management regimes at all the sites being successfully focussed on preserving species mesotrophic) plant species, which may explain the higher species richness. Ellenberg nutrient score. The decline in frequency in these high N deposition plots was
	5 5 5	richness. substantial (on average 22%) over a relatively short period of 15 yearsdespite the focussed
	6 6 6	Comparison between surveys management on these nature reserves, and indicates that local extinctions of these species
	7 7	Despite this, the frequency of rare and scarce species was significantly reduced, by a factor of The speciesrichness was lower in 2006-2009 than in 1990-1993 with an average loss of 2 could occur in the future unless N deposition declines.Both C. rotundifolia and B. media are
	8 9	Spatial variation 5, at sites with N deposition above 25 kg ha -1 y -1 . These relationships are difficult to test in a F o r F o r r o F regarded as poor competitors (Leishman, 1999; Dixon, 2002; Hutchings et al. 2003) and for
		R R R
		e v e v e v
		i e i e i e
		w w w
	19 17 18 19 19 18	O n l y Maximum summer temperature, along with soil depth, explained most of the variation in O n l y O n l y
	20 20 20	argued that their relatively small sample size (94 plots),and the correlation between N species richness and the frequency of different types of species; species richness was higher H' is a measure of diversity that accounts for both species richness and evenness of the
	21 21 21	deposition and climatic parameters, may have masked the significance of Ndeposition as an in plots with a higher maximum temperature. Plotswith deeper soils tended to occupy species present in the sample (Magurran, 2004). N deposition did not significantly affect
	22 22 22	explanatory variable. Our sample size is similar (106 plots) to that of Maskell et al. (2010), moderate to low slopes and had a higher organic content (r 2 =-0.14, p=0.022) anda higher changes in species richness but did significantly affect changes in evenness, indicating that
	23	but a more likely explanation for the lack of significant relationships with Ndeposition is that moisture content (r 2 = 0.35, p<0.001).The higher organic content and the fact that deeper changes in H' were driven primarily by changes in the evenness of the plots.
	24
	25 25

10

We found that N deposition did not explain spatial variation in species richness, diversity, or 11 the frequency of eutrophic species. This is consistent with a study in 1998 of 94 calcareous 12 grassland plots in the UK, in which there was no significantrelationship between N 13 deposition, or climatic variables, and species richness (Maskell et al. 2010), even though 14 significant relationships between N deposition and species richness were found in the same 15 survey in acid grasslands and heathlands. These findings contrast with results from 16 experimental studies in which clear N fertilization effects have been shown in calcicolous 17 habitats, with fast-growing eutrophic species outcompeting slow growing sensitive species 18 (e.g. Bobbink, 1991; Carroll et al. 2003; van den Berg et al. 2005). Maskell et al. (2010) 23 all of our sites are nature reserves, at which the management is aimed specifically at the 24 conservation of high species diversity and in which grazing regimes are implemented to 1 species diversity by reducing the number of competitive species (Collins et al. 1998; Olff& 2 8 rigorous way in multivariate models, as rare species, such as Epipactisatrorubens, 9 Triniaglauca and Pulsatillavulgaris, not only have a highly restricted geographic distribution 10 but are also typically highly localised on the sites where they do occur.Fisher and Stöcklin 11 (1997) and Bennie et al. (2006) both found a negative correlation between the initial 12 frequency of a plant species and the extinction rate in calcareous grasslands over a period of 13 several decades, over which Bennie et al. (2006) also found an increase in Ellenberg fertility 14 index, although there was not asignificant relationship between extinction rates and the 15 Ellenberg fertility index. Our results suggest a link between N deposition and the occurrence 16 of scarce and rare species, which may be due to local extinctions over a period of decades.

23 soils are more resistant to drought stressmay be important factorsin explaining the higher 24 species richness. Duckworth et al.(2000)found that temperature and soil organic matter are effect (Cohen 1988).Progression towards anundesirable reference dataset in high N 12 deposition plots showed the reduction in H' between 1990-1993 and 2006-2009 represented a 13 6.4% movement towards the H' value typical of British Neutral Grasslands. This implies that 14 a much longer period would be needed to reach the low values of H' typical of neutral 15 grasslands. However, many typical calcareous grassland species may be lost well before this 16 value of H' is reached, and we conclude that the change in H' at high N deposition sites does 17 represent a significant ecological change. 2 media, Gentianella amarella, Campanula rotundifolia and Galiumsterneri)have a low 8 C. rotundifolia, lower frequencies have been found at sites with high N deposition (Stevens et 9 al. 2006; Maskell et al. 2010). Linumcatharticum and G. amarella are respectively annual 10 and bi-annual species which may strongly depend on gap formation in the swards for 11 germination (e.g. Turnbull et al. 2005).These results are corroborated by an earlier study on 12 calcareous grasslands by Bennie et al. (2006) who found strong decreases in frequency over 13 the period 1950-2000 of C. rotundifolia, L. catharticum and B. media.

14 15

Table 2 :

 2 Summary of results from structural equation modelling, showing the decomposition of

	Page 33 of 51
	2
	3	correlations into standardised direct, indirect and total effects of seven explanatory variables
	4	on (a) the difference in Shannon diversity index and (b) the difference in the number of herb
	5	species between 2006-2009 and 1990-1993. Total effects are the sum of direct and indirect
	6	effects. Direct effect coefficients are equal to the path coefficients shown in Fig. 4a and b.
	7
	8
		a)	Difference in Shannon diversity between 2009 and 1993 (H2009-H1993)
			Direct	Indirect	Total
		Ndeposition	-0.302	-0.008	-0.310
		SOx deposition	-	-0.037	-0.037
		Mineral N soil	-0.022	-0.018	-0.040
		Soil pH	0.075	-	0.075
		Grazing pressure 0.063	-	0.063
		b)	Difference in herb species between 2009 and 1993 (n2009-n1993)
			Direct	Indirect	Total
		Ndeposition	-	-0.014	-0.014
		SOx deposition	-	-0.147	-0.147
		Mineral N soil	-	-0.072	-0.072
		Soil pH	0.298	-	0.298
	9
	10
	11
	12
	13
	14
	15
	16

1

Table 3 :

 3 Effect of N deposition class on the mean change in frequency of individual species

	2
	3	(expressed as the difference of frequency between 2006-2009 and 1990-1993), for those
	4	species with a change in frequency of more or less that 0.15 (15%) in the highest Ndeposition
	5	class.Shown are thecorrelation coefficients (Cor.) between the change in relative frequency of
	6	the species and modelled Ndeposition based on 2006-8 data and the level of significance of
	7	the correlation.
	8
	9
		Species	N deposition class (kgN ha -1 y -1 )
			0-15	15-25	25-35	Cor.	p
		Cynosuruscristatus	0.08	0.12	0.53	0.04	0.809
		Holcuslanatus	0.14	0.20	0.25	-0.08	0.611
		Polygala vulgaris	0.23	0.06	0.17	-0.03	0.812
		Linum catharticum	0.27	-0.02	-0.18	-0.34	0.026
		Briza media	0.17	0.05	-0.21	-0.37	0.015
		Gentianella amarella	-0.03	0.12	-0.24	-0.30	0.050
		Campanula rotundifolia	0.05	-0.15	-0.25	-0.61	<0.001
		Galiumsterneri	0.11	0.04	-0.26	-0.53	<0.001

1

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Global Change Biology

Acknowledgements

This study was supported by the UK Natural Environment Research Council under award 2 number NE/E00413.The original establishment of the plots in 1990-93 was supported by the 3 then UK Department of /the Environment under contract PECD 7/12/22. We thank 4 Countryside Council for Wales, Natural England, and the Scottish Natural Heritage for 5 granting permission to access andsample the sites.

A key finding from our study is that the effects of N deposition, and of other driving