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Within season and carry-over effects following exposure of grassland species mixtures to increasing background ozone
Environmental Pollution 159 (2011) 2420e2426
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Within season and carry-over effects following exposure of grassland species mixtures to increasing background ozone Felicity Hayes*, Gina Mills, Harry Harmens, Kirsten Wyness Centre for Ecology and Hydrology, Environment Centre Wales, Deiniol Road, Bangor, Gwynedd LL57 2UW, UK
a r t i c l e i n f o
a b s t r a c t
Article history: Received 11 January 2011 Received in revised form 22 June 2011 Accepted 25 June 2011
Few studies have investigated effects of increased background ozone in the absence of episodic peaks, despite a predicted increase throughout the northern hemisphere over the coming decades. In this study Leontodon hispidus was grown with Anthoxanthum odoratum or Dactylis glomerata and exposed in the UK to one of eight background ozone concentrations for 20 weeks in solardomes. Seasonal mean ozone concentrations ranged from 21.4 to 102.5 ppb. Ozone-induced senescence of L. hispidus was enhanced when grown with the more open canopy of A. odoratum compared to the denser growing D. glomerata. There was increased cover with increasing ozone exposure for both A. odoratum and D. glomerata, which resulted in an increase in the grass:Leontodon cover ratio in both community types. Carry-over effects of the ozone exposure were observed, including delayed winter die-back of L. hispidus and acceleration in the progression from flowers to seed-heads in the year following ozone exposure. Ó 2011 Elsevier Ltd. All rights reserved.
Keywords: Ozone Flowering Carry-over effects Competition Biomass Roots
1. Introduction Mean ozone concentrations in Europe have been increasing since industrial times, from a background of 10e15 ppb in the 1900’s to the current 35e40 ppb (Simmonds et al., 2004), due to increased precursor emissions from anthropogenic sources (Solberg et al., 2005; Volz and Kley, 1988). There is some evidence that episodic peaks of ozone are declining in Europe due to reductions in emissions of ozone precursors from local and regional sources (Szopa et al., 2006; Vingarzan, 2004). However, an increase in background ozone concentration throughout the northern hemisphere has been predicted due to hemispherical transport, particularly due to increased emissions from Asia (Jaffe and Ray, 2007; Vingarzan, 2004), with annual mean concentrations likely to reach 68 ppb by 2050 (Meehl et al., 2007) and may exceed 75 ppb over much of Europe by 2100 (Sitch et al., 2007). The predicted increases would include elevated ozone concentrations at nighttime, particularly at rural, high-altitude sites. The AOT40 index (The sum of the differences between the hourly mean ozone concentration (in ppb) and 40 ppb for each hour when the concentration exceeds 40 ppb, accumulated during daylight hours), which is one of the indices currently used to define critical levels for
* Corresponding author. E-mail address: [email protected] (F. Hayes). 0269-7491/$ e see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2011.06.034
crops, semi-natural vegetation and forests, assumes that only concentrations of ozone exceeding a threshold of 40 ppb are harmful for vegetation. However, some studies have shown that concentrations below this threshold can be harmful to some species of plants (e.g., Karlsson et al., 1995). In addition, exposure to low-moderate ozone concentrations below the 40 ppb threshold can result in a large total uptake by plants if the exposure continues through the night. Few studies have investigated the effects of increased background ozone concentrations in the absence of episodic peaks. When contrasting profiles of increased background versus increased peaks have been used, it has been demonstrated that increased background concentration can be as harmful to plants as an episodic ozone regime with similar total exposure (Dawnay and Mills, 2009; Hayes et al., 2010b; Nussbaum et al., 1995). This has been attributed, in part, to a discrepancy in the timing of ozone exposure with the time of maximum anti-oxidant defence in plants, which is generally in early-mid morning (Heath et al., 2009). Consequently there is a more biologically effective ozone dose during afternoon and evening exposure. It is also thought that some nocturnal uptake of ozone occurs for many plants due to night-time stomatal conductance, which although usually lower than day time conductance, can nevertheless result in considerable uptake of ozone into a plant when ozone concentrations are high (Musselman and Minnick, 2000). It has been demonstrated that in such cases the use of AOTX24 (accumulated over a 24 h period) was
F. Hayes et al. / Environmental Pollution 159 (2011) 2420e2426
a better fitting parameter to the response data than AOTX12 (used as a surrogate for accumulation during daylight hours) (Dawnay and Mills, 2009). Grassland communities have been identified as at risk from increasing ozone pollution as they contain species that are sensitive to ozone and are found in areas where the ambient ozone concentrations may already be high enough to cause damage (Bassin et al., 2007; Mills et al., 2007, 2011). Although several studies have investigated the impacts of ozone on component species of grasslands, comparatively few have been carried out on species mixtures and communities. These studies have shown negative effects of episodic peaks of ozone and the damaging effects demonstrated for grassland communities include visible injury (e.g., Bungener et al., 1999), premature leaf senescence (e.g., Franzaring et al., 2000), changes in biomass partitioning (e.g., Hayes et al., 2009) and reduced growth (e.g., Gimeno et al., 2004). These can lead to a change in species composition with longer-term ozone exposure (e.g., Hayes et al., 2010b). The response to ozone of some species varies according to the neighbouring plant species, for example the presence of Lolium perenne increased the proportion of senesced leaves in Trifolium repens (Hayes et al., 2010a). In a separate study, Poa pratensis showed reduced growth with increasing ozone exposure when grown with Veronica chamaedrys but not when grown with other species such as Achillea millefolium (Bender et al., 2006). In a study on a semi-natural grassland, the biomass of the forbs decreased in filtered air compared to ambient air conditions (Evans and Ashmore, 1992), in contrast to predictions based on glasshouse studies of individual plants. This was attributed to changes in light penetration through the canopy of the component grass species. Some studies suggest that the influence of the plant canopy on microclimate can alter the stomatal conductance and therefore ozone uptake of species within a plant community, for example, Lantinga et al. (1999) showed that PAR was dramatically reduced within a plant canopy, and PAR is one of the factors that can influence stomatal opening. A separate study demonstrated that a lower proportion of leaves showed visible injury symptoms in the centre of a T. repens/L. perenne plant canopy compared to those of the upper canopy and canopy edge, and this was also attributed to changes in microclimate, particularly PAR (Hayes et al., 2010a). It is therefore possible that species within a dense vegetation canopy could be protected from ozone exposure due to microclimatic conditions that do not favour ozone uptake. The response of flowering to ozone exposure has been studied for a few species and has been shown to involve changes in flower number, biomass and timing of development. A reduction in flower biomass with increasing ozone exposure has been demonstrated for some species, e.g., Trifolium cherleri, T. subterraneum and T. striatum (Gimeno et al., 2004). Similarly, ozone significantly reduced the number of flowers in Campanula rotundifolia (Rämö et al., 2007) and in some cases the reduced flowering persisted one month following ozone exposure (e.g., for T. striatum, Sanz et al., 2007). Spartina alterniflora showed delayed flowering and a reduction in the number of flower spikes produced with elevated ozone treatment (Taylor et al., 2002). Flower weights were significantly reduced for Eupatorium cannabinum and Plantago lanceolata when there had been no change in either flowering date or flower number (Franzaring et al., 2000). In contrast, increased female flower formation in response to ozone has been shown for Betula pendula (Darbah et al., 2007) and for Rubus cuneifolius an initial acceleration in flowering occurred in the second year of ozone exposure with the elevated ozone treatment, with increased flower numbers and an earlier time of peak production (Chappelka, 2002). Several studies have indicated that carry-over effects can occur following ozone exposure that may not manifest until the following
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year. However, the majority of studies on carry-over effects following ozone exposure are for trees. B. pendula saplings showed reduced contents of Rubisco, chlorophyll, reduced starch and nutrient content in leaves and decreased new shoot growth during a recovery period following ozone exposure (Oksanen and Saleem, 1999). A reduced number of overwintering buds with elevated ozone has also been shown for B. pendula, due to a lower carbon gain for bud formation (Oksanen, 2003; Riikonen et al., 2008). Buds have also been shown to be affected by ozone exposure in Fagus crenata seedlings, where those seedlings exposed to elevated ozone showed earlier leaf fall and a reduction in leaf non-structural carbohydrates, which resulted in later bud-break and a reduced number of leaves per bud in the following spring (Yonekura et al., 2004). Carry-over effects can also occur in the roots of plants, and for ponderosa pine seedlings (Pinus ponderosa) these effects included reduced new root production and lower concentrations of root starch in seedlings that had been previously exposed to ozone (Anderson et al., 1997). Investigations of carry-over effects for semi-natural vegetation have been more limited. In one study investigating effects for 33 species grown individually, re-growth of plants was affected in the spring following exposure to an episodic ozone regime. Species affected included upland grassland species such as Juncus effusus, Nardus stricta and Carex echinata; for the latter two species no effects were detected during the course of the ozone exposure (Hayes et al., 2006). For C. echinata this biomass reduction was attributed to a large decrease in flower biomass for those plants exposed to elevated ozone the previous year. In some other studies observed effects of ozone were enhanced during subsequent ozone exposures, where it was unclear whether this was due to a true carry-over effect or as a consequence of cumulative ozone exposure. For example, in clover, re-growth in subsequent ozone exposure periods was affected, whilst the biomass at the end of the first ozone exposure was not affected (Fumagalli et al., 2003; Nussbaum et al., 1995). Many semi-natural communities are dominated by perennial plant species, therefore cumulative/carryover effects may be important for the long-term viability of plant communities. The aims of this study were to investigate whether increasing background ozone concentrations are damaging to vegetation, whether the nature of the competing species modifies the response to ozone, and whether there are carry-over effects into the following growing season. This study used three species widely distributed in temperate grasslands; the forb Leontodon hispidus, and two grass species: Dactylis glomerata, a vigorous competitor and Anthoxanthum odoratum, a slower growing grass with a less dense canopy. All three of these species have been shown to respond to ozone at concentrations below 100 ppb (Hayes et al., 2007; Dawnay and Mills, 2009; Mills et al., 2009). Plants were grown in large containers in competitive mixtures of L. hispidus:A. odoratum and L. hispidus:D. glomerata, enabling the response of L. hispidus to be compared between the two community mixtures. Exposure to ozone was in computer-controlled solardomes with near-ambient climate conditions, allowing night-time concentrations to be maintained and therefore better mimicking ambient ozone profiles of upland regions. 2. Materials and methods 2.1. Ozone system and treatments Plants were exposed to ozone in solardomes (hemispherical greenhouses 3 m diameter, 2 m tall). Ozone was generated from oxygen concentrated from air (Workhorse 8, Dryden Aqua, UK) using an ozone generator (G11, Dryden Aqua, UK) and distributed to each solardome via PTFE tubing. The ozone system and treatments are described in Mills et al. (2009). Ozone was delivered to each solardome
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F. Hayes et al. / Environmental Pollution 159 (2011) 2420e2426
using mass flow controllers (Celerion, Ireland) controlled by computer software (Labview version 7). Ozone concentrations were continuously monitored in one solardome using a dedicated ozone analyser (Thermoelectron, Model 49C), allowing feedback to compensate for small variations in ozone production. In all solardomes the ozone concentration was measured for 5 min in every 30 min using two additional ozone analysers (Envirotech API 400A) of matched calibration. Eight ozone treatments were randomly allocated to the solardomes. The profiles used were based on a simulation of 1 week of ozone data from the nearby Marchlyn Mawr rural monitoring site (grid reference SH613619, 610 m.a.s.l.), using data from 31st May to 6th June 2006. This profile was repeated every week for 20 weeks with the treatments AA (simulated ambient), AA 20 ppb, AA þ 12 ppb, AA þ 24 ppb, AA þ 36 ppb, AA þ 48 ppb, AA þ 60 ppb, and AA þ 72 ppb (see Mills et al., 2009 for details). The ozone concentrations in the AA þ 36 and higher treatments were turned down to 50 ppb for access for plant measurements on Tuesdays, and occasionally on other days. 2.2. Plants and community set-up Communities were established using plug plants of Leontodon hispidus, Dactylis glomerata and Anthoxanthum odoratum (purchased from British Wildflower Plants, Norfolk, UK). L. hispidus was grown with either A. odoratum or D. glomerata. Fourteen litre pots (33.3 cm diameter 24.0 cm deep) were lined with perforated plastic sheeting to deter roots from growing through the drainage holes in the base of the pot, and filled with a compost mix in the ratio (by volume) 3 parts John Innes no 2: 1 part peat: 1 part gritty sand. Each pot contained four plants of L. hispidus and three plants of the appropriate grass species arranged in the same pattern. Communities (6 replicates per community type per solardome) were established on 23rd March 2007, and were moved into the solardomes (at 20 ppb) on 24th April. The ozone exposure treatments started on 9th May. Communities were cut back to 7 cm after 10 weeks on 18th July to simulate a conservation grassland cut (harvest data not presented), and the last day of exposure was after 20 weeks on the 25th September. Following exposure, communities were placed in a rabbit-proof plot in an adjacent field of pasture to allow natural die-back and overwintering. Starting on 13th May 2008 the L. hispidus/A. odoratum communities were reexposed to ozone (data not presented here), whereas the L. hispidus/D. glomerata communities remained in the outdoor field plot with assessments and harvests as detailed below. 2.3. Assessments of injury, leaf senescence and flowering Plants were assessed for visible injury and leaf senescence after exposure to ozone for 0, 1, 4, 7, 10, 14 and 19 weeks. A leaf was classified as senesced if >25% of the leaf was senesced, otherwise it was classified as healthy. Following the end of the ozone exposure on 25th September 2007, assessments were carried out on 28th November 2007, 30th January 2008 and 28th March 2008 for all communities, with additional assessments on 17th June, 15th July, 19th August and 9th September 2008 for the L. hispidus/D. glomerata communities only. During these assessments the numbers of flower buds, flower heads and seed-heads for each species in each mesocosm were also counted. 2.4. Stomatal conductance Stomatal conductance measurements were made on the abaxial surface of L. hispidus using a porometer (AP4, Delta-T), on days of stable meteorological conditions after exposure to the ozone regime for 17 weeks. Measurement of leaves in the upper canopy (mature leaves in full sunlight) and the inner canopy (younger leaves which were also more shaded) were taken, using five leaves for each canopy position of each community type in each solardome. 2.5. Plant cover and root biomass The area cover of each species was determined for each mesocosm by overlaying a grid of 1 cm squares over photographs taken from above on 25th September 2007. In each photograph the pot surface was covered by approximately 200 squares. Each species was scored as present if >50% of an individual square was occupied. On 18th February 2008 the root biomass was determined for three of the ozone treatments (AA, AA þ 24 and AA þ 60) for a subsample of approximately 20% of the soil taken as a vertical slice across the middle of each mesocosm, and with the same number and combination of component species (one L. hispidus and two grass) for each slice. The soil mix was removed from the roots by washing by hand. The mesocosms from the other five ozone treatments remained in the outdoor field plot. In July 2008 these communities were cut back to 7 cm to simulate a conservation grassland cut (data not included here). 2.6. Data analysis and statistics Datasets were analysed using linear regression analysis in Minitab (Version 14), with each datapoint in the regression representing the mean per treatment. Comparison
of regression lines was made using General Linear Model (Minitab, Version 14). Results were considered as significant when p < 0.05, and as trends at p < 0.1.
3. Results 3.1. Ozone exposure Ozone exposure over the 20 week period ranged from a seasonal 24 h mean of 21.4 ppbe102.5 ppb (Table 1), with the AOT40 (accumulated over 24 h) ranging from 0.07 ppm h to 214 ppm h. There was a linear relationship between 24 h mean ozone and 12 h mean ozone concentrations for the ozone treatments (r2 ¼ 0.9999). Mean PAR was 451 mmol m2 s1 during the day (07:00e18:00) and 28 mmol m2 s1 during the night-time (19:00e06:00). Further details of the ozone exposure, including the weekly mean ozone profiles for each of the ozone treatments, and climatic conditions in the solardomes are available in Mills et al. (2009). 3.2. Effects on plants during the ozone exposure No visible injury was observed during the course of the ozone exposure. After exposure to the ozone regime for 10 weeks, a linear increase in leaf senescence of L. hispidus with increasing ozone was shown when grown with A. odoratum (r2 ¼ 0.52, p ¼ 0.042), although this was not significant with D. glomerata (r2 ¼ 0.33, p ¼ 0.134; Fig. 1a). There was no significant difference between the slope of the regression lines indicating that there was no difference in the extent of leaf senescence on L. hispidus according to whether it was grown with A. odoratum compared to when grown with D. glomerata. The communities were cut back after 10 weeks, and nine weeks after this cut (after exposure to ozone for a total of 19 weeks), a linear increase in leaf senescence of L. hispidus with increasing ozone was again shown when grown with A. odoratum (r2 ¼ 0.77, p ¼ 0.004) and D. glomerata (r2 ¼ 0.42, p ¼ 0.081). Interestingly, L. hispidus showed a larger increase in leaf senescence with increasing ozone exposure when grown with A. odoratum than when grown with the densely growing coarse grass D. glomerata (Fig. 1b), with a significant difference between the slope of the two regression lines (p ¼ 0.022). Stomatal conductance measurements made after exposure to the ozone regime for 17 weeks revealed that although there were no significant differences in the stomatal conductance of leaves from the upper canopy of L. hispidus growing in the two community types, for leaves of the inner canopy stomatal conductance of L. hispidus grown with D. glomerata was approximately 30% lower than of those grown with A. odoratum (mean stomatal conductance was 176 mmol m2 s1 compared with 320 mmol m2 s1). This difference in stomatal conductance of L. hispidus between the two community types was consistent across all ozone treatments and
Table 1 Season mean ozone concentration, 12 h mean ozone concentration, AOT40 in daylight hours and AOT40 accumulated over 24 h for each treatment. AA ¼ simulated ambient ozone concentration. Ozone treatment AA AA AA AA AA AA AA AA
20 ppb þ þ þ þ þ þ
12 24 36 48 60 72
ppb ppb ppb ppb ppb ppb
Season mean ozone, ppb
12 h mean ozone (07:00e18:00), ppb
AOT40 (daylight), ppm h
AOT40 (24 h), ppm h
21.4 39.9 50.2 59.4 74.9 83.3 101.3 102.5
21.1 39.2 49.6 58.7 73.3 81.6 99.0 100.5
0.07 4.93 21.44 38.04 62.49 77.13 108.43 112.47
0.07 10.91 41.29 72.19 119.82 147.42 206.70 214.34
F. Hayes et al. / Environmental Pollution 159 (2011) 2420e2426
40
A
With A. odoratum With D. glomerata
r2
= 0.52 p=0.042
30
r2
= 0.33 p=0.134
20
10
0
0
50
100
150
Area covered, relative units
Senesced leaves, %
A
Seasonal mean O3 conc. (24 h, ppb)
r2 = 0.58 p=0.027 120 80 L. hispidus
40
r² = 0.31 p=0.154
A. odoratum
0 50
100
150
Seasonal mean O3 conc. (24 h, ppb)
40
With A. odoratum With D. glomerata
r2 = 0.77 p=0.004
B
30
20 r2 = 0.42 p=0.081
10
0
160
0
0
50
100
150
Seasonal mean O3 conc. (24 h, ppb)
Area covered, relative units
Senesced leaves, %
B
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160 120 80
At the end of the ozone exposure in September 2007, L. hispidus showed no change in cover per mesocosm with increasing ozone exposure when grown with either A. odoratum (r2 ¼ 0.31, p ¼ 0.154; Fig. 2a) or D. glomerata (r2 ¼ 0.33, p ¼ 0.139; Fig. 2b), while there was an increase in cover with increasing ozone exposure for A. odoratum (r2 ¼ 0.58, p ¼ 0.027; Fig. 2a) and D. glomerata (r2 ¼ 0.84, p ¼ 0.001; Fig. 2b). This combination of changes resulted in an increase in the grass:Leontodon cover ratio with both A. odoratum (r2 ¼ 0.54, p ¼ 0.039) and D. glomerata (r2 ¼ 0.31, p ¼ 0.151; Fig. 2c), with no significant difference in the slope of the two response relationships. 3.4. Winter die-back In the autumn the leaves of L. hispidus normally die back, however, an assessment of the mesocosms on 28th November 2007 showed a greater proportion of green, healthy leaves with
r² = 0.33 p=0.139
D. glomerata
0 0
50
100
150
Seasonal mean O3 conc. (24 h, ppb)
C Grass:Leontodon ratio
3.3. Species cover at the end of ozone exposure
L. hispidus
40
Fig. 1. Senescence of L. hispidus when grown in mixture with A. odoratum or D. glomerata after exposure to ozone for A) 10 weeks and B) 19 weeks. All communities were cut back after 10 weeks. Linear regression lines indicate changes in leaf senescence with increasing ozone exposure.
no interaction with ozone treatment was apparent (data not presented). No significant effects on timing of flowering on any of the species were found during the ozone exposure (data not presented). Similarly, 4 and 7 weeks after the end of ozone exposure there were no apparent carry-over effects of ozone on the number of flower heads of L. hispidus when grown with either A. odoratum or D. glomerata.
r² = 0.84 p=0.001
2.0
r² = 0.27 p=0.151
1.5 1.0 with D. glomerata
0.5
r² = 0.54 p=0.039
with A. odoratum
0.0
0
50
100
150
Seasonal mean O3 conc. (24 h, ppb) Fig. 2. Area covered by a) L. hispidus and A. odoratum and b) L. hispidus and D. glomerata, and c) the grass:Leontodon cover ratio after exposure to ozone for 19 weeks.
increasing ozone treatment for L. hispidus in both the A. odoratum and D. glomerata communities (Fig. 3). These leaves were mature leaves remaining since the ozone exposure, rather than a flush of new leaves that had formed subsequently. This increase was approximately 3-fold in the AA þ 72 treatment compared to AA 20 in both the A. odoratum communities (r2 ¼ 0.69, p ¼ 0.011) and the D. glomerata communities (r2 ¼ 0.53, p ¼ 0.04), with no significant difference in the slope with increasing ozone between
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F. Hayes et al. / Environmental Pollution 159 (2011) 2420e2426
1.6 'buds+flowers':seed-heads (ratio)
Green leaves (%)
40
r2 = 0.53 p=0.04
30
20 r2 = 0.69 p=0.011 10
r2 = 0.88 p=0.018
1.4 1.2 1 0.8 0.6 0.4 0.2 0
with A. odoratum
0
with D. glomerata 0 0
50
100
150
Seasonal mean O3 conc. (24 h, ppb) Fig. 3. Proportion of green leaves remaining on L. hispidus on 28th November, when grown with either A. odoratum or D. glomerata.
the community types. There was no significant difference in the proportion of green leaves of either A. odoratum or D. glomerata (data not presented). By 30th January 2008, following a sustained period of sub-zero overnight temperatures, L. hispidus in all mesocosms had completely died back. 3.5. Root weight In February 2008, just before the start of spring re-growth, the root biomass of L. hispidus was reduced by approximately 50% in the AA þ 60 treatment compared to AA in both the A. odoratum and D. glomerata communities. Using relative root biomass compared to the AA treatment to combine the root data, no difference was found when L. hispidus was grown with A. odoratum compared to D. glomerata (r2 ¼ 0.93; p ¼ 0.002 for the combined data; Fig. 4). Using the 95% confidence intervals for this response function, a small increase of 15 ppb above the simulated ambient treatment in this study (40 ppb) would be sufficient to induce a significant decrease of 10% in root biomass of L. hispidus. The root biomass of A. odoratum was reduced by 40% and 55% in the AA þ 24 and AA þ 60 treatments respectively compared to AA. For D. glomerata, the root biomass was reduced by approximately 30% in both the AA þ 24 and AA þ 60 treatments compared to AA.
1.2 r2=0.93 p=0.002
Root biomass (%)
1.0 0.8 0.6 0.4 0.2
with D. glomerata with A. odoratum
0.0 0
50
100
50
100
150
Seasonal mean O3 conc. (24 h, ppb)
150
Seasonal mean O3 conc. (24 h, ppb) Fig. 4. Relative weight of roots of L. hispidus (compared to the AA treatment) when grown in combination with D. glomerata or A. odoratum.
Fig. 5. Ratio of buds þ flowers heads to seed-heads for L. hispidus grown with D. glomerata in July 2008, after exposure to ozone in 2007.
3.6. Carry-over effects on flowering By 28th March 2008, plants had started to re-grow, however, there were no differences in the numbers of L. hispidus leaves per mesocosm between the different treatments in either community type (data not presented). On 19th May (when only the L. hispidus eD. glomerata communities had remained in the field plot) there were no flowers developing on the L. hispidus, and no significant difference in the number of D. glomerata panicles with increasing ozone treatment (r2 ¼ 0.25, p ¼ 0.393). By 17th June there was no difference in the number of flower heads/panicles on either L. hispidus or D. glomerata. On 15th July, although there were no differences in the total number of L. hispidus flower heads between the different ozone treatments, the ratio of buds and flower heads to seed-heads had changed, with the ratio decreasing with increasing ozone concentration (Fig. 5; r2 ¼ 0.88, p ¼ 0.018), indicating that the plants were going through the reproductive part of their life-cycle more quickly. There were no significant differences for D. glomerata. The same pattern was observed on 9th September, with the ratio of buds and flower heads to seed-heads decreasing for L. hispidus with increasing ozone concentration (r2 ¼ 0.42, p ¼ 0.238), and no panicles present for D. glomerata. 4. Discussion During the course of ozone exposure there was a linear increase in leaf senescence with increasing background ozone concentration. The extent of leaf senescence on L. hispidus was shown to be influenced by neighbouring species, with a larger increase in ozone-induced senescence observed in the more open canopy of A. odoratum compared to the denser canopy of D. glomerata after prolonged exposure to ozone. Reductions in ozone concentration within plant canopies have been demonstrated, with these reductions being up to 30% in a grassland canopy (Jäggi et al., 2006). Together with reduced windspeed influencing the boundary layer around plant leaves, this has been hypothesized to result in decreased ozone uptake by plants in more dense canopies. In the current study, there was evidence to support this theory as there was reduced stomatal conductance of inner canopy leaves of L. hispidus grown with D. glomerata compared to when grown with A. odoratum and this corresponded with the observed differences in leaf senescence of L. hispidus when grown with each of these species after prolonged exposure to ozone. At the end of the exposure, although there was no change in cover of L. hispidus with increasing ozone exposure biochemical changes must have occurred to result in the carry-over effects
F. Hayes et al. / Environmental Pollution 159 (2011) 2420e2426
observed. It is unlikely that the carry-over effects in this study were a consequence of the use of pots during the study, as the pots used were large and the plants had not become pot-bound by the end of the experiment. Previous studies have shown that elevated ozone tended to accelerate leaf abscission in the autumn for Betula papyrifera and Populus tremuloides at the Aspen FACE site, USA (Riikonen et al., 2008) and this has also been reported for several other tree species e.g., B. pendula (Uddling et al., 2006); Fagus sylvatica (Nunn et al., 2005). However, in the current study increasing ozone exposure corresponded with a delayed die-back of L. hispidus in autumn/winter, despite the enhanced leaf senescence that had occurred for these plants earlier in the growing season. There was also no difference according to which competing grass the L. hispidus was grown with, even though the extent of the earlier ozone-induced leaf senescence was greater on L. hispidus grown with A. odoratum. This could indicate a carry-over effect on hormonal signalling as leaf abscission is a highly regulated phase in leaf development with many genes encoding degradative enzymes having been identified (Buchanan-Wollaston, 1997; BuchananWollaston et al., 2003) and hormonal control by abscisic acid (ABA), ethylene and auxin being clearly demonstrated (Taylor and Whitelaw, 2001). Some support for this theory comes from previous studies that have indicated that prolonged ozone exposure disrupts hormonal control of stomatal functioning by ABA and ethylene in L. hispidus and the plants become insensitive to ABA (Mills et al., 2009; Wilkinson and Davies, 2009). ABA is also involved in the regulation of flowering, however the role is not well understood (Wilmowicz et al., 2008). Thus, it is possible that the accelerated flowering of L. hispidus in the current study, in conjunction with the delayed die-back in winter could indicate alterations in sensitivity to ABA or to the ABA signalling process. The observed accelerated flowering of L. hispidus with increasing background ozone concentration has implications for ecosystem services as this species is considered to be an important source of nectar for nectar feeders including butterflies, hoverflies, bees and other insects. These nectar feeders could be indirectly affected by previous ozone exposure as the timing of the food source availability is affected. In the current study, the acceleration in development of flowers in L. hispidus leading to a shorter timeperiod of flower duration was a carry-over effect and occurred some ten months following the end of ozone exposure in the previous growing season. Despite the differences in the extent of leaf senescence of L. hispidus when grown with A. odoratum or D. glomerata there was no difference in the relative root reduction with increasing ozone for L. hispidus between the two community types. Larger effects on roots than on shoots have been demonstrated for some species and a reduction in allocation to roots in response to ozone exposure has been shown for Cirsium arvense (Power and Ashmore, 2002), Cirsium dissectum (Franzaring et al., 2000) and some wetland species (Batty and Ashmore, 2003). A meta-analysis of ozone impacts on plant development by Grantz et al. (2006) showed that changes in allocation, with reduced allocation to roots with increasing ozone concentration commonly occur, although the extent of change is variable, and results from the current study support this because effects of ozone on root weight as a result of the ozone exposure were of a larger magnitude than changes in above-ground cover. A reduction in root biomass at the end of the ozone exposure suggests a decrease in available resources for re-growth in the following year and gives further indication of potential carry-over effects of ozone exposure. Based on a compilation of published data on the effects of episodic ozone regimes on above-ground biomass, L. hispidus has previously been identified as ozone sensitive (Hayes et al., 2007), with A. odoratum and D. glomerata classified as insensitive to ozone. However, other studies conducted within the solardomes have
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shown that all three species respond to ozone by enhancing leaf senescence, with thresholds for effects below 100 ppb (Dawnay and Mills, 2009; Mills et al., 2009). The current study indicates that although these species appeared insensitive to ozone based on changes in above-ground area covered (agreeing with Hayes et al., 2007), significant effects of ozone on these species occur below ground at lower concentrations. In addition the nature of the competing species can alter the response to ozone of L. hispidus with this species being more sensitive to ozone when grown with A. odoratum than when grown with D. glomerata in terms of enhanced leaf senescence. D. glomerata is a stronger competitor than A. odoratum with scores for ‘competitor’ sensu Grime (Grime et al., 1988) of 4 and 2 respectively. However, in this case it is likely that the structural characteristics of the species were most influential as D. glomerata formed a denser canopy which may have partially protected the lower growing L. hispidus from ozone exposure, even immediately following cut back of the communities. Although even with this partial protection due to the surrounding species, small increases in background ozone exposure were sufficient to cause significant carry-over effects on flowering, which could have large ecological consequences.
5. Conclusions This study has shown that as well as effects during ozone exposure, increases in background ozone concentrations predicted for the next few decades can be carried over beyond the exposure season in common temperate grassland species, with no thresholds for effects being detected. Although not tested here, the carry-over effects on leaf retention and flowering suggest a long-term effect on hormone signalling. The extent of ozone-induced enhanced leaf senescence during the season, plant cover at the end of the season and green leaf retention in L. hispidus was dependant on the strength of the competition from the grass species, with lower inner leaf conductances being measured in mixtures with the stronger competitor D. glomerata than in mixtures with A. odoratum.
Acknowledgments This work was funded by the UK Department for Environment, Food and Rural Affairs (Defra, contract number AQ3510) and the UK Natural Environment Research Council.
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Environmental Pollution journal homepage: www.elsevier.com/locate/envpol
Within season and carry-over effects following exposure of grassland species mixtures to increasing background ozone Felicity Hayes*, Gina Mills, Harry Harmens, Kirsten Wyness Centre for Ecology and Hydrology, Environment Centre Wales, Deiniol Road, Bangor, Gwynedd LL57 2UW, UK
a r t i c l e i n f o
a b s t r a c t
Article history: Received 11 January 2011 Received in revised form 22 June 2011 Accepted 25 June 2011
Few studies have investigated effects of increased background ozone in the absence of episodic peaks, despite a predicted increase throughout the northern hemisphere over the coming decades. In this study Leontodon hispidus was grown with Anthoxanthum odoratum or Dactylis glomerata and exposed in the UK to one of eight background ozone concentrations for 20 weeks in solardomes. Seasonal mean ozone concentrations ranged from 21.4 to 102.5 ppb. Ozone-induced senescence of L. hispidus was enhanced when grown with the more open canopy of A. odoratum compared to the denser growing D. glomerata. There was increased cover with increasing ozone exposure for both A. odoratum and D. glomerata, which resulted in an increase in the grass:Leontodon cover ratio in both community types. Carry-over effects of the ozone exposure were observed, including delayed winter die-back of L. hispidus and acceleration in the progression from flowers to seed-heads in the year following ozone exposure. Ó 2011 Elsevier Ltd. All rights reserved.
Keywords: Ozone Flowering Carry-over effects Competition Biomass Roots
1. Introduction Mean ozone concentrations in Europe have been increasing since industrial times, from a background of 10e15 ppb in the 1900’s to the current 35e40 ppb (Simmonds et al., 2004), due to increased precursor emissions from anthropogenic sources (Solberg et al., 2005; Volz and Kley, 1988). There is some evidence that episodic peaks of ozone are declining in Europe due to reductions in emissions of ozone precursors from local and regional sources (Szopa et al., 2006; Vingarzan, 2004). However, an increase in background ozone concentration throughout the northern hemisphere has been predicted due to hemispherical transport, particularly due to increased emissions from Asia (Jaffe and Ray, 2007; Vingarzan, 2004), with annual mean concentrations likely to reach 68 ppb by 2050 (Meehl et al., 2007) and may exceed 75 ppb over much of Europe by 2100 (Sitch et al., 2007). The predicted increases would include elevated ozone concentrations at nighttime, particularly at rural, high-altitude sites. The AOT40 index (The sum of the differences between the hourly mean ozone concentration (in ppb) and 40 ppb for each hour when the concentration exceeds 40 ppb, accumulated during daylight hours), which is one of the indices currently used to define critical levels for
* Corresponding author. E-mail address: [email protected] (F. Hayes). 0269-7491/$ e see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2011.06.034
crops, semi-natural vegetation and forests, assumes that only concentrations of ozone exceeding a threshold of 40 ppb are harmful for vegetation. However, some studies have shown that concentrations below this threshold can be harmful to some species of plants (e.g., Karlsson et al., 1995). In addition, exposure to low-moderate ozone concentrations below the 40 ppb threshold can result in a large total uptake by plants if the exposure continues through the night. Few studies have investigated the effects of increased background ozone concentrations in the absence of episodic peaks. When contrasting profiles of increased background versus increased peaks have been used, it has been demonstrated that increased background concentration can be as harmful to plants as an episodic ozone regime with similar total exposure (Dawnay and Mills, 2009; Hayes et al., 2010b; Nussbaum et al., 1995). This has been attributed, in part, to a discrepancy in the timing of ozone exposure with the time of maximum anti-oxidant defence in plants, which is generally in early-mid morning (Heath et al., 2009). Consequently there is a more biologically effective ozone dose during afternoon and evening exposure. It is also thought that some nocturnal uptake of ozone occurs for many plants due to night-time stomatal conductance, which although usually lower than day time conductance, can nevertheless result in considerable uptake of ozone into a plant when ozone concentrations are high (Musselman and Minnick, 2000). It has been demonstrated that in such cases the use of AOTX24 (accumulated over a 24 h period) was
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a better fitting parameter to the response data than AOTX12 (used as a surrogate for accumulation during daylight hours) (Dawnay and Mills, 2009). Grassland communities have been identified as at risk from increasing ozone pollution as they contain species that are sensitive to ozone and are found in areas where the ambient ozone concentrations may already be high enough to cause damage (Bassin et al., 2007; Mills et al., 2007, 2011). Although several studies have investigated the impacts of ozone on component species of grasslands, comparatively few have been carried out on species mixtures and communities. These studies have shown negative effects of episodic peaks of ozone and the damaging effects demonstrated for grassland communities include visible injury (e.g., Bungener et al., 1999), premature leaf senescence (e.g., Franzaring et al., 2000), changes in biomass partitioning (e.g., Hayes et al., 2009) and reduced growth (e.g., Gimeno et al., 2004). These can lead to a change in species composition with longer-term ozone exposure (e.g., Hayes et al., 2010b). The response to ozone of some species varies according to the neighbouring plant species, for example the presence of Lolium perenne increased the proportion of senesced leaves in Trifolium repens (Hayes et al., 2010a). In a separate study, Poa pratensis showed reduced growth with increasing ozone exposure when grown with Veronica chamaedrys but not when grown with other species such as Achillea millefolium (Bender et al., 2006). In a study on a semi-natural grassland, the biomass of the forbs decreased in filtered air compared to ambient air conditions (Evans and Ashmore, 1992), in contrast to predictions based on glasshouse studies of individual plants. This was attributed to changes in light penetration through the canopy of the component grass species. Some studies suggest that the influence of the plant canopy on microclimate can alter the stomatal conductance and therefore ozone uptake of species within a plant community, for example, Lantinga et al. (1999) showed that PAR was dramatically reduced within a plant canopy, and PAR is one of the factors that can influence stomatal opening. A separate study demonstrated that a lower proportion of leaves showed visible injury symptoms in the centre of a T. repens/L. perenne plant canopy compared to those of the upper canopy and canopy edge, and this was also attributed to changes in microclimate, particularly PAR (Hayes et al., 2010a). It is therefore possible that species within a dense vegetation canopy could be protected from ozone exposure due to microclimatic conditions that do not favour ozone uptake. The response of flowering to ozone exposure has been studied for a few species and has been shown to involve changes in flower number, biomass and timing of development. A reduction in flower biomass with increasing ozone exposure has been demonstrated for some species, e.g., Trifolium cherleri, T. subterraneum and T. striatum (Gimeno et al., 2004). Similarly, ozone significantly reduced the number of flowers in Campanula rotundifolia (Rämö et al., 2007) and in some cases the reduced flowering persisted one month following ozone exposure (e.g., for T. striatum, Sanz et al., 2007). Spartina alterniflora showed delayed flowering and a reduction in the number of flower spikes produced with elevated ozone treatment (Taylor et al., 2002). Flower weights were significantly reduced for Eupatorium cannabinum and Plantago lanceolata when there had been no change in either flowering date or flower number (Franzaring et al., 2000). In contrast, increased female flower formation in response to ozone has been shown for Betula pendula (Darbah et al., 2007) and for Rubus cuneifolius an initial acceleration in flowering occurred in the second year of ozone exposure with the elevated ozone treatment, with increased flower numbers and an earlier time of peak production (Chappelka, 2002). Several studies have indicated that carry-over effects can occur following ozone exposure that may not manifest until the following
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year. However, the majority of studies on carry-over effects following ozone exposure are for trees. B. pendula saplings showed reduced contents of Rubisco, chlorophyll, reduced starch and nutrient content in leaves and decreased new shoot growth during a recovery period following ozone exposure (Oksanen and Saleem, 1999). A reduced number of overwintering buds with elevated ozone has also been shown for B. pendula, due to a lower carbon gain for bud formation (Oksanen, 2003; Riikonen et al., 2008). Buds have also been shown to be affected by ozone exposure in Fagus crenata seedlings, where those seedlings exposed to elevated ozone showed earlier leaf fall and a reduction in leaf non-structural carbohydrates, which resulted in later bud-break and a reduced number of leaves per bud in the following spring (Yonekura et al., 2004). Carry-over effects can also occur in the roots of plants, and for ponderosa pine seedlings (Pinus ponderosa) these effects included reduced new root production and lower concentrations of root starch in seedlings that had been previously exposed to ozone (Anderson et al., 1997). Investigations of carry-over effects for semi-natural vegetation have been more limited. In one study investigating effects for 33 species grown individually, re-growth of plants was affected in the spring following exposure to an episodic ozone regime. Species affected included upland grassland species such as Juncus effusus, Nardus stricta and Carex echinata; for the latter two species no effects were detected during the course of the ozone exposure (Hayes et al., 2006). For C. echinata this biomass reduction was attributed to a large decrease in flower biomass for those plants exposed to elevated ozone the previous year. In some other studies observed effects of ozone were enhanced during subsequent ozone exposures, where it was unclear whether this was due to a true carry-over effect or as a consequence of cumulative ozone exposure. For example, in clover, re-growth in subsequent ozone exposure periods was affected, whilst the biomass at the end of the first ozone exposure was not affected (Fumagalli et al., 2003; Nussbaum et al., 1995). Many semi-natural communities are dominated by perennial plant species, therefore cumulative/carryover effects may be important for the long-term viability of plant communities. The aims of this study were to investigate whether increasing background ozone concentrations are damaging to vegetation, whether the nature of the competing species modifies the response to ozone, and whether there are carry-over effects into the following growing season. This study used three species widely distributed in temperate grasslands; the forb Leontodon hispidus, and two grass species: Dactylis glomerata, a vigorous competitor and Anthoxanthum odoratum, a slower growing grass with a less dense canopy. All three of these species have been shown to respond to ozone at concentrations below 100 ppb (Hayes et al., 2007; Dawnay and Mills, 2009; Mills et al., 2009). Plants were grown in large containers in competitive mixtures of L. hispidus:A. odoratum and L. hispidus:D. glomerata, enabling the response of L. hispidus to be compared between the two community mixtures. Exposure to ozone was in computer-controlled solardomes with near-ambient climate conditions, allowing night-time concentrations to be maintained and therefore better mimicking ambient ozone profiles of upland regions. 2. Materials and methods 2.1. Ozone system and treatments Plants were exposed to ozone in solardomes (hemispherical greenhouses 3 m diameter, 2 m tall). Ozone was generated from oxygen concentrated from air (Workhorse 8, Dryden Aqua, UK) using an ozone generator (G11, Dryden Aqua, UK) and distributed to each solardome via PTFE tubing. The ozone system and treatments are described in Mills et al. (2009). Ozone was delivered to each solardome
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using mass flow controllers (Celerion, Ireland) controlled by computer software (Labview version 7). Ozone concentrations were continuously monitored in one solardome using a dedicated ozone analyser (Thermoelectron, Model 49C), allowing feedback to compensate for small variations in ozone production. In all solardomes the ozone concentration was measured for 5 min in every 30 min using two additional ozone analysers (Envirotech API 400A) of matched calibration. Eight ozone treatments were randomly allocated to the solardomes. The profiles used were based on a simulation of 1 week of ozone data from the nearby Marchlyn Mawr rural monitoring site (grid reference SH613619, 610 m.a.s.l.), using data from 31st May to 6th June 2006. This profile was repeated every week for 20 weeks with the treatments AA (simulated ambient), AA 20 ppb, AA þ 12 ppb, AA þ 24 ppb, AA þ 36 ppb, AA þ 48 ppb, AA þ 60 ppb, and AA þ 72 ppb (see Mills et al., 2009 for details). The ozone concentrations in the AA þ 36 and higher treatments were turned down to 50 ppb for access for plant measurements on Tuesdays, and occasionally on other days. 2.2. Plants and community set-up Communities were established using plug plants of Leontodon hispidus, Dactylis glomerata and Anthoxanthum odoratum (purchased from British Wildflower Plants, Norfolk, UK). L. hispidus was grown with either A. odoratum or D. glomerata. Fourteen litre pots (33.3 cm diameter 24.0 cm deep) were lined with perforated plastic sheeting to deter roots from growing through the drainage holes in the base of the pot, and filled with a compost mix in the ratio (by volume) 3 parts John Innes no 2: 1 part peat: 1 part gritty sand. Each pot contained four plants of L. hispidus and three plants of the appropriate grass species arranged in the same pattern. Communities (6 replicates per community type per solardome) were established on 23rd March 2007, and were moved into the solardomes (at 20 ppb) on 24th April. The ozone exposure treatments started on 9th May. Communities were cut back to 7 cm after 10 weeks on 18th July to simulate a conservation grassland cut (harvest data not presented), and the last day of exposure was after 20 weeks on the 25th September. Following exposure, communities were placed in a rabbit-proof plot in an adjacent field of pasture to allow natural die-back and overwintering. Starting on 13th May 2008 the L. hispidus/A. odoratum communities were reexposed to ozone (data not presented here), whereas the L. hispidus/D. glomerata communities remained in the outdoor field plot with assessments and harvests as detailed below. 2.3. Assessments of injury, leaf senescence and flowering Plants were assessed for visible injury and leaf senescence after exposure to ozone for 0, 1, 4, 7, 10, 14 and 19 weeks. A leaf was classified as senesced if >25% of the leaf was senesced, otherwise it was classified as healthy. Following the end of the ozone exposure on 25th September 2007, assessments were carried out on 28th November 2007, 30th January 2008 and 28th March 2008 for all communities, with additional assessments on 17th June, 15th July, 19th August and 9th September 2008 for the L. hispidus/D. glomerata communities only. During these assessments the numbers of flower buds, flower heads and seed-heads for each species in each mesocosm were also counted. 2.4. Stomatal conductance Stomatal conductance measurements were made on the abaxial surface of L. hispidus using a porometer (AP4, Delta-T), on days of stable meteorological conditions after exposure to the ozone regime for 17 weeks. Measurement of leaves in the upper canopy (mature leaves in full sunlight) and the inner canopy (younger leaves which were also more shaded) were taken, using five leaves for each canopy position of each community type in each solardome. 2.5. Plant cover and root biomass The area cover of each species was determined for each mesocosm by overlaying a grid of 1 cm squares over photographs taken from above on 25th September 2007. In each photograph the pot surface was covered by approximately 200 squares. Each species was scored as present if >50% of an individual square was occupied. On 18th February 2008 the root biomass was determined for three of the ozone treatments (AA, AA þ 24 and AA þ 60) for a subsample of approximately 20% of the soil taken as a vertical slice across the middle of each mesocosm, and with the same number and combination of component species (one L. hispidus and two grass) for each slice. The soil mix was removed from the roots by washing by hand. The mesocosms from the other five ozone treatments remained in the outdoor field plot. In July 2008 these communities were cut back to 7 cm to simulate a conservation grassland cut (data not included here). 2.6. Data analysis and statistics Datasets were analysed using linear regression analysis in Minitab (Version 14), with each datapoint in the regression representing the mean per treatment. Comparison
of regression lines was made using General Linear Model (Minitab, Version 14). Results were considered as significant when p < 0.05, and as trends at p < 0.1.
3. Results 3.1. Ozone exposure Ozone exposure over the 20 week period ranged from a seasonal 24 h mean of 21.4 ppbe102.5 ppb (Table 1), with the AOT40 (accumulated over 24 h) ranging from 0.07 ppm h to 214 ppm h. There was a linear relationship between 24 h mean ozone and 12 h mean ozone concentrations for the ozone treatments (r2 ¼ 0.9999). Mean PAR was 451 mmol m2 s1 during the day (07:00e18:00) and 28 mmol m2 s1 during the night-time (19:00e06:00). Further details of the ozone exposure, including the weekly mean ozone profiles for each of the ozone treatments, and climatic conditions in the solardomes are available in Mills et al. (2009). 3.2. Effects on plants during the ozone exposure No visible injury was observed during the course of the ozone exposure. After exposure to the ozone regime for 10 weeks, a linear increase in leaf senescence of L. hispidus with increasing ozone was shown when grown with A. odoratum (r2 ¼ 0.52, p ¼ 0.042), although this was not significant with D. glomerata (r2 ¼ 0.33, p ¼ 0.134; Fig. 1a). There was no significant difference between the slope of the regression lines indicating that there was no difference in the extent of leaf senescence on L. hispidus according to whether it was grown with A. odoratum compared to when grown with D. glomerata. The communities were cut back after 10 weeks, and nine weeks after this cut (after exposure to ozone for a total of 19 weeks), a linear increase in leaf senescence of L. hispidus with increasing ozone was again shown when grown with A. odoratum (r2 ¼ 0.77, p ¼ 0.004) and D. glomerata (r2 ¼ 0.42, p ¼ 0.081). Interestingly, L. hispidus showed a larger increase in leaf senescence with increasing ozone exposure when grown with A. odoratum than when grown with the densely growing coarse grass D. glomerata (Fig. 1b), with a significant difference between the slope of the two regression lines (p ¼ 0.022). Stomatal conductance measurements made after exposure to the ozone regime for 17 weeks revealed that although there were no significant differences in the stomatal conductance of leaves from the upper canopy of L. hispidus growing in the two community types, for leaves of the inner canopy stomatal conductance of L. hispidus grown with D. glomerata was approximately 30% lower than of those grown with A. odoratum (mean stomatal conductance was 176 mmol m2 s1 compared with 320 mmol m2 s1). This difference in stomatal conductance of L. hispidus between the two community types was consistent across all ozone treatments and
Table 1 Season mean ozone concentration, 12 h mean ozone concentration, AOT40 in daylight hours and AOT40 accumulated over 24 h for each treatment. AA ¼ simulated ambient ozone concentration. Ozone treatment AA AA AA AA AA AA AA AA
20 ppb þ þ þ þ þ þ
12 24 36 48 60 72
ppb ppb ppb ppb ppb ppb
Season mean ozone, ppb
12 h mean ozone (07:00e18:00), ppb
AOT40 (daylight), ppm h
AOT40 (24 h), ppm h
21.4 39.9 50.2 59.4 74.9 83.3 101.3 102.5
21.1 39.2 49.6 58.7 73.3 81.6 99.0 100.5
0.07 4.93 21.44 38.04 62.49 77.13 108.43 112.47
0.07 10.91 41.29 72.19 119.82 147.42 206.70 214.34
F. Hayes et al. / Environmental Pollution 159 (2011) 2420e2426
40
A
With A. odoratum With D. glomerata
r2
= 0.52 p=0.042
30
r2
= 0.33 p=0.134
20
10
0
0
50
100
150
Area covered, relative units
Senesced leaves, %
A
Seasonal mean O3 conc. (24 h, ppb)
r2 = 0.58 p=0.027 120 80 L. hispidus
40
r² = 0.31 p=0.154
A. odoratum
0 50
100
150
Seasonal mean O3 conc. (24 h, ppb)
40
With A. odoratum With D. glomerata
r2 = 0.77 p=0.004
B
30
20 r2 = 0.42 p=0.081
10
0
160
0
0
50
100
150
Seasonal mean O3 conc. (24 h, ppb)
Area covered, relative units
Senesced leaves, %
B
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160 120 80
At the end of the ozone exposure in September 2007, L. hispidus showed no change in cover per mesocosm with increasing ozone exposure when grown with either A. odoratum (r2 ¼ 0.31, p ¼ 0.154; Fig. 2a) or D. glomerata (r2 ¼ 0.33, p ¼ 0.139; Fig. 2b), while there was an increase in cover with increasing ozone exposure for A. odoratum (r2 ¼ 0.58, p ¼ 0.027; Fig. 2a) and D. glomerata (r2 ¼ 0.84, p ¼ 0.001; Fig. 2b). This combination of changes resulted in an increase in the grass:Leontodon cover ratio with both A. odoratum (r2 ¼ 0.54, p ¼ 0.039) and D. glomerata (r2 ¼ 0.31, p ¼ 0.151; Fig. 2c), with no significant difference in the slope of the two response relationships. 3.4. Winter die-back In the autumn the leaves of L. hispidus normally die back, however, an assessment of the mesocosms on 28th November 2007 showed a greater proportion of green, healthy leaves with
r² = 0.33 p=0.139
D. glomerata
0 0
50
100
150
Seasonal mean O3 conc. (24 h, ppb)
C Grass:Leontodon ratio
3.3. Species cover at the end of ozone exposure
L. hispidus
40
Fig. 1. Senescence of L. hispidus when grown in mixture with A. odoratum or D. glomerata after exposure to ozone for A) 10 weeks and B) 19 weeks. All communities were cut back after 10 weeks. Linear regression lines indicate changes in leaf senescence with increasing ozone exposure.
no interaction with ozone treatment was apparent (data not presented). No significant effects on timing of flowering on any of the species were found during the ozone exposure (data not presented). Similarly, 4 and 7 weeks after the end of ozone exposure there were no apparent carry-over effects of ozone on the number of flower heads of L. hispidus when grown with either A. odoratum or D. glomerata.
r² = 0.84 p=0.001
2.0
r² = 0.27 p=0.151
1.5 1.0 with D. glomerata
0.5
r² = 0.54 p=0.039
with A. odoratum
0.0
0
50
100
150
Seasonal mean O3 conc. (24 h, ppb) Fig. 2. Area covered by a) L. hispidus and A. odoratum and b) L. hispidus and D. glomerata, and c) the grass:Leontodon cover ratio after exposure to ozone for 19 weeks.
increasing ozone treatment for L. hispidus in both the A. odoratum and D. glomerata communities (Fig. 3). These leaves were mature leaves remaining since the ozone exposure, rather than a flush of new leaves that had formed subsequently. This increase was approximately 3-fold in the AA þ 72 treatment compared to AA 20 in both the A. odoratum communities (r2 ¼ 0.69, p ¼ 0.011) and the D. glomerata communities (r2 ¼ 0.53, p ¼ 0.04), with no significant difference in the slope with increasing ozone between
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F. Hayes et al. / Environmental Pollution 159 (2011) 2420e2426
1.6 'buds+flowers':seed-heads (ratio)
Green leaves (%)
40
r2 = 0.53 p=0.04
30
20 r2 = 0.69 p=0.011 10
r2 = 0.88 p=0.018
1.4 1.2 1 0.8 0.6 0.4 0.2 0
with A. odoratum
0
with D. glomerata 0 0
50
100
150
Seasonal mean O3 conc. (24 h, ppb) Fig. 3. Proportion of green leaves remaining on L. hispidus on 28th November, when grown with either A. odoratum or D. glomerata.
the community types. There was no significant difference in the proportion of green leaves of either A. odoratum or D. glomerata (data not presented). By 30th January 2008, following a sustained period of sub-zero overnight temperatures, L. hispidus in all mesocosms had completely died back. 3.5. Root weight In February 2008, just before the start of spring re-growth, the root biomass of L. hispidus was reduced by approximately 50% in the AA þ 60 treatment compared to AA in both the A. odoratum and D. glomerata communities. Using relative root biomass compared to the AA treatment to combine the root data, no difference was found when L. hispidus was grown with A. odoratum compared to D. glomerata (r2 ¼ 0.93; p ¼ 0.002 for the combined data; Fig. 4). Using the 95% confidence intervals for this response function, a small increase of 15 ppb above the simulated ambient treatment in this study (40 ppb) would be sufficient to induce a significant decrease of 10% in root biomass of L. hispidus. The root biomass of A. odoratum was reduced by 40% and 55% in the AA þ 24 and AA þ 60 treatments respectively compared to AA. For D. glomerata, the root biomass was reduced by approximately 30% in both the AA þ 24 and AA þ 60 treatments compared to AA.
1.2 r2=0.93 p=0.002
Root biomass (%)
1.0 0.8 0.6 0.4 0.2
with D. glomerata with A. odoratum
0.0 0
50
100
50
100
150
Seasonal mean O3 conc. (24 h, ppb)
150
Seasonal mean O3 conc. (24 h, ppb) Fig. 4. Relative weight of roots of L. hispidus (compared to the AA treatment) when grown in combination with D. glomerata or A. odoratum.
Fig. 5. Ratio of buds þ flowers heads to seed-heads for L. hispidus grown with D. glomerata in July 2008, after exposure to ozone in 2007.
3.6. Carry-over effects on flowering By 28th March 2008, plants had started to re-grow, however, there were no differences in the numbers of L. hispidus leaves per mesocosm between the different treatments in either community type (data not presented). On 19th May (when only the L. hispidus eD. glomerata communities had remained in the field plot) there were no flowers developing on the L. hispidus, and no significant difference in the number of D. glomerata panicles with increasing ozone treatment (r2 ¼ 0.25, p ¼ 0.393). By 17th June there was no difference in the number of flower heads/panicles on either L. hispidus or D. glomerata. On 15th July, although there were no differences in the total number of L. hispidus flower heads between the different ozone treatments, the ratio of buds and flower heads to seed-heads had changed, with the ratio decreasing with increasing ozone concentration (Fig. 5; r2 ¼ 0.88, p ¼ 0.018), indicating that the plants were going through the reproductive part of their life-cycle more quickly. There were no significant differences for D. glomerata. The same pattern was observed on 9th September, with the ratio of buds and flower heads to seed-heads decreasing for L. hispidus with increasing ozone concentration (r2 ¼ 0.42, p ¼ 0.238), and no panicles present for D. glomerata. 4. Discussion During the course of ozone exposure there was a linear increase in leaf senescence with increasing background ozone concentration. The extent of leaf senescence on L. hispidus was shown to be influenced by neighbouring species, with a larger increase in ozone-induced senescence observed in the more open canopy of A. odoratum compared to the denser canopy of D. glomerata after prolonged exposure to ozone. Reductions in ozone concentration within plant canopies have been demonstrated, with these reductions being up to 30% in a grassland canopy (Jäggi et al., 2006). Together with reduced windspeed influencing the boundary layer around plant leaves, this has been hypothesized to result in decreased ozone uptake by plants in more dense canopies. In the current study, there was evidence to support this theory as there was reduced stomatal conductance of inner canopy leaves of L. hispidus grown with D. glomerata compared to when grown with A. odoratum and this corresponded with the observed differences in leaf senescence of L. hispidus when grown with each of these species after prolonged exposure to ozone. At the end of the exposure, although there was no change in cover of L. hispidus with increasing ozone exposure biochemical changes must have occurred to result in the carry-over effects
F. Hayes et al. / Environmental Pollution 159 (2011) 2420e2426
observed. It is unlikely that the carry-over effects in this study were a consequence of the use of pots during the study, as the pots used were large and the plants had not become pot-bound by the end of the experiment. Previous studies have shown that elevated ozone tended to accelerate leaf abscission in the autumn for Betula papyrifera and Populus tremuloides at the Aspen FACE site, USA (Riikonen et al., 2008) and this has also been reported for several other tree species e.g., B. pendula (Uddling et al., 2006); Fagus sylvatica (Nunn et al., 2005). However, in the current study increasing ozone exposure corresponded with a delayed die-back of L. hispidus in autumn/winter, despite the enhanced leaf senescence that had occurred for these plants earlier in the growing season. There was also no difference according to which competing grass the L. hispidus was grown with, even though the extent of the earlier ozone-induced leaf senescence was greater on L. hispidus grown with A. odoratum. This could indicate a carry-over effect on hormonal signalling as leaf abscission is a highly regulated phase in leaf development with many genes encoding degradative enzymes having been identified (Buchanan-Wollaston, 1997; BuchananWollaston et al., 2003) and hormonal control by abscisic acid (ABA), ethylene and auxin being clearly demonstrated (Taylor and Whitelaw, 2001). Some support for this theory comes from previous studies that have indicated that prolonged ozone exposure disrupts hormonal control of stomatal functioning by ABA and ethylene in L. hispidus and the plants become insensitive to ABA (Mills et al., 2009; Wilkinson and Davies, 2009). ABA is also involved in the regulation of flowering, however the role is not well understood (Wilmowicz et al., 2008). Thus, it is possible that the accelerated flowering of L. hispidus in the current study, in conjunction with the delayed die-back in winter could indicate alterations in sensitivity to ABA or to the ABA signalling process. The observed accelerated flowering of L. hispidus with increasing background ozone concentration has implications for ecosystem services as this species is considered to be an important source of nectar for nectar feeders including butterflies, hoverflies, bees and other insects. These nectar feeders could be indirectly affected by previous ozone exposure as the timing of the food source availability is affected. In the current study, the acceleration in development of flowers in L. hispidus leading to a shorter timeperiod of flower duration was a carry-over effect and occurred some ten months following the end of ozone exposure in the previous growing season. Despite the differences in the extent of leaf senescence of L. hispidus when grown with A. odoratum or D. glomerata there was no difference in the relative root reduction with increasing ozone for L. hispidus between the two community types. Larger effects on roots than on shoots have been demonstrated for some species and a reduction in allocation to roots in response to ozone exposure has been shown for Cirsium arvense (Power and Ashmore, 2002), Cirsium dissectum (Franzaring et al., 2000) and some wetland species (Batty and Ashmore, 2003). A meta-analysis of ozone impacts on plant development by Grantz et al. (2006) showed that changes in allocation, with reduced allocation to roots with increasing ozone concentration commonly occur, although the extent of change is variable, and results from the current study support this because effects of ozone on root weight as a result of the ozone exposure were of a larger magnitude than changes in above-ground cover. A reduction in root biomass at the end of the ozone exposure suggests a decrease in available resources for re-growth in the following year and gives further indication of potential carry-over effects of ozone exposure. Based on a compilation of published data on the effects of episodic ozone regimes on above-ground biomass, L. hispidus has previously been identified as ozone sensitive (Hayes et al., 2007), with A. odoratum and D. glomerata classified as insensitive to ozone. However, other studies conducted within the solardomes have
2425
shown that all three species respond to ozone by enhancing leaf senescence, with thresholds for effects below 100 ppb (Dawnay and Mills, 2009; Mills et al., 2009). The current study indicates that although these species appeared insensitive to ozone based on changes in above-ground area covered (agreeing with Hayes et al., 2007), significant effects of ozone on these species occur below ground at lower concentrations. In addition the nature of the competing species can alter the response to ozone of L. hispidus with this species being more sensitive to ozone when grown with A. odoratum than when grown with D. glomerata in terms of enhanced leaf senescence. D. glomerata is a stronger competitor than A. odoratum with scores for ‘competitor’ sensu Grime (Grime et al., 1988) of 4 and 2 respectively. However, in this case it is likely that the structural characteristics of the species were most influential as D. glomerata formed a denser canopy which may have partially protected the lower growing L. hispidus from ozone exposure, even immediately following cut back of the communities. Although even with this partial protection due to the surrounding species, small increases in background ozone exposure were sufficient to cause significant carry-over effects on flowering, which could have large ecological consequences.
5. Conclusions This study has shown that as well as effects during ozone exposure, increases in background ozone concentrations predicted for the next few decades can be carried over beyond the exposure season in common temperate grassland species, with no thresholds for effects being detected. Although not tested here, the carry-over effects on leaf retention and flowering suggest a long-term effect on hormone signalling. The extent of ozone-induced enhanced leaf senescence during the season, plant cover at the end of the season and green leaf retention in L. hispidus was dependant on the strength of the competition from the grass species, with lower inner leaf conductances being measured in mixtures with the stronger competitor D. glomerata than in mixtures with A. odoratum.
Acknowledgments This work was funded by the UK Department for Environment, Food and Rural Affairs (Defra, contract number AQ3510) and the UK Natural Environment Research Council.
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