Ozone :: Woodland and hedgerows


Effects and Implications

  • Visible leaf injury and/or premature leaf die-back, which could subsequently result in reduced growth during a growing season;
  • Reduced growth of sensitive species;
  • Potential alterations in numbers and timing of flowering and seed production;
  • Alterations of response to other environmental stresses such as drought stress
  • Enhanced susceptibility to pests and diseases.
  • Alterations to litter quality, which could affect decomposition

Habitat structure could be altered via changes in growth of species, particularly as the extent of growth reduction would vary between species. There may be seasons in which the structure is most influenced as changes in leaf die-back could result in a more open canopy.  In addition habitat function could be impacted via changes in timing of flowering and leaf-fall.

Overview: evidence, processes and main impacts

Effects of ozone have been studied using trees, shrubs and undergrowth species, although many studies have been short-term (< 3 years). Many studies on trees have used saplings rather than mature species, although there is some evidence that the response to ozone is similar in young and mature trees in terms of reduction in shoot growth (e.g. Braun et al., 2007) and the extent of visible injury (e.g. Baumgarten et al., 2000). Ozone is very reactive and the reactive oxygen (molecular) species that form as ozone enters a leaf can quickly cause damage to cell components, e.g. cell membranes. Reactive oxygen species are also produced by plant cells in many processes involved in growth and development and from abiotic stresses, e.g. high light and drought.  This is one of the reasons that some symptoms of ozone pollution are not specific to ozone.

Ozone exposure studies have shown that ozone-specific visible leaf injury symptoms can occur for some tree species, e.g. Fraxinus excelsior, Quercus robur (Gerosa et al., 2009) and Fagus sylvatica (Baumgarten et al. 2000), in addition to species of the under-storey vegetation, e.g. Ranunculus acris (Wagg et al., 2012). Various ozone-sensitive tree species show reductions in biomass in response to elevated ozone (e.g. Wittig et al., 2009 and references therein; LTRAP Convention, 2011 and references therein). In general, Angiosperms (trees with seed enclosed, usually in a fruit) are more sensitive to ozone than Gymnosperms (trees with naked seeds, usually in a cone; Wittig et al., 2009). Some studies have indicated that a change in biomass partitioning can occur in response to ozone. A decrease in the dry mass of roots of Betula pendula attributed to ozone has been shown at the end of the exposure (Riikonen et al. 2004) and could decrease the ability of a tree to take up water and nutrients. Wittig et al. (2009) reported a decline in root to shoot ratio in trees exposed to high ozone concentrations. It is thought that decreased partitioning to the roots may occur with increasing ozone exposure because the mature, lower leaves act as the main source of assimilate for root growth, and these are frequently the most damaged by ozone (Grantz et al. 2006, Cooley and Manning 1987, Okano et al. 1984). However, this has not been demonstrated for all species and some, e.g. Fagus sylvatica and Picea abies, showed no effect of ozone exposure on carbon allocation to roots (Andersen et al., 2010). Litter quantity and quality can also be affected by ozone exposure, e.g. for silver birch (Kasurinen et al., 2006), which can affect subsequent decomposition and may influence decomposers, detritivores and nutrient cycling. Holmes et al. (2006) showed that elevated ozone could reduce nitrogen availability in the soil via changes in litter production.

Elevated ozone has also been shown to have carry-over effects the following spring in some species, particularly those that form buds for overwintering as the quality of the buds can be affected by ozone (Riikonen et al., 2008). In addition, for grassland species it has been shown that ozone can affect seed production, viability (Black et al., 2000) and germination (Bender et al., 2006; Bergmann et al., 1999), and this may also occur for species of woodlands. There is a wide range in sensitivity to ozone of tree, shrub and undergrowth species, indicating that elevated ozone conditions could contribute to changes in species composition, particularly in the undergrowth. However, it can be difficult to predict and generalise about which groups of species are most ozone-sensitive.

It has been demonstrated that ozone can impair the functioning of leaf pores (stomata) responsible for gas exchange (e.g. water, carbon dioxide, ozone) between the plant and the atmosphere. A reduced response to drought has been shown for grassland species (e.g. Mills et al., 2009; Wagg et al., 2013), implying that soil drying during a prolonged drought would be further exacerbated by ozone pollution. It has also been suggested that changes in water-use corresponding to changes in stomatal opening in response to ozone could affect streamflow within a wooded catchment (McLaughlin et al., 2007). On the other hand, drought can override the stimulating ozone effects on fine-root dynamics and soil respiration in mature beech and spruce forests (Nikolova et al., 2010). At the leaf level, the impact of ozone was reduced because from early summer drought-driven stomatal closure pre-empted ozone-driven effects (Löw et al., 2006).

A concentration-based critical level was established for the protection of tree species based on production of roundwood for the forest industry, loss of carbon storage capacity in the living biomass of trees and other beneficial ecosystem services provided by trees. This is an AOT40 (accumulated ozone concentration over a threshold of 40 ppb in daylight hours) of 5 ppm.h (see critical levels table) and is based on the concentration of ozone in the atmosphere at the top of the tree canopy (LRTAP Convention, 2011). The applicability of this concentration to protect deciduous and evergreen tree species was verified based on exposure-response experiments (documented in LRTAP Convention., 2011).

More recently it has been shown that the impacts of ozone depend on the amount of the pollutant reaching the sites of damage within the leaf and stomatal (leaf pore) and flux-based critical levels have been developed. Flux-based critical levels for effects on trees are based on flux to the upper canopy leaves of individual species and are derived from experiments measuring effects of ozone on whole tree biomass. Although flux-effect relationships have been derived for several species, Norway spruce and combined beech and birch were selected as these were the most robust (see critical levels table). Critical levels were derived for the cumulative ozone flux responsible for either a 2% (for slow-growing Norway spruce) or a 4% (for combined beech and birch) reduction in annual growth (whole tree biomass) of young trees of up to 10 years of age, reflecting the age of the trees used in the experiments (Mills et al., 2011). Recently, the stomatal flux function for beech was validated by epidemiological studies with mature trees (Braun et al., 2010).

To date, no separate critical level has been established for under-storey vegetation. The critical level set for (semi-)natural vegetation is more appropriate for the ground flora. The flux-based critical levels for effects on (semi-)natural vegetation are based on flux to the upper canopy leaves of Trifolium (clover) species frequently found in grassland communities across Europe (see grasslands).

Pollutant type and risk

Ozone is a naturally occurring chemical in the troposphere. Natural sources of the precursors of ozone such as oxides of nitrogen and non-methane volatile organic compounds ensure that there is always a background concentration of ozone. Additional ozone is formed from complex photochemical reactions of precursors released due to anthropogenic emissions (particularly from transport). Ozone concentrations are usually highest in rural and upland areas downwind of major conurbations.

In the UK large areas of BAP priority woodland habitats coincide with areas where the ozone exposure is highest. It has been estimated that in England and Wales at least 75% of BAP priority ‘Upland oakwood’, ‘Wet woodland’ and ‘Lowland beech and yew’ are in areas where ozone concentrations are moderate to high (>4750 ppb.h, based on 1999-2003 values).  In Scotland at least 25% of BAP priority ‘Upland oakwood’, ‘Native pine woodlands’ and ‘Upland mixed ashwoods’ are in areas where ozone concentrations are moderate to high (Morrisey et al., 2007).

Indicators of ozone impacts

These can be difficult to identify in ‘field’ conditions. Often the symptoms of ozone injury are those of a general stress response, and in addition a rapid turnover of damaged leaves can make attribution to ozone pollution difficult.  Some species of trees and undergrowth species exhibit ozone-specific visible leaf injury symptoms in controlled studies. These ‘typical’ symptoms of visible leaf injury attributed to ozone have not been recorded in natural conditions in the UK, although this may be in part because there are few experienced observers. 

Examples of species specific responses

Some examples of specific responses are given in the table below. This does not represent a comprehensive review of all species impacts. Many additional examples for trees are given in Wittig et al. (2009).




Betula pendula

Reduced root biomass

Riikonen et al., 2004

Fagus sylvatica

Reduced shoot growth

Braun et al., 2007

Alnus glutinosa

Visible leaf injury

Gunthardt-Georg, 1996

Ranunculus acris

Reduced biomass

Wagg et al., 2012

Rhamnus cathartica

Visible leaf injury

Vanderheyden et al., 2001

What factors modify ozone impacts?

Climatic (e.g. temperature, humidity) and soil factors (e.g. water content) that influence stomatal uptake of ozone can influence ozone impacts in some species (see above section on stomatal fluxes). For some species drought conditions protected trees from injury caused by ozone exposure via reduced ozone uptake (e.g. Fagus sylvatica, Löw et al., 2006; Populus spp, Silim et al., 2009). However, a meta-analysis showed no conclusive evidence for a protective role of drought against ozone induced effects on growth and biomass (Wittig et al., 2009).

Critical levels:

Flux-based critical levels (Mills et al., 2011b, LRTAP Convention, 2011)



(per cent reduction)



Critical level

(mmol m-2 PLA4)

Norway spruce

Biomass (2%)



Birch and beech

Biomass (4%)



Conservation grasslands (based on clover)1

Biomass (10%)



Concentration-based critical levels (LRTAP Convention, 2011)






Critical level

(ppm h)

Forest trees

Growth reduction



(Semi-)natural vegetation communities dominated by perennials1

Growth reduction



1 Considered to be applicable for forest understorey.

2 POD1 = Phytotoxic Ozone Dose above a flux threshold of 1 nmol m-2 PLA s-1.

3 AOT40 = Accumulated ozone above a threshold of 40 ppb during daylight hours.

4 PLA = Projected leaf area.

Environmental limit: 

Habitat/ Ecosystem Type Critical Load/ Level Status Reliability Indication of exceedance Reference

AOT40 5000ppb over 6 months

UNECE 2010 expert judgement i.e. only limited or no data are avaliable for this type of receptor

AOT40 is the Accumulated concentration Over a Threshold of 40 ppb. If an hourly average ozone concentration exceeds 40 ppb the difference between the concentration and 40 ppb is added to a running total. The units are therefore ppb multiplied by hours. For forests, the AOT40 is summed for the daylight hours over a six month period. Daylight hours are defined as when solar radiation exceeds 50 W m-2. The daylight hours are when plant stomata are normally open. Most research has been done on beech, and there is little on birch or the other species typical of birch woodland. 

Flux-based critical levels, based on biomass reduction, are also available for local and regional assessment but are not yet incorporated into APIS. See ICP Vegetation Manual 2010.



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