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Air pollution and vegetation: ICP Vegetation annual report 2011/2012

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Air Pollution
and Vegetation
ICP Vegetation1
Annual Report
2011/2012
Harry Harmens1, Gina Mills1, Felicity Hayes1, David Norris1
and the participants of the ICP Vegetation
1 ICP Vegetation Programme Coordination Centre, Centre for Ecology and Hydrology,
Environment Centre Wales, Deiniol Road, Bangor, Gwynedd, LL57 2UW, UK
Tel: + 44 (0) 1248 374500, Fax: + 44 (0) 1248 362133, Email: hh@ceh.ac.uk
http://icpvegetation.ceh.ac.uk
September 2012
1 International Cooperative Programme on Effects of Air Pollution on Natural Vegetation and Crops.
Acknowledgements
We wish to thank the UK Department for Environment, Food and Rural Affairs (Defra) for the
continued financial support of the ICP Vegetation (Contract AQ0816). Contributions from Lisa
Emberson, Patrick Büker, Howard Cambridge, Steve Cinderby, Richard Falk and Alan Briolat (SEI-
York, UK), , Sally Wilkinson and Bill Davies (Lancaster University, UK), Stephen Arnold (University of
Leeds, UK), Stephen Sitch (University of Exeter, UK), Bill Collins and Chris Jones (Met Office, UK) as
sub-contractors are gratefully acknowledged. In addition, we wish to thank the United Nations
Economic Commission for Europe (UNECE) and the UK Natural Environment Research Council
(NERC) for the partial funding of the ICP Vegetation Programme Coordination Centre.
Contributions in kind from David Simpson (EMEP-MSC/West), Ilia Ilyin (EMEP-MSC-East), Per Erik
Karlsson (IVL, Sweden), Gerhard Soja (AIT, Austria) and Matthias Volk (ART, Switzerland) are also
gratefully acknowledged.
We thank Giacomo Gerosa, Antonio Ballerin-Denti and their colleagues at the Mathematics and
Physics Department, Università Cattolica del Sacro Cuore, Brescia (Italy) for their support in
organising the 25th ICP Vegetation Task Force meeting, and we thank the Swiss Federal Office for the
Environment (FOEN) for financial support for printing the WGE Impacts brochure.
Finallly, we wish to thank all of the ICP Vegetation participants for their continued contributions to the
programme and other bodies within the LRTAP Convention.
Front cover photo: Dr Laurence Jones (CEH, Bangor)
Executive Summary
Background
The International Cooperative Programme on Effects of Air Pollution on Natural Vegetation and Crops
(ICP Vegetation) was established in 1987. It is led by the UK and has its Programme Coordination
Centre at the Centre for Ecology and Hydrology (CEH) in Bangor. It is one of seven ICPs and Task
Forces that report to the Working Group on Effects (WGE) of the Convention on Long-range
Transboundary Air Pollution (LRTAP Convention) on the effects of atmospheric pollutants on different
components of the environment (e.g. forests, fresh waters, materials) and health in Europe and North-
America. Today, the ICP Vegetation comprises an enthusiastic group of over 200 scientists from 35
countries in the UNECE region with outreach activities to other regions such as Asia, Central America
and Africa. An overview of contributions to the WGE workplan and other research activities in the year
2011/12 is provided in this report.
25th ICP Vegetation Task Force meeting
The Programme Coordination Centre organised the 25th ICP Vegetation Task Force meeting, 31
January - 2 February 2012 in Brescia, Italy, in collaboration with the local hosts at the Ecophysiology
and Environmental Physics Laboratory - Mathematics and Physics Department, Università Cattolica
del Sacro Cuore. The meeting was attended by 73 experts from 21 countries, including Egypt and
South Africa. Participation at the annual Task Force meeting has been rising steadily over the 25
years from ca. 20 experts from 10 countries to the current level of over ca. 70 participants from more
than 20 countries. The Task Force discussed the progress with the workplan items for 2012 and the
medium-term workplan for 2013 - 2015 for the air pollutants ozone, heavy metals, nutrient nitrogen
and persistent organic pollutants (POPs). The meeting was preceded by a one-day ozone workshop
where the following two themes were discussed:
- Quantifying ozone impacts on Mediterranean forests;
- Mapping vegetation at risk from ozone at the national scale.
New scientific developments presented at the workshop support the use of the stomatal flux-based
method rather than the concentration-based method for ozone impact assessments on vegetation.
The workshop recommended to further develop the ozone flux-based method and provide further
field-based validation of ozone flux-effect relationships and critical levels via epidemiological studies.
A book of abstracts, details of presentations and the minutes of the 25th Task Force meeting and
ozone workshop are available from the ICP Vegetation web site (http://icpvegetation.ceh.ac.uk).
Reporting to the Convention and other publications
In addition to this report, the ICP Vegetation Programme Coordination Centre has provided a
technical report on ‘Effects of air pollution on natural vegetation and crops’
(ECE/EB.AIR/WG.1/2012/8) and contributed to the joint report (ECE/EB.AIR/WG.1/2012/3) and
impact analysis report (ECE/EB.AIR/WG.1/2012/13) of the WGE. It contributed to the Guidance
Document VII on heatlh and environmental improvements for the revision of the Gothenburg Protocol.
The ICP Vegetation also published the glossy report and summary brochure for policy makers on the
threat of ozone to food security, and published a glossy report on the impact of ozone on carbon
sequestration in Europe. The ICP Vegetation contributed to and facilitated printing of the colour
brochure of the WGE on ‘Impacts of air pollution on human health, ecosystems and cultural heritage’
in English, French and Russian. In addition, a colour leaflet was produced on ‘Mosses as biomonitors
of atmospheric heavy metal pollution in Europe’ in English and Russian. Three scientific papers have
been published or in press and the ICP Vegetation web site was updated regularly with new
information.
Contributions to the WGE common workplan
Further implementation of Guidelines on Reporting of Air Pollution Effects
The ICP Vegetation continued to monitor and model deposition to and impacts on vegetation for the
air pollutants ozone, heavy metals, nitrogen and POPs.
Final version of the impact analysis by the WGE
To support the revision of the Gothenburg Protocol, the WGE has conducted an analysis on the
impacts of air pollution on ecosystems, human health and materials under different emission
scenarios, including the application of recently developed effects indicators such as the phytotoxic
ozone dose (POD; flux-based approach). Results from the ICP Vegetation show that despite
predicted reductions in both ozone concentrations and stomatal fluxes in 2020, large areas in Europe
will remain at risk from adverse impacts of ozone on vegetation, even after implemation of maximum
technically feasible reductions, with areas at highest risk being predicted in parts of western, central
and southern Europe.
Ideas and actions to enhance the involvement of EECCA/SEE countries in Eastern Europe, the
Caucasus and Central Asia and on cooperation with activities outside the Air Convention
Working with the lead participant of the European moss survey in the Russian Federation, the ICP
Vegetation is actively encouraging the participation of more EECCA/SEE countries. For example,
Albania took part for the first time in the moss survey in 2010/11 and attended the ICP Vegetation
Task Force meeting for the first time in 2012. Together with the Stockholm Environment Institute (SEI)
in York (UK) the ICP Vegetation has produced a position paper on outreach activities to Malé
Declaration countries in South Asia. Suggestions were provided and discussed for further
collaboration in the near future at the third meeting of the Task Force of the Malé Declaration, 9-10
August 2012, Chonburi, Thailand. However, implementation of further collaboration is severly
hindered by the lack of available funds. The ICP Vegetation also has developed collaboration with
experts in Egypt, South Africa, Cuba and Japan, who have attended recent Task Force meetings of
the ICP Vegetation.
Progress with ICP Vegetation-specific workplan items in 2011/12
Supporting evidence for ozone impacts on vegetation
Since 2008, participants of the ICP Vegetation have been conducting biomonitoring campaigns using
ozone-sensitive (S156) and ozone-resistant (R123) genotypes of Phaseolus vulgaris (Bush bean,
French Dwarf bean). In 2011, the biomonitoring of ozone effects using bean was scaled down
compared to the previous two years, reflecting less interest from the participants. Nevertheless,
experiments were conducted with ozone-sensitive and ozone-resistant bean (Phaseolus vulgaris) at
nine sites across Europe and one in the USA. The data from the 2011 and previous biomonitoring and
ozone exposure experiments were combined in a database for dose-response analysis. The database
contains data from Belgium, France, Germany (3 sites), Greece (2 sites), Hungary (2 sites), Italy (3
sites), Japan, Slovenia (2 sites), Spain (3 sites), South Africa, UK (2 sites), Ukraine and the USA.
Visible leaf injury regularly occurred across the network, but there was no clear dose-response
relationship with ozone parameters. Similarly, there was no clear relationship between concentration-
based parameters and the ratio of the pod weight for the sensitive to that of the resistant bean. Flux-
effect relationships will be explored in the coming year.
Ozone impacts on carbon sequestration in Europe
Terrestrial vegetation, particularly forests, is an important sink for the greenhouse gases carbon
dioxide (CO2) and ozone. However, the air pollutant ozone has a negative impact on cell metabolism
and growth of ozone-sensitive plant species. Hence, this will result in a positive feedback to global
warming as less CO2 and ozone will be sequestered by vegetation, resulting in a further rise of their
concentrations in the atmosphere. The future impacts of ozone on carbon (C) sequestration in
European terrestrial ecosystems will depend on the interaction with and magnitude of the change of
the physical and pollution climate, represented by rising temperatures, increased drought frequency,
enhanced atmospheric CO2 concentration and reduced nitrogen deposition. For the first time we
applied the DO3SE (Deposition of Ozone for Stomatal Exchange) model to estimate the magnitude of
the impact of ambient ozone on C storage in the living biomass of trees. The Phytotoxic Ozone Dose
above a threshold value of Y nmol m-2 s
-1 (PODY) was calculated applying known flux-effect
relationships for various tree species. When applying a standard parameterisation for deciduous and
conifer trees, current ambient ground-level ozone was estimated to reduce C sequestration in the
living biomass of trees by 12.0 to 16.2% (depending on ozone, meteorological and climate input data)
in the EU27 + Norway + Switzerland in 2000. The flux-based approach indicated the highest ozone
impact on forests in central Europe, where moderate ozone concentrations coincide with a climate
conducive to high stomatal ozone fluxes and with high forest carbon stocks. A considerable reduction
was also calculated for parts of northern Europe, especially when applying climate region-specific
parameterisations. Under drought-free conditions (i.e. no limitation of soil water availability for tree
growth), the predicted reduction in C sequestration in the living biomass of trees increased from 12.0
to 17.3% in the year 2000, with the highest reductions predicted for the warmer and drier climates in
the southern half of Europe, particularly in the Mediterranean. Although a decline in stomatal ozone
flux was predicted in 2040, C sequestration in the living biomass of trees will still be reduced by
12.6% (compared to 16.2% in 2000).
The above results only describe ozone effects on the living biomass of trees and do not take into
account any effect of ozone on soil carbon cycling, impacts of potential changes of forest
management in the future or feedbacks to the climate system. Although the flux-response functions
used were derived for young trees (up to 10 years of age), there is scientific evidence from some
epidemiological studies that the functions are applicable to mature trees as assumed in this study.
There is a clear need to include the impacts of ozone on vegetation in global climate change
modelling.
Progress with European heavy metals and nitrogen in mosses survey 2010/11
Mosses have been collected for element analysis every five years since 1990 and the most recent
survey was conducted in 2010/11. A total of 26 countries will submit or have already submitted data
on heavy metals, of which 14 countries will also submit (or have submitted) data on nitrogen
concentrations in mosses. Nitrogen concentrations were reported for the first time in the 2005/6
European moss survey. As a pilot study, six countries have agreed to submit data on POPs
concentrations in mosses to further assess the suitability of mosses as biomonitors of atmospheric
POPs pollution. The final report on the 2010/11 European moss survey will be published in 2013.
Relationship between (i) heavy metal and (ii) nitrogen concentrations in mosses and their impacts on
ecosystems
A review of the scientific literature showed that little is known about the relationship between heavy
metal concentrations in mosses and the impacts of heavy metals on terrestrial ecosystems. Toxicity
effects of heavy metals are usually limited to areas close to pollution sources, with impacts often
declining exponentially with distance from the pollution source. For example, in agreement with an
observed gradient of reducing heavy metal concentrations in mosses away from a heavy metal
pollution source, an increase was observed in the abundance of soil mesofauna with distance from
the pollution source. However, in the European survey, mosses are not sampled close to pollution
sources and hence concentrations are often too low to be associated with an impact on terrestrial
ecosystems in the sampling areas. This does not mean that we should not be concerned about heavy
metal deposition in remote areas as metals will accumulate in the soil and might become a problem in
the future if bio-available concentrations reach critical limits.
A review of the scientific literature showed that little is also known about the relationship between
nitrogen concentrations in terrestrial mosses and impacts of nitrogen on terrestrial ecosystems. The
nitrogen concentration in the moss species used in the European moss survey tends to be a good
indicator of total atmospheric nitrogen deposition up to a deposition flux of ca. 15 kg ha-1 y-1. Above
this level, the nitrogen concentration in mosses tends to saturate, although the level at which
saturation occurs varies geogaphically. Empirical critical loads have been defined for various habitats
(ECE/EB.AIR/WG.1/2010/14), however, the effects indicators for exceedance have not been related
so far to nitrogen concentrations in mosses per se. For many terrestrial ecosystems with an empirical
critical load below 15 kg ha-1 y-1 nitrogen effects have been reported on moss species (e.g. changes in
moss species composition or abundance). Recent studies have shown that vegetation responses to
nitrogen deposition might depend more on the nitrogen form (ammonia or nitrate) than dose.
Vegetation tends to be more sensitive to ammonia (ECE/EB.AIR/WG.5/2007/3) than nitrate exposure.
Future activities of the ICP Vegetation
The ICP Vegetation Task Force has agreed to conduct a review and publish a glossy state of
knowledge report on ‘Ozone impacts on biodiversity and ecosystem services’ in 2013. Highlights from
this study will be submitted for inclusion in the WGE’s report on impacts of air pollution on biodiversity
and ecosystem services. As one of it’s core activities the ICP Vegetation will continue ozone stomatal
flux model development and flux map validation. Hence, we will continue to collate supporting
evidence for ozone impacts on vegetation and review the robustness of flux-effect relationships for
the establishment of new flux-based ozone critical levels for additional plant species. In 2013, the ICP
Vegetation will report on the outcome of the 2010/11 European moss survey for heavy metals,
nitrogen and POPs. The ICP Vegetation will also continue to explore opportunities for outreach
activities to other regions of the globe.
Contents
ACKNOWLEDGEMENTS
EXECUTIVE SUMMARY
1INTRODUCTION.................................................................................................................................... 10
1.1BACKGROUND ............................................................................................................................................ 10
1.2AIR POLLUTION PROBLEMS ADDRESSED BY THE ICP VEGETATION ............................................................. 10
1.2.1Ozone .............................................................................................................................................. 10
1.2.2Heavy metals ................................................................................................................................... 11
1.2.3Nitrogen .......................................................................................................................................... 12
1.2.4Persistent organic pollutants (POPs) .............................................................................................. 12
1.3WORKPLAN ITEMS FOR THE ICP VEGETATION IN 2012 ............................................................................... 12
2COORDINATION ACTIVITIES ........................................................................................................... 14
2.1ANNUAL TASK FORCE MEETING ................................................................................................................. 14
2.2REPORTS TO THE LRTAP CONVENTION ..................................................................................................... 15
2.3SCIENTIFIC PAPERS ..................................................................................................................................... 16
3ONGOING RESEARCH ACTIVITIES IN 2011/12 ............................................................................. 17
3.1CONTRIBUTIONS TO WGE COMMON WORKPLAN ITEMS .............................................................................. 17
3.1.1Further implementation of the Guidelines on Reporting of Air Pollution Effects ........................... 17
3.1.2Contributions to the completed version of the impact analysis by the WGE .................................. 17
3.1.3Ideas and actions to enhance the involvement of EECCA/SEE countries in the Eastern Europe,
the Caucasus and Central Asia and on cooperation with activities outside the Air Convention .... 20
3.2PROGRESS WITH COMMON ICP VEGETATION WORKPLAN ITEMS ................................................................. 20
3.2.1Supporting evidence for ozone impacts on vegetation .................................................................... 20
3.2.2Progress with European heavy metals and nitrogen in mosses survey 2010/11 ............................. 22
4IMPACTS OF OZONE ON CARBON SEQUESTRATION AND LINKAGES BETWEEN
OZONE AND CLIMATE CHANGE ..................................................................................................... 23
4.1INTRODUCTION ........................................................................................................................................... 23
4.1.1Background ..................................................................................................................................... 23
4.1.2Ozone pollution ............................................................................................................................... 23
4.1.3Vegetation as a sink for atmospheric CO2 and ozone ..................................................................... 23
4.1.4Ozone impacts in a changing climate .............................................................................................. 24
4.2IMPACTS OF OZONE ON C SEQUESTRATION IN THE LIVING BIOMASS OF TREES ............................................ 25
4.2.1First flux-based assessment for Europe for the current (2000) and future climate (2040) ............. 25
4.2.2Case study in northern and central Europe applying the AOT40 method ....................................... 28
4.2.3A global perspective of impacts on C storage in terrestrial ecosystems ......................................... 29
4.2.4Recommendations ........................................................................................................................... 30
5RELATIONSHIP BETWEEN (I) HEAVY METAL AND (II) NITROGEN
CONCENTRATIONS IN MOSSES AND THEIR IMPACTS ON ECOSYSTEMS ......................... 31
5.1HEAVY METALS .......................................................................................................................................... 31
5.2NITROGEN .................................................................................................................................................. 33
6FUTURE ACTIVITIES OF THE ICP VEGETATION ....................................................................... 37
6.1REVIEW OF OZONE IMPACTS ON BIODIVERSITY AND ECOSYSTEM SERVICES ................................................ 37
6.2MEDIUM-TERM WORKPLAN (2013-2015) OF THE ICP VEGETATION ........................................................... 37
REFERENCES ................................................................................................................................................... 39
ANNEX 1. PARTICIPATION IN THE ICP VEGETATION ........................................................................ 44
10
1 Introduction
1.1 Background
The International Cooperative Programme on Effects of Air Pollution on Natural Vegetation and Crops
(ICP Vegetation) was established in 1987, initially with the aim to assess the impacts of air pollutants
on crops, but in later years also on (semi-)natural vegetation. The ICP Vegetation is led by the UK
and has its Programme Coordination Centre at the Centre for Ecology and Hydrology (CEH) in
Bangor. The ICP Vegetation is one of seven ICPs and Task Forces that report to the Working Group
on Effects (WGE) of the Convention on Long-range Transboundary Air Pollution (LRTAP Convention)
on the effects of atmospheric pollutants on different components of the environment (e.g. forests,
fresh waters, materials) and health in Europe and North-America. The Convention provides the
essential framework for controlling and reducing damage to human health and the environment
caused by transboundary air pollution. So far, eight international Protocols have been drafted by the
Convention to deal with major long-range air pollution problems. ICP Vegetation focuses on the
following air pollution problems: quantifying the risks to vegetation posed by ozone pollution and the
atmospheric deposition of heavy metals, nitrogen and persistent organic pollutants (POPs) to
vegetation. The work of the ICP Vegetation contributed significantly to the recent revision of the
Gothenburg Protocol (finalised in May 2012), aiming to abate acidification, eutrophication and ground-
level ozone.
The ICP Vegetation comprises an enthusiastic group of over 200 scientists from 35 countries in the
UNECE region (Table 1.1). In addition, scientists from China, Cuba, Egypt, India, Japan, Pakistan
and South Africa participate as the ICP Vegetation stimulates outreach activities to other regions in
the world and invites scientists in those regions to collaborate with and participate in the programme
activities. The contact details for lead scientists for each group are included in Annex 1. In many
countries, several other scientists (too numerous to mention individually) also contribute to the
biomonitoring programmes, analysis, modelling and data synthesis procedures of the ICP Vegetation.
Table 1.1 Countries participating in the ICP Vegetation; in italics: not a Party to the LRTAP
Convention.
Albania
Austria
Belarus
Belgium
Bulgaria
China
Croatia
Cuba
Czech Republic
Denmark
Egypt
Estonia
Finland
France
FYR of Macedonia
Germany
Greece
Iceland
India
Italy
Japan
Latvia
Lithuania
Montenegro
Netherlands
Norway
Pakistan
Poland
Romania
Russian Federation
Serbia
Slovakia
Slovenia
South Africa
Spain
Sweden
Switzerland
Turkey
Ukraine
United Kingdom
USA
Uzbekistan
1.2 Air pollution problems addressed by the ICP Vegetation
1.2.1 Ozone
Ozone is a naturally occurring chemical present in both the stratosphere (in the ‘ozone layer’, 10 – 40
km above the earth) and the troposphere (0 – 10 km above the earth). Additional photochemical
11
reactions involving NOx, carbon monoxide and non-methane volatile organic compounds (NMVOCs)
released due to anthropogenic emissions (especially from vehicle sources) increase the concentration
of ozone in the troposphere. These emissions have caused a steady rise in the background ozone
concentrations in Europe and the USA since the 1950s (Royal Society, 2008). Superimposed on the
background tropospheric ozone are ozone episodes where elevated ozone concentrations in excess
of 50-60 ppb can last for several days. Ozone episodes can cause short-term responses in plants
such as the development of visible leaf injury (fine bronze or pale yellow specks on the upper surface
of leaves) or reductions in photosynthesis. If episodes are frequent, longer-term responses such as
reductions in growth and yield and early die-back can occur.
The negotiations concerning ozone for the Gothenburg Protocol (1999) were based on exceedance of
a concentration-based critical level of ozone for crops and (semi-)natural vegetation. However, since
then the biologically more relevant stomatal flux-based was developed, estimating the flux of ozone
from the exterior of the leaf through the stomatal pores to the sites of damage (Emberson et al., 2000;
Pleijel et al., 2007). During 2009/10, flux-based critical levels of ozone for vegetation were reviewed at
an LRTAP Convention workshop in Ispra, November 2009 and new/revised flux-based critical levels
were agreed at follow-on discussions at the 23rd ICP Vegetation Task Force meeting, February 2010
(Harmens et al., 2010a; LRTAP Convention, 2010; Mills et al., 2011b). They include policy-relevant
indicators for i) agricultural crops to protect security of food supplies; ii) forest trees to protect against
loss of carbon storage in living trees and loss of other ecosystem services such as soil erosion,
avalanche protection and flood prevention; iii) grassland (productive grasslands and grassland of high
conservation value) to protect against for example loss of vitality and fodder quality. The flux-based
approach is now the preferred method for assessing the risk of ozone impacts on vegetation, as
described in Annex 1 of the revised Gothenburg Protocol. Particulate matter (PM2.5), and thereby also
black carbon as a component of PM2.5, has now been included in the Gothenburg Protocol. The
recently revised Gothenburg Protocol requires that EU member states jointly cut their emissions of
sulphur dioxide by 59%, nitrogen oxides (a precursor of ozone) by 42%, ammonia by 6%, volatile
organic compounds (a precursor of ozone) by 28% and particles by 22% between 2005 and 2020.
Once the national emission reduction obligations have been implemented in 2020, the revised
Protocol is expected to result in significant reductions in human health impacts and adverse impacts
on the environment from air pollution. Despite these emission reductions, air pollution will still pose
significant risk to human health and the environment after 2020.
The ozone sub-group of the ICP Vegetation contributes models, state of knowledge reports and
information to the LRTAP Convention on the impacts of ambient ozone on vegetation; dose-response
relationships for species and vegetation types; ozone fluxes, vegetation characteristics and stomatal
conductance; flux modelling methods and the derivation of critical levels and risk assessment for
policy application.
1.2.2 Heavy metals
Concern over the accumulation of heavy metals in ecosystems and their impacts on the environment
and human health, increased during the 1980s and 1990s. Currently some of the most significant
sources include:
Metals industry (Al, As, Cr, Cu, Fe, Zn);
Other manufacturing industries and construction (As, Cd, Cr, Hg, Ni, Pb);
Electricity and heat production (Cd, Hg, Ni);
Road transportation (Cu and Sb from brake wear, Pb and V from petrol, Zn from tires);
Petroleum refining (Ni, V);
Phosphate fertilisers in agricultural areas (Cd).
The heavy metals cadmium, lead and mercury were targeted in the 1998 Aarhus Protocol as the
environment and human health were expected to be most at risk from adverse effects of these
12
metals. This Protocol is currently under review. Atmospheric deposition of metals has a direct effect
on the contamination of crops used for animal and human consumption (Harmens et al., 2005).
The ICP Vegetation is addressing a short-fall of data on heavy metal deposition to vegetation by
coordinating a well-established programme that monitors the deposition of heavy metals to mosses.
The programme, originally established in 1980 as a Swedish initiative, involves the collection of
naturally-occurring mosses and determination of their heavy metal concentration at five-year intervals.
European surveys have taken place every five years since 1990, with the latest survey having been
conducted in 2010/11. The results of this recent survey will be published in 2013 and will also include
data on nitrogen and POPs concentrations in mosses. Spatial and temporal trends (1990 – 2005) in
the concentrations of heavy metals in mosses across Europe have been described by Harmens et al.
(2008; 2010b). Detailed statistical analysis showed that spatial variation in the cadmium and lead
concentrations in mosses is primarily determined by the atmospheric deposition of these metals,
whereas it is less clear which factor primarily determines the mercury concentration in mosses
(Harmens et al., 2012; Holy et al., 2010; Schröder et al., 2010b).
1.2.3 Nitrogen
In recent decades, concern over the impact of nitrogen on low nutrient ecosystems such as
heathlands, moorlands, blanket bogs and (semi-)natural grassland has increased. The empirical
critical loads for nitrogen were reviewed and revised recently (Bobbink and Hettelingh, 2011;
ECE/EB.AIR/WG.1/2010/14). In 2009, the WGE gathered evidence on the impacts of airborne
nitrogen on the environment and human health with the aim of drawing attention to the current threat
of atmospheric nitrogen deposition to the environment and human health
(ECE/EB.AIR/WG.1/2009/15). Details on the contribution of the ICP Vegetation can be found in
Harmens et al. (2009). Previously, plant communities most likely to be at risk from both enhanced
nitrogen and ozone pollution across Europe were identified (Harmens et al., 2006). In 2005/6, the total
nitrogen concentration in mosses was determined for the first time at almost 3,000 sites to assess the
application of mosses as biomonitors of nitrogen deposition at the European scale (Harmens et al.,
2011b; Schröder et al., 2010a). The European nitrogen in moss survey was repeated in 2010/11.
There are many groups within Europe studying atmospheric nitrogen fluxes and their impact on
vegetation (e.g. Nitrogen in Europe (NinE), ECLAIRE, COST 729). The ICP Vegetation maintains
close links with these groups to provide up-to-date information on the impacts of nitrogen on
vegetation to the WGE of the LRTAP Convention.
1.2.4 Persistent organic pollutants (POPs)
POPs are organic substances that possess toxic and/or carcinogenic characteristics, are degrading
very slowly, bioaccumulate in the food chain and are prone to long-range transboundary atmospheric
transport and deposition. In 1998, the Aarhus Protocol on POPs was adopted and a list of 16
substances was targeted to eliminate any discharges, emissions and losses in the long term. In 2009,
seven new substances were included. In 2001, the Stockholm Convention on POPs was established
as a global treaty under the United Nations Environment Programme (UNEP), and new substances
were added in 2009. Mosses are known to accumulate POPs (Harmens et al., in press) and in the
2010/11 ICP vegetation European moss survey six countries have determined the concentration of
selected POPs (polycyclic aromatic hydrocarbons (PAHs) in particular) in mosses to assess spatial
patterns of POPs deposition to vegetation.
1.3 Workplan items for the ICP Vegetation in 2012
For the first time the Executive Body of the LRTAP Convention agreed on a biannual workplan at its
29th meeting in December 2011 (see ECE/EB.AIR/109/Add.2). Here we will report on the following
workplan items for the ICP Vegetation in 2012:
13
Supporting evidence for ozone impacts on vegetation;
Impacts of ozone on carbon sequestration, including linkages between ozone and climate
change;
Progress with European heavy metals and nitrogen in mosses survey 2010/11;
Relationship between (i) heavy metal and (ii) nitrogen concentrations in mosses and their
impacts on terrestrial ecosystems.
In addition, the ICP Vegetation was requested to report on the following common workplan items of
the WGE:
Further implementation of the Guidelines on Reporting of Monitoring and Modelling of Air
Pollution Effects;
Final version of the impact analysis by the WGE;
Ideas and actions to enhance the involvement of EECCA/SEE countries in the Eastern
Europe, the Caucasus and Central Asia and on cooperation with activities outside the Air
Convention.
The remaining items agreed in the biannual workplan for 2012-2013 will be reported in 2013 (see
Section 6.2).
Progress with most of the above workplan activities is described in Chapter 3. In Chapter 4, the
impacts of ozone on carbon sequestration are described and Chapter 5 provides a review on
available knowledge on the relationship between i) heavy metal and ii) nitrogen concentrations in
mosses and impacts on terrestrial ecosystems. Finally, new activities of the ICP Vegetation are
described in Chapter 6, including the medium-term workplan for 2013 – 2015 (up-dated at the 25th ICP
Vegetation Task Force Meeting, 31 January – 2 February 2012, Brescia, Italy).
14
2 Coordination activities
2.1 Annual Task Force meeting
The Programme Coordination Centre organised the 25th ICP Vegetation Task Force meeting, 31
January – 2 February 2012 in Brescia, Italy, in collaboration with the local host at the Ecophysiology
and Environmental Physics Laboratory - Mathematics and Physics Department, Università Cattolica
del Sacro Cuore. The meeting was attended by 73 experts from 21 countries, including 19 Parties to
the LTRAP Convention and guests from Egypt and South Africa. A book of abstracts, details of
presentations and the minutes of the 25th Task Force meeting are available from the ICP Vegetation
web site (http://icpvegetation.ceh.ac.uk).
The Task Force discussed the progress with the workplan items for 2012 (see Section 1.3) and
updated the medium-term workplan for 2013 - 2015 (see Section 6.2) for the air pollutants ozone,
heavy metals, nutrient nitrogen and POPs. In addition, the ozone expert groups established in 2011
(Harmens et al., 2011a) reported on progress and activities conducted since the 24th ICP Vegetation
Task Force meeting in 2011. To support the reporting on impacts of ozone on biodiversity and
ecosystems services in 2013, an additional temporary expert group was established to review this
theme. The Task Force took note of the conclusions and recommendations from the one-day ozone
workshop on 31st January 2012 in Brescia, Italy.
At the workshop presentations were given and discussions were held on the following two themes:
- Quantifying ozone impacts on Mediterranean Forests;
- Mapping vegetation at risk from ozone at the national scale.
The following conclusions were drawn at the workshop:
New scientific developments support the current text and conclusions in the Modelling and
Mapping Manual, i.e. the flux-based method provides better indicators than the concentration-
based method for ozone impact assessments on vegetation;
Current flux-based critical levels for beech/birch were validated by epidemiological studies on
mature trees in Switzerland.
The following recommendations were made at the workshop:
Further field-based validation of ozone flux-effect relationships and critical levels for
vegetation is required via epidemiological studies;
Validation of DO3SE with eddy covariance flux data would make a valuable contribution;
The ozone flux-based method should be further developed by:
- Expanding the number of species with flux-effect relationships;
- Standardising the method of up-scaling for mature trees;
- Further qualifying and quantifying uncertainties;
- Developing a standard protocol for estimating the maximal stomatal conductance (gmax);
- Including effects of biogenic volatile organic compounds (BVOC);
- Further developing the Ball–Berry photosynthesis model in DO3SE;
- Further stimulating the cooperation with ICP Forests and make use of their available data;
- Adding a new technical annex to chapter 3 of the Modelling and Mapping Manual with flux
parameterisations for additional species.
Some of these recommendations are already being addressed in the European Framework 7 project
‘ECLAIRE’ (Effects of Climate Change on Air Pollution and Response Strategies for European
Ecosystems; http://www.eclaire-fp7.eu/) which includes contributions from several ICP Vegetation
participants and other LRTAP Convention bodies.
At the Task Force meeting, participants of the European moss survey reported on progress with data
analysis and submission of the 2010/11 survey on heavy metals, nitrogen and POPs. They are keen
15
to conduct the next European moss survey in 2015 (depending on available national funds). The chair
reiterated that continued and additional participation from countries in Southern-Eastern Europe
(SEE), Eastern Europe, Caucasus and Central Asia (EECCA), with potential outreach to other parts of
Asia, is highly desirable.
The Task Force acknowledged and encouraged further fruitful collaborations with the bodies and
centres under the Steering Body to EMEP, in particular EMEP/MSC-West, EMEP/MSC-East, the Task
Force on Integrated Assessment Modelling and the Task Force on the Hemispheric Transport of Air
Pollution, and bodies under the Working Group of Strategies and Review, in particular the Task Force
on Reactive Nitrogen. In addition, the Task Force encouraged further development of outreach
activities to other regions in the world (see Section 3.1.3).
Over the years participation in the ICP Vegetation and attendance of the Task Force meetings has
been rising (Figure 2.1). Originally named as the ICP Crops, focussing on the impacts of ozone on
crops, the programme started to incorporate impacts on (semi-)natural vegetation later on and
therefore gained its current name in the mid-1990s. In 2001, the ICP Vegetation took over the
coordination of the European moss survey on heavy metals from the Nordic Council of Ministers and
therefore widened its scope and further enhanced particitation in its activities.
Figure 2.1 Participation in ICP Vegetation Task Force meetings since 1987.
The 26th Task Force meeting will be hosted by IVL - Swedish Environmental Research Institute, and
will be held in Halmstad, Sweden, from 28 - 31 January 2013.
2.2 Reports to the LRTAP Convention
The ICP Vegetation Programme Coordination Centre has reported progress with the 2012 workplan
items in the following documents for the 31st session of the WGE
(http://www.unece.org/index.php?id=24661):
- ECE/EB.AIR/WG.1/2012/3: Joint report of the ICPs, Task Force on Health and Joint Expert
Group on Dynamic Modelling;
- ECE/EB.AIR/WG.1/2012/8: Effects of air pollution on natural vegetation and crops
(technical report from the ICP Vegetation);
- ECE/EB.AIR/WG.1/2012/13: 2012 impact assessment: effects indicators as tools to evaluate air
pollution abatement policies (see Section 3.1.2);
The ICP Vegetation also contributed to Guidance Document VII on health and
environmentalImprovements from the WGE, submitted to the 30th meeting of the Executive Body, 30
April – 4 May 2012 (informal document 3) for the revision of the Gothenburg Protocol.
0
10
20
30
40
0
10
20
30
40
50
60
70
80
1987
1989
1991
1993
1996
1998
2000
2002
2004
2006
2008
2010
2012
No. of countries
No. of participants
Yea r
No. of participants
No. of cou ntri es
16
In addition, the Programme Coordination Centre for the ICP Vegetation has:
- published the final glossy report and a summary brochure on ‘Ozone pollution: A hidden threat
to food security’ (Mills and Harmens, eds, 2011);
- published a glossy report on ‘Ozone pollution: Impacts on carbon sequestration in Europe’
(Harmens and Mills, eds, 2012), see Chapter 4;
- published the current annual report on line;
- contributed to the colour brochure of the WGE on ‘Impacts of air pollution on human health,
ecosystems and cultural heritage’ and facilitated printing of the brochure in three languages
(English, French and Russian) with a contribution in kind from the Swiss Federal Office for the
Environment (FOEN);
- contributed to the final report on ‘Impacts of air pollution on ecosystems, human health and
materials under different Gothenburg Protocol scenarios to support the revision of the
Gothenburg Protocol’ (see also ECE/EB.AIR/WG.1/2012/13);
- published a leaflet on ‘Mosses as biomonitors of atmospheric heavy metal pollution in Europe’
(also available in Russian).
2.3 Scientific papers
The following papers have been published or are in press:
Grünhage, L., Pleijel, H,, Mills, G., Bender, J., Danielsson, H., Lehmann, Y., Castell, J.F., Bethenod, O. (2012).
Updated stomatal flux and flux-effect models for wheat for quantifying effects of ozone on grain yield, grain mass
and protein yield. Environmental Pollution 165: 147-157.
Harmens, H., Norris, D. A., Cooper, D.M., Mills, G., Steinnes E., Kubin, E., Thöni, L., Aboal, J.R., Alber, R.,
Carballeira, A., Coșkun, M., De Temmerman, L., Frolova, M., Frontasyeva, M., Gonzáles-Miqueo,L., Jeran, Z.,
Leblond S., Liiv, S., Maňkovská, B., Pesch, R., Poikolainen, J., Rühling, Å., Santamaria, J. M., Simonèiè, P.,
Schröder, W., Suchara, I., Yurukova, L., Zechmeister, H. G. (2011). Nitrogen concentrations in mosses indicate
the spatial distribution of atmospheric nitrogen deposition in Europe. Environmental Pollution 159: 2852-2860.
Harmens, H., Ilyin, I., Mills, G., Aboal, J.R., Alber, R., Blum, O., Coşkun, M., De Temmerman, L., Fernández,
J.A., Figueira, R., Frontasyeva, M., Godzik, B., Goltsova, N., Jeran, Z., Korzekwa, S., Kubin, E., Kvietkus, K.,
Leblond, S., Liiv, S., Magnússon, S.H., Maňkovská, B., Nikodemus, O., Pesch, R., Poikolainen, J., Radnović, D.,
Rühling, Å., Santamaria, J.M., Schröder, W., Spiric, Z., Stafilov, T., Steinnes, E., Suchara, I., Tabor, G., Thöni, L.,
Turcsányi, G., Yurukova, L., Zechmeister, H.G. (2012). Country-specific correlations across Europe between
modelled atmospheric cadmium and lead deposition and concentrations in mosses. Environmental Pollution 166:
1-9.
Harmens, H., Foan, L., Simon, V., Mills, G. (in press). Terrestrial mosses as biomonitors of atmospheric POPs
pollution: A review. Environmental Pollution.
17
3 Ongoing research activities in 2011/12
In this chapter, progress made with the common WGE and ICP Vegetation workplan for 2012 is
summarised. New ICP Vegetation workplan items in 2012 are described in detail in Chapter 4 and 5.
3.1 Contributions to WGE common workplan items
3.1.1 Further implementation of the Guidelines on Reporting of Air Pollution Effects
Table 3.1 provides an overview of the monitoring and modelling effects reported by the ICP
Vegetation according to the Guidelines (ECE/EB.AIR/2008/11).
Table 3.1 Monitoring and modelling effects reported by the ICP Vegetation.
Parameter Ozone Heavy metals Nitrogen POPs
Growth and yield reduction
Leaf and foliar damage
Exceedance critical levels
Climatic factors
Concentrations in mosses
X
X
X
X
X
X
X
3.1.2 Contributions to the completed version of the impact analysis by the WGE
Background
To support the revision of the Gothenburg Protocol (finalised in May 2012), the WGE has conducted
an analysis on the impacts of air pollution on ecosystems, human health and materials under different
emission scenarios. The objectives of this analysis were to:
Provide information on effects of air pollution on ecosystems, human health and materials to
support decisions for the revision of the Gothenburg Protocol;
Demonstrate application of new science and effects indicators, developed since 1999, to
illustrate the potential impact of policy/decisions on the environment, human health and
materials;
Illustrate effectiveness of emission scenarios to improve the environment and human health.
Here we report on the analysis conducted by the ICP Vegetation regarding the impacts of ozone on
vegetation. This is an update of the interim analysis reported last year (Harmens et al., 2011a). The
update reflects the discussions during the various phases of the negotiations of the Gothenburg
Protocol revision. The updated analysis on the risk of ozone impacts on forests is based on scenarios
published by IIASA in August 2011 (described in CIAM report 4/2011, Amann et al., 2011). The
analysis for the risk of ozone impacts on crops were not updated and are therefore based on
scenarios published by IIASA in October 2010 (described in CIAM report 1/2010, Amann et al., 2010).
The updated scenarios and projections applied for forests were:
- NAT2000: historical data for the year 2000 based mainly on national information;
- COB2020: Cost Optimised Baseline for the year 2020. This dataset is generated assuming
that only current (2011) legislation still applies in 2020 (comparable to NAT2020 described
previously and applied for crops);
- MID2020: assuming a higher ambition level for environmental targets than COB2020;
- MTFR2020: data based on a scenario assuming that all technically feasible technologies are
implemented by 2020.
18
The baseline activity data on energy use, transport, and agricultural activities were issued from
different sources, including national submissions to IIASA and from specialized sectorial energy,
transport and agricultural models (e.g., PRIMES, TREMOVE and CAPRI). They were then used as
input data for the GAINS model with which scenarios were optimised so that emissions control
scenarios would achieve environmental targets for human health and environmental impacts
(acidification, eutrophication, effect of ground-level ozone) as discussed in the 48th session of the
WGSR. MTFR represents the reduction that would be obtained if the most stringent regulations were
implemented. Any decision leading to some emission reduction will lead to a situation between the
baseline and the MTFR scenario.
For the development of the 1999 Gothenburg Protocol, AOT402 was used to indicate the risk to
vegetation of adverse impacts of ozone. Since then, a biologicially more relevant impact indicator has
been developed, the Phytotoxic Ozone Dose above a threshold Y (PODY), which gives a better
correlation between the locations where ozone damage was reported in Europe between 1990 and
2006 and maps of ozone flux (PODY) than maps of AOT40 (Hayes et al., 2007; Mills et al., 2011a).
Recently, new or revised flux-based critical levels were developed for crops (potato, tomato, wheat),
trees (beech/birch, Norway spruce) and white clover as a representative species of grasslands and
(semi-)natural vegetation (Harmens et al., 2010a; LRTAP Convention, 2010; Mills et al., 2011b).
Crop yield and economic losses based on ozone flux-effects indicators
Using the flux-based approach and NAT scenarios, economic losses due to ozone for wheat were
estimated to be 3.2 billion euros in EU27+Switzerland+Norway in 2000 reducing to 1.96 billion euros
in 2020 (Table 3.2). Although the percentage wheat yield reduction is predicted to decline in 2020,
only a very small reduction in the proportion of EMEP grid squares exceeding the critical level is
predicted. Proportional reductions in yield and economic value for tomato, an important crop for
southern areas, were similar to those for wheat for NAT2020 compared to NAT2000. Further details
can be found in a report describing the impacts of ozone on food security (Mills and Harmens, 2011).
Table 3.2 Predicted impacts of ozone pollution on wheat and tomato yield and economic value,
together with critical level exceedance in EU27+Switzerland+Norway in 2000 and 2020 under the
current legislation scenario (NAT scenario). Analysis was conducted on a 50 x 50 km EMEP grid
square using crop values in 2000 and an ozone stomatal flux-based risk assessment.
Crop Wheat Tomato
Emission scenario NAT2000 NAT2020 NAT2000 NAT2020
Economic losses
(billion Euro)
3.20 1.96 1.02 0.63
Percentage of
EMEP grid squares
exceeding critical
level*
84.8 82.2 77.8 51.3
Mean yield loss (%)* 13.7 9.1 9.4 5.7
* Calculated for the grid squares where the crop is grown.
Mapping risk of ozone impacts on forest growth: application of flux-based methodology
Comparison of ozone risk maps for forests applying the different scenarios and projections shows that
despite the predicted reductions in stomatal fluxes in the future, large areas in Europe will remain at
risk from adverse impacts of ozone on forest growth, with areas at highest risk being predicted in
parts of western, central and southern Europe (Figure 3.1). In Figure 3.2, the proportion of grid
squares in each category illustrated on the maps is shown for the four scenarios. Although for the
2020 scenarios there is a decrease in the proportion of grid squares in the highest categories, there
2 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.
19
remains a high proportion of grid squares in the middle to high categories (25 – 26% and 11 – 16% for
a POD1 of 24 – 28 mmol m-2 and 28 – 32 mmol m-2 respectively), indicating a continuing risk of
damage. Hence, additional measures to reduce the emissions of ozone precursors will be required to
protect large areas in Europe from adverse impacts of ozone on forests in 2020.
Figure 3.1. The risk of adverse ozone impacts in on biomass production in forest using the generic
deciduous tree flux model (POD1) for a) 2000 (NAT2000) and 2020: b) COB2020, c) MID2020, and d)
MTFR2020.
For further details we refer to the full impacts report produced by the WGE (see
http://www.unece.org/fileadmin/DAM/env/documents/2012/EB/n_14_Report_WGE.pdf), with a
summary of the results being reported in ECE/EB.AIR/WG.1/2012/13. In collaboration with the other
ICPs and Task Force on Health, the ICP Vegetation Programme Coordination Centre also produced a
glossy brochure on ‘Impacts of air pollution on human health, ecosystems and cultural heritage’,
summarising the results, conclusions and recommendations (see http://icpvegetation.ceh.ac.uk). The
brochure is also available in French and Russian.
(a) (b)
(c) (d)
20
Figure 3.2 Proportion of grid squares within specified categories of POD1 calculated using the generic
deciduous tree flux model as calculated with the different datasets and scenarios.
3.1.3 Ideas and actions to enhance the involvement of EECCA/SEE countries in the
Eastern Europe, the Caucasus and Central Asia and on cooperation with
activities outside the Air Convention
Working with the lead participant of the European moss survey in the Russian Federation, the ICP
Vegetation is actively encouraging the participation of more EECCA/SEE countries. For example,
Albania took part for the first time in the moss survey in 2010/11 and attended the ICP Vegetation
Task Force meeting for the first time in 2012. Every year we try to find funds for experts in
EECCA/SEE countries to participate in our annual Task Force meeting. In 2012, a short leaflet was
produced on the results of the 2005/6 European moss survey, which was translated into Russian for
distribution in EECCA countries. The glossy brochure on ‘Impacts of air pollution on human health,
ecosystems and cultural heritage’ (see Section 3.1.2) was translated into Russian and widely
distributed within the Convention and by contacts in EECCA countries.
Via the Stockholm Environment Institute (SEI) in York (UK), which hosts the secretariat of the Global
Atmospheric Pollution Forum, the ICP Vegetation has continued to encourage collaboration with
South Asia, particularly with Malé Declaration countries. For example, the ICP Vegetation has
produced a position paper on outreach activities with the Malé Declaration, which was presented at
the Third Meeting of the Task Force on Future Development of Malé Declaration on Control and
Prevention of Air Pollution and its Likely Transboundary Effects for South Asia (Malé Declaration),
held on 9-10 August 2012 in Chonburi, Thailand. Suggestions were provided and discussed for the
near-term collaboration (next three years) between the two regional atmospheric pollution
programmes. The Malé Declaration showed a general interest in intensifying the collaboration with the
LRTAP Convention as outlined in the position paper submitted by the ICP Vegetation and SEI, but
expressed the need for funding to make this collaboration happen. Furthermore, the ICP Vegetation
has developed collaboration with experts in Egypt, South Africa, Cuba and Japan, who have attended
recent Task Force meetings of the ICP Vegetation. Further collaboration is often hindered by the lack
of available funds.
3.2 Progress with common ICP Vegetation workplan items
3.2.1 Supporting evidence for ozone impacts on vegetation
Since 2008, participants of the ICP Vegetation have been conducting biomonitoring campaigns using
ozone-sensitive (S156) and ozone-resistant (R123) genotypes of Phaseolus vulgaris (Bush bean,
French Dwarf bean) that had been selected at the USDA-ARS Plant Science Unit field site near
Raleigh, North Carolina, USA. The bean lines were developed from a genetic cross reported by Dick
21
Reinert (described in Reinert and Eason (2000)). Individual sensitive (S) and tolerant (R) lines were
identified, the S156 and R123 lines were selected, and then tested in a bioindicator experiment
reported in Burkey et al. (2005). A trial of this system occurred in central and southern parts of Europe
during the summer of 2008. This was extended in 2009 and included again in the ozone biomonitoring
programme for 2010 and 2011.
In 2011, the biomonitoring of ozone effects using bean was scaled down compared to the previous
two years, reflecting less interest from the participants. Nevertheless, experiments were conducted
with ozone-sensitive and ozone-resistant bean (Phaseolus vulgaris) at nine sites across Europe and
one in the USA. As in previous years, bean seeds of the strains S156 and R123 were kindly provided
by Kent Burkey (USA). Seeds of both varieties and an updated experimental protocol (ICP
Vegetation, 2011) were sent out to participants who recorded the occurrence of visible injury to leaves
and quantified the reduction in pod yield of the sensitive compared to the resistant variety for plants
exposed to ambient ozone. Some participants carried out stomatal conductance measurements to
contribute to the development of a flux-effect model.
The data from the 2011 and previous biomonitoring and ozone exposure experiments conducted in
2008, 2009 and 2010 were combined in a database for dose-response analysis. The database
contains data from Belgium, France, Germany (3 sites), Greece (2 sites), Hungary (2 sites), Italy (3
sites), Japan, Slovenia (2 sites), Spain (3 sites), South Africa, UK (2 sites), Ukraine and the USA.
Over 3000 stomatal conductance measurements have been made on the bean plants and used to
generate an ozone flux model using the Emberson et al. (2000) approach. Over the course of the
bean biomonitoring experiment, hourly accumulated ozone fluxes ranged from 4.4 (Bangor, UK) to
18.9 (Seibersdorf, Austria) mmol m-2. Visible leaf injury regularly occurred across the network (Figure
3.3), but there was no clear dose-response relationship with concentrarion-based ozone parameters.
Similarly, there was no clear relationship between concentration-based parameters and the ratio of
the pod weight for the sensitive to that of the resistant bean. Flux-effect relationships will be explored
in the coming year.
Figure 3.3 Locations where ozone injury has been detected on bean between 2008 and 2011.
Overall, the bean biomonitoring system does seem to provide a good indication of the occurrence of
ozone concentrations that are high enough to visibly damage plants. As such it is very valuable for
use in countries as proof or otherwise that ozone levels are causing damage. However, we are
concerned that differences between the sensitive and resistant biotypes are not strong enough for
continued application as a biomonitor for yield effects across all climate regions in Europe.
Japan
USA
22
3.2.2 Progress with European heavy metals and nitrogen in mosses survey 2010/11
The European moss biomonitoring network was originally established in 1990 to estimate
atmospheric heavy metal deposition at the European scale. The network provides a time-integrated
measure of heavy metal and potentially nitrogen deposition from the atmosphere to terrestrial
ecosystems (Harmens et al., 2010b; 2011b). It is easier and cheaper than conventional precipitation
analysis as it avoids the need for deploying large numbers of precipitation collectors with an
associated long-term programme of routine sample collection and analysis. Therefore, a much higher
sampling density can be achieved than with conventional precipitation analysis.
Mosses have been collected for element analysis every five years since 1990 and the most recent
survey was conducted in 2010/11. A total of 26 countries will submit or have already submitted data
on heavy metals, of which 14 countries will also submit (or have submitted) data on nitrogen
concentrations in mosses (Table 3.3). As a pilot study, six countries have agreed to submit data on
POPs concentrations in mosses. A recent review has shown that POPs concentrations in mosses can
also be a useful indicator for the atmospheric deposition of selected organic compounds such as
polycyclic aromatic hydrocarbons (PAHs) and polychlorinated biphenyls (PCBs) (Harmens et al.,
2011a, in press). In addition, some countries will determine the sulphur concentration in mosses.
Table 3.3 Countries (regions) participating in the European moss survey 2010/11. All twenty six
countries will report on heavy metals, 14 countries will report on nitrogen and six countries will report
on selected POPs concentration in mosses.
Country N POPs
Albania
Austria
Belarus
Belgium
Bulgaria
Croatia
Czech Republic
Denmark (Faroe Islands)
Estonia
Finland
France  
FYR of Macedonia
Iceland
Italy (Bolzano)
Kosovo*
Norway
Poland  
Romania
Russian Federation
Serbia
Slovakia
Slovenia  
Spain (Galicia, Navarra, Rioja) (Navarra, Rioja) (Navarra)
Sweden
Switzerland  
Ukraine (Donetsk)
* For this study considered as a separate region.
23
4 Impacts of ozone on carbon sequestration and
linkages between ozone and climate change
4.1 Introduction
4.1.1 Background
Since the industrial revolution, concentrations of carbon dioxide (CO2) in the atmosphere have been
rising, initially slowly but in recent decades more rapidly (IPPC, 2007). This is primarily due to an
increase in fossil fuel burning associated with population growth and enhanced social and economic
development. In recent decades deforestation, especially in the tropics, has also contributed
considerably to the rise in atmospheric CO2 as tropical forests are a major sink for CO2. The rise in
atmospheric CO2 concentrations has resulted in a rise in the surface temperature of the earth (global
warming). In addition to CO2 increasing, the atmospheric concentrations of other gases contributing to
global warming (greenhouse gases) such as nitrous oxide, methane, halocarbons and ozone have
risen too. Depending on future scenarios, the earth’s surface temperature is predicted to rise by a
further ca. 2 – 4oC by the end of the 21st century. Currently, ozone is considered to be the third most
important greenhouse gas, after CO2 and methane (IPPC, 2007). In contrast to CO2 and halocarbons,
ozone is a short-lived greenhouse gas, so any reductions in ground-level ozone production will reduce
atmospheric ozone concentrations within months and hence reduce its contribution to global warming.
Long-lived greenhouse gases will stay in the atmosphere for a long time, so even when emissions are
kept constant at the 2000 level, a further rise in surface temperature of 0.5oC is predicted by the end
of the 21st century. We have investigated how ozone pollution is currently, and likely in the future to
continue to be, reducing carbon (C) sequestration in the living biomass of trees (and other
vegetation), thereby potentially exacerbating global warming. Here we provide a summary of this
study, further details can be found in Harmens and Mills (2012).
4.1.2 Ozone pollution
As well as being a greenhouse gas, ozone is also an important atmospheric pollutant and has
adverse effects on human health and the environment. Ozone is a naturally occurring chemical that
can be found in both the stratosphere (as the so-called "ozone layer", 10 - 40 km above the Earth)
and the troposphere (at “ground level”, 0 - 10 km above the Earth). At ground level there is always a
background concentration of ozone resulting from natural sources of the precursors and stratospheric
incursions. Of concern for human health and vegetation (including C sequestration and food
production) is the additional tropospheric ozone which is formed from complex photochemical
reactions from fossil fuel burning in industrial and transport activities. As a result of these emissions,
there has been a steady rise in the background ozone concentration in Europe since the 1950s to the
current 30 – 40 ppb (Royal Society, 2008). Background ozone concentrations in Europe are still rising
and predicted to rise until at least 2030, in part due to hemispheric transport of the precursors of
ozone from other parts (developed and developing areas) of the world. Background concentrations in
Europe have now reached levels where they have adverse impacts on vegetation. During periods of
hot dry weather and stable air pressure, ozone episodes occur where concentrations rise above 60
ppb for several days at a time.
4.1.3 Vegetation as a sink for atmospheric CO2 and ozone
Atmospheric gases such as CO2, ozone and water vapour are exchanged through microscopic
stomatal pores on leaves. This for instance enables plants to fix CO2 for photosynthesis and hence
growth, and to transpire for the adjustment of the internal water balance. The more open the stomata
are, the more CO2 and ozone will enter the plant and the more water will transpire. Ozone entering
the plant has the potential to damage plant cells by forming reactive oxygen species, which can lead
to detrimental effects on photosynthesis and growth and/or ultimately to cell death. The magnitude of
these damaging effects depends on the plant species and genotype, concentration of ozone, duration
24
of exposure, climate and soil conditions. Plants are able to detoxify a certain amount of ozone, but
above this amount damage to vegetation is likely to occur, either as acute damage due to exposure to
‘high’ ozone concentrations that usually occur during ozone episodes or as chronic damage due to
prolonged exposure to elevated background ozone concentrations. Hence, terrestrial vegetation is
considered an important sink for the greenhouse gases CO2 and ozone. However, if ozone
concentrations are high enough to reduce photosynthesis (i.e. CO2 fixation) and/or above-ground
plant growth, then less CO2 and ozone will be taken up by the vegetation, leading to a positive
feedback to atmospheric CO2 and ozone concentrations and therefore global warming (Sitch et al.,
2007).
4.1.4 Ozone impacts in a changing climate
The future impacts of ozone on C sequestration in European terrestrial ecosystems will depend on the
interaction with and magnitude of the change of the physical and pollution climate, represented by
rising temperatures, increased drought frequency, enhanced atmospheric CO2 concentration and
reduced nitrogen deposition. Ecosystems are inherently complex, and for any one aspect of
functioning, there are multitudes of driving factors. Exposure studies on the interaction between ozone
and other pollutants (nitrogen) and climate change often show the following:
Elevated CO2 concentrations – Elevated ozone and CO2 often affect plant physiology and soil
processes in opposite directions. Hence, the overall response and resulting impact on C
sequestration might well be cancelled out when both gases are enriched in the atmosphere
(Harmens and Mills, 2012).
Warming – A rise in temperature stimulates ozone formation and directly affects the stomatal
uptake of ozone since this process is temperature dependent. Warming can also indirectly
affect the uptake of ozone via impacts on relative humidity, plant development and soil water
availability, all of which influence the stomatal gas exchange (Emberson et al., 2000). Some
studies have shown that atmospheric ozone concentrations modify the response of plant
species and genotypes to warming (Kasurinen et al., 2012).
Enhanced drought – It has often been postulated that drought will protect vegetation from
ozone damage as the stomatal pores shut down more during periods of drought to prevent
water loss. However, the interactions between ozone and drought (mediated via plant
hormones) are more complex than first thought and drought might not protect ozone sensitive
species from adverse impacts of ozone (Mills et al., 2009; Wilkinson and Davies, 2009, 2010).
Nitrogen deposition – Relatively few studies have investigated the impacts of both ozone and
nitrogen on vegetation. Evidence suggests that ozone and nitrogen can have both synergistic
and antagonistic effects on species and ecosystem processes, and that they may interact in
unpredictable ways to affect plant communities (Harmens et al., 2006).
Relatively few studies have investigated the interactive impacts of two or more drivers of change. The
outcome of such studies often indicates complex interactions and non-linearity in responses. There is
an urgent need for more field-based, larger scale experiments where vegetation is exposed to
multiple drivers of climate change for several years (at least one decade) to further investigate the
overall impact of a combination of drivers of change on terrestrial ecosystems. Modelling studies to
predict future impacts of change should also be based on a multifactorial approach. So far, the
impacts of ozone on vegetation and feedbacks to the climate have hardly been considered in global
climate modelling. Recent modelling studies have shown that the indirect impact of ozone on global
warming via its impacts on vegetation might be contributing as much to global warming as its direct
effect as a greenhouse gas (Sitch et al., 2007).
25
4.2 Impacts of ozone on C sequestration in the living biomass of trees
4.2.1 First flux-based assessment for Europe for the current (2000) and future
climate (2040)
The DO3SE (Deposition of Ozone for Stomatal Exchange) model (Emberson et al., 2000) was applied
by Büker, Emberson and colleagues (SEI-York) to estimate the magnitude of the impact of ambient
ozone on C storage in the living biomass of trees. The Phytotoxic Ozone Dose above a threshold
value of Y nmol m-2 s-1 (PODY) was calculated applying known flux-effect relationships for various tree
species (LRTAP Convention, 2010; Mills et al., 2011b).
The following input data were used (Harmens and Mills, 2012):
i) Ozone and meteorological data provided by EMEP for the year 2000 (Simpson, pers.
comm.), and ii) ozone and climate data provided by the Rossby Centre regional Atmospheric
climate model (RCA3) for current (2000) and future (2040) years (Kjellström et al., 2005).
Land cover data to identify the distribution of forest tree species: i) for EMEP data the
species-specific JRC land cover data (http://forest.jrc.ec.europa.eu/distribution) and for ii)
RCA data the UNECE Long-Range Transboundary Air Pollution (LRTAP) Convention
harmonised land cover data were used (Cinderby et al., 2007).
Forest C stock data were derived from the European forests inventory dataset (Forest
Europe, 2011).
In addition, for the year 2000 using EMEP ozone and meteorological data, the application of generic
parameterisations for trees in DO3SE (POD1) were compared with the application of climate region
specific parameterisations (Emberson et al., 2007), with a mixture of POD1 and POD1.6, (Karlsson et
al., 2007) and a deactivated soil moisture deficit (SMD; Büker et al., 2011) module (POD1), i.e. no
limitation of soil moisture on stomatal conductance and hence ozone flux (no influence of drought).
Reductions in C stock due to ozone were calculated from the potential C stock present in both years
should there have been no impact of ozone on the C stock. This calculation assumes that the trees
were exposed to the same ozone flux as in 2000 or 2040 during the build-up of the C stock.
The 2040 scenario runs used the GEA-LOW-CLE emissions generated by IIASA for the year 2050
(http://cityzen-project.eu), together with RCA meteorology for 2040-2049. Thus, both emissions and
meteorology were changed. The GEA-LOW-CLE emission scenario is based on the illustrative
scenario of the GEA Efficiency pathway group in terms of energy demand and use, and the
implementation of a stringent climate policy corresponding to a maximum of 2 oC rise in global
temperature target. In addition, this scenario assumes global implementation of extremely stringent
pollution policies (SLE) until 2030. These stringent air quality control strategies are much more
ambitious than the currently planned legislations, but are still lower than the so called Maximum
Feasible Reduction (MFR) which describes the technological frontier in terms of possible air quality
control strategies by 2030.
Table 4.1 Estimated percentage reduction of C storage in the living biomass of trees due to ozone in
2000 and 2040 in the EU27 + Norway + Switzerland. SMDoff = soil moisture deficit module switched
off; PODY = Phytotoxic Ozone Dose above a threshold value of Y nmol m-2 s
-1. The reduction is
calculated from the potential C stock present in both years should there have been no impact of
ozone on the C stock.
Modelled
input data
Yea
r
Parameterisation
DO3SE model
PODY
Reduction C
storage (%)
EMEP 2000 Generic POD1 12.0
2000 Climate region-specific POD1/1.6 13.7
2000 SMDoff POD1 17.3
RCA3 2000 Generic POD1 16.2
2040 Generic POD1 12.6
26
Figure 4.1 PODY in 2000 calculated from EMEP input data and applying the following
parameterisations in DO3SE: (a) generic parameterisation (Y = 1 nmol m-2 PLA s-1), (b) climate region
specific parameterisation (Y is a mixture of 1 and 1.6 nmol m-2 PLA s-1), and (c) generic
parameterisation with soil moisture module switched off (i.e. no soil water limitations). For
comparison, (d) AOT40 in 2000 is also shown (Harmens and Mills, 2012).
Figure 4.2 PODY in (a) 2000 and (b) 2040, calculated from RCA input data and applying the generic
parameterisation in DO3SE (Y = 1 nmol m-2 PLA s-1; Harmens and Mills, 2012).
(b)
(a) (b)
(b) (a)
(a) (b)
(c) (d)
27
Figure 4.3 Absolute reduction (Mt C) in C storage in the living biomass of trees due to ozone in 2000,
applying PODY calculated from EMEP input data and applying the following parameterisations in
DO3SE: (a) generic parameterisation (Y = 1 nmol m-2 PLA s-1), (b) climate region specific
parameterisation (Y is a mixture of 1 and 1.6 nmol m-2 PLA s-1), and (c) generic parameterisation with
soil moisture module switched off (i.e. no soil water limitations). The reduction is calculated from the
potential C stock present in both years should there have been no impact of ozone on the C stock
(Harmens and Mills, 2012).
Figure 4.4 Absolute reduction (Mt C) in C storage in the living biomass of trees due to ozone applying
PODY in (a) 2000 and (b) 2040, calculated from RCA input data and applying the generic
parameterisation in DO3SE (Y = 1 nmol m-2 PLA s-1). The reduction is calculated from the potential C
stock present in both years should there have been no impact of ozone on the C stock (Harmens and
Mills, 2012).
(a) (b)
(c)
(a) (b)
28
The main results are (Table 4.1, Figures 4.1 - 4.4):
When applying the flux-based methodology and a generic parameterisation for deciduous and
conifer trees, a reduction of C sequestration in the living biomass of trees by 12.0 (EMEP
input data) to 16.2% (RCA input data) was calculated. The flux-based approach indicates a
high risk of ozone impacts on forests in Atlantic and Continental Central Europe, and also a
considerable risk in southern parts of northern Europe (in comparison with the concentration
based approach).
The climate-region specific parameterisation for 2000 revealed slightly higher C reductions
(13.7%) due to ozone compared to the generic parameterisation (12.0%) for calculating
PODY.
The deactivation of the soil moisture deficit module of the DO3SE model, which simulates
drought-free stomatal ozone uptake conditions throughout Europe, led to an increase in C
reduction, especially in the warmer and drier climates in Central and Mediterranean Europe.
Although a decline in stomatal ozone flux was predicted in 2040, C sequestration in the living
biomass of trees will still be reduced by 12.6% (compared to 16.2% in 2000). The decline in
stomatal ozone flux in 2040 is mainly a result of a predicted reduction in atmospheric ozone
concentrations across Europe.
Whilst the spatial patterns and temporal trends indicated above can be postulated with a considerable
degree of certainty, the absolute values of C reductions have to be interpreted with care. It should be
remembered that these are for effects on annual increment in living tree biomass only, and do not
take into account any effect on soil C processes, including any direct or indirect ozone effects on
below-ground processes that affect the rate of C turnover in the soil. Furthermore, the response
functions used were derived for young trees (up to 10 years of age). However, there is scientific
evidence from epidemiological studies that the functions are applicable to mature trees within forests
too (Braun et al., 2010).
4.2.2 Case study in northern and central Europe applying the AOT40 method
A more detailed study based on relative growth rates of trees was conducted by Karlsson (IVL,
Sweden) to assess the impacts of current ambient atmospheric ozone concentration (in comparison
to pre-industrial ozone levels in the range of 10 -15 ppb, i.e. AOT40 = 0) on C sequestration in the
living biomass of trees in temperate and boreal forests (Harmens and Mills, 2012). As exposure-
response relationships based on AOT40 are more commonly reported in the literature, the AOT40
method was applied here. The AOT40 values per country used here were annual means for the
growing season of trees for the period 2000-2005 and were provided by EMEP. However, it should be
noted that the AOT40 approach might underestimate the risk of ozone impacts on vegetation in
northern European countries in particular (e.g. Hayes et al., 2007; Mills et al., 2011a; see also above).
Using data from forests inventories on forest types, age classes and structure, growth and harvest
rates and combining these with AOT40-based dose response relationships for young trees, calculated
yearly growth increment values were converted to C stock changes. The estimated percent reduction
in the change of the living biomass C stock across forests in ten countries was 10%. However, for
different countries these values ranged between 2 and 32% (Table 4.2).
The most important factor that determines the changes in the forest living biomass C stock is the gap
between growth and harvest rates. If this gap is small, then a certain growth reduction caused by
ozone will have a relatively large impact on the C stock change, and vice versa. By far the most
important countries for C sequestration in the living biomass C stocks in northern and central Europe
are Sweden, Finland, Poland and Germany. Ozone-induced growth reductions will also result in an
economic loss for forest owners.
29
Table 4.2 Estimated reductions in annual C sequestration due to current ambient ozone exposure as
compared to pre-industrial ozone levels in northern and central Europe.
Country Decline (%) Country Decline (%) Country Decline (%)
Czech Republic 32.0 Finland 2.2 Lithuania 13.8
Denmark 5.8 Germany 12.3 Norway 1.8
Estonia 4.5 Latvia 8.8 Poland 12.8
All countries 9.8 Sweden 8.6
4.2.3 A global perspective of impacts on C storage in terrestrial ecosystems
The JULES (Joint UK Land Environment Simulator) model has been run with ozone fields and
observed climatology over the period 1901-2040 to assess the impacts of ozone on the global C and
water cycle (Sitch et al., 2007). In JULES, the plant damage due to ozone directly reduces plant
photosynthesis, and thereby indirectly, leaf stomatal conductance. With elevated near surface ozone
levels, the model simulates decreased plant productivity, and as less CO2 is required for
photosynthesis, reduced stomatal conductance. Therefore, the plant is able to preserve water
supplies. However, some recent studies have shown that ozone impairs stomatal functioning such
that ozone might enhance rather than reduce stomatal conductance (Mills et al., 2009; Wilkinson and
Davies, 2009; 2010). As no direct effect of ozone on stomatal functioning is currently incorporated into
JULES, the indirect effect of ozone on stomata via photosynthesis was switched off (‘fixed stomata’)
in the current study to investigate the consequences for the global C and water cycle. In JULES, the
ozone flux-based method was applied (Sitch et al., 2007).
Table 4.3 Simulated future percentages changes (% Δ) in carbon (C) and water cycle (runoff)
variables globally for three time periods: 1901-2040, 1901-2000 and 2000-2040. GPP = Gross
Primary Productivity, Veg = vegetation, Gs = stomatal conductance (Scenario: SRES A2).
1901-2040 % Δ GPP % Δ VegC % Δ SoilC % Δ TotalC % Δ Runof
f
% Δ Gs
Control -15.4 -10.9 -9.7 -10.0 12.6 -13.3
Fixed
stomata
-17.9 -11.8 -10.5 -10.9 1.4 -1.6
2000-2040
Control -6.9 -5.0 -4.1 -4.4 4.5 -5.0
Fixed
stomata
-8.1 -5.5 -4.6 -4.8 0.6 -0.5
1901-2000
Control -9.2 -6.2 -5.8 -5.9 7.7 -8.7
Fixed
stomata
-10.7 -6.7 -6.2 -6.4 0.8 -1.1
Applying ozone stomatal flux response relationships in JULES, the model predicted that the reduction
in C stored in vegetation is 6.2% globally and almost 4% in Europe in 2000 compared to 1900, and is
predicted to rise to 10.9% globally and ca. 5 to 6% in Europe by 2040 (Table 4.3) due to a predicted
rise in atmospheric ozone concentrations in the future emission scenario applied. As expected, results
from the control run suggest a large indirect effect of ozone (via photosynthesis) on stomatal
conductance and runoff. Unsurprisingly, stomatal conductance and river runoff changed little through
time in the fixed stomata simulation, where the indirect effect of ozone on stomata via photosynthesis
was switched off. However, despite the difference in stomatal conductance response between
simulations, the differences in the response of the C cycle are rather modest. It can be concluded that
in the absence of a direct effect of ozone on stomatal conductance, ozone-vegetation impacts act to
increase river runoff and freshwater availability substantially due to a reduced water loss from soil via
transpiration from vegetation. However, such an increase might not occur if ozone has adverse
impacts on stomatal functioning, reducing their responsiveness to environmental stimuli.
In addition, Sitch and colleagues analysed the impacts of ozone and drought interactions on plant
productivity in Europe by applying the climate of the year 2003, which was a very dry year across the
30
whole of Europe. Large reductions in plant productivity were simulated under drought conditions. The
net impact of ozone is to further reduce plant productivity under drought. In the absence of a direct
effect of ozone on stomatal conductance, ozone acts to partially offset drought effects on vegetation
(Harmens and Mills, 2012).
4.2.4 Recommendations
Policy More stringent reductions of the emissions of precursors of ozone are required across the
globe to further reduce both peak and background concentrations of ozone and hence reduce the
threat from ozone pollution to C sequestration. It would be of benefit to better integrate policies and
abatement measures aimed at reducing air pollution and climate change as both will affect C
sequestration in the future. Improved quantification of impacts of ozone within the context of climate
change is urgently required to facilitate improved future predictions of the impacts of ozone on C
storage in the living biomass of trees. Stringent abatement policies aimed at short-lived climate
forcers such as ozone provide an almost immediate benefit for their contribution to global warming.
Research There is an urgent need for more field-based, larger scale experiments where vegetation is
exposed to multiple drivers of climate change for several years (at least one decade) to further
investigate the overall impact of a combination of drivers of change on C sequestration in terrestrial
ecosystems. Further development of the ozone flux-based method and establishment of robust flux-
effect relationships are required for additional tree species, in particular for those species representing
the Mediterranean areas. Field-based ozone experiments should also include the assessment of
ozone impacts on below-ground processes and soil C content. Further epidemiological studies on
mature forest stands are required for the validation of existing and new ozone flux-effect relationships.
Experiments are needed on the interacting effects of climate change and ozone, including quantifying
impacts of reduced soil moisture availability, rising temperature and incidences of heat stress, impacts
of rising CO2 concentrations and declining nitrogen deposition. Impacts of other drivers of change on
existing flux-effect relationships should be investigated. Further development of climate region-
specific parameterisations for flux models is needed to improve the accuracy of predictions. Existing
flux models (e.g. DO3SE) will have to be further developed to include more mechanistic approaches
for the accurate prediction of combined effects of ozone, other pollutants and climate change, on
various plant physiological processes and hence C sequestration.
There is an urgent need to further include ozone as a driver of change in global climate change
modelling to quantify its impact (either directly or indirectly via impacts on vegetation) on global
warming. Such modelling should further investigate the mechanisms of interactions between ozone
and other drivers of global warming. Finally, there is a need to quantify the economic impacts of
ozone on forest growth in order to establish the economic consequences for the wood industry. For
enhanced C storage in the living biomass in the future, the ozone-sensitivity of tree species and
varieties should be considered as a factor in future breeding and forests management programmes.
31
5 Relationship between (i) heavy metal and (ii)
nitrogen concentrations in mosses and their
impacts on ecosystems
5.1 Heavy metals
Some heavy metals (Co, Cu, Fe, Mo, Mn, Ni, Zn) play an essential role in cell metabolism, whereas
others (e.g. Cd, Hg, Pb) are not known to be essential for life. Essential heavy metals are needed in
small quantities only (micronutrients), and both essential and non-essential heavy metals are
potentially toxic when they are available in excess for uptake by organisms in the environment (e.g.
Woolhouse, 1983; Sánchez, 2008; Boyd, 2010). Most likely, the ability of heavy metal ions to bind
strongly to O, N and S atoms is the basis of their toxicity (Borovok, 1990). It has been shown that
metals can modify chemical communication between individuals, resulting in ‘info-disruption’ that can
affect animal behaviour and social structure and hence intraspecies and interspecies interactions,
however ‘info-disruption’ by metals in terrestrial habitats is not well studied (Boyd, 2010). The
atmospheric deposition of heavy metals increased since the industrial revolution, but in recent
decades the deposition of heavy metals in Europe has began to decline (Travnikov et al., 2012) due
to the use of cleaner fuel in combustion processes (e.g. less coal and more gas, unleaded petrol),
implementation of air pollution abatement policies (applying the latest technology to filter emissions at
the source) and closing down many heavy polluting sources in parts of Europe.
Terrestrial mosses primarily receive heavy metals from atmospheric deposition as they lack a root
system. In the last three decades, mosses have been applied successfully as biomonitors of heavy
metal deposition across Europe (Harmens et al., 2008; 2010b). Detailed literature reviews on the
application of mosses as biomonitors of heavy metals have been conducted by Burton (1990), Tyler
(1990), Onianwa (2001) and Zechmeister et al. (2003). Although the heavy metal concentration in
mosses provides no direct quantitative measurement of deposition, this information can be derived by
using regression or correlation approaches relating the measured heavy metal concentrations in
mosses to deposition data (e.g. Berg and Steinnes, 1997; Berg et al., 2003; Zechmeister et al., 2004).
Recently, Bouquete et al. (2011) recommended that the results of moss biomonitoring studies should
be regarded as qualitative or semi-qualitative, rather than attempting to provide absolute data, which
may not be temporally representative, and may have a high degree of uncertainty associated with
them.
Here we discuss whether there is any field-based evidence for a relationship between heavy metal
concentrations in terrestrial mosses and impacts of heavy metals on terrestrial ecosystems. Many
studies have demonstrated that the highest metal concentrations in mosses are often found within
500–2000 meters of emission sources, showing a significant decreasing gradient over this distance
(Türkan et al., 1995; Fernández et al., 2000, 2007; Salemaa et al., 2004; Santameria et al., 2010),
although values higher than background levels have been obtained at distances of over 20 km from
the industry (Zechmeister et al., 2004; Schintu et al., 2005). Heavy metal deposition near roads
generally declines to background levels within 250 m distance from the road, however elevated
deposition can be observed up to 1000 m from very busy roads (Zechmeister et al., 2005).
In the scientific literature there is a lack of a direct comparison between heavy metal concentrations in
mosses and impacts on ecosystems. Impacts of heavy metals on trerrestrial ecosystems are often
most pronounced in areas close to pollution sources (such as heavy metal industry and mines), with
impacts declining with distance from the pollution source. Similarly, heavy metal concentrations in
mosses tend to decline in a gradient away from pollution sources (Zechmeister et al., 2003), often
exponentially (Zechmeister et al., 2004). Some studies have made an indirect comparison, for
example, Santamaria et al. (2012) reported that the declining gradient of heavy metal concentrations
in mosses away from a pollution source (González-Miqueo et al., 2010) coincided with an increase of
the abundance of soil mesofauna.
32
The European moss survey aims to provide an indication of the deposition of heavy metals away from
pollution sources, primarily in rural areas, and the contribution of long-range transport to heavy metal
deposition to vegetation. In agreement with the decline in the annual deposition of heavy metals in
recent decades across Europe (Travnikov et al., 2012), the heavy metal concentration in mosses has
also declined (Harmens et al., 2010b). Heavy metal concentrations in mosses tend to reflect the
accumulated heavy metal deposition over a growing period of two to three years. Hence, they provide
no indication of historical accumulation of heavy metals in soils over a longer period. However,
temporal trends can be determined by repeated sampling of mosses in time (Harmens et al., 2010b).
Although deposition of heavy metals to above-ground plant parts can lead to uptake via the leaves
(Harmens et al., 2005), the risk of heavy metal toxicity to terrestrial ecosystems is often expressed as
a function of the free metal ion concentration in soil solution. The LRTAP Convention has developed
the critical loads approach based on established critical limits of heavy metals in soil solution (UBA,
2004). These critical limits are based on no-observed effect concentration (NOEC), often determined
for single metals in standardised laboratory conditions for specific indicator species of toxicity. Little is
known about the toxicity of metal mixtures in soil solutions and hence the NOEC for metal mixtures.
Exceedance of the critical loads provide an indication of the risk of adverse impacts of heavy metals
on terrestrial ecosystems. Hettelingh and Sliggers (2006) and the Task Force on Heavy Metals (2006)
concluded that available information on the metals chromium, nickel, copper, zinc, arsenic and
selenemium suggests that none of these metals achieve high enough concentrations as a result of
long-range atmospheric transport and deposition to cause adverse effects on terrestrial ecosystems.
However, although the area of exceedance of the critical loads for these heavy metals is small, even
small exceedances may result in effects in the future due to the accumulative nature of heavy metals
in soils. These results support the focus of the 1998 Aarhus Protocol on Heavy Metals on the metals
cadmium, lead and mercury.
In 2000, the European ecosystem area at risk of adverse impacts of cadmium, lead and mercury was
estimated to be <1, 42 and 77% respectively (Hettelingh and Sliggers, 2006) whereas in 2010 it was
estimated to be <1, 15 and 71% (Slootweg et al., 2010). However, hardly any field-based evidence is
available to validate the crital load exceedance calculations for terrestrial ecosystems. Tipping et al.
(2010) suggested that the critical loads calculations for mercury might overestimate the level of critical
load exceedance. In the UK, there was almost no exceedance of the critical load for mercury in 2010.
This is a vast improvement from the area of the critical load exceedance in 1970, which was 13% for
mercury in rural areas of the UK. Although mercury concentrations in mosses were determined at
sites across Europe, data from less sites and countries is available for mercury than for cadmium and
lead. Schröder et al. (2010b) showed that the correlation between metal deposition rates and
concentrations in mosses is rather weak for mercury and considerably lower than for cadmium and
lead, indicating that mosses might not be that suitable as biomonitors of mercury deposition or air
concentrations. This might be related to the specific chemistry of mercury pollution (Harmens et al.,
2010b).
In 2000 and 2010, the highest areas of critical load exceedance for cadmium were estimated in
Bulgaria and Macedonia (Hettelingh and Sliggers, 2006). Although these countries also have high
levels of cadmium concentrations in mosses (Harmens et al., 2010b), high levels in mosses were also
observerd in other countries such as Belgium and Slovakia, where hardly any critical load
exceedance was estimated. For lead the highest areas of critical load exceedance was calculated for
the European part of the Russian Federation in 2000 and 2010, however, data on the lead
concentration in mosses is scarce for this region.
One should bear in mind that ecosystems are exposed to different stressors and that it is difficult to
disentangle impacts of single stressors in the field. De Zwart et al. (2010) made a first attempt to
estimate the loss of species due to cadmium and lead depositions in Europe. One of the endpoints for
the critical loads of cadmium and lead is the ecotoxicological effect of metal ions in soil solution on
soil micro-organisms, plant and invertebrates. Depositions will (eventually) result in a concentration in
soil solution in equilibrium with each other, depending on ecosystem properties like leaching, uptake
33
and soil characteristics such as pH, organic matter and clay content. Toxicity data for soil dwelling
organisms and terrestrial plants are comparatively scarce. Hence, De Zwart et al. (2010) applied
publically available data on acute median lethal or effective concentrations (LC50 or EC50) based on
aquatic toxicity tests. There is no indication that the sensitivity of organisms living in the soil is
intrinsically different from the sensitivity of organisms living in surface waters, provided that the
evaluation is based on the truly bioavailable fraction of the metals. Based on this approach, De Zwart
et al. (2010) concluded that toxicity effects of cadmium and lead are close to zero in the vast majority
of ecosystems across Europe. There is also little evidence of adverse effects at current levels of metal
deposition on vegetation in the UK (RoTAP, 2012).
In summary, according to current knowledge, the relatively low levels of heavy metals in mosses
(compared to previous decades) due to long-range transport are unlikely to indicate any adverse
impact of heavy metals on ecosystems. A straightforward relationship between heavy metal
concentrations in mosses and calculated critical load exceedances is not to be expected as the heavy
metal concentration in mosses reflect atmospheric depositon of heavy metals whereas critical load
exceedances for soil solution is not only determined by heavy metal deposition but also affected by
soil characteristics. It should be noted that despite a general decline in heavy metal deposition across
Europe in recent decades, metals accumulate in soils and might therefore become a problem in the
future if bio-available concentrations reach critical limits in soil solution (UBA, 2004). The risk,
estimated by the use of critical limits, is more important for assessing current threats from heavy
metals to biota than critical loads, which are relevant at ‘steady state’, and which may not be achieved
for centuries. Changes in soil composition as a result of changes in climate, or mechanical
disturbance, may release the stored material in a bioavailable form, and this is one of the largest
uncertainties when considering the impact of future climate on heavy metals in the environment
(RoTAP, 2012).
5.2 Nitrogen
Nitrogen is a macronutrient and essential for the growth of the majority of living organisms. However,
species differ in their requirement for nitrogen intake for a healthy growth: some species have
adapted to living in a low nitrogen environment, whereas others have adapted to living in a high
nitrogen environment. In response to the rising demand for food and energy, increasing
anthropogenic nitrogen emissions have resulted in atmospheric nitrogen deposition becoming an
important and dominant source of nitrogen for some ecosystems (Erisman et al., 1998; Galloway et
al., 2008; Sutton et al., 2011). Global anthropogenic nitrogen depositions are now around the same
order of magnitude as nitrogen input from natural sources, leading to a more than doubling of the
nitrogen pool available to terrestrial organisms in less than a century (Vitousek et al., 1997). Reactive
nitrogen compounds are mainly present in the atmosphere in oxidised or reduced forms. The main
anthropogenic sources for oxidised forms of nitrogen are combustion processes in transport, industry
and energy production, estimated to contribute up to 70% of oxidised nitrogen emissions (Bragazza et
al., 2005). Emission sources of reduced forms of nitrogen are primarily related to agricultural activities
such as animal husbandry (manure) and the application and production of fertilizers. Nitrogen emitted
into the atmosphere is subject to short and long range atmospheric transport (Galloway et al., 2008).
Reactive nitrogen can be redistributed from emission hot-spots (i.e. agricultural and densely
populated regions) to remote regions with undisturbed ecosystems naturally adapted to very low
nitrogen inputs and availability. Enhanced nitrogen deposition may result in acidification and
eutrophication of ecosystems, potentially leading to changes in plant diversity (Bobbink et al., 2010;
Stevens et al., 2011). In large parts of Europe the critical loads of eutrophication for ecosystems,
including those where mosses play an important role, are exceeded and are predicted to remain
exceeded in the near future (Hettelingh et al., 2011).
Several studies have shown that mosses have the potential to be indicators of atmospheric nitrogen
deposition (Harmens et al., 2011b, and references therein). However, sometimes the relationship
between atmospheric nitrogen deposition and the nitrogen concentration in mosses is weak (e.g.
Stevens et al., 2011) or shown to be species-specific (Arroniz-Crespo et al., 2008; Salemaa et al.,
34
2008). In 2005, ectohydric moss species were sampled for the first time at the European scale to
indicate spatial patterns of atmospheric nitrogen deposition across Europe (Harmens et al., 2011b).
Detailed statistical analysis of the European moss data (Schröder et al., 2010a) revealed that the total
nitrogen concentration in mosses is significantly and best correlated with EMEP modelled air
concentrations and atmospheric nitrogen deposition rates in comparison to other predictors that might
contribute to the spatial variation of nitrogen concentrations in mosses. The variation in the total
nitrogen concentration in mosses was best explained by the variation in ammonium (NH4+)
concentration in air, followed by nitrogen dioxide (NO2) concentrations in air. An apparent asymptotic
relationship was found between EMEP modelled total atmospheric nitrogen deposition and the total
nitrogen concentration in mosses (Figure 5.1; Harmens et al., 2011b). Factors potentially affecting
this relationship were discussed in more detail in the same study. Saturation appears to start at
nitrogen deposition rates of ca. 15 kg ha-1 y-1, which might indicate the threshold of adverse impacts of
nitrogen on the moss species sampled. For many habitats in Europe a nitrogen deposition of 15 kg
ha-1 y
-1 is within the range or even above the empirical critical load for nitrogen (Bobbink and
Hettelingh, 2011).
Figure 5.1 Relationship between EMEP modelled average annual total nitrogen deposition for 2003-
2005 and averaged nitrogen concentration in mosses in 2005/6 for EMEP grid cells were at least five
moss sampling sites were present (Harmens et al., 2011b).
Although many studies have reported separately on the nitrogen concentration in mosses and on the
impacts of elevated nitrogen deposition on terrestrial ecosystems, we are not aware of studies
reporting on the relationship between both. Based on a survey of 153 acid grasslands from Atlantic
Europe, Stevens et al. (2011) reported on a positive albeit weak relationship between the nitrogen
concentration in the moss species Rhytidiadelphus squarrosus and atmospheric nitrogen deposition.
Such a relationship was not observed for two vascular plant species. Nevertheless, Stevens et al.
(2011) concluded that R. squarrosus was not a good indicator of atmospheric nitrogen deposition in
acid grasslands. In the same study, grass species richness as a proportion of total species richness
increased whereas forb species richness decreased with increasing nitrogen deposition, indicating a
change in species composition. Over a period of 14 years, Zechmeister et al. (2007) observed that a
few moss species (Hypnum cupressiforme, Leucodon sciuroides) responded to ambient nitrogen
deposition levels by an increment in their population coverage. However, most moss species
remained stable in their overall abundance. Although species turnover rates were rapid, observed
changes in species composition could only to some extent be attributed to effects of airborne
pollution. The moss communities as a whole did not show directional changes attributable to the
observed levels of nitrogen deposition and the decrease of sulphur deposition. Thus, the substantial
exceedance of critical loads for eutrophication effects did not lead to acute injuries. If at all, such
injuries tended to be chronic injuries of individuals within the moss population.
Within the LRTAP Convention, the critical load approach has been developed to identify areas at risk
from adverse affects of air pollution (UBA, 2004; Hettelingh et al., 2007). Modelled critical loads of
nitrogen are based on the acceptable nitrogen concentration in soil solution, i.e. the critical value at
35
which nitrogen starts to leach from the soil. Applying the mass balance method, the critical nitrogen
load from deposition can then be calculated. In addition, empirical nitrogen critical loads for vegetation
have been defined (Bobbink and Hettelingh, 2011), based on the effects of elevated nitrogen
deposition on vegetation. Compared to modelled critical loads, empirical critical loads are generally
higher for the most sensitive ecosystems (Hetteling et al., 2007). Nevertheless, mapped exceedances
of empirical and modelled critical loads show a good resemblance. Areas in western Europe are
particularly at risk from critical load exceedance, as shown for example for modelled critical loads in
Figure 5.2. Although the same areas also have high concentrations of nitrogen in mosses, in parts of
continental and eastern Europe the nitrogen concentrations in mosses are relatively higher than the
critical load exceedance. One should bear in mind that whereas nitrogen concentrations in mosses
are well correlated with atmospheric deposition of nitrogen (Schröder et al., 2010a), soil
characteristics significantly affect the exceedance of nitrogen critical loads.
a) b)
Figure 5.2 a) Mean concentration of nitrogen in mosses per EMEP grid square in 2005/6 and b) the
average accumulated exceedance (AAE) of modelled critical loads of nitrogen (Nut N) in 2005. The
size of the coloured squares reflects the area exceeded. Source AAE data: ICP Modelling and
Mapping, Coordination Centre for Effects.
In the UK, critical loads for effects of nitrogen deposition on major sensitive habitats are exceeded for
58% of their area (RoTAP, 2012). There is strong evidence that nitrogen deposition has significantly
reduced the number of plant species per unit area (species richness) in a range of habitats of high
conservation value over large areas of the UK. The observed loss of plant species richness is
primarily due to a decline in frequency of species adapted to low nutrient habitats. In cases where
overall species richness has not changed, species characteristic of low nutrient habitats have been
replaced by species adapted to higher nutrient availability, with undesirable implications for habitat
conservation. Graminiod cover tends to have increased at the expense of forb cover. Moss (and
lichen) species richness and community composition is more dynamic than that of vascular plants,
with a replacement of more sensitive species by a more nitrogen pollution-tolerant community. There
is no evidence of further declines in species richness over the last 20 years in areas of high nitrogen
deposition, where much of the decline may have preceded the 1980s. However, there is evidence that
current nitrogen deposition in many parts of the UK is associated with further declines in the
frequency of sensitive plant species. Taken together, the data from field surveys and experimental
studies provide a strong body of coherent evidence that exceedance of critical loads of nitrogen
deposition is associated with adverse effects on terrestrial biodiversity at a UK scale. Whereas field
36
surveys (either repeated in time or over a gradient of nitrogen deposition) might be able to show a
relationship between nitrogen deposition and impacts on terrestrial ecosystems (Maskell et al., 2010),
causality of such a relationship can only truly be tested where only atmospheric nitrogen deposition is
manipulated as a driver of change. In field surveys it is difficult to disentangle the impact of nitrogen
deposition from other changes, such as natural succession, land-use, management history, climate
change, recovery from acidification (RoTAP, 2012). One should remember that nitrogen deposition
has both eutrophying and acidifying effects and observed impacts of enhance nitrogen deposition
might be through accelerated soil acidification rather than eutrophication as such (RoTAP, 2012;
Stevens et al., 2011).
Recent studies have shown that vegetation responses to nitrogen deposition might depend more on
the nitrogen form (ammonia or nitrate) than dose. Vegetation tends to be more sensitive to ammonia
than nitrate exposure (Leith et al., 2005; Pitcairn et al., 2006; Sheppard et al., 2008; 2009; Verhoeven
et al., 2011). Ammonia is more likely than wet depositon (ammonium and nitrate) to cause changes in
vegetation for a given rate of nitrogen deposition. To reflect these new findings, the critical levels for
ammonia were reduced in 2007, with lower critical levels being set for mosses and lichens than for
herbaceous plant species (ECE/EB.AIR/WG.5/2007/3; Cape et al., 2009). However, one should bear
in mind that the critical load for total nitrogen deposition makes no distinction between the forms in
which nitrogen is deposited. In the UK, the critical level for ammonia for lower plants is exceeded over
69% of the land area, and that for higher plants is exceeded over 19% of the UK (RoTAP, 2012).
In gradient studies in areas with high nitrogen deposition, Pitcairn et al. (2006) found that the nitrogen
concentration in mosses responds differently to wet and dry deposited nitrogen and appears to
respond more to concentrations of nitrate and ammonium in precipitation than to total nitrogen
deposition at wet deposition sites. Regional studies in the UK have shown maximum nitrogen
concentrations in mosses of 1.6% in areas dominated by wet deposition, despite relatively large
inputs of nitrogen, whereas in gradient studies around livestock farms dominated by dry deposition,
tissue nitrogen values of up to 4% were measured (Pitcairn et al., 2006). Nordin et al. (2006) found
that moss species in boreal forests take up predominantly ammonium, whereas biomass production
tended to be higher with nitrate fertilization. This resulted in a higher nitrogen concentration in the
mosses after ammonium exposure only. However, similar Spearman rank correlation coefficients
were found between the total nitrogen concentration in mosses and EMEP modelled air
concentrations or atmospheric deposition rates of different nitrogen forms (Schröder et al., 2010a).
In summary, nitrogen concentrations in mosses can provide a good indication of terrestrial
ecosystems being at risk from adverse impacts of enhanced atmospheric nitrogen deposition and can
serve as an early warning system (Harmens et al., 2011b). This could be true particularly for areas
that have traditionally been exposed to low atmospheric nitrogen deposition and are currently being
exposed to rising levels of nitrogen pollution. However, in large areas in Europe it might not be
possible anymore to establish a relationship between the total nitrogen concentration in mosses and
impacts of atmospheric nitrogen deposition on terrestrial ecosystems (if such a relationship could be
established at all) due to the historic rise in nitrogen deposition that has changed ecosystem
properties already. A combination of bioindicators is likely to be best to establish the current state of
terrestrial ecosystems, in particular for areas of high conservation value (Nordin et al., 2009).
37
6 Future activities of the ICP Vegetation
6.1 Review of ozone impacts on biodiversity and ecosystem services
The ICP Vegetation will review the potential (and where available, quantified) impacts of ozone in
Europe on the provisioning, regulating, supporting and cultural services involving vegetation. This will
include a review of current knowledge of whether ambient ozone impacts on plant biodiversity. We will
incorporate results from scientific publications and national reports to provide an up-to-date synthesis
of current knowledge. Highlights from this study will be submitted for inclusion in the WGE’s report on
impacts of air pollution on biodiversity and ecosystem services.
We plan to include the following chapters in the glossy report from this study:
1. Introduction
2. Sensitivity of European vegetation to ozone and the potential for impacts on biodiversity
3. Impacts on provisioning services, including food (see Mills and Harmens, 2011) and timber
production
4. Impacts on regulatory services including pollination, C sequestration (see Harmens and
Mills, 2012) and climate, air quality and water resources
5. Impacts on supporting services, including nutrient cycling, water cycling and primary
production
6. Impacts on cultural services, including leisure, recreation and amenity
7. Contributions from ICP Vegetation participants on nationally-funded research on this
subject
8. Conclusion and research recommendations.
6.2 Medium-term workplan (2013-2015) of the ICP Vegetation
As one of it’s core activities the ICP Vegetation will continue ozone stomatal flux model developments
and flux map validation. Hence, we will continue to collate supporting evidence for ozone impacts on
vegetation and review the robustness of flux-effect relationships for the establishment of new flux-
based ozone critical levels for additional plant species. In 2013, the ICP Vegetation will report on the
outcome of the 2010/11 European moss survey for heavy metals, nitrogen and POPs. The ICP
Vegetation will also continue to explore opportunities for outreach activities to other regions of the
globe.
The following medium-term workplan was adopted at the 25th Task Force Meeting of the ICP
Vegetation (Brescia, Italy, 31 January – 2 February 2012):
2013 (see ECE/EB.AIR.109/Add.2):
Report on supporting evidence for ozone impacts on vegetation;
Report on the European heavy metals and nitrogen in mosses survey 2010/11;
Report on the pilot study of mosses as biomonitors of POPs.
2014:
Report on supporting evidence for ozone impacts on vegetation;
Update of chapter 3 of the Modelling and Mapping Manual by inclusion of a new annex
describing further technical developments;
Report on ozone impacts on vegetation in a changing climate;
Report on heavy metal and nitrogen concentrations in mosses in EECCA/SEE countries;
Report on preparations for the moss survey 2015/16.
38
2015:
Report on supporting evidence for ozone impacts on vegetation;
Report on air pollution impacts on vegetation in EECCA/SEE countries;
Report on interacting effects of co-occurring ozone and nitrogen pollutants on vegetation;
Report on progress with the moss survey 2015/16.
Common workplan items of the WGE for 2013 have been described in the biannual workplan for the
LRTAP Convention (see ECE/EB.AIR.109/Add.2) and include:
i) Report on the further implementation of the Guidelines on Reporting of Monitoring and
Modelling of Air Pollution Effects;
ii) Report on ideas and actions to enhance the involvement of EECCA/SEE countries in the
Eastern Europe, the Caucasus and Central Asia and on cooperation with activities outside
the Air convention;
iii) Report on impacts on biodiversity and ecosystems services.
Common workplan items beyond 2013 will be decided at the WGE meeting in September 2013.
These and the ICP Vegetation-specific workplan items for 2014 and 2015 are subject to approval by
the Executive Body of the LRTAP Convention in December 2013.
39
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44
Annex 1. Participation in the ICP Vegetation
In many countries, several other scientists (too numerous to include here) also contribute to the work
programme of the ICP Vegetation.
Name/Country Institute Email
Ozone
Heavy
metals
Nitrogen
Albania
Pranvera Lazo
University of Tirana
Faculty of Natural Sciences
Tirana
pranveralazo@gmail.com
Austria
Gerhard Soja AIT Austrian Institute of
Technology GmbH
Konrad Lorenz-Str. 24
3430 Tulln
gerhard.soja@ait.ac.at
Harald Zechmeister Dept. of Conservation Biology,
Vegetation- and Landscape
Ecology
University of Vienna
Althanstraße 14
A 1090 Vienna
Harald.Zechmeister@univie.ac.at
 
Belarus
Yulia Aleksiayenak International Sakharov
Environmental University, Minsk
beataa@gmail.com
Belgium
Ludwig De Temmerman
Karine Vandermeiren
Nadia Waegeneers
Ann Ruttens
Veterinary and Agrochemical
Research Centre
CODA-CERVA
Leuvensesteenweg 17
B-3080 Tervuren
ludwig.detemmerman@var.fgov.be
kavan@var.fgov.be
nawae@var.fgov.be
anrut@var.fgov.be
 
Bulgaria
Lilyana Yurukova
Institute of Botany
Bulgarian Academy of Sciences
Acad. G.Bonchev Str., Block 23
1113 BG, Sofia
yur7lild@bio.bas.bg
 
Savka Miranova Department of Atomic Physics
Plovdiv University Paisii
Hilendarski
Tsar Assen Str. 24
4000 Plovdiv
savmar@pu.acad.bg
Croatia
Zdravko Spiric Oikon Ltd.
Institute for Applied Ecology
Trg senjskih uskoka 1-2
10020 Zagreb
zspiric@oikon.hr
 
Czech Republic
Ivan Suchara
Julie Sucharová
Silva Tarouca Research Institute
for Landscape and Ornamental
Gardening
Kvetnove namesti 391
CZ-252 43 Pruhonice
suchara@vukoz.cz
sucharova@vukoz.cz
 
Denmark (Faroe Islands)
Maria Dam
Katrin Hoydal
Environment Agency
Traðagøta 38
FO-165 Argir
mariad@us.fo
katrinh@us.fo
Estonia
Siiri Liiv Tallinn Botanic Garden
Kloostrimetsa tee 52
11913 Tallinn
siiri@tba.ee
 
Finland
Eero Kubin
Juha Piispanen
Jarmo Poikolainen
Jouni Karhu
Finnish Forest Research Institute
Muhos Research Station
Kirkkosaarentie 7
FIN-91500 Muhos
Eero.Kubin@metla.fi
Juha.Piispanen@metla.fi
Jarmo.Poikolainen@metla.fi
Jouni.Karhu@metla.fi
 
45
Name/Country Institute Email
Ozone
Heavy
metals
Nitrogen
Sirkku Manninen
Department of Biological and
Environmental Sciences, P.O.
Box 56, 00014 University of
Helsinki
sirkku.manninen@helsinki.fi
Former Yugoslav Republic of Macedonia
Trajce Stafilov
Viktor Urumov
Institute of Chemistry, Faculty of
Science, SS. Cyril and
Methodius University
Arhimedova 5, Skopje
trajcest@ pmf.ukim.mk
urumov@pmf.ukim.mk
France
Jean-François Castell
Olivier Bethenod
Laila Mahdhi
UMR EGC/AgroParisTech-INRA
78850 Thiverval-Grignon
castell@grignon.inra.fr
bethenod@grignon.inra.fr
lmahdhi@bcgn.grignon.inra.fr
Laurence Galsomiès ADEME, Deptartment Air
27 rue Louis Vicat
75737 Paris Cedex 15
laurence.galsomies@ademe.fr
 
Jean-Paul Garrec
Didier le Thiec
INRA-Nancy
F-54280 Champenoux
garrec@nancy.inra.fr
lethiec@nancy.inra.fr
Sabastien Leblond Muséum National d'Histoire
Naturelle France, 57 rue Cuvier
Case 39, 75005 Paris
sleblond@mnhn.fr
 
Yves Jolivet UHP, Nancy University jolivet@scbiol.uhp-nancy.fr
Matthieu Baggard
Anne Repellin
Université Paris Est Créteil matthieu.bagard@u-pec.fr
repellin@u-pec.fr
Valérie Simon
Louis Foan
Institut National Polytechnique
de Toulouse
4 Allée Emile Monso
31432 Toulouse Cedex 4
Valerie.Simon@ensiacet.fr
louise.foan@ensiacet.fr
From 01-09-12:
louise.foan@gmail.com
POPs
Germany
Jürgen Bender
Hans-Joachim Weigel
Institute of Biodiversity
Johann Heinrich von Thünen-
Institute (vTI), Bundesallee 50
D-38116 Braunschweig
juergen.bender@vti.bund.de
hans.weigel@vti.bund.de
Ludger Grünhage Institute for Plant Ecology
Justus-Liebig-University,
Heinrich-Buff-Ring 26-32
D-35392 Giessen
Ludger.Gruenhage@bot2.bio.uni-
giessen.de
Andreas Fangmeier
Andreas Klumpp
Jürgen Franzaring
Universität Hohenheim
Institut fűr Landschafts- und
Pflanzenökologie
Schloss Mittelbau (West)
70599 Stuttgart-Hohenheim
afangm@uni-hohenheim.de
aklumpp@uni-hohenheim.de
franzari@uni-hohenheim.de
Winfried Schröder
Roland Pesch
Hochschule Vechta, Institute für
Umweltwissenschaften
Postfach 1553
D-49364 Vechta
wschroeder@iuw.uni-vechta.de
rpesch@iuw.uni-vechta.de
 
Willy Werner
Stephanie Boltersdorf
University Trier, Department of
Geobotany, Behringstr. 5
54286 Trier
werner@uni-trier.de
Stefanie.Boltersdorf@gmx.de
Greece
Dimitris Velissariou Technological Educational
Institute of Kalamata
Antikalamos 241 00, Kalamata
d.velissariou@teikal.gr
Costas Saitanis Agricultural University of Athens
Laboratory of Ecology &
Environmental Sciences
Iera Odos 75
Botanikos 11855, Athens
saitanis@aua.gr
Eleni Goumenaki Technological Education Institute
Crete, 71004 Heraklion
Crete
egoumen@staff.teicrete.gr
46
Name/Country Institute Email
Ozone
Heavy
metals
Nitrogen
Iceland
Sigurður Magnússon Icelandic Institute of Natural
History, Hlemmur 3,
125 Reykjavík
sigurdur@ni.is
Italy
Stanislaw Cieslik
Ivano Fumagalli
European Commission, Joint
Research Centre - Institute for
Environment and Sustainability
Via E. Fermi, 2749,
I-21027 Ispra (VA)
stanislaw.cieslik@jrc.it
ivan.fumagalli@jrc.it
Gianfranco Rana
Marcello Mastrorilli
CRA-Research Unit for
Agriculture in Dry Environments
via C. Ulpiani, 5 70125 Bari
gianfranco.rana@entecra.it
marcello.mastrorilli@entecra.it
Luigi Postiglione
Massimo Fagnano
Dip. Di Ingegneria agraria ed
Agronomia del Territorio
Università degli studi di Napoli
Federico II, Via Università 100
80055 Portici (Naples)
postigli@unina.it
fagnano@unina.it
Cristina Nali
Alessandra Francini-
Ferrante
Dipartimento Coltivazione e
Difesa delle Specie Legnose “G.
Scavamuzzi”
Via del Borghetto 80
56124 Pisa
cnali@agr.unipi.it
afrancini@agr.unipi.it
Fausto Manes
Marcello Vitale
Elisabetta Salvatori
Lina Fusaro
Dipartimento di Biologia
Vegetale, Università di Roma
“La Sapienza”, Piazzale Aldo
Moro 5, I-00185 Rome
fausto.manes@uniroma1.it
marcello.vitale@uniromal.it
elisabetta.salvatori@uniroma1.it
lina.fusaro@uniroma1.it
Renate Alber Environmental Agency of
Bolzano, Biological Laboratory
Via Sottomonte 2
I-39055 Laives
Renate.Alber@provinz.bz.it
 
Alessandra de Marco
Gaia.Righini
Mihaela Mircea
ENEA, CR Casaccia
Via Anguillarese 301
00060 S. Maria di Galeria, Rome
alessandra.demarco@enea.it
gaia.righini@enea.it
mihaela.mircea@enea.it
Giacomo Gerosa
Angelo Finco
Riccardo Marzuoli
Universita’ Cattolica del S.c. di
Brescia, Via Pertini 11
24035 Curno
giacomo.gerosa@unicatt.it
angelo.finco@unicatt.it
riccardo.marzuoli@unicatt.it
Valerio Silli University of L'Aquila
Department of Environmental
Sciences
Valerio3001@gmail.com
Silvano Fares Agricultural Research Council
Research Centre for the Soil-
Plant System
Via della Navicella 2-4
00184 Rome, Italy
silvano.fares@entecra.it
Latvia
Olgerts Nikodemus Faculty of Geography and Earth
Sciences, University of Latvia
19 Raina blvd, Riga, LV 1586
nikodemu@lu.lv
 
Guntis Brumelis
Guntis Tabors
Faculty of Biology
University of Latvia
4 Kronvalda blvd
Riga, LV 1842
moss@lu.lv
guntis@lu.lv
Marina Frolova
Latvian Environment, Geology
and Meteorology Agency
Maskavas Str. 165
Riga, LV 1019
marina.frolova@lvgma.gov.lv
 
Inara Melece University of Latvia inaramelece@inbox.lv
Lithuania
Kestutis Kvietkus
Darius Valiulis
Institute of Physics
Savanoriu Ave 231
LT-02300 Vilnius
kvietkus@ktl.mii.lt
Valiulis@ar.fi.lt
47
Name/Country Institute Email
Ozone
Heavy
metals
Nitrogen
Montenegro
Slobodan Jovanovic Faculty of Natural Sciences
Laboratory for Nuclear
Spectrometry, Dz. Vasingtona 2
MNE-2000 Podgorica
bobo_jovanovic@yahoo.co.uk
Netherlands
Aart Sterkenburg RIVM Lab for Ecological Risk
Assessment, P.O. Box 1,
NL-3720 BA Bilthoven
aart.sterkenburg@rivm.nl
Norway
Eiliv Steinnes
Torunn Berg
Department of Chemistry
Norwegian University of Science
and Technology
NO-7491 Trondheim
eiliv.steinnes@chem.ntnu.no
torunn.berg@chem.ntnu.no
Hilde Uggerud Norwegian Institute for Air
Research (NILU)
htu@nilu.no
Poland
Barbara Godzik, Grażyna
Szarek-Łukaszewska,
Pawel Kapusta
W. Szafer Institute of Botany
Polish Academy of Sciences
Lubicz Str. 46, 31-512 Krakow
b.godzik@botany.pl
p.kapusta@botany.pl
 
Klaudine Borowiak Department of Ecology and
Environmental Protection
August Cieszkowski Agricultural
University of Poznan, ul.
Piatkowska 94C, 61-691 Poznan
klaudine@owl.au.poznan.pl
Romania
Adriana Lucaciu National Institute of Physics and
Nuclear Engineering
Horia Hulubei, Atomistilor 407,
MG-6, 76900 Bucharest
lucaciuadriana@yahoo.com
Raluca Mocanu Faculty of Chemistry
Al. I. Cuza University, B-dul
Caroll, nr. 11. code 00506 Lasi
ralucamocanu2003@yahoo.com
Antoaneta Ene Dunarea de Jos
University of Galati
aene@ugal.ro
Russian Federation
Marina Frontasyeva
Elena Ermakova
Yulia Pankratova
Konstantin Vergel
Frank Laboratory of Neutron
Physics, Joint Institute for
Nuclear Research, Joliot Curie
6 141980 Dubna
marina@nf.jinr.ru
eco@nf.jinr.ru
pankr@nf.jinr.ru
verkn@mail.ru
Natalia Goltsova Biological Research Institute
St.Petersburg State University
St Peterhof
198504 St. Petersburg
Natalia.Goltsova@pobox.spbu.ru
Serbia
Miodrag Krmar
Dragan Radnovich
Faculty of Science
University Novi Sad
Trg Dositeja Obradovica 4
21000 Novi Sad
miodrag.krmar@dbe.uns.ac.rs
dragan.radnovic@dbe.uns.ac.rs
Slovakia
Blanka Maňkovská Institute of Landscape Ecology
Slovak Academy of Science,
Štefánikova str. 3,
814 99 Bratislava, Slovakia
bmankov@stonline.sk
 
Slovenia
Franc Batic
Boris Turk
Klemen Eler
University of Ljubljana,
Biotechnical Faculty, Agronomy
Department, Jamnikarjeva 101,
1000 Ljubljana
franc.batic@bf.uni-lj.si
boris.turk@bf.uni-lj.si
klemen.eler@bf.uni-lj.si
Zvonka Jeran Jožef Stefan Institute
Dep. of Environmental Sciences,
Jamova 39, 1000 Ljubljana
zvonka.jeran@ijs.si
 
48
Name/Country Institute Email
Ozone
Heavy
metals
Nitrogen
Spain
J. Angel Fernández
Escribano
Alejo Carballeira Ocaña
J.R. Aboal
Ecologia
Facultad De Biologia
Univ. Santiago de Compostela
15782 Santiago de Compostela
bfjafe@usc.es
bfalejo@usc.es
bfjaboal@usc.es
 
Victoria Bermejo, Rocio
Alonso, Ignacio González
Fernández, Susana Elvira
Cozar, Héctor Calvete
Sogo
Departamento de Impacto
Ambiental de la Energía
CIEMAT, Ed 70
Avda. Complutense 22
28040 Madrid
victoria.bermejo@ciemat.es
rocio.alonso@ciemat.es
ignacio.gonzalez@ciemat.es
susana.elvira@ciemat.es
hector.calvete@ciemat.es
Vicent Calatayud
Esperanza Calvo
Fundacion CEAM
Parque Tecnologico
C/Charles R Darwin 14
Paterna, E-46980 Valencia
vicent@ceam.es
espe@ceam.es
Jesús Santamaria
Juan Jose Irigoyen
Raúl Bermejo-Orduna
Laura Gonzalez Miqueo
Sheila Izquieta
Departmento de Quimica y
Edafologia
Universidad de Navarra
Facultad de Ciencias
Irunlarrea No 1
31008 Pamplona I, Navarra
chusmi@unav.es
jirigo@unav.es
rberord@unav.es
lgonzale2@alumni.unav.es
sizquieta@alumni.unav.es
 
Javier Martínez Abaigar
Encarnación Núñez Olivera
Rafael Tomás Las Heras
CCT, Madre de Dios 51
Universidad de La Rioja
26006 Logroño, La Rioja
javier.martinez@unirioja.es
 
J. María Infante Olarte Gobierno de La Rioja
Dirección General de Calidad
Ambiental y Agua
Prado Viejo, 62 bis
26071 Logroño, La Rioja
dg.calidadambiental@larioja.org
 
Sweden
Per-Erik Karlsson
Gunilla Pihl Karlsson
Helena Danielsson
IVL Swedish Environmental
Research Institute
PO Box 5302,
SE-400 14 Göteborg
pererik.karlsson@ivl.se
gunilla.pihl.karlsson@ivl.se
helena.danielsson@ivl.se
Håkan Pleijel Environmental Science and
Conservation,
Göteborg University
PO Box 464, S-40530 Göteborg
hakan.pleijel@dpes.gu.se
Åke Rühling Humlekärrshultsvägen 10, S-572
41 Oskarshamn
ake.ruhling@telia.com
Switzerland
Jürg Fuhrer
Seraina Bassin
Matthias Volk
Verena Blanke
Agroscope Research Station
ART, Reckenholzstr. 191
CH-8046 Zurich
juerg.fuhrer@art.admin.ch
seraina.bassin@art.admin.ch
matthias.volk@art.admin.ch
verena.blanke@art.admin.ch
Sabine Braun Institute for Applied Plant Biology
Sangrubenstrasse 25
CH-4124 Schönenbuch
sabine.braun@iap.ch
Lotti Thöni FUB-Research Group for
Environmental Monitoring
Alte Jonastrasse 83
CH-8640 Rapperswil-Jona
lotti.thoeni@fub-ag.ch
 
Tobias Walser ETH Zurich
Institute of Environmental
Engineering
HIF C13 ETH-Hoenggerberg
CH-8093 Zurich
tobias.walser@ifu.baug.ethz.ch
Turkey
Mahmut Coskun Canakkale Onsekiz Mart
University, Health Service
Vocational College
17100 Çanakkale
coskunafm@yahoo.com
 
49
Name/Country Institute Email
Ozone
Heavy
metals
Nitrogen
Ukraine
Oleg Blum National Botanical Garden
Academy of Science of Ukraine
Timiryazevska St. 1, 01014 Kyiv
blum@nbg.kiev.ua
United Kingdom
Harry Harmens
(Chairman), Gina Mills
(Head of Programme
Centre), Felicity Hayes,
Laurence Jones, David
Norris, Jane Hall,
David Cooper
Centre for Ecology and
Hydrology
Environment Centre Wales
Deiniol Road
Bangor
Gwynedd LL57 2UW
hh@ceh.ac.uk
gmi@ceh.ac.uk
fhay@ceh.ac.uk
lj@ceh.ac.uk
danor@ceh.ac.uk
jrha@ceh.ac.uk
cooper@ceh.ac.uk
 
Lisa Emberson,
Steve Cinderby
Patrick Büker
Howard Cambridge
Stockholm Environment Institute,
Biology Department
University of York
Heslington, York YO10 5DD
l.emberson@york.ac.uk
sc9@york.ac.uk
pb25@york.ac.uk
hmc4@york.ac.uk
Sally Power
Emma Green
Nathan Callaghan
Department of Environmental
Science and Technology,
Imperial College,
Silwood Park Campus
Ascot, Berkshire SL5 7PY
s.power@imperial.ac.uk
emma.r.green@imperial.ac.uk
doctornathancallaghan@
googlemail.com
Sally Wilkinson
Bill Davies
Lancaster Environment Centre
Lancaster University
Lancaster LA1 4YQ
s.wilkinson4@lancaster.ac.uk
w.davies@lancaster.ac.uk
Mike Ashmore
University of York
Department of Biology
Heslington, York YO10 5DD
ma512@york.ac.uk
Mike Holland EMRC, 2 New Buildings
Whitchurch Hill
Reading RG8 7PW
mike.holland@emrc.co.uk
US
A
Filzgerald Booker
Kent Burkey
Edwin Fiscus
US Department of Agriculture
ARS, N.C. State University
3908 Inwood Road
Raleigh, North Carolina 27603
fbooker@mindspring.com
Kent.Burkey@ars.usda.gov
edfiscus01@sprynet.com
Uzbekistan
Natalya Akinshina
Azamat Azizov
National University of
Uzbekistan, Department of
Applied Ecology, Vuzgorodok,
NUUz, 100174 Tashkent
nat_akinshina@mail.ru
azazizov@rambler.ru
Outside UNECE region:
China
Zhaozhong Feng Temporary address:
Göteborg University
zhaozhong.feng@dpes.gu.se
Cuba
Jesús Ramirez Institute of Meteorology, Cuba jramirez_cu@yahoo.com
Egypt
Samia Madkour University of Alexandria,
Damanhour
samiamadkour@yahoo.co.uk
India
Dinesh Saxena Department of Botany
Bareilly College, Bareilly
dinesh.botany@gmail.com
Japan
Yoshihisa Kohno Central Research Institute of
Electric Power Industry (CRIEPI
kohno@criepi.denken.or.jp
Pakistan
Sheikh Saeed Ahmad Fatima Jinnah Women University
Rawalpindi
drsaeed@fjwu.edu.pk
South Africa
Gert Krüger
Elmien Heyneke
Jacques Berner
School of Environmental
Sciences, North-West University,
Potchefstroom, 2520
Gert.Kruger@nwu.ac.za
12605654@nwu.ac.za
jacques.berner@nwu.ac.za
50
AirPollutionandVegetation
ICPVegetation
AnnualReport2011/2012
This report describes the recent work of the International Cooperative Programme
on effects of air pollution on natural vegetation and crops (ICP Vegetation), a
research programme conducted in 35 countries in the UNECE region, with
outreach activities to other regions. Reporting to the Working Group on Effects of
the Convention on Long-range Transboundary Air Pollution, the ICP Vegetation is
providing information for the review and revision of international protocols to
reduce air pollution problems caused by ground-level ozone, heavy metals,
nitrogen and persistent organic pollutants (POPs). Progress and recent results
from the following activities are reported:
Contributions to revision of the Gothenburg Protocol.
Impacts of ozone on carbon sequestration.
Ozone biomonitoring programme.
European heavy metal and nitrogen in mosses survey 2010/2011, including
a pilot study on POPs.
Relationship between i) heavy metals and ii) nitrogen concentrations in
mosses and impacts on terrestrial ecosystems.
For further information please contact:
Harry Harmens
Centre for Ecology and Hydrology
Environment Centre Wales
Deiniol Road
Bangor
Gwynedd LL57 2UW
United Kingdom
Tel: +44 (0) 1248 374500
Fax: +44 (0) 1248 362133
Email: hh@ceh.ac.uk
ISBN: 978-1-906698-35-5
... Heagle et al. (1999) e Britz & Robinson (2001 observaram que diferentes cultivares de soja utilizadas na agricultura norte-americana quando expostas ao ozônio, diminuíram a produção de biomassa e grãos podendo ocorrer prejuízos econômicos de até US$ 4,00 bilhões ao ano (Ashmore, 2005). Ao avaliar a sensibilidade ao ozônio das espécies de maior importância econômica cultivadas na China, Europa e Estados Unidos, Chameides et al. (1999) e Aunan et al. (2000), Mills et al. (2000) e Ashmore (2003), respectivamente, concluíram que a soja é uma das espécies mais sensíveis. ...
Article
Full-text available
Oobjetivo deste trabalho foi avaliar o crescimento inicial, acúmulo de biomassa, trocas gasosas e defesas antioxidativas de soja 'Tracajá', cultivada na Região Amazônica, exposta ao ozônio sob condições controladas. Sementes germinadas em vasos foram levadas para duas câmaras, uma com ar filtrado (AF) e outra com ar filtrado mais 30ppb de ozônio (AF+O3). Aos 10 e 20dias após a semeadura, as trocas gasosas, crescimento em altura e acúmulo de biomassa foram medidos; aos 20dias após a semeadura, as defesas antioxidativas (ácido ascórbico e superóxido dismutase) foram analisadas. Aos 10dias após a semeadura, a fotossíntese líquida, condutância estomática, transpiração, altura, área foliar e biomassa foram 16, 27, 11, 22, 29 e 18% menores, respectivamente, no tratamento AF+O3. Aos 20dias após a semeadura, além dessas variáveis, comprimento da raiz, diâmetro do caule e razão raiz:parte aérea foram 10, 15 e 12%menores, respectivamente, apesar da concentração de ácido ascórbico e da atividade da superóxido dismutase terem aumentado. Plântulas de soja 'Tracajá' apresentam baixa tolerância à concentração de 30ppb de ozônio.
... It has long been established that ozone is damaging to vegetation, with many studies reporting effects of this pollutant at ambient concentrations throughout Europe (Skarby et al ., 1998;Bungener et al ., 1999a;Mills et al ., 2000). The development of a critical levels approach to establishing thresholds for response to ozone over the past decade has led to an intensification in research into effects on a wide range of plant species. ...
Article
Summary • Species of fen and fen-meadow communities, well supplied with water and nutrients, are characterised by high rates of growth, stomatal conductance values and specific leaf areas, all factors which have been associated with high sensitivity to ozone. We therefore examined the effects of ozone on 12 characteristic fen and fen-meadow species. • Plants received either filtered air or ozone; AOT40 exposures ranged from 9200 to 14 300 ppb h. Eight of the 12 species exhibited foliar injury in response to ozone exposure, with the first signs of injury on Vicia cracca, following an AOT40 exposure of only 1950 pbb h. • Ozone exposure significantly reduced plant photosynthetic rate, stomatal conductance and biomass production in four species. Cirsium arvense exhibited the greatest biomass response to ozone (32% and 58% reduction in above- and below-ground weight, respectively). Species with higher levels of visible injury tended to show greater reductions in biomass. There was a significant positive association between stomatal conductance and the magnitude of ozone effects on root biomass. • The widespread occurrence of either visible injury or growth reductions amongst the species screened, and the magnitude of effects on the most sensitive species, indicate that species of fens and fen-meadows may be more sensitive to ozone than other seminatural ecosystems which have been the focus of recent ozone studies.
Technical Report
Full-text available
This report presents the latest version of the Rossby Centre regional atmospheric model, RCA3, with focus on model improvements since the earlier version, RCA2. The main changes in RCA3 relate to the treatment of land surface processes. Apart from the changes in land surface parameterizations several changes in the calculation of radiation, clouds, condensate and precipitation have been made. The new parameterizations hold a more realistic description of the climate system. Simulated present day climate is evaluated compared to observations. The new model version show equally good, or better, correspondence to observational climatologies as RCA2, when forced by perfect boundary conditions. Seasonal mean temperature errors are generally within ±1oC except during winter in north-western Russia where a larger positive bias is identified. Both the diurnal temperature range and the annual temperature range are found to be underestimated in the model. Precipitation biases are generally smaller than in the corresponding reanalysis data used as boundary conditions, showing the benefit of a higher horizontal resolution. The model is used for the regionalization of two transient global climate change projections for the time period 1961- 2100. The radiative forcing of the climate system is based on observed concentrations of greenhouse gases until 1990 and on the IPCC SRES B2 and A2 emissions scenarios for the remaining time period. Long-term averages as well as measures of the variability around these averages are presented for a number of variables including precipitation and near-surface temperature. It is shown that the changes in variability sometimes differ from the changes in averages. For instance, in north-eastern Europe, the mean increase in wintertime temperatures is followed by an even stronger reduction in the number of very cold days in winter. This kind of performance of the climate system implies that methods of inferring data from climate change projections to other periods than those actually simulated have to be used with care, at least when it comes to variables that are expected to change in a non-linear way. Further, these new regional climate change projections address the whole 21st century.
Article
Full-text available
Mites and springtails are important members of soil mesofauna and have been proven to be good bioindicators of airborne pollutants. We studied the surrounding area of a steel mill located in a mountain valley of North Spain. Previous studies had documented the existence of a pollution gradient in this area due to the emissions of the factory, thus providing an interesting site to investigate the potential effects of pollutants (heavy metals and nitrogen) on soil biodiversity. The density of Acari and Collembola significantly decreased with the increase in concentration of Cr, Mn, Zn, Cd and Pb. Mites appeared to be more sensitive to heavy metal pollution than springtails. Likewise, the density of these microarthropoda was lower in those soils exhibiting higher nitrogen content. The species composition of the community of Acari and Collembola changed according to heavy metal pollution. Significant differences in abundance, species richness and diversity were observed between the communities of the sampling sites. Some species were exclusive of the less polluted sites, while other appeared in the most contaminated ones. This different response of soil mesofauna to pollutants suggests that some mite or springtail species could be used as bioindicators of heavy metal pollution.
Article
Identification of genetic control of ozone (O3) sensitivity is desirable for selection of plant cultivars which are indicator of O3 stress. A cross was made between two cultivars of snap bean (Phaseolus vulgaris L.), 'Oregon 91' (P1) and 'Wade Bush' (P2) an O3-sensitive and O3-insensitive cultivar, respectively. Ten genetic populations (generations), 'Oregon 91' (P1), 'Wade Bush' (P2), F1, F2, backcrosses to both parents, and all reciprocal crosses, were field planted in each of two summers and evaluated for injury to O3. Ozone responses for the reciprocal crosses were not significantly different for any generation, so injury ratings from the reciprocal crosses were combined for each generation to provide six populations (P1, P2, F1, F1, BC1, and BC2) for analysis. When components of genetic variation were estimated from the six generations, additive genetic variance was the most important component in the total genetic variance available, although dominance variance was also a significant component. There was a inconsistency in the magnitude and the direction of the factors contributing to the dominance effects and also a large environmental component making up the phenotypic variance. Estimates of broad-sense heritability and narrow-sense heritability were 60% and 44%, respectively. Results suggest that O3-sensitive and O3-insensitive selections could be screened and evaluated in an ambient O3 environment. Several generations will be necessary, however, to develop 'Bush Blue Lake' type selections that vary only in sensitivity to O3.
Article
The DO3SE (Deposition of O3 for Stomatal Exchange) model is an established tool for estimating ozone (O3) deposition, stomatal flux and impacts to a variety of vegetation types across Europe. It has been embedded within the EMEP (European Monitoring and Evaluation Programme) photochemical model to provide a policy tool capable of relating the flux-based risk of vegetation damage to O3 precursor emission scenarios for use in policy formulation. A key limitation of regional flux-based risk assessments has been the assumption that soil water deficits are not limiting O3 flux due to the unavailability of evaluated methods for modelling soil water deficits and their influence on stomatal conductance (gsto), and subsequent O3 flux.
Technical Report
This report describes the status of the impact assessment of nitrogen, sulphur and heavy metal depositions in Europe and the progress made regarding the relation between nitrogen deposition and loss of biodiversity. Part 1 Progress CCE The Centre for Integrated Assessment Modelling (CIAM) prepared Baseline (BL) and Maximum Feasible Reduction (MFR) scenarios with resulting nitrogen and sulphur depositions. Chapter 1 reports the impacts regarding exceedances of acidification and nitrogen critical loads, including results of the so-called “ex-post analysis”. In addition, results from dynamic modelling were used to analyse the delays in responses of soil chemistry to changes in depositions. Conclusions include that ‘environmental improvements’ achieved under MFR in comparison to BL are considerable for all indicators. However, it should also be noted that MFR does not lead to non-exceedance of critical loads and requirements for sustainable soil chemistry (i.e. non-violation of the chemical criterion) for all ecosystem areas in Europe. Knowledge on nitrogen impacts has been further extended within the effects community by assessing the interaction between N and carbon (C). This is also reflected in an extension of the widely-used VSD model to include interactions between N- and C-pools and -fluxes, in a new model version, named VSD+. VSD+ has been applied by National Focal Centres (NFCs) of the International Collaborative Programme Modelling and Mapping (ICP M&M) on sites for which measurements are available. Another step forward in the assessment of impacts is the use of models that predict the abundance of species, based on abiotic conditions. NFCs were urged to familiarize themselves with one of them, the VEG model. Also data on species abundance in combination with abiotic data to feed vegetation models has been requested from NFCs for assimilation into a European database. These issues have been bundled into the 2009-2010 call for data of the CCE, of which the results are discussed in Chapter 2. In total 14 NFCs have responded to (part of) the call. A workshop on the review and revision of empirical critical loads and dose-response relationships was held under the Convention on Long-range Transboundary Air Pollution, in Noordwijkerhout, from 23-25 June 2010. The newly agreed critical loads and recommendations on their use are summarized in Chapter 3. Part 2 Indicators and Assessment of Change of Plant Species Diversity This part elaborates the progress in the model development to link soil chemistry to vegetation effects. This is in line with the long-term strategy of the Convention which includes the encouragement of the assessment of air pollution effects with respect to the change of biodiversity. For this the VSD+ model has been linked to the VEG model. Results of this model combination have been evaluated and compared to results of the ForSAFE-VEG model combination. An important aspect for vegetation modelling that was missing in VSD+ is the modelling of the light that plants receive below the forest canopy. How this and the model coupling have been implemented together with results of the comparison can be found in Chapter 4. In recent years, discussions took place on selecting an appropriate effect indicator to quantify changes in biodiversity with respect to (nitrogen) deposition. The arguments and proposed approaches are brought together in a framework which can help to focus this discussion. The reader can find this rationale and the framework in Chapter 5. Part 3 Heavy Metals The Protocol on Heavy Metals (HM) was signed in 1998 and entered into force in 2003. Currently the process for the revision of the Heavy Metals Protocol is underway. To support additional information to the negotiations on the proposed amendments on the HM Protocol, a research project has been commissioned by the Netherlands to TNO, EMEP MSC-E and the CCE. In this project four scenarios were compared for which emissions, costs of emission reductions, depositions and exceedances of critical loads have been calculated. The description of the emission scenarios, including the potential measures and their costs can be found in Chapter 6. The depositions that result from the emissions are reported and discussed in Chapter 7, including the estimation of re-suspension of metals from soils into the air. Chapter 8 gives the exceedances of the critical loads for the given scenarios and discusses the implications of re-suspension on the critical load concept. In chapter 9 the toxicological effects of metal concentrations in the soil solution on soil microorganisms, plants and invertebrates are tentatively addressed using the CCE background database of a European ecosystem. Conclusions of Part 3 include that the costs of revision of the HM protocol for UNECE Europe are estimated to be 1.3 and 11.6 billion €for Option 2 and Option 1, respectively. The reduction of emissions is not only beneficial regarding heavy metal pollution, the measures taken in Option 2 and Option 1 may also bring about considerable reductions of PM2.5 emissions in Europe. For the additional Hg emission reduction measures another 2.6 billion € should be added for both options. Depositions of heavy metals are reduced, but not to the same extend as the reductions in emissions, due to the process of re-suspension. While the emission reductions are reflected in the lowering of critical load exceedances everywhere, still large parts of Europe’s nature remain at risk. Uncertainty analysis requires further assessment of the state of implementation of the current protocol and of the origins of re-suspended deposition. Part 4 Finally, consists of the national reports sent by the NFCs describing their submissions to the 2009-2010 call for data, which was adopted by the Working Group on Effects at its 28th session (Geneva, 23-25 September 2009). Key words Acidification, air pollution effects, biodiversity, critical loads, dose−esponse relationships, dynamic modelling, ecosystem services, eutrophication, exceedance, LRTAP Convention, heavy metals
Article
Worldwide there is concern about the continuing release of persistent organic pollutants (POPs) into the environment. In this study we review the application of mosses as biomonitors of atmospheric deposition of POPs. Examples in the literature show that mosses are suitable organisms to monitor spatial patterns and temporal trends of atmospheric concentrations or deposition of POPs. These examples include polycyclic aromatic hydrocarbons (PAHs), polychlorobiphenyls (PCBs), dioxins and furans (PCDD/Fs), and polybrominated diphenyl ethers (PBDEs). The majority of studies report on PAHs concentrations in mosses and relative few studies have been conducted on other POPs. So far, many studies have focused on spatial patterns around pollution sources or the concentration in mosses in remote areas such as the polar regions, as an indication of long-range transport of POPs. Very few studies have determined temporal trends or have directly related the concentrations in mosses with measured atmospheric concentrations and/or deposition fluxes.