Nutrients or pollutants? Nitrogen deposition to

Transcription

Nutrients or pollutants? Nitrogen deposition to European forests
Werner Eugster and Matthias Haeni
Institute of Agricultural Sciences, Department of Environmental Systems Science, ETH
Zürich, LFW C55.2, Universitätsstrasse 2, CH–8092 Zürich, Switzerland
Abstract
Forests take up gaseous, liquid, and particulate substances that are present in the air.
Although considered nutrients on the one hand, nitrogen inputs exceeding the critical load
that can be absorbed by an ecosystem act as pollutants. This chapter outlines the effects of N
deposition to forest ecosystems and discusses recent progress that has been made to more
accurately quantify dry deposition, which at many a forest location in Europe is larger than
wet and occult deposition.
To quantify the effects of N deposition on tree growth, a good measure for net ecosystem
production (NEP) is needed. Eddy covariance (EC) flux measurements are one established
way to quantify NEP. While EC flux measurements are costly and remain restricted in their
application to a few suitable locations, dendrometer measurements with high temporal
resolution show a similar seasonal and annual signal to NEP. Such measurements are
becoming increasingly important to quantify ecosystem biomass accumulation, which can be
related to N deposition rates. The policy relevance of such activities emerges from the
UNECE's Gothenburg protocol to abate acidification, eutrophication and ground-level ozone,
but also the quantification of natural sinks under the Kyoto protocol profits from such
measurements.
1. Introduction
The effects of N deposition on forest ecosystems are described in the conceptual framework of
Aber et al. (1989, 1998), which was further refined by Emmet (2007). N deposition affects all
ecosystem compartments and alters many biogeochemical processes. Its effect is modified by
site conditions, such as nutrient availability in the soil, species composition of the forest
stand and management history, in combination with the changing climate conditions and
pollutant loads.
In this chapter we will focus on N deposition loads to European forest ecosystems. The
emphasis is on the most recent scientific developments aiming at quantifying N dry
deposition, which tends to be the dominant pathway of N deposition in temperate climates
where the majority of days do not see rainfall, and where wet deposition is hence unable to
represent total ecosystem-scale N inputs. For conventional N deposition measurements it is
EUGSTER and HAENI / Manuscript Final Version
—1—
will appear with Elsevier (Matyssek et al.)
however indispensable to consult the standard protocols of ICP Forests (2010). It is not the
intent of this chapter to duplicate the wealth of existing literature on non-dry deposition
processes, but for the sake of completeness a very brief summary of all relevant pathways will
be given to put dry deposition into the context of total N deposition. Effects of N deposition
are always a combination of several input pathways and cannot be cleanly teased apart. But
with respect to dry deposition, above-ground N uptake processes must be more strongly
investigated than would be required in cases where the primary pathway of anthropogenic N
inputs is wet deposition.
N input must be seen in a long-term perspective and on an integrative ecosystem scale. At
short timescales and low N deposition rates the nutritional effect may be dominant, primarily
increasing tree growth. At longer timescales and higher N deposition rates, this effect can
become detrimental. This transition may be slow and cannot easily be generalized. Shortterm studies with seedlings are best suited to study specific physiological processes, but the
results obtained in this way cannot readily be transferred to mature trees and whole forest
ecosystems.
N is the most important nutrient for plant growth and it is often the limiting factor for net
primary production in temperate forest ecosystems (Vitousek and Field, 2001). Elaborated
methods have been established to measure wet deposition within ICP Forests, and canopy
balance models are used to estimate the total N deposition. At many European locations,
however, dry deposition of reduced and oxidized N-containing gases is an even larger
component of N deposition than wet deposition (see e.g. Flechard et al., 2011, Eugster et al.,
1998). Newer developments now allow for a better quantification of dry deposition rates of
gases and particles. After a short summary on effects of N deposition to forest ecosystems
this chapter focuses on these recent developments that may in the future complement
existing N deposition measurements in Europe, but also briefly addresses the other
components of total N deposition measurements which must remain an essential part of
long-term monitoring efforts even at sites where gaseous dry deposition is the primary source
of N deposition.
2. Effects of nitrogen deposition to forest ecosystems
N uptake (Fig. 1) is typically via the roots of the trees, but it has been shown by many studies
that canopy N uptake (CNU) in forests also play an important role (e.g. Schaefer and Reiners,
1990; Sievering et al., 2007). Earlier studies have mostly focused on foliar uptake by saplings
and young plants in experimental treatments, and the results obtained in this way were then
scaled up to whole canopies of more mature trees. It has been recognized that neglecting
CNU in some of the widely used canopy budget models is probably a wrong assumption (ICP
Forests, 2010, p. 39). This indicates where the most urgent present-day research needs are:
—2—
to obtain realistic and accurate estimates of ecosystem-scale total N deposition and to assess
all pathways of N uptake that may have detrimental effects on ecosystems.
Figure 1 about here
It is agreed that N inputs below the critical loads are – in the long term – not detrimental to
ecosystem structure and functions. Nitrogen deposition rates above a critical threshold will –
again in the long term – lead to an impairment of the functions, depending on the site quality
and other stressors present. Having said that, it should however also be recalled that the
effect of N deposition on species composition in forests starts at rather low levels, and the
apparently positive effect of added nutrients may have a strong negative effect on biodiversity
(e.g. Vitousek et al. 1997, Bobbink et al. 1998, 2010, Kozovits and Bustamante, this volume).
This may be a key issue to keep in mind in natural and semi-natural, extensively managed
forest ecosystems in Europe. For Central Europe Ellenberg (1990) showed that the vast
majority of endangered vascular plant species growing in forests and woodlands are
associated with N-poor environments (up to 40% of the species that grow in the N-poor class
2 on the nominal scale by Ellenberg et al. 1991).
Foliar uptake leads to the fact that estimates of N input via throughfall and bulk deposition
collection with subsequent calculation of total deposition with models may substantially
underestimate actual N inputs to forest ecosystems (Eilers et al. 1992). Nave and Curtis
(2011) estimated that more than 50% of atmospheric N deposition is actually intercepted and
retained in forest canopies. Moreover, Eilers et al. (1992) found that above-ground uptake of
N may cause nutritional disturbances by changes in the N:Ca and N:Mg ratios in the needles,
which may lead to Mg-deficiency symptoms of Norway spruce (Eilers et al. 1992). Sparks
(2009) clearly showed that direct foliar uptake of N is a direct addition of N to plant
metabolism and could potentially more readily influence plant growth (Fig. 1, right) compared
to soil-deposited N (Fig. 1, left). Present-day ecosystem process models however still do not
explicitly simulate this more direct N pathway (see Sparks 2009), and hence cannot always
adequately represent the effects of N deposition on forest growth.
Nihlgard (1985) drew the attention of the scientific community to N as a possible cause for
forest dieback, and since then an ever increasing number of studies have addressed the
question how N deposition – mostly in combination with other pollutants – affects tree growth
and health status. Krupa (2003) provided a concise overview over the different pathways and
effects on terrestrial ecosystems, and even more detailed information on specific aspects of
pollutant fluxes to different forest tree species can be found in the book by Elling et al.
(2007).
2.1 What are the typical detrimental effects of nitrogen deposition?
N deposition affects all compartments of forests. The narrowing of the carbon to nitrogen
(C/N) ratio of the forest floor indicates accumulation of N and is a measure to assess the risk
—3—
of nitrate release in the seepage water. In general, a C/N below 25 can be seen as the
threshold for enhanced nitrate leaching via the seepage water (e.g. Gundersen et al. 1998;
Aber et al. 2003, Dise et al. 2009). At the same time the composition of the understory is
altered in many regions leading to a reduced biodiversity (Bobbink et al. 2010; Nordin et al.
2006).
The N to nutrient ratios (N/Mg, N/K, N/P) in needles and leaves tend to become larger,
indicating a shortage of other nutrients which will lead to an enhanced sensitivity to natural
stress factors, like a lowered frost resistance and enhanced susceptibility to insects (e.g.
Schulze 1989; Gundersen 1998; Flückiger and Braun 1999).
There is increasing evidence that in Central Europe the phosphorous (P) nutrition is also
declining (Khanna et al. 2007; Prietzel et al. 2008), leading to – at least regional – reduced
growth (Braun et al. 2010). The pathways are not fully understood, but a reduction of P
uptake via roots due to N mediated inhibition of mycorrhiza is discussed (Aber et al. 1989;
Braun et al. 2010; see Kraigher et al., this volume). The linkages between C, N and P must be
taken into account for the modelling of carbon sequestration (Goll et al. 2012; see also
Bytnerowicz et al., this volume).
To a certain degree N acts as a fertilizer. If the natural N limitation is removed, the trees react
with enhanced growth. This is probably one reason for the observed increased growth of trees
in Europe in the last decades (e.g. Spieker et al. 1996). However, this does not imply that the
effect is also positive at the ecosystem level and over longer time periods; biodiversity of a
forest, for example, is normally determined primarily by the understory vegetation where
many N-sensitive species may disappear while the few tree species thrive (see review by
Bobbink et al. 2010). De Shrijver et al. (2008) for example emphasize that the effects of N
deposition on soil acidification, groundwater and surface water quality, biodiversity, and
ecosystem services must be considered in addition to the simple nutritional effect that N
deposition may exert on tree biomass increment alone.
The input of N above the demand for tree growth is termed N saturation, which means that
no further increase of growth results from additional N input (De Vries et al. 2006, 2008,
Laubhann et al. 2009). There are indications, that the enhanced growth is today halted on
many sites (Nellemann and Thomsen 2001). The excess N cannot be taken up or stored in
soil or vegetation and percolates through the soil, leading to acidification. This in turn means
a depletion of base cations (Mg, K, Ca) or an increase of the amount of acidic cations (Al, Mn,
H) in the soil. If the acidification potential of nitrogen is higher than the natural weathering
rate of the soil forming bedrock, the soils acidify.
To integrate all these aspects, an ecosystem-specific critical threshold is defined above which
ecosystem damage is likely to occur. This is taken into account in the calculation of the
critical loads, which are the basis for the air pollution abatement policy in the UNECE region
—4—
(UNECE 2005). The site-specific calculated critical load for nutritional N is for most forests in
the range of 10 to 20 kg N ha–1 yr–1 (Spranger et al. 2004, p. 125); the empirical critical load
for e.g. coniferous woodland is about 5 to 15 kg N ha–1 yr–1 (Bobbink and Hettelingh 2010).
On the European scale 74% of the total terrestrial area (forests and other ecosystems) of the
EU27 countries (52% for all European countries) is at risk of nutrient N eutrophication
(reference year 2000, Slootweg et al. 2010, p. 17).
Knowledge of the effects of N on the ecosystem level is essential. To find and verify causeeffect relations and to establish critical thresholds, the input of N in forests and the
interrelations with other pollutants should be known as accurately as possible.
3. The components of nitrogen deposition
The IPC Forests (2010) manual lists all relevant components of N deposition:
1. Wet-only deposition
2. Bulk deposition
3. Dry deposition
4. Occult deposition
5. Throughfall deposition
6. Stemflow deposition
7. Total atmospheric deposition
These components are non-additive and basically reflect the difficulty of separating one
component clearly from another. However, this would not be that critical, if a good
measurement method existed to accurately quantify total atmospheric deposition. Its
definition according to IPC Forests (2010) is the sum of wet-only deposition and dry
deposition, but not including occult deposition. A more integrative definition would define
total deposition as wet (rainfall and snow, Section 3.1), occult (fog, Section 3.2), and dry
deposition (Section 3.3) as we will do in this text. In areas with substantial snowfall one could
argue that snow deposition is a separate pathway for N input, but for the sake of simplicity
we will consider snow deposition a component of wet deposition, since for the ecosystem this
input becomes typically only relevant when the snow melts.
Despite the difficulties in terminology and how to measure the components it should be clear
that total N input by all relevant pathways must be quantified accurately to relate
measurable effects to the N deposition rates in order to specify and quantify detrimental
effects (see Section 2) on ecosystems.
—5—
3.1 Wet deposition by rainfall and snow
3.1.1 Wet-only deposition
Wet deposition is the most important pathway of N deposition in tropical rainforests without
a dry period. In temperate forests the contribution of wet deposition to total N deposition
however depends (1) on the probability of rain at a given site, and (2) on the concentrations of
N compounds in rainfall. At localities where rain is rare it is clear that wet deposition may be
substantially less than 50% of total deposition. For example, in western Switzerland Eugster
et al. (1998) found wet deposition rates on the order of 7–8 kg N ha–1 yr–1, which contributed
roughly 14% to 33% to total N deposition. Wet deposition is often a larger-scale phenomenon
since very local rains are restricted to short periods in summer, and even then the ions in the
rainwater are often transported over larger distances and hence small-scale variability in wet
deposition is small compared to the variability of the other components.
3.1.2 Bulk deposition
If bulk deposition is measured in place of wet-only deposition, then the deposition estimates
tend to be higher. Even higher than bulk deposition are N deposition values determined from
throughfall data in forest canopies. Throughfall is believed to reflect the composition of both
incident precipitation (wet deposition) and dry deposition from N containing components that
are deposited on branches and foliage during dry periods, which are then washed off by
precipitation and can be collected as throughfall (e.g. Thimonier et al. 2005). If CNU takes
place, or leaching from the canopy, then some assumptions must be made to be able to
derive dry deposition from the difference between throughfall deposition and bulk deposition
measurements (Thimonier et al. 2005). Namely, gaseous components can be taken up via
stomata (see Cieslik, this volume) and hence will not get deposited on leaf surfaces and thus
also will not be washed off. It may however hold for aerosol deposition of solid dry fine dust
particles. The N input from aerosols is however a much smaller component than gaseous dry
deposition (see Section 3.3).
3.1.3 Deposition by snow
Very little attention has been given to N deposition by snow so far. In northern countries and
in mountain ranges this however may be a relevant additional pathway of N input. The
scavenging of aerosol particles by falling snow flakes was found to be somewhat greater than
that for liquid precipitation (Kyrö et al. 2009) and corresponded to rain intensities around 3
mm h–1 (rather intense downpour) even when snowfall remained below 0.2 mm h–1. This
means that snow deposition of N should be carefully considered in polluted areas where the
increased efficiency for scavenging aerosols may constitute a significant pathway for N
deposition. Because snow deposition occurs during the dormant season of the vegetation, it
may not immediately affect the ecosystem, but during spring snow melt a peak influx of
—6—
deposited ions may occur. For lakes and streams this was termed “acid shock” (Jeffries et al.
1979) which in a similar way can also be expected for N inputs deposited by snow.
3.2 Occult deposition by fog
In temperate forests, water inputs by fog are rather small, but N concentrations in fog water
are one to two orders of magnitude larger than in rain water, and hence deposition amounts
can be considerable at localities with frequent fog. At the Lägeren mixed forest in
Switzerland, for example, the N input by fog was quantified at 4.3 kg N ha–1 yr–1 which is
16.4% of total N deposition (Burkard et al. 2003). The tree species composition is dominated
by beech, followed by Norway spruce, ash, sycamore, silver fir and 8 less abundant species.
This is 62% of wet deposition (6.9 kg N ha–1 yr–1) or 28% of dry deposition (15 kg N ha–1 yr–1).
Also at the Waldstein site in the Fichtelgebirge, Germany, fog deposition contributed a similar
amount of N to total deposition as was measured as wet-only deposition (Thalmann et al.
2002).
Although these are important N additions, an experimental treatment of selected trees at the
Lägeren site with a mist that reflects N concentration in fog water primarily showed an
increased growth effect without clear signs of induced stress symptoms (Wortman et al.
2012). Elemental concentrations in Lägeren beech leaves showed that both N/P and N/K
ratios were 6% and 19% higher, respectively, for N-amended leaves in comparison with
control leaves. This supports the expectation of CNU (Fig. 1) which is enhanced by occult
deposition compared to the shorter contact time and lower N concentrations brought about
by rainwater. On the other hand, occult deposition to temperate forests should not be
overestimated: often the high frequencies of fog occurrence coincide with the cold period of
the year when trees are dormant, and hence it remains to be investigated in more detail how
important the timing of fog and its N input is to forest ecosystems.
3.3 Dry deposition of gases and aerosol particles
3.3.1 Methods for field measurement
Gaseous dry deposition is most directly measured by the eddy covariance (EC) method
(Aubinet et al. 2012). This method uses a three-dimensional anemometer – mostly one
working with ultrasonic impulses to determine wind speed and air temperature – in
combination with a fast gas analyzer that allows measuring every tenth of a second or so to
systematically probe turbulent transport of the gas of interest. It is then assumed that the
deposition rate equals the flux rate measured with EC at a certain height above the forest
canopy. Such measurements are feasible for most gases such as NO (nitric oxide), NO2
(nitrogen dioxide), NOx (the sum of NO and NO2), HNO3 (nitric acid) and PAN (peroxyacetile
nitrate).
—7—
The dry deposition of aerosol particles can also be measured with EC, but it is still a
challenge: while instruments exist to measure the number flux of particles, it is not yet
possible to measure the exact chemical composition at a time interval suitable for EC. Also
gaseous fluxes of NH3 are not easily measured directly with EC, as Ferrara et al. (2012) have
shown. Hence many scientists chose a variant of the EC method that allows the deployment
of relatively slow sensors via the relaxed eddy accumulation (REA) technique: the gas of
interest or aerosol components are sampled in two reservoirs, one which collects air
whenever the vertical wind speed is positive and above a given threshold value, while the
other collects air from down-moving wind gusts with a wind speed above that threshold. The
concentration difference between the two reservoirs can then be analysed with an analyser
that requires considerably longer analysis time than that required by EC (which is on the
order of fractions of a second in most cases). The deposition flux is then derived from that
concentration gradient times a turbulent mixing coefficient (which relates the gradient to the
corresponding flux), and a correction factor (β) to correct for the effect of the REA wind speed
thresholds on the flux estimates. In the case of aerosols the REA method was used in several
combinations to obtain defensible flux estimates.
All these methods are expensive, require high scientific and technical skills to operate, and
can only be carried out at very few forest sites that meet all these requirements. Hence,
probably the most important advance in the methodology of dry deposition flux
measurements in the past decade was made during the NitroEurope project
(http://www.nitroeurope.eu) with its development of a low-cost Delta denuder system for
sampling gaseous NH3, NO2, HNO3, HONO, SO2, HCl and aerosol NH4+, NO3–, SO42–, Cl–, Na+,
Ca2+, and Mg2+ (Fig. 2). This allows obtaining monthly mean concentrations of the most
important N gases and aerosol components, from which deposition can be estimated via
transfer models. In this way an independent estimate of dry deposition can be made that
does not require any assumptions on how the difference between throughfall and bulk
precipitation may actually be related to dry deposition.
Figure 2 about here
3.3.2 Contribution to total N deposition
Aerosol dry deposition was only estimated to be on the order of 2.0 to 2.3 kg N ha–1 yr–1 in
parts of Switzerland (Eugster et al. 1998). Fluxes derived from Delta denuders as presented
by Flechard et al. (2011) in their Table 3 were used for additional analyses presented here.
These numbers suggest that aerosol dry deposition (NH4+ and NO3–) to European forests
ranges between 0.3 and 6.6 kg N ha–1 yr–1 with a median of 3.0 kg N ha–1 yr–1 and the gaseous
dry deposition (NH3, NO2 and HNO3) ranges between 1.2 and 25.8 kg N ha–1 yr–1. On average,
gaseous dry deposition was a factor 1.5 to 4.5 larger than aerosol deposition (median 2.6).
The most challenging finding by Flechard et al. (2011) however might be that these estimates
appear rather sensitive to deposition model assumptions, as can be seen in Fig. 3.
—8—
Figure 3 about here
The four deposition models used to quantify dry deposition based on mean concentration
measurements were: CBED used in the United Kingdom; CDRY used by Environment
Canada; EMEP used in Europe under the Convention on Long Range Transboundary Air
Pollution (see also Tuovinen et al., this volume); and IDEM used in The Netherlands. All four
models use a surface-vegetation-atmosphere transfer scheme that quantifies a deposition
velocity vd for each gaseous and particulate component of dry deposition. The deposition rate
is then computed as vd times the mean concentration. Differences in the models to quantify
vd hence reflect differences in degree of complexity in how the canopy is modelled, and
differences in how atmospheric conditions (turbulence, stability) are treated by the model and
how stomatal uptake is computed (see also Cieslik, this volume).
A principal component analysis of total N dry deposition and all individual components (NH 3,
HNO3, NO2, NH4+ and NO3– ) obtained from the four deposition models is shown in Fig. 4.
Despite the large differences between models in their relative contributions to total N dry
deposition seen in Fig. 3 the PCA in Fig. 4 clearly reveals that in absolute numbers the model
differences are not that important: 96.7% of total variance in the dry deposition estimates for
forest ecosystems in Europe can be explained by the first principal component. Total N
deposition from the CBED model from the UK (Smith et al. 2000) almost perfectly represents
the first principal component (Fig. 4). This means that the CBED total is the best
representation for the whole set of deposition estimates. It is more strongly influenced by
NH3 deposition than the other three models and assumes much lower NO2 deposition rates
than the Dutch IDEM model (Bleeker et al., 2004; Erisman et al., 1994) and – even more
pronounced – the Canadian CDRY model (Zhang et al., 2003). Despite these functional
differences, the overall statistical average N dry deposition as a function of altitude (Fig. 5) is
surprisingly similar for the CBED and IDEM models. Deposition rates are however
substantially larger than what EMEP and CDRY predict from exactly the same measured
concentration data obtained from low-cost Delta denuder systems.
Figures 4 and 5 about here
4. Recent developments to assess effect on tree growth
At sites with long-term EC flux measurements the net CO2 flux can be directly related to N
deposition. For other sites where EC flux measurements are not feasible, some progress has
recently been made to use point-dendrometer data with high temporal resolution to monitor
growth rates over the season, which in the long term should provide similarly important data
to link N deposition to tree growth.
Stem radius variations of trees are a measure for biomass accumulation at the monthly,
seasonal and annual time scale (Zweifel et al. 2010). Figure 6 shows the concept of how tree
growth can be quantified from these measurements: on the annual time scale the time traces
—9—
of continuous high-resolution stem diameter measurements (Fig. 6a) and independent
cumulative net CO2 flux measurements performed by EC (Fig. 6b) are strikingly similar.
Figure 6 about here
4.1 Dendrometric measurements
The extra noise seen in the dendrometer data reflects the diurnal water cycle in the tree (see
Zweifel et al. 2010) which is unrelated to tree growth at short (daily, weekly) timescales,
whereas the signal at longer timescales is related to environmental factors, mostly
temperature trend and water availability, including bark and the degradation of dead phloem
cells, in turn affecting plant water balance and plant growth (see also Zweifel and Häsler,
2000, 2001; Cocozza et al., 2012). An important additional insight into tree growth is shown
during the winter where the natural shrinking of the bark is similar to the respiration losses
measured by EC outside the growing season. This signal adds value on top of conventional
simple tree-ring width measurements which only reflect growing-season conditions.
4.2 Net ecosystem productivity measurements
The CO2 sources (respiration) and CO2 sinks (assimilation) are large budget components of
opposite directions that sum to a relatively small NEP over a year (Barford et al., 2001;
Körner et al., 2005; Zweifel et al., 2010). As the net budget of these components is directly
measured with EC, it is generally assumed that the accuracy of the annual estimates of NEP
should be better than the accuracy of the difference between the estimates of the two large
components gross primary production (GPP) and total ecosystem respiration (TER)
determined separately (Buchmann and Schulze, 1999).
By installing dendrometers at EC research sites within Europe, we envisage gaining
dendrometric growth data of various tree species under different climates. These will then be
related to NEP measured by EC as a proof of concept that NEP at annual and possibly even
monthly resolution (during the growing season) could be obtained from representative
dendrometer measurements.
5. Policy Relevance of the Knowledge on Nitrogen Deposition
The reduction of N deposition is a matter of the UNECE Convention on Long Range
Transboundary Air Pollution (CLRTAP), which came into force in 1979. The protocol on
reduction of nitrous oxides (1988, Sofia) was the first based on the effect based approach, in
that it used the critical loads to calculate the required deposition reduction to reach
ecosystem protection. The Gothenburg protocol (UNECE 1999), the “Protocol to Abate
Acidification, Eutrophication and Ground-level Ozone”, is also effect based and regards also
the interactions between acidification, eutrophication and ozone. Since 2008 the Task Force
on reactive N under the Working Group on Strategies and Review of the CLRTAP is mandated
— 10 —
to develop a better understanding of the integrated, multi-pollutant nature of reactive N. The
main aim is the development of integrated solutions for N reductions. This aim requires
adequate research networks, and the development of supersites for integrated forest
ecosystem research (Mikkelsen et al., this volume) is an important step in this direction.
In Switzerland the effects of N are monitored at 9 Level II sites of ICP Forests and on cantonal
monitoring sites (IAP). At more than 15 of these sites, selected to form an N and water deficit
gradient (TreeNet-CH1), the long-term net ecosystem productivity (NEP) is measured with
point dendrometers. At two of these sites NEP is measured also with the eddy covariance
method (EC). An initial study by Zweifel et al. (2010) has shown excellent agreement between
EC-derived NEP and tree diameter increments (Section 4). This unique network started in
2011 and has since then delivered high-resolution dendrometer data and ancillary
meteorological data. It is envisaged to combine the TreeNet-CH network with the Swiss
Fluxnet network (Lägeren and Davos) and the regular forest inventories (National Forest
Inventory, NFI) to be able to compile quarterly if not even monthly interpolated maps of
actual tree growth throughout a growing season. This would allow us to see effects of N
inputs in combination with climate anomalies in due time. At the annual time scale such
information is expected to provide a more robust basis to estimate the forest CO2 budget, as
is expected in the national inventory reporting under the Kyoto Protocol. This will allow
estimating the mitigation potential of forests and its relationship with N inputs and other
environmental factors.
Acknowledgments
We thank Dr. Sabine Augustin from the Swiss Federal Office for the Environment for her
valuable contribution to this chapter, in particular to the policy relevance aspects. The
authors were supported by funds received from the Swiss Staatssekretariat für Bildung und
Forschung (contract SBF Nr. C10.0101) to participate in COST Action FP0903 “Climate
Change and Forest Mitigation and Adaptation in a Polluted Environment”, and by the Swiss
National Science Foundation (grant PDFMP3_132562). We also thank the Swiss Federal
Office for the Environment for funding the Swiss TreeNet network, and Dr. Roman Zweifel for
scientific collaboration and technical support with dendrometer measurements.
1
http://www.treenet.info
— 11 —
References
Aber, J. D., C. L. Goodale, S. V. Ollinger, M.-L. Smith, A. H. Magill, M. E. Martin, R. A. Hellett, J.
L. Stoddard (2003) Is nitrogen deposition altering the nitrogen status of northeastern forests?
BioScience, 53, 375–389.
Aber, J. D., K. J. Nadelhoffer, P. Steudler and J. M. Melillo (1989) Nitrogen saturation in northern
forest ecosystems. BioScience, 39, 378–386.
Aber, J. D., W. McDowell, K. J. Nadelhoffer, A. Magill, G. Berntson, M. Kamakea, S. McNulty, W.
Currie, L. Rustad and I. Fernandez (1998) Nitrogen saturation in northern forest ecosystems –
Hypotheses revisited. BioScience, 48, 921–934.
Aubinet, M., T. Vesala, and D. Papale (eds.) (2012) Eddy Covariance – A Practical Guide to
Measurement and Data Analysis. Dordrecht Heidelberg London New York: Springer. 449 pp.
Barford, C., S. Wofsy, M. Goulden, J. Munger, E. Pyle, S. Urbanski, L. Hutyra, S. Saleska, D.
Fitzjarrald, and K. Moore (2001) Factors controlling long-and short-term sequestration of
atmospheric CO2 in a mid-latitude forest. Science 294 (5547), 1688–1691.
Bleeker, A., G. J. Reinds, A. T. Vermeulen, W. de Vries, and J. W. Erisman (2004) Critical loads
and resent deposition thresholds of nitrogen and acidity and their exceedances at the level II
and level I monitoring plots in Europe. Technical report, ECN report ECN-C-04-117, Petten,
The Netherlands.
Bobbink, R. and J. P. Hettelingh (eds) (2010) Review and revision of empirical critical loads and
dose-response relationships. Coordination Centre for Effects, National Institute for Public
Health and the Environment (RIVM), The Netherlands, 243 pp.
Bobbink, R., K. Hicks, J. Galloway, T. Spranger, R. Alkemade, M. Ashmore, M. Bustamante, S.
Cinderby, E. Davidson, F. Dentener, B. Emmett, J.-W. Erisman, M. Fenn, F. Gilliam, A. Nordin,
L. Pardo, W. De Vries (2010) Global assessment of nitrogen deposition effects on terrestrial
plant diversity: a synthesis. Ecological Applications, 20, 30–59.
Bobbink, R., M. Hornung and J. G. M. Roelofs (1998) The effects of air borne nitrogen pollutants
on species diversity in natural and semi-natural European vegetation. Journal of Ecology, 86,
717–738.
Braun, S., V. F. D. Thomas, R. Quiring and W. Flückiger (2010) Does nitrogen deposition increase
forest production? The role of phosphorus. Environmental Pollution, 158, 2043–2052.
Buchmann, N. and E. D. Schulze (1999) Net CO2 and H2O fluxes of terrestrial ecosystems. Global
Biogeochemical Cycles 13, 751–760.
Burkard, R., P. Bützberger, and W. Eugster (2003) Vertical fogwater flux measurements above an
elevated forest canopy at the Lägeren research site, Switzerland. Atmospheric Environment, 37
(21), 2979–2990.
Cocozza, C., A. Giovannelli, B. Lasserre, C. Cantini, F. Lombardi, and R. Tognetti (2012) A novel
mathematical procedure to interpret the stem radius variation in olive trees. Agricultural and
Forest Meteorology, 161, 80–93.
De Schrijver, A., K. Verheyen, J. Mertens, J. Staelens, K. Wuyts, and B. Muys (2008) Nitrogen
saturation and net ecosystem production. Nature, 451, E1.
— 12 —
De Vries, W., G. J. Reinds, P. Gundersen, and H. Sterba (2006) The impact of nitrogen deposition
on carbon sequestration in European forests and forest soils. Global Change Biology, 12, 1151–
1173.
De Vries, W., S. Solberg, M. Dobbertin, H. Sterba, D. Laubhahn, G. J. Reinds, G.-J. Nabuurs, P.
Gundersen, and M. A. Sutton (2008) Ecologically implausible carbon response? Nature, 451,
E1–E3.
Dise, N. B., J. J. Rothwell, V. Gauci, C. van der Salm and W. de Vries (2009) Predicting dissolved
inorganic nitrogen leaching in European forests using two independent databases. Science of
the Total Environment, 407, 1798–1808.
Eilers, G., R. Brumme, and E. Matzner (1992) Above-ground N-uptake from wet deposition by
Norway spruce (Picea abies Karst.) Forest Ecology and Management, 51, 239–249.
Ellenberg, H., H. E. Weber, R. Düll, V. Wirth, W. Werner, and D. Paulissen (1991) Zeigerwerte von
Pflanzen in Mitteleuropa. Vol. 18. Series Scripta Geobotanica. Göttingen: Erich Goltze. 248 pp.
Ellenberg, jun., H. (1990) Ökologische Veränderungen in Biozönosen durch Stickstoffeintrag. In:
Ammoniak in der Umwelt – Kreisläufe, Wirkungen, Minderung. Kuratorium für Technik und
Bauwesen in der Landwirtschaft, Darmstadt, pp. 44.1–44.24.
Elling, W., U. Heber, A. Polle, and F. Beese (2007) Schädigung von Waldökosystemen:
Auswirkungen anthropogener Umweltveränderungen und Schutzmassnahmen. München:
Elsevier Spektrum Akademie-Verlag. 422 pp.
Emmett, B. (2007) Nitrogen saturation of terrestrial ecosystems: some recent findings and their
implications for our conceptual framework. Water, Air and Soil Pollution, 7, 99–109.
Erisman, J. W., A. van Pul, and P. Wyers (1994) Parametrization of surface resistance for the
quantification of atmospheric deposition of acidifying pollutants and ozone. Atmospheric
Environment, 28, 2595–2607.
Eugster, W. (2008) Fog research. Die Erde 139 (1–2), 1–10.
Eugster, W., S. Perego, H. Wanner, A. Leuenberger, M. Liechti, M. Reinhardt, P. Geissbühler, M.
Gempeler, and J. Schenk (1998) Spatial Variation in Annual Nitrogen Deposition in a Rural
Region in Switzerland. Environmental Pollution, 102 (S1), 327–335.
Ferrara, R., B. Loubet, P. Di Tommasi, T. Bertolini, V. Magliulo, P. Cellier, W. Eugster, and G.
Rana (2012) Eddy covariance measurement of ammonia fluxes: Comparison of high frequency
correction methodologies. Agricultural and Forest Meteorology, 158, 30–42.
Flechard, C. R., E. Nemitz, R. I. Smith, D. Fowler, A. T. Vermeulen, A. Blecker, J. W. Erisman, D.
Simpson, L. Zhang, Y. S. Tang, and M. A. Sutton (2011) Dry deposition of reactive nitrogen to
European ecosystems: a comparison of inferential models across the NitroEurope network.
Atmospheric Chemistry and Physics, 11, 2703–2728.
Flückiger, W. and S. Braun (1999) Nitrogen and its effects on growth, nutrient status and parasite
attacks in beech and Norway Spruce. Water, Air and Soil Pollution, 116, 99–110.
Goll, D. S., V. Brovkin, B. R. Parida, C. H. Reick, J. Kattge, P. B. Reich, P. M. van Bodegom and U.
Niinemets (2012) Nutrient limitation reduces land carbon uptake in simulations with a model
of combined carbon, nitrogen and phosphorus cycling. Biogeosciences, 9, 3547–3569.
Gundersen, P. (1998) Effects of enhanced nitrogen deposition in a spruce forest at Klosterhede,
Denmark, examined by NH4NO3 addition. Forest Ecology and Management, 101, 251–268.
— 13 —
Gundersen, P., I. Callesen and W. de Vries (1998) Nitrate leaching is related to forest floor C/N
ratios. Environmental Pollution, 102, 403–407.
ICP Forests (2010) Manual on methods and criteria for harmonized sampling, assessment,
monitoring and analysis of the effects of air pollution on forests, Part XIV: Sampling and analysis
of Deposition. Hamburg, vTI, 66 pp., http://www.icp-forests.org/pdf/FINAL_Depo.pdf
Jeffries D., C. Cox and P. Dillen (1979) Depression of pH in lakes and streams in central Ontario
during snowmelt. Journal of the Fisheries Research Board, Canada, 36, 640–646.
Khanna, P. K., J. Bauhus, K. J. Meiwes, M. Kohler, S. Rumpf, E. Schönfelder (2007) Assessment
of changes in the phosphorous status of forest ecosystems in Germany – Literature review and
analysis of existing data. Report to the German Federal Ministry of Food, Agriculture and
Consumer Protection.
Körner, C., R. Asshof, O. Bignucolo, S. Hattenschwiler, S. G. Keel, S. Pelaez-Riedl, S. Pepin, R. T.
W. Siegwolf, and G. Zotz (2005) Carbon flux and growth in mature deciduous forest trees
exposed to elevated CO2. Science, 309 (5739), 1360–1362.
Krupa, S. V. (2003) Effects of atmospheric ammonia (NH3) on terrestrial vegetation: a review.
Environmental Pollution, 124, 179–221.
Kyrö, E.-M., T. Grönholm, H. Vuollekoski, A. Virkkula, M. Kulmala, and L. Laakso (2009) Snow
scavenging of ultrafine particles: field measurements and parameterization. Boreal
Environment Research, 14, 527–538.
Laubhann, D., H. Sterba, G.-J. Reinds, and W. De Vries (2009) The impact of atmospheric
deposition and climate on forest growth in European monitoring plots: An individual tree
growth model. Forest Ecology and Management, 258, 1751–1761.
Nave, L. E. and P. S. Curtis (2011) Uptake and partitioning of simulated atmospheric N inputs in
Populus tremuloides – Pinus strobus forest mesocosms. Botany, 89, 379–386.
Nellemann, C. and M. G. Thomsen (2001) Long-term changes in forest growth: potential effects of
nitrogen deposition and acidification. Water, Air, and Soil Pollution, 128, 197–205.
Nihlgard, B. (1985) The ammonium hypothesis – an additional explanation to the forest dieback in
Europe. Ambio, 14, 1–8.
Nordin, A., J. Strengom and L. Ericson (2006) Responses to ammonium and nitrate additions by
boreal plants and their natural enemies. Environmental Pollution, 141, 167–174.
Prietzel J., K. E. Rehfuess, U. Stetter and H. Pretzsch (2008) Changes in soil chemistry, stand
nutrition, and stand growth at two Scots pine (Pinus sylvestris L.) sites in Central Europe
during 40 years after fertilization, liming, and lupine introduction. European Journal of Forest
Research, 127, 43–61.
Schaefer, D. A., and W. A. Reiners (1990) Throughfall chemistry and canopy processing
mechanisms. In: Lindberg, S. E., A. L. Page and S. E. Norton (eds.) Acidic precipitation, Vol. 3:
Sources, Deposition, and Canopy Interactions. New York: Springer, pp. 241–284.
Schulze, E.-D. (1989) Air pollution and forest decline in a spruce (Picea abies) forest. Science, 244,
776–783.
Sievering, H., T. Tomaszewski, and J. Torizzo (2007) Canopy uptake of atmospheric N deposition
at a conifer forest: Part I – Canopy N budget, photosynthetic efficiency and net ecosystem
exchange. Tellus, 59B, 483–492.
— 14 —
Slootweg, J., M. Posch, and J.–P. Hettelingh (eds) (2010) Progress in the modelling of critical
thresholds and dynamic modelling, including impacts on vegetation in Europe. CCE Status
Report 2010, Coordination Centre for Effects, http://www.rivm.nl/cce
Smith, R. I., D. Fowler, M. A. Sutton, C. Flechard, and M. Coyle (2000) Regional estimation of
pollutant gas deposition in the UK: model description, sensitivity analyses and outputs.
Atmospheric Environment, 34, 3757–3777.
Sparks, J. P. (2009) Ecological ramifications of the direct foliar uptake of nitrogen. Oecologia, 159,
1–13.
Spieker, H., K. Mielkäinen, M. Köhl and J. Skovsgaard (eds.) (1996) Growth Trends in European
Forests: Studies from 12 Countries. EFI Research Report No. 5, Berlin: Springer-Verlag, 372 pp.
Spranger, T., U. Lorenz, and H.-D. Gregor (eds.) (2004) Manual on methodologies and criteria for
modelling and mapping critical loads & levels and air pollution effects, risks and trends. Berlin:
Federal Environmental Agency, Texte, 52 (4), 266 pp.
Thalmann, E., R. Burkard, T. Wrzesinsky, W. Eugster, and O. Klemm (2002) Ion fluxes from fog
and rain to an agricultural and a forest ecosystem in Europe. Atmospheric Research, 64, 147–
158.
Thimonier, A., M. Schmitt, P. Waldner, and B. Rihm (2005) Atmospheric deposition on Swiss longterm forest ecosystem research (LWF) plots. Environmental Monitoring and Assessment, 104,
81–118.
UNECE (1999) The 1999 Gothenburg Protocol to Abate Acidification, Eutrophication and Groundlevel Ozone. http://www.unece.org/fileadmin/DAM/env/lrtap/full
%20text/1999%20Multi.E.Amended.2005.pdf
UNECE (2005) Development, modelling and mapping of critical loads and their input data.
http://www.unece.org/fileadmin/DAM/env/documents/2005/eb/wg1/EB.AIR.WG.1.2005.10.
e.pdf
Vitousek, P. M., C. B. Field (2001) Input/output balances and nitrogen limitation in terrestrial
ecosystems. In: Schulze E.-D. et al., Global Biogeochemical Cycles in the Climate System. San
Diego: Academic Press, pp. 217–225.
Vitousek, P. M., J. D. Aber, R. W. Howarth, G. E. Likens, P. A. Matson, D. W. Schindler, W. H.
Schlesinger, and D. G. Tilman (1997) Human alteration of the global nitrogen cycle: sources
and consequences. Ecological Applications, 7 (3), 737–750.
Wortman, E., T. Tomaszewski, P. Waldner, P. Schleppi, A. Thimonier, W. Eugster, N. Buchmann,
and H. Sievering (2012) Atmospheric nitrogen deposition and canopy retention influences on
photosynthetic performance at two high nitrogen deposition Swiss forests. Tellus B, 64, 1–14.
Zhang, L., J. R. Brook, and R. Vet (2003) A revised parameterization for gaseous dry deposition in
air-quality models. Atmospheric Chemistry and Physics, 3, 2067–2082.
Zweifel, R. and R. Häsler (2000) Frost-induced Reversible Shrinkage of Bark of Mature Subalpine
Conifers. Agricultural and Forest Meteorology, 102, 213–222.
Zweifel, R. and R. Häsler (2001) Dynamics of Water Storage in Mature Subalpine Picea abies:
Temporal and Spatial Patterns of Change in Stem Radius. Tree Physiology, 21, 561–569.
Zweifel, R., W. Eugster, S. Etzold, M. Dobbertin, N. Buchmann, and R. Häsler (2010) Link between
continuous stem radius changes and net ecosystem productivity of a subalpine Norway spruce
— 15 —
forest in the Swiss Alps. New Phytologist, 187, 819–830.
— 16 —
Figures
Fig. 1: Schematic showing the similarities and differences between nutrient and pollutant
inputs to ecosystems by rainfall (left) and by fog deposition (right). Reprinted from Eugster
(2008).
— 17 —
Fig. 2: Delta denuder system for sampling gaseous NH3, NO2, HNO3, HONO, SO2, HCl and
aerosol NH4+, NO3–, SO42–, Cl–, Na+, Ca2+, Mg2+ (a) on site operational at Lägeren mixed forest in
Switzerland; (b) denuder sample train which is exchanged every month; (c) interior of housing
with door open. Numbers in (c) are: 1, durable casing; 2, sampling inlet (funnel); 3,4,
sampling train; 5,7, tube connection; 6, gas volume flow meter ; 8, in-line large particle filter;
9, air pump with heater; 10–13, electrical installation. Modified after
http://www.nitroeurope.eu/sites/nitroeurope.eu/files/neu_data/Disseminationdeliverables/NEU_delta_ws_tfmm_Helsinki_may2006_presentation.pdf
— 18 —
Fig. 3: Relative contributions of reactive N species to total inorganic N dry deposition to 29 European forest ecosystems (F) in comparison
with 9 semi-natural grasslands (SN), 8 cut or grazed grasslands (G), and 9 croplands (C). Deposition was calculated with measured monthly
mean concentrations of NH3, HNO3, NO2, NH4+ and NO3– using four different deposition models (DBED, CDRY, EMEP, IDEM). Location names
start with an abbreviation for the country followed by a 3-character location acronym. Reprinted from Flechard et al. (2011).
— 19 —
Fig. 4: Principal component analysis (covariance-based PCA) of dry deposition estimates for
29 European forest ecosystems based on data published by Flechard et al. (2011). Deposition
was calculated with measured monthly mean concentrations of NH3, HNO3, NO2, NH4+ and
NO3– using four different deposition models (DBED, CDRY, EMEP, IDEM). The graph shows
the first two principal components explaining 98.4% of total variance (96.7% and 1.7% for
the first and second component, respectively). This indicates that total deposition (bold
vectors) strongly depends on assumptions made by the deposition model. With the British
CBED gaseous deposition via NH3 is by far the most relevant dry deposition pathway,
whereas the Canadian CDRY model more strongly emphasizes NO2 dry deposition. Unnamed
grey arrows represent all other deposition pathways that do not lead to substantial
contributions to total dry deposition by any of the models used by Flechard et al. (2011).
— 20 —
Fig. 5: Altitude dependence of dry deposition (log scale) of total inorganic N to 29 European
forest ecosystems determined from four deposition models that use measured monthly mean
concentrations. All models show a local maximum for sites with an altitude around 800 m
a.s.l., indicating the specific relevance of N deposition to montane forest ecosystems. Data
were taken from Flechard et al. (2011).
— 21 —
Fig. 6: The time traces of dendrometer measurements DR on single trees (a) and of
ecosystem-scale eddy covariance measurements of net ecosystem productivity NEP (b) are
very similar in a subalpine Norway spruce forest at Davos, Switzerland. Both panels show the
full year 2003 with the last months of 2002 and the first months of 2004. Using such
relationships it should become possible to see effects of pollutants and nitrogen deposition at
the whole-tree scale and at time resolutions well below the annual resolution of tree ring
widths that have been used so far. Reprinted from Zweifel et al. (2010). Current research
focuses on how the relationship between DR (a) and NEP (b) holds for other forest types in
different climatic zones of Europe.
— 22 —

Nutrients or pollutants? Nitrogen deposition to

Transcription

Similar documents

Deposition Boot Camp: The Dos and the Don`ts

Use of MEM for Protection of Karst Cave – Example from Križna Jama

Weathering, Erosion, or Deposition?

Overview and abstracts of lectures

Progress in Nitrogen Deposition Monitoring and Modelling

Constructive Processes

Chicago to target absent teachers

Presentation - National Atmospheric Deposition Program