Download Chapter 2 The effects of nitrogen deposition on the structure and

2 The effects of nitrogen deposition
structure and functioning of ecosystems
on
the
Bobbink, R., D. Bal, H.F. van Dobben, A.J.M. Jansen, M. Nijssen, H. Siepel, J.H.J.
Schamineé, N.A.C. Smits & W. de Vries
2.1 I NTRODUCTION
Human activities threaten the structure and functioning of natural and semi-natural ecosystems
in many ways, and thus also the associated diversity of plant and animal species.
One of the main anthropogenic threats in our temperate climate zone is the increased air
pollution by both reduced and oxidized nitrogen compounds in the form of NHx and NOy (e.g.
Sala et al. 2000; Galloway & Cowling 2002; Bobbink et al. 2010a). Also in the Netherlands it is
recognized that high nitrogen deposition is a major limiting factor to maintain or restore a good
state of conservation in sensitive natural areas.
As such, nitrogen (N) is not a problem. On the contrary, it is one of the essential building blocks
for life on earth. The problem lies in the extent to which this element is added to our
environment in reactive form. For centuries, only organic fertilizer or other organic materials were
used to increase the agricultural production, while later also guano or ‘Chile saltpeter' (recovered
from bird droppings) were used. This situation changed after the invention of synthetic
conversion of the inert molecular nitrogen (N2) into the reactive ammonia by Fritz Haber in 1909
and the industrial scaling of that by Carl Bosch (both got the Nobel Prize). This 'Haber-Bosch
process’ made large-scale fertilizer production possible and its use very much increased after
1920. In the course of the twentieth century, increasing amounts of fertilizer were used to
increase agricultural production. Cattle breeding also increasingly intensified as a result. The
downside was that more and more reactive nitrogen disappeared from the agricultural system to
ground or surface water or through air emissions.
This chapter first covers some basic principles with regard to emission, transport and deposition
of nitrogen compounds. This includes the influence of vegetation on the deposition process,
including the edge effect. This is followed by a discussion of the main effects of nitrogen
deposition that are currently still relevant. These are N-eutrophication, acidification of soil or
water and the negative effects of reduced nitrogen. The effects of nitrogen deposition on fauna
are then discussed. These effects relate to the invertebrates in particular, as these tend to be
highly dependent on the structure and composition of the vegetation. It will be made clear that
the relationships in ecosystems are complex. There usually are no simple dose-response
relationships. The focus will be on the ecological and chemical processes, so that the significance
of the recovery strategies discussed in chapter 3 can be better understood.
The intermezzo directly following this chapter will specifically cover complex processes in damp
and wet ecosystems (biogeochemical mechanisms in wet ecosystems). These processes can pose
other significant impediments for habitat restoration. This intermezzo explains hydrological
recovery measures, which are not primarily directed against the effects of nitrogen.
39
Part I - November 2012 version -
2.2 E MISSION , TRANSPORT AND DEPOSITION OF NITROGEN COMPOUNDS
The discharge (emission) of air pollution in Western Europe has significantly increased in the
course of the twentieth century. Sulphur dioxide (SO2) was the main component of air pollution
until the late seventies of the last century. After that nitrogen compounds became increasingly
important, both relatively and absolutely. Nitrogen oxides (NOx: especially NO2 and NO) arise
mainly from the combustion of fossil fuels in industry, power plants, heating systems and traffic.
At the moment, the main source of nitrogen oxides is (freight) traffic. Ammonia gas (NH3) is
mainly liberated by volatilization from manure and urine during grazing, in the shed or in
storage, and previously when the manure was spread on the field. Other sources are industry -
where ammonia is liberated during some manufacturing processes or when the NOx-installations
malfunction -, car traffic and the storage of effluent.
2.2.1 CHEMICAL CONVERSIONS IN THE ATMOSPHERE
Nitrogen oxides and ammonia end up in the atmosphere after emission. There, these substances
undergo chemical reactions under the influence of sunlight and the presence of other materials.
The atmospheric conversions can, among other things, lead to the formation of nitrate (NO3-),
ammonium (NH4+) and nitric acid (HNO3, gas phase). NH4+ and NO3- are the dominant N-
containing ions in the water-particles in clouds or fog and in precipitation. Ammonia (NH3) reacts
rapidly with acids, producing ammonium salts. Here, small solid particles (so-called aerosols) are
formed, such as (NH4)2SO4 or NH4NO3, whereby nitric acid (HNO3) and sulphuric acid (H2SO4) are
neutralized by NH3. The conversion into aerosols is of importance for the distance across which
the substances are transported. In short, nitrogen compounds are found in the atmosphere as
gas, ion and aerosol.
2.2.2 TRANSPORT AND DEPOSITION
Once airborne, the emitted gas is carried along by the wind and spread quickly. This can be
compared with a plume of smoke from a campfire: the plume is carried along by the wind and
dissipated simultaneously. After a few hundred meters, the smoke is no longer visible, while it
can still be smelled. Dilution of the concentrations of substances occurs rapidly in the air. All
kinds of chemical conversions take place there as well (see above). The various gases and
aerosols can thus cover great distances in the atmosphere under the influence of air currents. A
complex of different factors determines how far the different components get. Especially
important are the emission level, the atmospheric conditions (speed of air currents, turbulence,
etc.), the rate of chemical reactions, the deposition rate of the particular compound, and the
nature and roughness of the earth's surface with its vegetation. Eventually all those substances
will end up on the earth’s surface. This process is called deposition, and it can occur in various
ways (Diagram 2.1).
40
Part I - November 2012 version -
Diagram 2.1 Schematic representation
of the deposition of air pollutants
(Erisman 1990).
Gassen = gases
Aerosolen = aerosols
Aanhechting = attachment
Absorptie = absorption
Wolken = clouds
Neerslag = precipitation
Natte depositie = wet deposition
Droge depositie = dry deposition
Vegetatie = vegetation
Doorval =throughfall
Stamafstroom = stemflow
Bodem = soil
The direct deposition or absorption of gases or aerosols from the atmosphere to the earth's
surface (soil, water or vegetation) is called dry deposition. A measure of the speed of this process
is the dry deposition velocity (Vd): the higher it is, the faster the gas or the particle is removed
from the atmosphere. Thus, the transportation distance of NH3 is short due to the high Vd of this
gas, while the transportation distance of the ammonium aerosol is much higher due to its lower
Vd. A large proportion of the NO2 is discharged by traffic at a low level. However, due to the low
deposition velocity of NO2 and the low conversion rates into gases HNO3 and HNO2, both of which
have a high deposition velocity, NO2 is nevertheless often transported over long distances. For
more background information on the (dry) deposition process, see e.g. Erisman (1992), Asman et
al. (1998) and Fowler (2002).
In addition to the direct deposition of gases and particles on the surface, the dry deposition just
described, two other processes play a part in the removal of nitrogen from the atmosphere. These
are the solution in clouds or rain and subsequent precipitation (wet deposition; see Diagram 2.1)
and the deposition of cloud or fog drops directly onto the vegetation or soil (clouds or fog
deposition, also called occult deposition). In the Netherlands, the latter process contributes
almost nothing (<1-2%) to the total deposition of N compounds from the atmosphere. Wet
deposition provides about 25-30% of the total N deposition. The rest is dry deposition (De Haan
et al. 2008).
The total nitrogen deposition in the Netherlands increased significantly after 1950 until the end
of the eighties of the last century (Diagram 2.2). This was caused by increased intensive stock
farming and the use of fossil fuels. From 1993, a decrease (30-40 %) in the deposition of reduced
N (ammonia and ammonium) was realised through various measures. From 2003 up to the
present the values have remained roughly constant. The deposition of oxidized N (nitrogen oxide)
41
Part I - November 2012 version -
also increased after 1945, but levelled off earlier than the deposition of reduced N. After 1980 to
1985 a very gradual decline has been observed. For about three to four decades reduced N has
been the predominant form (> 75%) of nitrogen deposition in Dutch nature areas (De Haan et al.
2008).
Diagram 2.2 The average annual deposition of reduced and oxidized N in the Nederland in the period
1900-2004 reduced from historical emission sources (according to Noordijk 2007 from De Haan et al.
2008).
At present, more than half of the nitrogen deposition in the Netherlands comes from domestic
sources, the rest comes from abroad or from "natural" sources (Diagram 2.3). Agriculture is the
economic sector, which by far contributes the most to nitrogen deposition from the Netherlands
(46% in 2004).
42
Part I - November 2012 version -
Diagram 2.3 Contribution (%) of the various sources to the total nitrogen deposition in the Netherlands
in 2004 (From De Haan et al. 2008).
Afkomstig uit Nederland = from the Netherlands
Landbouw = agriculture
Overig = other
Afkomstig uit buitenland = from abroad
Natuurlijke achtergrond = natural background
The deposition of ammonia from an agricultural source is high in the vicinity of that source in
comparison with the deposition at a greater distance from that source. Starting from a point
source of ammonia at a 3 m height (for example, a shed), about 20% of the released ammonia
comes down within a distance of one kilometre from the source. On this scale, it almost
exclusively concerns dry deposition of ammonia. Wet deposition only plays a significant part after
10 kilometres. After a 100 km transport about 60% of the released ammonia has been deposited.
There is, however, also transport over very long distance: after a 1,000 km about 20% of the
original ammonia is still in the atmosphere in some form (e.g. Kros et al. 2008).
Due to the spatial distribution of the sources and the various transmission and conversion
processes in the atmosphere, the deposition of N-compounds is not the same everywhere
(Diagram 2.4). Even in a small country like the Netherlands there are major differences: for
example, the total deposition of NOy (the sum of dry and wet deposition of NO + NO2 + HNO3) is
clearly higher in urban areas (including in the west of the country), while the total deposition of
NHx (the sum of dry and wet deposition of NH4+ en NH3) is higher in rural areas, where the
highest values are found in the Peel region, the Gelderland Valley, Twente and the Achterhoek.
43
Part I - November 2012 version -
Diagram 2.4 Geographic spread of the total nitrogen deposition (mol/ha) in the Netherlands in 2006.
This map has been made using the Operational Priority Substances (model) of the RIVM/PBL.
2.2.3 THE INFLUENCE OF VEGETATION ON THE DEPOSITION PROCESS
The nature of the vegetation also plays a part in the dry air deposition. In general, a "rougher"
vegetation has a more filtering effect. The rougher surface provides more turbulence in the air
just above the vegetation. Due to this turbulence, the gases or aerosols travel a more winding
road as it were, instead of a straight road as is the case above a smooth surface. As a result, the
particles are found above the interface of the vegetation and the air for a longer period of time,
and with that the chance of deposition or absorption, thus dry deposition, increases. The above
example relates to both the dry deposition of NHy as that of NOx (and SOx).
In general, the higher a vegetation is, the more dry deposition takes place. The aerodynamic
roughness and with that the filtering effect and dry deposition velocity increase in the following
order: open water / bare soil < low grassland < high grassland < heath < shrubs < broad-leaved
forest < pine forest. In short, forests have greater aerodynamic roughness than low vegetation,
so that there is a higher dry deposition of nitrogen than when the same area would have been
covered by grassland or heath land. In other words, in this way, with equal concentrations of NHx
and NOx, a lot of extra N can be found in forests. Nevertheless, the dry deposition within one
physiognomic group, such as forests, is not at all always the same. The fact is that the dry
deposition is not only determined by the height of the vegetation, but also influenced by other
characteristics of the forest that affect the aerodynamic roughness. Examples include the
structure of the crowns (homogeneous or not), tree density, tree species, differences in leaf
characteristics or the total leaf area (Van Ek & Draaijers 1991, Draaijers 1993, Wuyts et al.
2008b).
44
Part I - November 2012 version -
Although it was believed until the nineties of the last century that low vegetation, such as
grassland, would barely capture extra dry deposition compared to bare soil or open water, field
research in several countries, including the Netherlands, has shown that dry deposition in low
vegetations is not negligible. With increasing Leaf Area Index (LAI), which is to say, the total leaf
area of the vegetation per unit of soil surface, the dry deposition of S-and N-compounds in
grassland significantly increases (Heil et al. 1988; Bobbink et al. 1990). The increase in dry
deposition occurred at LAI values above 1.5-2, while in highly productive grasslands with an LAI
of 4-6 the dry deposition is almost as high as in some forests. In fact, therefore, the larger leaf
area low vegetation has, the more air turbulence and filtration occurs, resulting in a greater dry
deposition of substances from the atmosphere. A portion of the nitrogen compounds deposited
from the air is immediately absorbed by plants above ground through the leaves, particularly in
the nutrient-poor to moderate nutrient-rich conditions.
EDGE EFFECT
Research in both coniferous and broad-leaved forests in the 1980s showed that transitions in the
landscape, such as from low vegetation (e.g. grassland or heathland) to forest, also greatly affect
the dry deposition process of air pollutants (e.g. Ivens 1990; Draaijers 1993). The dry N
deposition in an approximately 10-20 m wide woodland fringe is significantly higher (on average
1.5 x) than that in the centre of the forest and it decreased exponentially to the regular
deposition values in the centre of the forest. The dry deposition is measurably increased in an
edge zone, which is comprised of the first five edge heights. This enhanced deposition of N- and
S-compounds was most prominently found in woodland fringes that were exposed to the
prevailing wind directions (Draaijers 1993). Recently, a similar study was conducted in Flanders. It
showed that the nitrogen deposition in woodland fringe zones is 0.5 to almost 4 times higher
than that in the centre of the forest (median values: 1.5 for NH4 and 1.6 for NO3) (De Schrijver et
al. 2007; Wuyts et al. 2008a).
The zone in which elevated levels were detected was between 15 to over 100 metres wide
(median value: 50 m), within which most was captured in the first section. Apart from that, an
increased N-capture also takes place in the short vegetation right next to the edge of the forest
(Wuyts et al. 2009). Because forests in the Netherlands are often very fragmented and small, at
least half of the forests can be considered as woodland fringe zones with enhanced deposition
(Draaijers 1993). Moreover, in almost all N-deposition models, including those for the PAS
(AERIUS), transitions between forests and low vegetation are not taken into account. After all,
fixed parameters for each vegetation class (e.g. forest or heath) are used for calculations. Because
of the edge effect the atmospheric nitrogen deposition on forests is therefore often
underestimated (Draaijers 1993).
In addition, the following aspects were quantified in the recent Flemish research (Wuyts 2009):
- It makes a big difference whether there is a sharp transition from low vegetation to forest or a
woodland fringe that is gradually increasing in height (in the form of a zone of saplings or a
gradual edge vegetation): the additional deposition in the woodland fringe is significantly greater
when there is a sharp transition compared to when there is a gradual transition, and the zone in
which the edge effect occurs is wider when there is a sharp transition compared to when there is
a gradual one (see also Table 3 in Wuyts et al. 2009).
- The width of the zone in which the edge effect occurs is smaller when there is a dense forest
45
Part I - November 2012 version -
compared to when there is sparse forest, but the extent of the edge effect (i.e. the amount of
additional capture in the edge zone) is larger in the case of a dense coniferous forest compared
to in a sparse coniferous forest (Chapter 6 in Wuyts 2009).
When taking management and restoration measures in the context of the PAS, this 'deposition
behaviour’ can possibly be taken into account. To reduce the capture of nitrogen, the following
measures concerning structure are imaginable:
• to capture nitrogen outside sensitive habitats by maintaining or developing dense, high forests
(preferably coniferous forests) with a sharp woodland fringe in the zone between the nitrogen
sources and sensitive habitats (the forest then serves as a buffer zone);
• to develop edge and mantle vegetation as a gradual transition between low vegetation and the
woodland fringe zones that need to be protected (this reduces deposition in the wood fringe
zone);
• if a gradual transition cannot be realized, then concentrate the quality improvement of forests
on areas that are not part of the woodland fringe zone, or - if the woodland fringe zone is
valuable – thin this zone so that relatively little additional nitrogen is captured in this zone.
2.3 Different effects of nitrogen deposition
Six different effects of nitrogen deposition can be distinguished (see Chapter 1). The current
chapter will elaborate on the three main effects. These are eutrophication by gradual increase of
N-availability, acidification of soil and water and the negative effects of the increased availability
of reduced N (ammonium).
2.3.1 N-EUTROPHICATION
The increased uptake of nitrogen by the vegetation usually causes an increase in biomass
production of plant parts aboveground in particular, seeing that the productivity of many
terrestrial ecosystems in areas with a temperate climate, such as in Western Europe, is in principle
limited by nitrogen (e.g. Vitousek & Howarth 1991; LeBauer & Treseder 2008). Due to this
increase in productivity the plant litter production is also increased, and gradually also the release
of nitrogen compounds from organic material (Diagram 2.5). This process is called
mineralization. Because the nitrogen cycle is closed in many natural situations of our climate
zone, this means that the nitrogen inventory in the system is ever increasing due to Ndeposition, and so making more nitrogen available due to the higher mineralization (Aerts &
Bobbink 1999).
Only after a prolonged high nitrogen supply is there increased leaching from the upper soil layers
to the groundwater. Also, by this, various ecosystems will no longer have become nitrogenlimited but phosphorus-limited. Furthermore, the succession accelerates due to the increased
nitrogen availability, in particular in systems that are still developing. In this way, accelerated tree
growth can occur, such as a strong increase of birch in peat moors, but also more and rapid
shrubs formation in the dry dunes.
46
Part I - November 2012 version -
Diagram 2.5 Relationship between the nitrogen mineralization rate in the soil of a species-rich Nardus
grassland (squares) and a lime-rich grassland (circles) in Derbyshire (England) which were irrigated
with different nitrogen amounts during a period of 3 years (Morecroft et al. 1994).
SHIFTS IN COMPETITION
In time, the increase in the availability of nitrogen will cause shifts in the competition between
plant species and thus to changes in the dominance of species and the species composition of
the vegetation. Nitrogen-loving, fast-growing plant species thus gradually displace the
characteristic, less competitive species and will eventually dominate the vegetation. Recently it
has been shown experimentally that the reduced light penetration in grassland vegetation is the
determining factor (Hautier et al. 2009). A large part of the species in semi-natural and natural
ecosystems is adapted to a low nitrogen availability in soil (Ellenberg 1988). As a result thereof,
the wealth of plant species will in the long term decline in nutrient-poor to moderately nutrient-
rich systems. Low-growing herbs, such as rosette plants and short-lived species, nitrogen-fixing
plants such as papilionaceous flowers, species with traditionally small populations and lichens
will decrease, whereas nitrogen-loving species, including various gramineae, will dominate (e.g.
Bobbink et al. 1998, Clark et al. 2007). Examples of species which make up grassy or rugged
ecosystems are Wavy Hair-grass (Deschampsia flexuosa)) and Purple Moor Grass (Molinia
caerulea) in heathlands, Heath False Brome (Brachypodium pinnatum) in calcareous grasslands,
Bushgrass (Calamagrostis epigejos) and European beachgrass (Ammophila arenaria) in the dunes
and blackberries (including Rubus fruticosus aggr.), Wavy Hair-grass and European elderberry
(Sambucus nigra) in forests (Diagram 2.6 and 2.7) (e.g. Dise et al. 2011). In short, rare species are
becoming increasingly rare and common species more common. The vegetation composition also
becomes more homogeneous (‘homogenisation’), because the micro variability in light
penetration and nutrient concentrations (nitrate and ammonium) are greatly reduced by the above
47
Part I - November 2012 version -
processes. This also usually means a decrease in habitat specific species, and a consequent lower
local quality of the habitat type.
Diagram 2.6 Image of calcareous grassland vegetation without an extra N-donation (left) and after a 3
years’ treatment with 100 kg N ha-1 yr-1 (right) (pictures: R. Bobbink). Through N-eutrophication, the
growth of the grass Heath False Brome (Brachypodium pinnatum) increased very dramatically under the
mowing management of that time, causing the disappearance of many low herbs from the vegetation
(Bobbink 1991).
48
Part I - November 2012 version -
Diagram 2.7 Cover of Rubus fruticosus aggr. in Swiss forest observation plots in relation to the total
nitrogen deposition: the cover is usually substantially higher when the nitrogen deposition is higher
than 20-25 kg N per ha per year (from Flückiger & Braun 2004).
In addition to the effects on higher plants, an excessive supply of nitrogen also causes a
deterioration of fungi, in particular of the ectomycorrhizal species (Arnold 1991). In the
Netherlands, these are mainly forest mushrooms that live in symbiosis with the roots of trees.
The ectomycorrhizas are, inter alia, of great importance for the nutrient uptake of trees. Many
species of this group have become rare or have almost disappeared from our forests. Also, the
charge of ectomycorrhizal species on tree roots decreases due to increasing nitrogen supplies.
This potentially has adverse effects on the growth and development of the trees themselves
(Bobbink et al. 2003).
Another effect that occurs is the change in microclimate. Because of earlier plant growth the soil
warms up more slowly in the spring, slowing down the development of, for example, butterflies,
which made it through winter as an egg or caterpillar. This can lead to a mismatch with the
vegetation growth (food for caterpillars), to lower survival and smaller populations (Wallis de Vries
& Van Swaay 2006). See paragraph 2.4.1.
DECREASE IN SPECIES RICHNESS
The increase in the availability of nitrogen can not only lead to changes in the species
composition of vegetation, but can often also cause a decrease in species richness. For example,
a recent comparison of plant species with various abiotic factors in dry low-nutrient grasslands
(habitat type 6230) in nine countries across Europe revealed that species richness is mainly
related to the total N deposition, and that this relationship is primarily negative: as N deposition
increases, the number of species decreases (Figure 2.8; Stevens et al. 2010). Dorland & Van Loon
2011 analyse the data in this figure in more detail. They propose a sigmoid curve, in which an
average of 20 species in cases of low deposition falls to an average of 11 species in cases of high
deposition. This decrease mainly takes place between 12 and 29 kg N/ha/yr (after which the
number of species stabilises). The extinct species are primarily those which are characteristic of a
particular habitat type, such that the vegetation remaining in cases of high deposition often no
longer meets the definition of that habitat type. This means that loss of quality may ultimately
also lead to surface loss.
Habitat types that are naturally acidic, are often characterised by species that are highly resistant
to acidification and toxicity caused by ammonium and aluminium, and these types are therefore
particularly sensitive to the long-term effects of eutrophication, since the nitrogen availability
was originally very low.
49
Part I - November 2012 version -
Figure 2.8 The relationship between the number of plant species in low-nutrient grasslands (Violion
Caninae) as they appear across Europe and the N deposition. The negative relationship observed is
highly significant (Stevens et al. 2010).
Critical loads reflect the sensitivity of ecosystems: the lower the critical load, the more sensitive
the system. In section 1.2.1 explains how critical loads come about. In many areas of the
Netherlands, the current nitrogen deposition exceeds the critical load of the existing habitat type,
this means that the critical load has been exceeded (also known as 'exceedance'). The figure
above indicates the likely effects on the species richness when the critical load is exceeded.
Figure 2.8 shows this for an individual habitat type. As mentioned above, the decrease in species
occurs at around 12 kg N/ha/yr, which corresponds with the critical load established for this
habitat type (model outcome 11.6 kg N/ha/yr, consistent with the empirical range of 10-15 kg
N/ha/yr; see recovery strategy H6230 in Part II).
50
Part I - November 2012 version -
There is also experimental evidence that the species richness in nature reserves will actually
reduce as the excess of the critical loads increases. This can be seen in figure 2.9, in which a
significant negative correlation was found between the exceedance of critical loads and the
species richness ratio in experiments with the addition of N in European nature reserves with a
variety of non-forest vegetation (Bobbink & Hettelingh 2011). The excess is calculated as the sum
of the experimental N-dose plus the background deposition minus the critical load. Values under
1 indicate that the species richness is lower in cases of excess than in the control study. So 0.75
means that species richness has declined by 25%.
Figure 2.9 Overview of the relationship between plant species richness and excessive (in kg N/ha/yr)
critical loads. Each point is the average result of an N-addition experiment in grasslands, montane
heath and subarctic and subalpine (heath) vegetation, where the species richness in the N-treatment is
divided by those in the control plots (species richness ratio). (Bobbink et al. 2010c).
DISTURBED NITROGEN CYCLE AND NITROGEN LEACHING
In addition to the changes described above in the competitive position, the speed of the nitrogen
cycle in the system gradually increases through nitrogen eutrophication (Aerts & Bobbink 1999).
The vegetation's biomass production is higher and includes more litter production both above
and below ground, often with higher nitrogen concentrations. This initially serves to increase the
rate of degradation of the organic material (mineralisation). However, there is also evidence that
the decomposition of litter in fact slows down in the long term, due to increased N concentrations
and depending on which stage the formation of humus has reached (Berg & Matzner 1997;
Hagedoorn et al. 2003). Mechanisms for this are provided in Janssens et al. 2010. At the same
time, the immobilisation of ammonium and nitrate in several systems decreases through
heterotrophic micro-organisms, eventually causing the excess of nitrogen to be made available
51
Part I - November 2012 version -
more quickly for the vegetation (Tietema et al. 1993; Nadelhoffer et al. 2004). This has also been
found in the case of dune ecosystems, where the phenomenon was most apparent in acidic soils
(Kooijman & Besse 2002, Kooijman et al. 2009).
Changes in the vegetation composition may in themselves also result in additional acceleration of
the nitrogen cycle, due, for example, to the fact that the litter from the now more dominant
species contains higher levels of nitrogen and thus breaks down more easily (Aerts & Chapin
2000). In the case of a consistently high nitrogen deposition, plant growth is no longer limited by
nitrogen after a certain point, but rather by another element (phosphorous, potassium or
magnesium), or for example, a lack of water (Aerts & Bobbink 1999). In such cases, there is no
additional growth due to an excessive supply of nitrogen - although a part of the additional
nitrogen will be absorbed by the vegetation. The latter leads to even higher nitrogen
concentrations in the vegetation which affect the amount of nitrogen in the litter and the
mineralisation of nitrogen.
The aforementioned processes cause the C/N ratio of the soil's top layer to gradually decrease.
The continuous accumulation of nitrogen, increased speed of the nitrogen cycle and saturation of
the ecosystem with nitrogen steadily increase the risk of inorganic nitrogen leaching into the
(shallow) groundwater. A system like this is referred to as being saturated with nitrogen. This
term is used mainly for forests and forests saturated with nitrogen characteristically show
reduced levels of the C/N ratio from the organic horizon, increased nitrate leaching, increased
emission of N2O and the termination of the growth stimulation by additional nitrogen. In some
cases, a decline in tree growth may also occur. In terrestrial systems, nitrogen almost always gets
washed away in the form of nitrate, since ammonium in the soil is not very mobile, and only a
very limited amount disappears into the groundwater. It is only in wet systems, including
peatlands, that ammonium leaching to the groundwater can also be of quantitative importance
(Kros et al. 2008).
In the deciduous and coniferous forests of Europe, nitrate leaching into the groundwater is
strongly related to the total nitrogen deposition that enters the forest (see Dise & Wright 1995;
De Vries et al. 2007; Dise et al. 2009). In cases of nitrogen deposition below 8-10 kg N/ha/yr,
almost no nitrate leaches into the groundwater in forests. Furthermore, increased nitrogen
deposition causes the leaching to increase significantly. The identified correlation is clear,
although a greatly increased supply and saturation means the variation in nitrogen leaching is
quite large (Figure 2.10). This is primarily related to the C/N ratio of the topsoil and the
significant differences in climate (temperature, humidity) in the gradient studied in Europe.
Increased nitrogen deposition can increase the amount of nitrogen (in the form of nitrate) that
enters the groundwater system and thus also the outgoing surface water or groundwater-fed
vegetation. During this transport process, nitrate can seep into pyrite layers by way of
denitrification. Unfortunately, this causes the sulphide to be converted into sulphate, which can
have serious consequences for these wet ecosystems, following transportation by groundwaterfed systems (for details on this process, see the Intermezzo following this chapter).
52
Part I - November 2012 version -
Figure 2.10 The leaching of nitrogen compounds into the groundwater under NW European
forests in relation to the supply of nitrogen from the atmosphere (Dise & Wright 1995).
The altered nitrogen chemistry of the vegetation can also have serious consequences for
herbivores and thus the rest of the food chain in the ecosystem (see section 2.3.5).
In addition to the increased nitrogen concentrations, the proportions of this element in relation to
other nutrients (especially phosphorous, magnesium and/or potassium) are disrupted, and this
nutrient imbalance can, in turn, have serious consequences for the growth and development of
plants and trees (see, for example, Nihlgard 1985; Roelofs et al. 1985). This phenomenon is
exacerbated in situations where soil acidification occurs, with cations such as magnesium and
potassium leaching at an accelerated pace and the relationship between nitrogen and cations can
become even more unbalanced (see section 2.2.2). This process was probably one of the reasons
that a decline in vitality was observed among various species in the Netherlands in the 1980s
(see, for example Van Dijk 1993). Very strong reductions in the S deposition - and to a lesser
extent, in the N-deposition - have put a stop to this decline in vitality, and some species have
even improved again.
2.3.2 ACIDIFICATION
In addition to eutrophication, acidification of the soil or water is one of the main effects of the
atmospheric deposition of N compounds. Understanding the processes that lead to soil
acidification is thus essential when implementing effective restorative measures to combat the
effects of N deposition. A significant proportion of the negative effects of N-deposition on
biodiversity is due to soil acidification, particularly in the case of ecosystems that were originally
weakly buffered (pH 4.5-6.5) (see, for example Bobbink et al. 1998).
53
Part I - November 2012 version -
BUFFERING CAPACITY OF THE SOIL
A soil bearing vegetation does not only consist of an aqueous solution (soil moisture), but also a
solid phase (mineral soil, organic matter). Micro-organisms, soil fauna and underground plant
components are also present - in short, a complex system. It is therefore insufficient to simply
measure the pH of the soil moisture in order to determine acidification, as shown by the
following.
As an example of what can happen during the soil acidification process, a hydrochloric acid
solution3is slowly added to a calcareous soil (CaCO) (Figure 2.11). It appears that the pH in the
water, which is collected at the bottom of the soil, remains between 7 and 8 for quite some time.
The soil water does not get more acidic initially. The fact that the pH does not drop, in spite of
the addition of acid, is because buffer reactions - such as lime dissolving - may occur. As such, a
better definition of soil acidification is the reduction in the buffer capacity - also known as the
acid neutralising capacity (ANC) of the soil. During the process, the ground loses a pH-buffering
substance (see below in reaction 1) which is only available in finite amounts. Once that is nearly
completely finished, the pH will fall.
Figure 2.11 The schematic course of the acidity in a well-drained calcareous soil column with the
continuous addition of a strong acid (source: Ulrich 1981).
The buffering capacity of the soil can be reduced both by natural processes and by the deposition
of acidifying pollutants (nitrogen and sulphur). The deposition of NH3 also contributes to this
process through nitrification in which acid (H+) is produced (see below). Acid-consuming
processes occur in the soil which in turn increases the buffer capacity. This process is also called
alkalinisation, as a hydroxide ion (OH-) or bicarbonate (HCO3-) is released. Denitrification and
sulphate reduction are two of the most important soil processes that can consume acid. For
details about these processes, see the Intermezzo. These processes ensure that the pH is usually
54
Part I - November 2012 version -
high under reduced conditions in wet soils and remains relatively high in cases of high
deposition.
BUFFERING MECHANISMS AND SOIL ACIDIFICATION
There are a number of mechanisms that can be used to buffer the supply of acidifying substances
to soils. The way in which this buffering takes place depends on the base material (the type of
soil), and whether or not there is an inflow of ground water (see below). If a lime-rich, dry soil
with a neutral pH value (pH = 7) is continuously irrigated with an acid solution, the following pHchanges will occur in sequence due to various buffer mechanisms (Figure 2.11). A detailed
description of this is provided in Ulrich (1981).
1)
Carbonate erosion: in calcareous soils (pH> 6.8), buffering takes place through the reaction
of the acid with the lime that is present (CaCO3 - solid phase):
CaCO3 + H+ → Ca2+ + HCO3-
(1)
During this process, calcium and bicarbonate dissolve which leaches ions into the groundwater.
The original concentration of H+ in the soil does not change initially, such that the pH remains
constant. Compared with other buffering mechanisms, the reaction of acid with lime is rather
quick. If the calcium content is reduced to less than about 0.3%, then the buffering capacity
almost disappears through carbonate erosion and the pH suddenly drops quickly (De Vries et al.
1994) Measurement data and model simulations suggest that such calcium levels can reduce the
pH in the topsoil (the upper 10cm) of calcareous dune soils from approximately 6.5 to 3.0 within
a few decades. The low erosion rate and almost negligible availability of exchangeable cations
and Al hydroxides cause the rapid decline in pH predicted in these soils. In other soils, however,
the decrease in pH is slowed down by mechanisms which are discussed below.
2) One buffer mechanism that is present in lime-free (or decalcified soil) with a pH range
between 4.2 and 6.5 is that of the cation exchange through the soil's absorption complex (Figure
2.12). This complex is composed of clay minerals and/or organic components that are negatively
charged on the outside, which causes cations (Ca2+, Mg2+, K+, Na+) to this complex to be
weakly adsorbed. When additional H+ comes into the ground, the hydrogen ions can displace the
cations in the complex, which causes these cations to end up in the soil solution. The hydrogen
ions themselves are then adsorbed on the complex, and no longer dissolve meaning the pH does
not change.
Figure 2.12 Schematic representation of the acid buffering by cation exchange (changed according to
De Graaf et al. 1994).
55
Part I - November 2012 version -
Cation exchange is a fast-running buffer process, but the capacity is quite limited. The capacity is
gradually reduced through displacement by hydrogen ions. The term 'base saturation' is normally
used in such cases to indicate what percentage of the soil's adsorption complex is occupied by
so-called base cations (Ca2+, Mg2+, K+, Na+). The term 'basic' is used here because the respective
cations are derived from strong hydroxides, also known as bases.
3)
A reaction which is much slower but almost always provides a very large buffer capacity is
the weathering of silicate minerals. This reaction occurs in lime-free soils (pH <6.5). Primary
silicate minerals dissolve to produce secondary silicates. One example is the erosion of alkali
feldspar:
2 KAlSi3O8 + 2 H+ + 9 H2O → 2 K+ + Al2Si2O5(OH)4 + 4 H4SiO4
(alkali feldspar)
(kaolinite)
(silicic acid)
(2)
This process is generally very slow. This low speed means this mechanism contributes to the
increased supply of acidifying deposition but only slightly contributes to the actual buffering in
the soil. Ecologically speaking, this is not of very important in the short term (years, decades).
However, this process is quite important in soils where the cation exchange plays an even smaller
role.
4) The next buffer mechanism in lime-free, acid soils (pH <4.5) is the weathering of the
aluminium contained in the soil. We refer to the aluminium buffer range in which the following
reaction occurs:
Al(OH)3 + 3 H+ → Al3+ + 3 H2O
(3)
When this reaction is triggered, an increasing amount of AI3+ enters the soil solution, whereas Al
was previously almost exclusively present in its non-dissolved form in the soil. In the same way
as H+, Al3+ can also be bound to the soil complex, but this process cannot prevent an increase in
free Al3+. It is important to note that (dissolved) Al3+ is toxic for many plant and animal species. If
high levels of Al3+ are found in the soil moisture, this means that the buffer capacity of the soil
has already been consumed to a large extent (Figure 2.11).
At this point, it is useful to mention the iron buffer range (not shown in Figure 2.11). If the pH
levels are <3.8, amorphous iron oxides dissolve in case of acid buffering reactions in the
presence of dissolved organic matter, while if pH values are <3.0, iron plays an increasingly
dominant role in soil buffering and a (very) high amount of Fe3+ dissolves (Verstraten et al. 1989).
This last process (pH <3.0) occurs seldom in practice in the Netherlands, but was prominent, for
example, in the 1980s/1990s in the extremely overburdened 'black triangle' of Europe.
5)
In systems with bicarbonate in the soil water - caused, for example, by regional seepage -
buffering by bicarbonate may occur:
HCO3- + H+ → H2CO3
(4)
Buffering by bicarbonate is the most important buffer mechanism in surface water, but is
relatively insignificant for the majority of terrestrial situations. The amount of bicarbonate in the
soil water is generally quite low and so the capacity of this buffering is also low compared to the
56
Part I - November 2012 version -
previous four buffering mechanisms unless there is a regular supply of bicarbonate through
groundwater seepage or flooding of the surface water occurs. The extraction of groundwater and
the regulation of the surface water has often indirectly caused acidification through the loss of
seepage or flooding. Besides bicarbonate, calcium (or other base cations) is also supplied through
seepage or surface water, which allows the base saturation to be recharged through the
aforementioned cation exchange. This is discussed in the next section.
SENSITIVITY TO ACIDIFICATION: A WORLD OF DIFFERENCE
It should be made clear that the precise course of the reduction in buffer capacity (acidification of
the soil) and the effects of this on the pH of the soil moisture are highly dependent on the soil
material, and its position in the landscape. Buffering is not present everywhere in the landscape
in equal measure. We discuss the situation in the Dutch countryside in sequence, going from high
to low levels.
In locations where permanent infiltration (downward movement of groundwater) occurs, the pH
depends on the weatherability of the mineral fraction of the soil. The pH levels in lime-free
cover-sand areas are low (5.0 or less) and buffering primarily takes place through cation
exchange or aluminium buffering. Higher pH values (weak acid to acid) may occur in loamy soils.
These types of soils are found in areas where loess, boulder clay or loamy cover-sand is on the
ground surface or close beneath it. Buffering takes place here through cation exchange.
Situations like this can also be found on the transitions to lower areas where cover-sands are
covered with river loam or clay during occasional floods. Calcareous soil (chalk, marine clay,
calcareous river clay and calcareous loess) can have neutral or alkaline pH values. Buffering
occurs here through carbonate erosion.
In moist to wet soils where seepage or flooding occurs (periodically), this water flow is usually an
important supply of buffer substances (cations, bicarbonate), which can greatly increase the
buffering through cation exchange with the adsorption complex. Regular flooding with surface
water that is rich in sludge and alkali causes the soil to have a soil adsorption complex with high
capacity that is periodically saturated with bases. Outside the inundation period, pH levels are in
the weak acidic range, because the base saturation is reduced due to the exchange of H+ for base
cations. If the periodic supply lapses or the intervening periods become longer, the base
saturation will decline even further and reduce the cation exchange. The pH is subsequently
buffered to a lower level.
The presence of bicarbonate-rich groundwater in the root zone buffers the pH level into the
neutral to weakly acidic range. Groundwater is usually rich in bicarbonates due to the fact that it
has passed through calcareous layers as it flowed to seepage areas (see equation 1). This means
it is enriched with cations, namely Ca2+, such that the adsorption complex becomes saturated
with bases in periods of seepage or capillary rise. In periods where no seepage occurs, the
available bicarbonate gets used and the cation exchange with the adsorption complex takes over
the buffering effect. Local high-level seepage from canals, water supply ditches or polders is
often rich in bicarbonate and cations and is therefore similar to 'natural' bicarbonate-rich
groundwater. In such cases, the ionic composition is different, which is evident from the
increased levels of Cl-, Na+ and K+, for example.
57
Part I - November 2012 version -
Local groundwater systems with bicarbonate- and calcium-rich groundwater also arise in areas
where the soil contains some calcareous material at a certain depth. Areas that are supplied by
the local base-rich water - such as those described in the above situation with regional seepage are buffered in the neutral to weakly acidic range by bicarbonate and cation exchange. Places
where the soil contains calcareous material at a certain depth appear in areas such as South
Limburg, the central and western parts of North Brabant, and in the Achterhoek, Twente and
Salland. If lime is lacking in the flow-through,the pH in local seepage areas will be buffered at a
(much) lower level.
All this means soils are very sensitive to acidification in some areas, and in other areas, hardly at
all. A quantification of the sensitivity to acidification based on the above considerations is
provided by De Vries et al. (1989). This analysis shows that, in the Netherlands, the Pleistocene
sandy soils in the centre, south and east of the country and the dunes along the coast are
particularly between susceptible and very susceptible to accelerated soil acidification through
atmospheric deposition (Figure 2.13).
Figure 2.13 Overview of
acidification of sensitive soils in
the Netherlands. The indicated
dark sections are sensitive to
soil acidification through
atmospheric deposition (source
De Vries et al. 1989).
If you compare this map with the deposition in the Netherlands, you immediately notice that the
deposition of NH3 is highest in these areas which are sensitive to acidification. Furthermore, this
national scale hardly shows where soils are being buffered by the supply of groundwater and are
therefore (much) less sensitive to soil acidification.
58
Part I - November 2012 version -
EFFECTS OF SOIL ACIDIFICATION: A COMPLEX OF FACTORS
During soil acidification, all kinds of conversions take place in the soil, in which all kinds of
cations can be dissolved alongside the decrease in the buffer capacity (ANC). Because annual
rainfall in the Netherlands exceeds evaporation, a downward transport of water occurs on average
over the year. These water flows allow these cations to leach into deeper layers or to the
groundwater. Due to the electrical neutrality in the downward transport of cations, mobile anions
such as bicarbonate (HCO3-) and chloride (Cl-) (in neutral soils) or SO42-, NO3- or organic acid
residues
may wash out (in acidic soils). Soil acidification can seriously affect the nature of an ecosystem
and thus its biodiversity: pH levels can decrease significantly, base cation shortages may occur
and an excess toxic metals, especially Al3+ can be released (Figure 2.14). The decrease in pH
levels (<4.5) may increasingly inhibit the nitrification, increasing the ammonium-nitrate ratio
and, finally, the rate of degradation of organic material (decomposition) can be greatly reduced
which causes litter accumulation in acidified ecosystems to be very common. Because several
species cannot tolerate the combination of a low pH and high concentrations of NH4+ and
released Al,3+ acidification almost always leads to a loss of species. A detailed overview of the
effects of acidification is provided in De Vries 2008.
Soil acidification
(decrease in acid neutralisation capacity)
Increase in H+-ions
+
H -concentration
Leaching of base
Release of toxic
Inhibition of nitrification
cations
metals (Al, Cd)
and decomposition
base cations
Figure 2.14 Overview of factors that (can) change in the soil due to acidification (according to Bobbink
& Lamers 1999).
ACIDIFICATION OF SURFACE WATERS
Surface waters with low alkalinity have a low buffering capacity. Deposition of N and S will
therefore respectively lead directly or indirectly to acidification. Additional ammonium will be
nitrified in these waters (at pH> 4.0). During this process, H+ ions will be formed which cause the
pH levels to drop. Experimental studies have shown that a two-year treatment 19 kg N/ha/yr may
already lead to major changes (Schuurkes et al. 1987). When the pH drops below 5 as a result of
these acidification processes, acidic-intolerant freshwater species will disappear (Arts et al.
59
Part I - November 2012 version -
1990ab). A proportion of the species that are characteristic of weakly buffered and very weakly
buffered fens, such as Shoreweed( Littorella uniflora), can still remain below a pH of 5, but
submerged peat moss may grow over these freshwater plants. Besides the excessive growth of
peat mosses, a (temporary) proliferation of Bulbous Rush (Juncus bulbosus) also often occurs,
that specifically responds to ammonium (and not to the acidifying effect of S; Schuurkes et al.
1987). Under these circumstances, peat mosses and bulbous rush make the most of the high
availability of nitrogen and carbon and can therefore quickly build up biomass and become very
dominant (Schuurkes et al. 1986). As such, all characteristic plants gradually disappear from
acidified fens due to changes in water chemistry and the massive growth of bulbous rush and
toothed sphagnum.
ACIDIFICATION AND FAUNA
Besides the effects on plants, acidity of the soil can also affect the fauna (see also section 2.4).
Soil acidification causes increased levels of cations (Ca, K, Mg) to leach from the upper soil layer
to deeper layers. Due mainly to the shortage of Ca2+ snails, woodlice and millipedes (all with a
calcium-dominated exoskeleton) disappear into strongly acidified soils. In those circumstances,
insects provide too little calcium. This causes songbird eggshells - which depend on these prey to become thinner and their breeding success to decrease significantly, especially in areas where
no other sources of calcium are available. This phenomenon has been clearly demonstrated for
the great tit, but is probably also the case for other birds in these acidified ecosystems (Graveland
et al. 1994).
In the case of sparrowhawks (Accipiter nisus) in nutrient-poor forests, it was found that a mineral
deficiency in oak leaves was leading to the absence of a specific protein that can transport and
store vitamin B2 in caterpillars. Although these effects were not observed in the caterpillars that
eat the leaves and the great tits (Parus major) (with relatively short lives), which mainly feed on
these caterpillars, a vitamin B2 deficiency was found to occur in the long-living sparrowhawk.
This resulted in a sharp decrease in breast muscle and low levels of vitamin B2 in their eggs,
which was in turn linked to fatal embryonic abnormalities (Van den Burg 2000). This may also
play a role for other long-living species.
2.3.3 THE NEGATIVE EFFECTS OF REDUCED NITROGEN
Nitrate and ammonium are the inorganic forms of nitrogen which occur in soil and water, and
which may be used as a source of nitrogen by plants. A wide range of ratios of nitrate and
ammonium occur in semi-natural and natural ecosystems: nitrate is the dominant form of
nitrogen in good to moderately buffered conditions (pH> 5.0), while ammonium is inherently very
dominant under acidic conditions (pH <4.5). Plant physiology shows that plant species from
calcareous or slightly acidic habitats have adapted to nitrate as a source of nitrogen, or a
combination of nitrate and ammonium, despite the fact that this is ammonium under acidic
conditions (see, for example, Gigon & Rorison 1972; Kinzel 1982).
60
Part I - November 2012 version -
AMMONIUM TOXICITY
When plant species from growth areas where there was originally virtually no ammonium present
as a source of nitrogen, are exposed to increased concentrations of ammonium, a complex of
negative phenomena are often observed (see, for example Britto & Kronzucker 2002; Stevens et
al. 2011). The physiology of these changes can lead to a significant reduction in growth and
development in sensitive species (Figure 2.15).
The negative effects of increased ammonium concentrations and/or an increased ammoniumnitrate ratio on growth and development under laboratory conditions in the Netherlands
demonstrated for typical vascular plants and mosses from different habitats: low-nutrient
grasslands, species-rich heaths, weakly and very weakly buffered fens, blue grasslands and
floating fens (Schuurkes et al. 1986; Roelofs et al. 1996; De Graaf et al. 1998; Lucassen et al.
2003; Paulissen et al. 2004 & Van den Berg et al. 2005). Based on some 300 vegetation samples
with geochemical measurements, it was recently found that the presence of typical Red List
species in the Pleistocene sand landscape in the Netherlands is strongly correlated with a low
ammonium-nitrate ratio and/or low ammonium concentrations in the soil (Kleijn et al. 2008; De
Graaf et al. 2009). For example, only two of the Red List species were present with high
ammonium-nitrate ratios, while all other remaining Red List species were found in cases where
there were (very) low ratios. The opposite was true, however, in the case of the species found in
heathland (Figure 2.16). Very recently, studies were carried out on mini-ecosystems with species
and soils from low-nutrient environments or with the addition of N in floating fens in Ireland and
showed that the negative effects of ammonium may also be prominent in the field (Van den Berg
et al. 2008; Verhoeven et al. 2011).
Figure 2.15 Picture of the effect of high ammonium concentration on the growth of leopard's bane
(Arnica montana) after three months of growth in a continuous-flow culture at pH 4 (photo: MCC de
Graaf). The plant on the left only has only had (100 µmol/l) in nitrate nutrition and the other plants
61
Part I - November 2012 version -
beside it have also had ammonium (1000 µmol/l). For the other results from this study, see De Graaf et
al. (1998).
Figure 2.16 The relationship between the ammonium-nitrate ratio in the soil and the soil pH for
general plant species (blue diamonds) or Red List species (red squares) from Dutch heathlands. The
standard error (SE) is shown in addition to the mean value (Kleijn et al. 2008).
A CRUCIAL ROLE: THE NITRIFICATION RATE OF THE SOIL
Generally speaking, it can be stated that, for many years, reduced nitrogen has been the major
component of total nitrogen deposition in the vast majority of the Netherlands, and that this
occurs at an increased rate in areas with a lot of intensive livestock production (see section 2.1).
However, this does not automatically mean that the vegetation is exposed to the same ratio of
reduced and oxidised nitrogen as the ratio present in the on-site deposition. The ratio of reduced
versus oxidised nitrogen that enters the plant is only more or less equal to the ratio present in
the air and in the deposition in case of the direct uptake of nitrogen compounds from the
atmosphere by aboveground plant components, or the whole organism in the case of mosses and
lichens, although reduced nitrogen (ammonium and ammonia) is usually absorbed more easily
above ground than oxidised nitrogen (nitrate and nitric oxide). It is also relevant to mention that
conditions with limited nitrogen can lead to a substantial part of the nitrogen deposition being
absorbed by aboveground plant parts.
In many cases, the ammonium that enters into the soil or water is rapidly converted into nitrate
by microorganisms, a process that is known as nitrification. These process releases two protons
(H+) per nitrogen molecule. The nitrification rate is strongly influenced by the abiotic conditions,
in which the pH levels and the amount of oxygen are particularly important. Decreasing pH values
from 5 to 4 can cause the nitrification to decrease further, despite the fact that a considerable
62
Part I - November 2012 version -
level of nitrification can occur in forests even down to very low pH values (3.0-3.5) (Roelofs et al.
1984). Under reduced, oxygen-free conditions, such as those found in permanent wet soil,
nitrate is no longer formed from ammonium, except in the root zone of plants that lose oxygen
through their roots. Under (alternating) wet conditions, however, nitrate can be denitrified and
disappear into the atmosphere as N2 (or sometimes N2O). Bicarbonate is formed during this
process and the soil gets 'deacidified' (see Intermezzo). All this means that nitrate is still the
dominant form of nitrogen as regards the uptake of nitrogen by the vegetation in strong or
moderately buffered conditions and areas that are not too wet, even in areas with a high
proportion of reduced nitrogen in the air supply. Obviously, this does not apply to mosses and
lichens that do not absorb their nitrogen from the soil, but rather directly from the atmosphere or
rainwater.
All in all, it can be said that the negative effects of reduced nitrogen may or may not play a
prominent role in the degradation of natural and semi-natural ecosystems due to increased
nitrogen supply (Bobbink & Lamers 2002; Stevens et al. 2011).
In highly buffered conditions (pH ≥ 7), most of the plant species present are only adapted to
nitrate supply, but are hardly exposed to the negative effects of reduced nitrogen due to the high
nitrification rate in the soil or water. Obviously this does not apply to mosses and lichens that
were (originally) present in these buffered systems and are exposed to the increased reduced
nitrogen values. In systems that are already acidic, such as bogs, heaths and some forests (pH ≤
4.2) ammonium has always been the only source of inorganic nitrogen, which causes the
characteristic plant species to adapt to the ammonium supply. The consequence of this is that the
aforementioned effects of the increased availability of reduced nitrogen in these systems have
much lower effects, or no effect at all.
The negative effects are particularly important in those situations where there were originally
several plant species that were adapted to nitrate as the dominant form of nitrogen, and not to
ammonium supply, and where the soil is only moderately to weakly buffered. The most serious
consequences of more ammonium in the plant's nutrition are therefore mainly found in what were
previously weakly to moderately buffered conditions (pH 4.5 to 6.7), for example, as was
originally the case in low-nutrient grasslands, very weakly buffered fens, formerly species-rich
heathlands and forests on loamy soil. In these systems, the soil or water layer can often be
significantly sensitive to acidification. Prolonged exposure to acid deposition, can cause the pH to
be so low (pH <4.5) that virtually no nitrification occurs and as a result, high levels of ammonium
accumulate in both the soil and plant (Bobbink et al. 1998; Stevens et al. 2011). Sensitive plant
species are also threatened by the other effects of soil acidification, such as elevated levels of
aluminium and reduced amounts of base cations which cause the susceptibility to ammonium to
increase even more (see section 2.3). Several mosses and lichens also experience direct damage,
particularly those species that are not adapted to a high uptake and assimilation of reduced
nitrogen (Sutton et al. 2009).
2.4 Effects on the habitats of fauna
The previous sections dealt with the effects of nitrogen deposition on soil and water and the
vegetation dependent on them. This section outlines what the effects are for animal species.
Studies analysing the overall effects of nitrogen deposition on ecosystems analysis almost always
fail to study the effects on the fauna. If these are mentioned, it is usually in relation to changes in
63
Part I - November 2012 version -
the herbivory, and, even then it is only the effects on plants rather than those on the herbivores,
that are addressed (e.g. Gilliam 2006). In a few cases, scholars have indicated that there is a gap
in our knowledge concerning in wildlife research (Adams 2003).
The direct effects of elevated nitrogen deposition on fauna (section 2.4.5) are rarely dealt with
and almost only play a role in aquatic environments. After all, aquatic animals have large and/or
thin permeable skin surfaces (gills or tegument) that come into direct contact with the
environment for oxygen uptake. This greater exposure means that more immediate effects are to
be expected. A large majority of the effects of nitrogen deposition are indirect in nature and
come about through chemical processes in soil and water, the resulting changes in vegetation or
cumulatively by working their way into the food chain in the case of predators.
Due to the fact that these effects occur indirectly and that determining the effect on animals is
more difficult to determine than that on the soil, water and vegetation, there is very little
scientific evidence available for these effects and the underlying mechanisms. Moreover, the
effects of nitrogen deposition interfere with other factors such as depletion, acidification
(especially sulphur compounds), changes in land use, climate change, the prevention of invasive
alien species and elevated levels of CO2 and O3 in the air (see, for example Rabalais 2002, Fenn
et al. 2003ab). This interference has both mitigating and enhancing effects.
Nevertheless, the effects of nitrogen deposition on wildlife can be demonstrated or shown to be
(very) likely using both field measurements and experimental studies. Unlike the influence of the
nitrogen cycle on vegetation (e.g. Krupa 2003, Bobbink et al. 2010) and on aquatic fauna
(Camargo & Alonso 2006), an overview of the mechanisms by which nitrogen deposition can
affect terrestrial fauna has never been published in a scientific journal. This is why all processes
found in literature studies have been summarised in the diagram shown in Figure 2.17. In the
current text, the various processes are developed and supported with citations from the
literature. The restorative strategies for habitat types (Part II) and living areas that fall outside
these habitat types (Part III) refer to this format of processes as regards the effects on species of
the Birds and Habitats Directive ('protected species') and typical species (as a quality aspect of
habitat types).
A large majority of the effects of nitrogen deposition are indirect in nature and come about
through changes in vegetation or water and can have a cumulative effect in the food chain on
species in the higher trophic levels. Below we present an overview of the various processes
behind the effects of nitrogen deposition on fauna given the relationship between these
processes and changes in soil and vegetation described in the preceding sections. These mainly
involve the indirect effects of increased biomass production of vegetation due to increased
nitrogen availability. These processes are shown in a diagram in figure 2.17. The processes are
then discussed and supported by citations from the literature published on this subject.
64
Part I - November 2012 version -
air Increased N deposition soil/water Lowering of pH in water and soil Increase in N available Availability of materials vegetation Increased biomass production (including litter and organic soil) Disturbed nutrient balance in plants Higher and more homogeneous vegetation structure Decrease in herbs and low grasses and decrease in flower density fauna 3) Decrease in quantity of food plants (in time and/or space) 1) Cooler and damper microclimate (in time and/or space) 4) Decrease in quality of food plants (in time and/or space)
2) Decrease in opportunities to 6) Reduction in availability of prey animals and 5) Physiological problems reproduce host species (including impaired growth and (size, quality and accessibility) osmosis, oxygen deficiency) Figure 2.17 Simplified diagram of the impact of nitrogen deposition on animals. Almost all effects have an indirect
impact through changes in the soil, surface water, vegetation and litter. Direct effects of the acidifying effects of
nitrogen deposition almost exclusively occur through physiological problems in aquatic environments (dotted line).
65
Part I - November 2012 version -
2.4.1 COOLER AND DAMPER MICROCLIMATE (1)
An increased production of plant biomass creates a higher and denser layer of vegetation (living
and dead standing vegetation and a layer of litter) which radiation from the sun is less able to
penetrate and less air circulation takes place just above the soil. Heat generation just above the
soil is thus inhibited and the overall temperature sum in, or directly on, the soil is lower. The
lower temperature and lack of air circulation creates a damper microclimate. Very little research
has been carried out into the relationship between nitrogen deposition and the effects of a
changed microclimate on fauna. However, the effects listed below are very plausible based on
fundamental ecological research on animal species.
A lower temperature sum results in the slower development of invertebrates, making the total
development time longer. The development time for the immature stage (egg, larva or nymph)
can become so long that the entire life cycle cannot be completed within one season, as
demonstrated in the case of locusts by Van Wingerden et al. (1991 en 1992). Research carried out
by Schirmel et al. (2011) also showed that a warm and dry microclimate plays an important role in
the distribution and diversity of locusts in dry grey dunes.
Larger invertebrate species need a higher overall temperature sum to complete their cycle and
they deteriorate earlier and more rapidly than small invertebrates. In the grasslands examined,
this leads to the disappearance of these larger species, often in favour of smaller species that can
complete their life cycle. When the vegetation structure is replaced in winter (through
interventions or natural processes), but closes up again through growth in the course of the
season, animal species that need heat for their development in the spring are much less affected
by this fallowing than species that require heat in the (late) summer, as has been shown for ants
on calcareous grasslands (Van Noordwijk et al. 2011). Those species that are characteristic of
open vegetation are resistant to drought and high temperatures. Fallowing causes these species
to disappear (such as ground beetles from dune grasslands; Nijssen et al. 2011), probably
because they are displaced by less adapted, but more competitive species.
Due to a lower temperature, and a dampening of the air flow, dense vegetation also has a damp
micro-climate that can lead to yeast infections in hibernating caterpillars and thus high mortality
(Bink 1992, Wallis de Vries & Van Swaay 2006). High, moist vegetation can lead to hypothermia in
precocial chicks and therefore an even higher need for food. Frequent or prolonged heating by
the parents does not lead to a solution, since this leads to a reduction in the necessary feeding
time (Schekkerman 2008).
Interventions to remove the higher vegetation and thick layer of litter do not always lead to an
improvement for invertebrates. Grazing leads to shorter, but denser vegetation, which does not
sufficiently restore the temperature loss on and in the soil (Wouters et al. 2012). Grazing and a
higher frequency of mowing, turf cutting or chopping can all lead to an increase in the frequency
of disturbance, making this a limiting factor for - usually larger - species with a long
development time. The frequency of the disturbance, in conjunction with the population's
restorative capacity or the species' recolonisation capacity determines the effect on its local
presence (Siepel 1996, Lindberg & Bengtsson 2006, Van Noordwijk et al. 2011).
2.4.2 DECREASE IN OPPORTUNITY TO REPRODUCE (2)
As a result of higher vegetation and the overgrowing of open (or mosaic) vegetation, suitable
locations for reproducing become physically inaccessible or inappropriate. This concerns both a
nesting opportunity for ground-nesting birds, places to store eggs in the soil, spawning grounds
66
Part I - November 2012 version -
for fish in open water and mating opportunities. No research has been conducted into the specific
relationship between nitrogen deposition and decreased reproduction opportunities for fauna.
Based on correlational research into the occurrence of species and overgrowth, however, it
appears the aforementioned effects are quite plausible for ants and wasps (Peeters et al. 2004),
butterflies (Bink 1992), different kinds of locusts (Lensink 1963, Kleukers et al. 1997) and nesting
birds (Van Turnhout et al. 2010). In part, this process interferes with the microclimate becoming
cooler and damper and in practice it is almost impossible to determine in many cases whether or
not a location is physically inaccessible for reproducing or if the site has become climatically
unsuitable.
2.4.3 DECREASE IN QUANTITY OF FOOD PLANTS (3)
Plants are a vital food source for many animal species, both in the form of fresh biomass, dead
biomass (litter and humus), flowers (nectar and pollen) and seeds. In this study, a distinction is
made between food plants that act as host plants (biomass gets eaten) or those that serve as
nectar plants (flowers get visited), wherever necessary. Problems with the availability of food
plants for animal species can occur in several ways: (a) The food plants disappear totally or
decrease in terms of density or size to the extent that the animal population cannot be provided
with sufficient food; (b) The food plants are not available within the activity period of the species
in question (mismatch in time), (c) the physical distance between the food plants and other
essential landscape elements (such as nesting, hibernation and mating areas, etc.) are
irreconcilable. This last point regarding the fragmentation of biotopes within a habitat is
discussed in Part III (Landscape ecological embedding of restorative strategies).
A change in the composition of vegetation from herbaceous to more grasses due to the increased
availability of nitrogen (Bobbink 1991; Stevens et al. 2004; Clark & Tilman 2008) generally leads
to a deterioration of the animal species that use the herbaceous or low, narrow-leaved grasses as
specific host plants. Virtually no research has been conducted into the specific relationship
between nitrogen deposition and the reduction of food plants, but correlative research on the
decline of animal species and the food plants on which they depend, makes this process very
plausible. Ockinger et al. (2006) identify a correlation between the decrease in butterfly species
and nitrogen deposition and suggest a decrease in (the availability of) food plants. The scarce
amount of species-specific research is primarily concerned with pearl butterflies, such as Marsh
Fritillary (Fowles & Smith 2006) and Niobe Fritillary (Salz & Fartmann 2009) and ‘checkerspot
butterflies’ in North America (Weiss 1999).
Flower visiting insects are dependent on flowering plants for their energy (nectar) and nutrients
(pollen - used as a protein for reproduction, for example). Change in vegetation composition from herbaceous to more grasses due to increased nitrogen availability - can cause a decline in
flower visiting insects. This applies both to species that are specialised in one or a few plant
species for foraging purposes (Biesmeijer et al. 2006; Fründ et al. 2010) as to species that
produce several generations per year, or that build up new colonies every year (e.g. bumblebees)
and are therefore dependent on several different plant species for much of the season.
Bumblebees that have a wider range of flowers available to them, produce larger colonies
(Goulson et al. 2002). In addition, several species of bumblebee - related to tongue length and
body size - have different uses for their habitat which decreases the competition for food. This
results in several species appearing in a landscape with a diverse range of flowers (Sowig 1989;
Westphal et al. 2006; Kleijn & Raemakers 2008).
67
Part I - November 2012 version -
The diversity, density and quality of flowering plants and the availability of nesting areas (see
process 2) also affect the interaction between flowering plants and flower visiting insects that
pollinate these plants (e.g. Hoover et al. 2012).
2.4.4 DECREASE IN QUALITY OF FOOD PLANTS (4)
Plants are a vital food source for many animal species, both in the form of fresh biomass, dead
biomass (litter and humus), flowers (nectar and pollen) and seeds. In addition to problems in
terms of the quantity of these plants (see process 3), changes in the quality of food plants for the
animal species can also occur. It is divided into the following processes: (a) a shift in the nutrient
balance in the food plant, which means animals cannot access a sufficient amount of essential
nutrients; (b) a shift in the relationship between nutrients and antifeeding substances, so that the
plant is no longer edible for a particular species of animal; (c) shift of the nutrient balance and/or
of the relationship between nutrients and antifeeding substances in the time, leading to the food
plant no longer being available within the period of activity of the respective animal species
(mismatch in time); (d) shifts in the food quality of a plant which only a few species can benefit
from, thus shifting competitive conditions.
Little research has been conducted into the changes in the quality of plant food due to nitrogen
deposition. The effects of changes in plant quality on both individuals and populations of
herbivorous animals can be significant (Awmack & Leather 2002). There is growing evidence that
both the eutrophic and acidifying effects of nitrogen deposition affect the uptake of macro and
micronutrients by plants, and thus the food quality of these plants for herbivores (Throop &
Lerdau 2004, Nijssen et al. 2011). An interference occurs which causes elevated CO2
concentrations in the atmosphere and climate change (Hoover et al. 2012).
The way in which this change enters the ecosystem through herbivores depends both on the
species of plant and the species of herbivore (Throop & Lerdau 2004) and the interaction between
them (Throop 2005). Furthermore, most studies report a positive influence of nitrogen deposition
on plant quality, expressed in terms of the total N-content of the tissue, sometimes broken down
into different types of amino acids. Because herbivores are food for predators, these effects also
have an impact higher up in the food chain (see process 6). Plants growing under an increased
supply of nitrogen store this nitrogen in the tissue Increased nitrogen deposition can thus change
the N: P ratio in plant tissue. This also changes the nutritional value of the plant. This occurs
more frequently in systems which are or have become P-limited, through the removal of organic
biomass by mowing and cutting sods (Güsewell 2004). In a fertilisation experiment, Ohlson et al.
(1995) found that the nitrogen concentration in grass leaves, herbs and dwarf shrubs increased in
the undergrowth of forests, where the addition of ammonium had a greater effect than that of
nitrate. The increase of nitrogen in the plant tissue was mainly caused by an increase of the
amino acids glutamine, asparagine and arginine.
Grey-hair grass (Corynephorus canescens) was shown by Nijssen & Siepel 2010 to store
additional nitrogen in non-protein components which are likely to have inhibitory effects on
feeding. Mottled grasshoppers (Myrmeleotettix maculatus typical species from the grey dunes)
which were bred on nitrogen-rich grey-hair grass showed higher mortality and lower growth than
those bred with grey-hair grass without any added nitrogen. Field studies have also shown that
mottled grasshoppers in areas with high nitrogen deposition have a lower body weight than
mottled grasshoppers from areas with low nitrogen deposition. Aeolian deposition of fresh
mineral-rich driftsand actually leads to an improvement in the food quality of grey-hair grass and
68
Part I - November 2012 version -
consequently to a higher survival and growth of mottled grasshoppers. Furthermore, that aeolian
deposition of sand is inhibited by nitrogen deposition, due to the increased production of algae
and higher plants.
Depending on the plant species, an increase in absorbed nitrate and/or ammonium can also lead
to a higher concentration of proteins and amino acids in plants, so that these plants become
more attractive as a food source for insects. This is particularly true for plants that do not
possess the ability to create secondary plant substances containing N as a defence. This reaction
leads to the formation of an infestation on herbivores and also to shifts in vegetation
composition, resulting in a deterioration in the quality of the habitat (see, for example Berdowski
1987).
2.4.5 PHYSIOLOGICAL PROBLEMS (5)
The direct toxic effects of nitrogen compounds (NH4+, NH3, NO2-, HNO2, NO3-) on animals are
known only in aquatic environments (e.g. Thurston et al. 1981, Berendzen et al. 2001), where
NH3 is the most toxic, and NO3- the least toxic (Camargo & Alonso 2006). In addition, the direct
effects of nitrogen deposition occur as a result of the associated acidification. These effects do
interfere, given that the pH affects both the (im)mobility of heavy metals and the toxicity of these
metals, and of NO3 for wildlife (Schuurkes et al. 1986, Thurston et al. 1981, Leuven et al. 1992,
Gerhardt 1993). Interference also occurs through the indirect effects of nitrogen deposition, such
as temporary oxygen shortages due to increased biomass growth in vegetation. The effects of
nitrogen deposition only really play a role in naturally (moderate) nutrient-poor, not to weakly
buffered systems, or in systems that have largely lost their natural buffering capacity through the
long-term effects of acidification by sulphur compounds.
The effect of various nitrogen compounds in aquatic systems and the effects on the fauna in
these systems has been addressed in various publications (including Rabalais 2002, Camargo &
Alonso 2006). The direct effects of acidification concern physiological problems, such as a
disturbed ion balance and osmosis or problems in the oxygen levels due to the dysfunctioning of
oxygen-carrying proteins (Camargo et al. 2005). These physiological problems can directly cause
mortality or impaired development in the immature stage.
Several aquatic animal groups decline significantly or disappear entirely due to acidification. Fish
and snails do not occur at pH levels below 5.0-5.2. The disappearance of these groups as a result
of acidification is described in Leuven et al. (1986), Leuven & Oyen (1987) en Økland (1992).
Other groups of invertebrates (such as amphipods and water fleas) cannot withstand strong
acidification. The critical pH value varies greatly by animal species and both enhancing and
mitigating effects have been demonstrated due to the presence of heavy metals and minerals often mobilised as a result of acidification - at the level of this critical value (Vangenechten et al.
1989, Økland 1992, Gerhardt 1993). Nitrate and ammonium in higher concentrations are toxic
for aquatic fauna (Berendzen et al. 1987), wherein the toxicity of NH3 increases as the pH
decreases (Thurston et al. 1981). In the breeding waters of amphibians, strong acidification
causes the direct mortality of embryos, but weaker acidification can lead to a lower hatching
success of eggs and a higher degree of damage to the eggs by fungi (Leuven et al. 1986b, Freda
1986).
An additional amount of nitrate or ammonium in water can lead to an increased growth in algae,
as long as no other substances such as phosphorous play a limiting role (e.g. Loeb et al. 2010).
The large amount of algae can lead to lower oxygen levels. This situation is particularly prevalent
in smaller stationary (or slowly flowing) waters; in rivers or large stationary waters, currents and
69
Part I - November 2012 version -
the wind ensure better mixing and oxygen uptake from the air. Aquatic insects with high oxygen
requirements may suffer from periodic oxygen deficiencies (Jeppesen et al. 2000), but this also
applies to certain species of fish (De Nie 1997).
2.4.6 REDUCTION IN AVAILABILITY OF PREY ANIMALS AND HOST SPECIES (6)
Predators and parasitic species are dependent on other animal species to serve as prey or as a
host to ensure food supply, or completion of their life cycle. When these animal species
experience the negative effects of increased nitrogen deposition, problems may also arise for
these predators and parasites. These problems can be divided into the following processes: (a)
The disappearance or strong decrease of animal species which causes (a population of) predators
or parasites to not have sufficient food or hosts: (b) a decrease in the accessibility of prey animals
or host species due to higher and/or more dense vegetation; (c) a reduction in the food quality of
prey species; (e) an increase in the physical distance between the prey animals or hosts and other
essential elements in the landscape (such as nesting, hibernation and mating areas, etc). This
fragmentation of biotopes within a habitat is discussed in Part III (Landscape ecological
embedding of restorative strategies). Very little research has been conducted into the specific
relationship between nitrogen deposition and the decreasing availability of prey or host species.
However, correlational research makes the processes described very plausible.
When nitrogen deposition precipitates into ammonium form, this results in acidification of the
soil and waters (lower pH value), a reduction in the amount of available calcium and a disturbance
of the mineral balance. These effects enter indirectly into the food chain. In waters, such strong
acidification leads to effects such as a reduction in fish stocks, which has consequences for fisheating birds and also for birds with a combined diet of fish and insects. Snails, woodlice,
millipedes (all with a calcium-dominated exoskeleton) have disappeared in highly acidified soils
and insects produce too little calcium in those circumstances, causing the eggshells of songbirds
- which are dependent on these prey - to weaken (Graveland et al. 1994). In the case of black
terns, it has been demonstrated that acidic waters do not constitute a suitable habitat due to their
lack of fish, which can lead to weak bone structure in chickens (Beintema 1997). The calcium
deficiency in invertebrate herbivores is partly reinforced by a lower concentration of their host
plants because of the reduced presence of ectomycorrhizal species on the roots. Even birds that
are (partially) dependent on aquatic insects, may experience the effects of acidified waters. In the
case of dippers, it has been demonstrated that they have a declining population trend in acidified
waters, which seems to be directly related to a reduced supply of calcium, resulting in thinner
eggshells, smaller eggs and young, and thus a lower rate of reproductive success (Tyler &
Ormerond 1992; the fact that weaker animals, whose lower competitiveness levels mean they lose
out on the most suitable breeding sites, do not breed by these waters could, however not be
excluded). It is expected that other insectivorous birds are also affected by this.
In the case of sparrowhawks in nutrient-poor forests, it has been shown that a mineral deficiency
in oak leaves has led to the absence of a specific protein that can transport and store vitamin B2
in caterpillars. Although such effects were not observed in the caterpillars that eat the leaves and
the (relatively short-living) great tits that mainly feed on these caterpillars, the longer-living
Sparrowhawk showed a vitamin B2 deficiency. This resulted in a sharp decrease in breast muscle,
low levels of vitamin B2 in eggs and associated fatal embryonic abnormalities (Van den Burg
2000). This may also play a role in other long-living bird species.
A cumulative effect of the above is a decrease in the food supply for insectivores. This means that
(at times sometimes small) effects of nitrogen deposition can be passed onto species from higher
70
Part I - November 2012 version -
trophic levels. A reduction in the food supply is caused by both a lower abundance, lower
diversity and reduced visibility or accessibility to potential prey, and by a shift from large to small
invertebrate species. The consequences of a reduced average-sized prey for insectivores (birds
and bats) are described by scholars such as Siepel 1990, Beintema et al. 1991 en Schekkerman &
Beintema 2007. The combined effect of smaller prey and a less diverse supply of prey (which may
cause a temporary lack of food) in relation to nitrogen-related fallowing in dune areas, is
described by Kuper et al. (2000).
2.5 Literature
Adams, M.B. 2003. Ecological issues related to N deposition to natural ecosystems: research
needs. Environment International 29: 189– 199.
Aerts, R. & F.S. Chapin 2000. The mineral nutrition of wild plants revisited: A re-evaluation of
processes and patterns. Advances in Ecological Research 30: 1-67.
Aerts, R. & R. Bobbink 1999. The impacts of atmospheric nitrogen deposition on vegetation
processes in non-forest ecosystems. In: Langan, S.J. (ed.) The impact of nitrogen
deposition on natural and semi-natural ecosystems. Kluwer, Dordrecht, pp. 85-122.
Arnolds, E. 1991. Decline of ectomycorrhizal fungi in Europe. Agric. Ecosys. Environ. 35:
209-244.
Arts, G.H.P., G. van der Velde, J.G.M. Roelofs & C.A.M. van Swaay 1990b. Successional
changes in the soft-water macrophyte vegetation of (sub)atlantic, sandy, lowland
regions during this century. Freshwat. Biol. 24: 287-294.
Arts, G.H.P., J.G.M. Roelofs & M.J.H. de Lyon 1990a. Differential tolerances among soft-water
macrophyte species to acidification. Can. J. Bot. 68: 2127-2134.
Asman, W.A.H., M.A. Sutton & J.K. Schjorring 1998. Ammonia: emission, atmospheric
transport and deposition. New Phytologist 139: 27-48.
Awmack, C.S. & S.R Leather 2002. Host plant quality and fecundity in herbivorous insects.
Annu. Rev. Entomol. 47: 817–44.
Beintema, A.J. 1997. European black terns (Chlidonias niger) in trouble: examples of dietary
problems. Colon Waterbirds 20:558–565.
Beintema, A.J., J.B. Thissen, D. Tensen & G.H. Visser 1991. Feeding ecology of charadriiform
chicks in agricultural grassland. Ardea 79: 31-44.
Berdowski, J J M. 1987. The catastrophic death of Calluna vulgaris in Dutch heathland.
Dissertatie Utrecht, 132 p.
Berendzen, N., R. Schulz & M. Liess 2001. Effects of chronic ammonium and nitrite
contamination on the macroinvertebrate community in running water microcosms. Wat.
Res. Vol. 35: 3478–3482.
Berg B. & E. Matzner 1997. Effect of N deposition on decomposition of plant litter and soil
organic matter in forest systems. Environmental Reviews 5: 1-25.
Biesmeijer, J.C., S.P.M. Roberts, M. Reemer, R. Ohlemüller, M. Edwards, T. Peeters, A.P.
Schaffers, S.G. Potts, R. Kleukers, C.D.Thomas, J.Settele & W.E. Kunin 2006. Parallel
declines in pollinators and insect-pollinated plants in Britain and the Netherlands.
Science 313: 351-354.
Bink, F.A. 1992. De ecologische atlas van de dagvlinders van Noordwest Europa. Schuyt & Co,
512 pp. (Translation: The ecological atlas of the diurnal butterflies of North West Europe.)
71
Part I - November 2012 version -
Bobbink, R. & Hettelingh J.P. (eds.) 2011. Review and revision of empirical critical loads and
dose response relationships . Proceedings of an expert workshop, Noordwijkerhout,
23-25 June 2010. CCE/RIVM, Bilthoven.
Bobbink, R. & L.P.M. Lamers 2002. Effects of increased nitrogen deposition. Air pollution and
plant life 2nd edition (eds J.N.B. Bell, M. Treshow), pp. 201-235. John Wiley & Sons,
Ltd, Oxford.
Bobbink, R. & Lamers, L.P.M. 1999. Effecten van stikstofhoudende luchtverontreiniging op
vegetaties; een overzicht. Rapport R13 Technische Commissie Bodembescherming,
Den Haag. (Translation: The effects of nitrogen pollution on vegetation: an overview.)
Bobbink, R. 1991. Effects of nutrient enrichment in Dutch chalk grassland. Journal of Applied
Ecology 28: 28-41.
Bobbink, R., H. Tomassen, M. Weijters & J-P. Hettelingh 2010c. Revisie en update van
kritische N-depositiewaarden voor Europese Natuur. De Levende Natuur 111: 254258. (Translation: Revision and update of critical N-deposition values for European Nature.)
Bobbink, R., Heil, G.W. & Scheffers, M. 1990. Atmosferische depositie van NOx op
bermvegetaties langs autosnelwegen. Vakgroep Botanische Oecologie en
Evolutiebiologie, Rijksuniversiteit te Utrecht, 64 pp. (Translation: The atmospheric deposition
of NOx on roadside vegetation along motorways.)
Bobbink, R., Hornung, M. & Roelofs, J.G.M. 1998. The effects of air-borne nitrogen pollutants
on species diversity in natural and semi-natural vegetation - a review. Journal of
Ecology 86: 717-738 .
Bobbink, R., K. Hicks, J. Galloway et al. 2010a. Global assessment of nitrogen deposition
effects on plant terrestrial biodiversity: a synthesis. Ecological Applications 20: 30-59.
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 plant terrestrial biodiversity: a synthesis. Ecological Applications 20: 30-59.
Bobbink, R., M. Ashmore, S. Braun, W. Fluckiger, I.J.J. van den Wyngaert 2003. Empirical
nitrogen critical loads for natural and semi-natural ecosystems: 2002 update. In: B.
Achermann & R. Bobbink (eds.) Empirical critical loads for nitrogen. Environmental
Documentation No. 164 Air, pp. 43-170. Swiss Agency for Environment, Forest and
Landscape SAEFL, Berne.
Britto, D.T. & H.J. Kronzucker 2002. NH4+ toxicity in higher plants: a critical review. Journal
of Plant Physiology 159: 567-584.
Camargo, J.A. & A. Alonso 2006. Ecological and toxicological effects of inorganic nitrogen
pollution in aquatic ecosystems: A global assessment. Environment International 32:
831–849.
Camargo, J.A., A. Alonso & A. Salamanca 2006. Nitrate toxicity to aquatic animals: a review
with new data for freshwater invertebrates. Chemosphere 58: 1258-1267.
Clark, C., Cleland, E.E., Collins, S.C., Fargione, J., Gough, L., Gross, K.L., Pennings, S., Suding,
K.N., Grace, J. 2007. Environmental and plant community determinants of species loss
following nitrogen enrichment. Ecology Letters 10: 596-607.
Clark, C.M. & D. Tilman 2008. Loss of plant species after chronic low-level nitrogen
deposition to prairie grassland. Nature 451: 712-715.
De Graaf, M.C.C. P.J.M. Verbeek, M.J.R. Cals & J.G.M. Roelofs 1994. Effectgerichte
maatregelen tegen verzuring en eutrofiering van matig mineraalrijke heide en
schraallanden. Eindrapport Monitoring eerste fase. Vakgroep oecologie, KUN,
72
Part I - November 2012 version -
Nijmegen, 248 p. (Translation: Effect-oriented measures against the acidification and
eutrophication of moderately mineral-rich heathlands and low-nutrient soils.)
De Graaf, M.C.C., Bobbink, R., Smits, N.A.C., Van Diggelen, R. & Roelofs, J.G.M. 2009.
Biodiversity, vegetation gradients and key biogeochemical processes in the heathland
landscape. Biological Conservation 142: 2191-2201.
De Graaf, M.C.C., R. Bobbink, J.G.M. Roelofs & P.J.M. Verbeek 1998. Differential effects of
ammonium and nitrate on three heathland species. Plant Ecology 135: 185-196.
De Haan, B.J., Kros, J., Bobbink, R., van Jaarsveld, J.A., De Vries, W. & Noordijk, H. 2008.
Ammoniak in Nederland. Rapport Planbureau voor de leefomgeving, 500125003,
Bilthoven. (Translation: Ammonia in the Netherlands.)
De Nie, H.W. 1997. Atlas van de Nederlandse zoetwatervissen. Media publishing,
Doetinchem, 151 pp. (Translation: Atlas of Dutch freshwater fish.)
De Schrijver, A., R. Devlaeminck, J. Mertens, K. Wuyts, M. Hermy & K. Verheyen 2007. On the
importance of incorporating forest edge deposition for evaluating exceedance of
critical pollutant loads. Applied Vegetation Science 10: 293-298.
De Vries, W. 2008. Verzuring: oorzaken, effecten, kritische belastingen en monitoring van de
gevolgen van ingezet beleid. Wageningen, Alterra Wageningen UR, Rapport 1699,
88pp. (Translation: Acidification: causes, effects, critical loads and monitoring of the effects of the
established policy.)
De Vries, W., A. Breeuwsma & F. de Vries 1989. Kwetsbaarheid van de Nederlandse bodem
voor verzuring. Een voorlopige indicatie in het kader van de Richtlijn "Ammoniak en
Veehouderij". Wageningen, DLO-Staring Centrum, Rapport 29, 74 pp. (Translation:
Vulnerability of Dutch soil to acidity. A preliminary indication under the 'Ammonia and Livestock'
Directive.)
De Vries, W., C. van der Salm, G.J. Reinds & J.W. Erisman 2007. Element fluxes through
intensively monitored forest ecosystems in Europe and their relationships with stand
and site characteristics. Environmental Pollution 148: 501–513.
De Vries, W., Klijn, J. & J. Kros 1994. Simulation of the long-term impact of atmospheric
deposition on Dune ecosystems in the Netherlands. Journal of Applied Ecology 31: 5973.
Dise, N.B, J.J. Rothwell, V. Gauci, C. van der Salm & W. de Vries 2009. Predicting dissolved
inorganic nitrogen leaching in European forests using two independent databases.
Science of the total Environment 407: 1798-1808.
Dise, N.B. & R.F. Wright 1995. Nitrogen leaching from European forests in relation to
nitrogen deposition. Forest Ecology and Management 71: 153-161.
Dise, N.B., M. Ashmore, S. Belyazid 2011. Nitrogen as a threat to European terrestrial
biodiversity. In: The European Nitrogen Assessment, ed. M. A. Sutton, C. M. Howard, J.
W. Erisman et al.Cambridge University Press.
Dorland, E. & A. van Loon 2011. Verkenning kwantificering processen ten behoeve van
herstelstrategieën PAS. KWR 2011.008. KWR, Nieuwegein. (Translation: Exploration of
quantification processes for programmed nitrogen approach restorative strategies.)
Draaijers, G. 1993. The variablility of atmospheric deposition to forests. The effects of
canopy structure and forest edges. PhD Thesis, University of Utrecht.
Ellenberg, H. 1988. Vegetation Ecology of Central Europe. Cambridge University Press,
Cambridge.
73
Part I - November 2012 version -
Erisman, J.W. 1990. Estimates of the roughness length at Dutch Air Quality Monitoring
stations and on a grid basis over the Netherlands, Report 723001003, National
Institute of Public Health and Environmental Protection, the Netherlands.
Erisman, J.W. 1992. Atmospheric deposition of acidifying compounds in the Netherlands.
Dissertatie University of Utrecht, The Netherlands.
Fenn,M.E., J.S. Baron, E.B. Allen, H.M. Rueth, K.R. Nydick, L. Geiser, W.D. Bowman, J.O.
Sickman, T. Meixner, D.W. Johnson & P. Neitlich 2003b. Ecological effects of nitrogen
deposition in the western United States. Bioscience 53: 404-420.
Fenn,M.E., R. Haeuber, G.S. Tonnesen, J.S. Baron, S. Grossman-Clarke, D. Hope, D.A. Jaffe, S.
Copeland, L. Geiser, H.M. Rueth & J.O. Sickman 2003a. Nitrogen emissions, deposition,
and monitoring in the western United States. Bioscience 53: 391-403.
Flückiger, W. & Braun, S. 2004. Wie geht es unserem Wald? Ergebnisse aus
Dauerbeobachtungsflächen von 1984 bis 2004, Bericht 2. Institut für Angewandte
Pflanzenbiologie, Schönenbuch, 67 pp. (Translation: What is the state of our forests? Results
from permanent observation plots between 1984 and 2004)
Fowler, D. 2002. Pollutant deposition and uptake by vegetation. In: In: J. N. B. Bell and M.
Treshow (eds), Air pollution and plant life, pp. 201-235. John Wiley & Sons,
Chichester.
Fowles, A.P. & R.G. Smith 2006. Mapping the Habitat Quality of Patch Networks for the Marsh
Fritillary Euphydryas aurinia (Rottemburg, 1775) (Lepidoptera, Nymphalidae) in Wales.
Journal of Insect Conservation, Volume 10: 161-177.
Freda, J. 1986. The influence of acidic pond water on amphibians: A review . Water, Air, & Soil
Pollution Volume 30: 439-450.
Freda, J. 1986. The influence of acidic pond water on amphibians: A review Water, Air, & Soil
Pollution Volume 30, Numbers 1-2.
Fründ, J., K.E. Linsenmair & N. Blüthgen 2010. Pollinator diversity and specialization in
relation to flower diversity. Oikos 119: 1581-1590.
Galloway, J.N. & E.B. Cowling 2002. Reactive nitrogen and the world: 200 years of change.
Ambio 31: 64-71.
Gerhardt, A. 1993. Review of impact of heavy metals on stream invertebrates with special
emphasis on acid conditions Water, Air, & Soil Pollution. Volume 66: 289-314.
Gigon, A. & I.H. Rorison 1972. The response of some ecologically distinct plant species to
nitrate- and to ammonium-nitrogen. Journal of Ecology 60: 93-102.
Gilliam, F.S. 2006. Response of the Herbaceous Layer of Forest Ecosystems to Excess
Nitrogen Deposition. Journal of Ecology 94: 1176-1191.
Goulson, D., W. Hughes, L. Derwent & J. Stout 2002. Colony growth of the bumblebee,
Bombus terrestris, in improved and conventional agricultural and suburban habitats.
Oecologia 130: 267-273.
Graveland, J., R. van der Wal, J.H. van Balen en A.J. van Noordwijk 1994. Poor reproduction in
forest passerines from decline of snail abundance on acidified soils. Nature 368, 446448.
Grennfelt, P & H. Hultberg 1986. Effects of nitrogen deposition on the acidification of
terrestrial and aquatic ecosystems. Water, air & Soil Pollution 30: 945-963.
Guérold F., J.P. Boudot, G. Jacquemin, D. Vein, D. Merlet & J. Rouillet 2000. Macroinvertebrate
community loss as a result of headwater stream acidification in the Vosges Mountains
(N-E France). Biodiversity Conservation 9: 767–783.
74
Part I - November 2012 version -
Güsewell, S. 2004. N: P Ratios in Terrestrial Plants: Variation and Functional Significance. New
Phytologist 164: 243-266.
Hagedoorn F., D. Spinnler & R. Siegwolf 2003. Increased N deposition retards mineralisation
of old soil organic matter. Soil Biology and Biochemistry 35: 1683-1692.
Hautier, Y., Niklaus, P.A. & Hector, A. 2009. Competition for light causes plant biodiversity
loss after eutrophication. Science 324: 636-638.
Heil, G.W., Werger, M.J.A., De Mol, W., Van Dam, D., & Heijne, B. 1988. Capture of
atmospheric ammonium by grassland canopies. Science 239: 764-765.
Hoover, S.E.R., J.J. Ladley, A.A. Shchepetkina, M. Tisch, S.P. Gieseg & J.M. Tylianakis 2012.
Warming, CO2, and nitrogen deposition interactively affect a plant-pollinator
mutualism. Ecology Letters 15: 227–234.
Ivens, W. 1990. Atmospheric deposition onto forests. PhD Thesis, University of Utrecht, The
Netherlands.
Janssens, I.A., W. Dieleman & S. Luyssaert 2010. Nitrogen deposition retards forest soil
respiration. Nature geoscience 3: 315 – 322.
Jeppesen, E., J.P. Jensen, M. Sondergaard, T. Lauridsen & F. Landkildehus 2000. Trophic
structure, species richness and biodiversity in Danish lakes: changes along a
phosphorous gradient. Freshwater Biology 45: 201-218.
Kinzel, S. 1982. Pflanzenökologie und Mineralstoffwechsel. Ulmer, Stuttgart. (Translation: Plant
ecology and mineral metabolism)
Kleijn, D. & I. Raemakers 2008. A retrospective analysis of pollen host plant use by stable
and declining bumble bee species. Ecology 89: 1811-1823.
Kleijn, D., Bekker, R.M., Bobbink, R., De Graaf, M.C.C. & Roelofs, J.G.M. 2008. In search for
key biogeochemical factors affecting plant species persistence in heathland and acidic
grasslands: a comparison of common and rare species. Journal of Applied Ecology 45:
680-687.
Kleukers, R.M.J.C., E.J. van Nieukerken, B. Odé, L.P.M. Willemse & W.K.R.E. van Wingerden,
1997. De sprinkhanen en krekels van Nederland (Orthoptera). Nederlandse Fauna 1.
Nationaal Natuurhistorisch Museum, KNNV-Uitgeverij & EIS-Nederland, Leiden.
(Translation: The grasshoppers and crickets of the Netherlands (Orthoptera).)
Kooijman, A. M. & M. Besse 2002. The higher availability of N and P in lime-poor than in
lime-rich coastal dunes in the Netherlands. Journal of Ecology 90: 394-403.
Kooijman, A. M., H. Noordijk, A. van Hinsberg & C. Cusell 2009. Stikstofdepositie in de
duinen - een analyse van N-depositie, kritische niveaus, erfenissen uit het verleden en
stikstofefficiёntie in verschillende duinzones. Universiteit van Amsterdam & Planbureau
voor de Leefomgeving. 56 p. (Translation: Nitrogen deposition in the dunes - an analysis of Ndeposition, critical levels, legacies of the past and nitrogen efficiency in different dune areas.)
Kros, J., B.J. de Haan, R. Bobbink, J.A. van Jaarsveld, J.G.M. Roelofs & W.de Vries 2008.
Effecten van ammoniak op de Nederlandse natuur. Wageningen, Alterra-rapport 1698,
132 p. (Translation: Effects of ammonia on Dutch nature.)
Krupa, S.V. 2003. Effects of atmospheric ammonia (NH3) on terrestrial vegetation: a review.
Environmental Pollution 124: 179-221.
Kuper, J., G.J. Van Duinen, M. Nijssen, M. Geertsma & H.Esselink 2000. Is the decline of the
Red-backed Shrike (Lanius collurio) in the Dutch coastal dune area caused by a
decrease in insect diversity? Ring 22 : 11-25.
LeBauer, D.S. & K.K. Treseder 2008. Nitrogen limitation of net primary productivity in
terrestrial ecosystems is globally distributed. Ecology 89: 371-379.
75
Part I - November 2012 version -
Lensink, B.M. 1963. Distribution ecology of some Acrididae (Orthoptera) in the dunes of
Voorne, Netherlands - Tijdschr. Entomol. 106: 357-443.
Leuven, R.S.E.W & F.G.F. Oyen 1987. Impact of acidification and eutrophication on the
distribution of fish species in shallow and lentic soft waters of The Netherlands: an
historical perspective. Journal of Fish Biology 31: 753-774.
Leuven, R.S.E.W, H.L.M. Kersten, J.A.A.R. Schuurkes, J.G.M. Roelofs & G.H.P. Arts 1986a.
Evidence for recent acidification of lentic soft waters in The Netherlands. Water Air &
Soil Pollution 30: 387-392.
Leuven, R.S.E.W., C. den Hartog, M.M.C. Christiaans & W.H.C. Heijligers 1986b. Effects of
water acidification on the distribution pattern and the reproductive success of
amphibians. Experientia 42: 495-503.
Leuven, R.S.E.W., G. van der Velde & H.L.M. Kersten 1992. Interrelation betweens pH and
other Physio-chemical factors of Dutch soft waters. Arch. Hydrobiology 126: 27-51.
Lindberg, N. & J. Bengtsson 2006. Recovery of forest soil fauna diversity and composition
after repeated summer droughts. Oikos 114: 496-506.
Loeb, R., P.F.M. Verdonschot, F. Kragt & H. van Grinsven 2010. Sturen op fosfor of stikstof
voor verbetering ecologische kwaliteit van zoete wateren. H2O 22: 32-34. (Translation:
Steering towards phosphorus or nitrogen for the improvement of the ecological quality of freshwaters)
Lucassen, E.C.H.E., Bobbink, R., Smolders, A.J.P., Van der Ven, P.J.M., Lamers, L.P.M. &
Roelofs, J.G.M. 2003. Interactive effects of low pH and high ammonium levels
responsible for the decline of Cirsium dissectum (L.) Hill. Plant Ecology 165: 45–52.
Morecroft, M.D., E.K. Sellers & J.A. Lee 1994. An experimental investigation into the effects of
atmospheric nitrogen deposition on two semi-natural grasslands. Journal of Ecology.
82: 475-483.
Nadelhoffer, K.J., B.P. Colman, W.S. Currie, A. H. Magill & J.D. Aber 2004. Decadal scale fates
of 15N tracers added to oak and pine stands under ambient and elevated N inputs at
the Harvard Forest (USA). Forest Ecology and Management 196: 89-107.
Nihlgård, B. 1985. The ammonium hyphothesis - an additional explanation to the forest
dieback in Europe. Ambio 14: 2-8.
Nijssen, M., M.J.P.M. Riksen, L. Sparrius, L. Kuiters, A. Kooiman, R.J. Bijlsma, P. Jungerius, A.
van den Burg, H. van Dobben, R. Ketner-Oostra, C. van Swaay, C. van Turnhout & R. de
Waal 2011. Effectgerichte maatregelen voor het herstel en beheer van stuifzanden OBN stuifzandonderzoek 2006-2010. Rapport nr. 2011/OBN144-DZ. Directie Kennis
en Innovatie, Ministerie van Economische Zaken, Landbouw en Innovatie, Den Haag.
(Translation: Targeted measures for the restoration and management of drifting sands - OBN shifting
sands research 2006-2010.)
Ockinger, E. O. Hammarstedt, S.G. Nilsson & H.G. Smith 2006. The relationship between local
extinctions of grassland butterflies and increased soil nitrogen levels. Biological
Conservation 128: 564 –573.
Ohlson, M., A. Nordin & T. Näsholm 1995. Accumulation of Amino Acids in Forest Plants in
Relation to Ecological Amplitude and Nitrogen Supply. Functional Ecology 9: 596-605.
Økland, J. 1992. Effects of acidic water on freshwater snails: results from a study of 1000
lakes throughout Norway. Environmental pollution 78: 127-130.
Paulissen, M.P.C.P., Van der Ven, P.J.M., Dees, A.J. & Bobbink, R. 2004. Differential effects of
nitrate and ammonium on three fen bryophyte species in relation to pollutant nitrogen
input. The New Phytologist 164: 451-458.
76
Part I - November 2012 version -
Peeters, T.M.J., C. van Achterberg, W.R.B. Heitmans, W.F. Klein, V. Lefeber, A.J. van Loon, A.A.
Mabelis, H. Nieuwenhuijsen, M. Reemer, J. de Rond, J. Smit & H.H.W. Velthuis 2004. De
wespen en mieren van Nederland (Hymenoptera: Aculeata). - Nederlandse Fauna 6,
NNM Naturalis Leiden, KNNV Uitgeverij Utrecht & EIS-Nederland Leiden, 507 p.
(Translation: The wasps and ants of the Netherlands)
Rabalais, N.N. 2002. Nitrogen in Aquatic Ecosystems. Ambio 31: 102-112.
Roelofs J.G.M., L.G.M. Clasquin, I.M.C. Driessen & A.J. Kempers 1984. De gevolgen van
zwavel en stikstofhoudende neerslag op de vegetatie in heide- en heideven-milieus.
In: E.H. Adema and J. van Ham, Editors, Proc. Symp. Zure Regen, 's-Hertogenbosch 17–
18 November 1983, Zure Regen, oorzaken, effecten en beleid, Pudoc, Wageningen.
(Translation: The effects of sulphur and nitrogen deposition on the vegetation in heath and moorland
pool environments.)
Roelofs, J.G.M., A.J. Kempers, L.F.M. Houdijk & J. Jansen 1985. The effect of air-borne
ammonium sulphate on Pinus nigra var. maritima in the Netherlands. Plant and Soil 84:
45-56.
Roelofs, J.G.M., Bobbink, R., Brouwer, E. & de Graaf, M.C.C. 1996. Restoration ecology of
aquatic and terrestrial vegetation of non-calcareous sandy soils in the Netherlands.
Acta Botanica Neerlandica 45: 517-541.
Sala, O.E., F.S. Chapin & J.J. Armesto 2000. Global biodiversity scenarios for the year 2100.
Science 287: 1770–1774.
Salz, A. & T. Fartmann 2009. Coastal dunes as important strongholds for the survival of the
rare Niobe fritillary (Argynnis niobe). Journal of Insect Conservation 13: 643-654.
Schekkerman, H. & A.J. Beintema 2007. Abundance of invertebrates and foraging success of
Black-tailed Godwit Limosa limosa chicks in relation to agricultural grassland
management. Ardea 95: 39-54.
Schekkerman, H. 2008. Precocial problems. Shorebird chick performance in relation to
weather, farming, and predation. Thesis RU Groningen/ Alterra Scientific Contributions
24, 228 pp.
Schirmel, J., J. Mantilla-Contreras, I. Blindow & T. Fartmann 2011. Impacts of succession and
grass encroachment on heathland Orthoptera. Journal of Insect Conservation 15: 633–
642.
Schuurkes, J.A.A.R., I.C.C. Heck, P.L.G.M. Hesen, R.S.E.W. Leuven & J.G.M. Roelofs 1986.
Effects of sulphuric acid and acidifying ammonium deposition on water quality and
vegetation of simulated soft water ecosystems. Water, Air, and Soil Pollution 31: 267-
272.
Schuurkes, J.A.A.R., Kok, C.J. & Den Hartog, C. 1986. Ammonium and nitrate uptake by
aquatic plants from poorly buffered and acidified waters. Aquatic Botany 24: 131-146.
Schuurkes, J.A.A.R., M.A. Elbers, J.J.F. Gudden & J.G.M. Roelofs 1987. Effects of simulated
ammonium sulphate and sulphuric acid rain on acidification, water quality and flora of
small-scale soft water systems. Aquat. Bot. 28: 199-225.
Siepel, H. 1990. The influence of management on food size in the menu of insectivorous
animals. In: Sommeijer, M.J. & J. van der Blom (eds) Experimental and applied
entomology, Proc. Neth. Entomol. Soc. Amsterdam Vol I. p 69-74.
Siepel, H. 1996. The importance of unpredictable and short-term environmental extremes
for biodiversity in oribatid mites. Biodiversity Letters 3: 26-34.
77
Part I - November 2012 version -
Sowig, P. 1989. Effects of flowering plant's patch size on species composition of pollinator
communities, foraging strategies, and resource partitioning in bumblebees
(Hymenoptera: Apidae). Oecologia 78: 550-558.
Stevens, C.J., Dise, N.B., Mountford, J.O. & D.J. Gowing 2004. Impact of nitrogen deposition
on the species richness of grasslands. Science 303: 1876 – 1879.
Stevens, C.J., Duprè, C., Dorland E. et al. 2010. Nitrogen deposition threatens species
richness of grasslands across Europe. Environmental Pollution 158: 2940-2945.
Stevens, C.T., Manning, P., van den Berg, L.J.L. et al. 2011. Ecosystem responses to reduced
and oxidised nitrogen inputs in European terrestrial habitats. Environmental Pollution
159: 665-676.
Sutton, M.A., Reis, S. & Baker, S.M.H. (eds) 2009. Atmospheric ammonia- Detecting emission
changes and environmental impacts. Springer.
Throop, H.L. & M.T. Lerdau 2004. Effects of nitrogen deposition on insect herbivory:
implications for community and ecosystem processes. Ecosystems 7: 109 – 133
Throop, H.L. 2005. Nitrogen deposition and herbivory affect biomass production and
allocation in an annual plant. Oikos 111: 91 – 100.
Thurston, R.V., R.C. Russo & G.A. Vinogradov 1981. Ammonia Toxicity to Fishes. Effect of pH
on the Toxicity of the Un-ionized Ammonia Species. Environmental Science &
Technology, Volume 15: 837–840.
Tietema, A., L. Riemer, J.M. Verstraten, M.P. van der Maas, A.J. van Wijk & I. Van Voorthuysen.
1993. Nitrogen transformations in four acid forest soils subject to increased nitrogen
deposition. Forest Ecology and Management 57: 29-44.
Tyler, S.J. & S.J. Oremerond 1992. A review of the likely causal pathways relating the reduced
density of breeding dippers Cinclus cinclus to the acidification of upland streams.
Environmental pollution 78: 49-55.
Ulrich, B. 1981. Okologische Gruppierung von Boden nach ihrem chemischen Bodenzustand.
Z. Pflanzenernähr. Bodenk. 144: 289-305. (Translation: Ecological soil groupings based on
their chemical soil condition)
Van den Berg, L.J., C.J. Peters, M.R. Ashmore & J.G.M. Roelofs 2008. Reduced nitrogen has a
greater effect than oxidised nitrogen on dry heathland vegetation. Environmental
Pollution 154: 359-369.
Van den Berg, L.J.L., Dorland, E., Vergeer, P., Hart, M.A.C., Bobbink, R. & Roelofs, J.G.M.
2005. Decline of acid-sensitive plant species in heathland can be attributed to
ammonium toxicity in combination with low pH. New Phytologist 166: 551-564.
Van den Burg, A.B. 2000. The causes of egg hatching failures in wild birds
studied in the Barn Owl Tyto alba and the Sparrowhawk Accipiter nisus. Dissertatie
University of Nottingham.
Van Dijk, H.F.G. 1993. Excess nitrogen deposition: a stress factor in Dutch plantation forests.
PhD thesis, Univeristy of Nijmegen, The Netherlands.
Van Ek, R. & G.J.P. Draaijers 1991. Atmospheric deposition in relation to forest stand
structure. Report Dept. Of Physical Geography, University of Utrecht, The Netherlands.
Van Noordwijk, C.G.E. 2011. Using life-history analysis to improve restoration management.
Proceedings of the Netherlands Entomological Society Meeting 22: 79-89.
Van Noordwijk, T., P. Boer, A.A. Mabelis, W.C.E.P. Verberk & H. Siepel 2011 (subm.). Lifehistory strategies as a tool to identify restoration constraints: A case-study on ants in
chalk grasslands. Ecological Indicators.
78
Part I - November 2012 version -
Van Turnhout, C.A.M., R.S.E.W. Leuven, A.J. Hendriks, G. Kurstjens, A.J. van Strien, R.P.B.
Foppen & H. Siepel 2010, in press. Ecological strategies successfully predict the effects
of river floodplain rehabilitation on breeding birds. River Research and Applications.
Van Wingerden, W.K.R.E., A.R. van Kreveld & W. Bongers 1992. Analysis of species
composition and abundance of grasshoppers (Orth. Acrididae) in natural and fertilized
grasslands. Journal of Applied Entomology 113: 138–152.
Van Wingerden, W.K.R.E., J.C.M. Musters, & F.I.M. Maaskamp 1991. The influence of
temperature on the duration of egg development in west European grasshoppers
(Orthoptera: Acrididae). Oecologia 87: 417– 423.
Vangenechten, J.H.D, H. Witters & O.L.J. Vandenborght 1989. Laboratory studies on
invertebrate survival and physiology in acid waters. In: Morris et al., 1989. Acid toxicity
and aquatic animals. Cambridge University Press, 282 p.
Verhoeven, J.T.A., Beltman, B., Dorland, E., Robat, S.A. & R. Bobbink 2011. Differential effects
of ammonium and nitrate deposition on fen phanerogams and bryophytes. Applied
Vegetation Science 14: 149-157.
Verstraten, J.M., T. Tietema & J.C.R. Dopheide 1989. Bodemverzuring: principes en
voorbeelden. K.N.A.G. Geografisch Tijdschrift XXIII: 251. (Translation: Soil acidification:
principles and examples.)
Vitousek, P.M. & R.W. Howarth 1991. Nitrogen limitation on land and in the sea: How can it
occur? Biogeochemistry 13: 87-115.
Wallis de Vries, M.F. & C.A.M. van Swaay 2006. Global warming and excess nitrogen may
induce butterfly decline by microclimatic cooling. Global change Biology 12: 1601626.
Weiss, S. B. 1999. Cars, Cows, and Checkerspot Butterflies: Nitrogen Deposition and
Management of Nutrient-Poor Grasslands for a Threatened Species. Conservation
Biology 13: 1467-1486.
Westphal, C., I. Steffan-Dewenter & T. Tscharntke 2006. Bumblebees experience landscapes
at different spatial scales: possible implications for coexistence. Oecologia 149: 289300.
Wouters, B., M. Nijssen, J. Vogels & R. Verdonschot 2009. Eindrapport Voorbereidingsplan
Kustduinen. Rapport Stichting Bargerveen, Nijmegen. (Translation: Final report: preparatory
plan for coastal dunes.)
Wuyts K. 2009. Patterns of throughfall deposition, nitrate seepage, and soil acidification in
contrasting forest edges. Ph.D. thesis, Ghent University, Belgium, 202p. ISBN-number:
978-90-5989-283-5.
Wuyts, K., A. De Schrijver, F. Vermeiren & K. Verheyen 2009. Gradual forest edges can
mitigate edge effects on throughfall deposition if their size and shape are well
considered. Forest Ecology and Management 257: 679-687.
Wuyts, K., A. De Schrijver, J. Staelens, L. Gielis, J. Vandenbruwane & K. Verheyen 2008b.
Comparison of forest edge effects on throughfall deposition in different forest types
Environmental Pollution 156: 854-861.
Wuyts, K., A. De Schrijver, J. Staelens, M. Gielis, G Geudens, & K. Verheyen 2008a. Patterns of
throughfall deposition along a transect in forest edges of silver birch and Corsican
pine. Canadian Journal of Forest Research 38: 449-461.
79
Part I - November 2012 version -