Long-term forest soil acidification, nutrient leaching and vegetation development: Linking modelling and surveys of a primeval spruce forest in the Ukrainian Transcarpathian Mts.

1. Ecological Modelling 244 (2012) 28–37
Contents lists available at SciVerse ScienceDirect
Ecological Modelling
journal homepage: www.elsevier.com/locate/ecolmodel
Long-term forest soil acidification, nutrient leaching and vegetation
development: Linking modelling and surveys of a primeval spruce forest
in the Ukrainian Transcarpathian Mts.
J. Hruˇskaa,∗
, F. Oulehlea,b
, P. ˇSamonilc
, J. ˇSebestad
, K. Tahovskáe
, R. Hlebd
, J. Houˇskad
, J. ˇSikla
a
Czech Geological Survey, Klárov 3, 118 21 Prague 1, Czech Republic
b
Centre for Ecology and Hydrology, Deiniol Road, Bangor LL57 2UW, United Kingdom
c
Department of Forest Ecology, The Silva Tarouca Research Institute for Landscape and Ornamental Gardening, Lidická 25/27, 657 20 Brno, Czech Republic
d
Faculty of Forestry and Wood Technology, Mendel University in Brno, Zemˇedˇelská 1, 613 00 Brno, Czech Republic
e
Department of Ecosystem Biology, Faculty of Science, University of South Bohemia, Braniˇsovská 31, 370 05 ˇCeské Budˇejovice, Czech Republic
a r t i c l e i n f o
Article history:
Received 6 March 2012
Received in revised form 19 June 2012
Accepted 22 June 2012
Available online 28 July 2012
Keywords:
Soil acidification
MAGIC model
Nitrogen leaching
Acid deposition
Transcarpathian Mts.
Norway spruce
Vegetation changes
a b s t r a c t
The biogeochemical model MAGIC was applied to simulate long-term (1880–2050) soil and stratified soil
solution (30 and 90 cm depth) chemistry at a spruce dominated site in the western Ukraine (Pop Ivan,
1480 m a.s.l.) to evaluate the effects of acid deposition on soil acidification in a less polluted region of
Europe.
Since 2008, sulphur (S) deposition of 9 kg ha−1
year−1
and nitrogen (N) deposition of 8.5 kg ha−1
year−1
have been measured at Pop Ivan. The recent deposition of S and N is about 30% and 50% of those values
estimated for the early 1980s, respectively. Acidic deposition caused the depletion of base cations (Ca, Mg,
Na, K) from the soil cation exchange complex, which resulted in a decrease of calcium and magnesium
saturation between 1935 and 2008 in the top mineral soil (0–30 cm) and deeper mineral soil (30–80 cm)
by 67% and 88%, respectively. Base cation leaching acted as the major buffer mechanism against incoming
acidity, therefore the measured inorganic aluminium (Al) concentration in soil solutions is ca. 10 ␮mol L−1
and the subsequent molar (Ca + Mg + K)/Al ratio above 1. Recovery of the soil solution pH and Al is expected
within the next 40 years, whereas the soil base saturation will only increase slowly, from 6% to 9.8% in the
top soil and from 5.5% to 11% in the deeper mineral soil. Since the 1960s, modelled inorganic N leaching
(as NO3) has started to increase following the trend in N deposition. Modelling and experimental evidence
suggest that N availability from mineralization and deposition exceeds the rate of microbial and plant
immobilization. Thus, soil N accumulation since the 1960s has been limited. A significant increase in
nitrophilous species as well as a decrease of herb layer diversity was observed between 1936 and 1997.
© 2012 Elsevier B.V. All rights reserved.
1. Introduction
Emissions of potentially acidifying compounds have accelerated
due to anthropogenic activities since the mid 19th century. Conse-
quently, the transport and deposition of sulphur (S) and nitrogen
(N) have resulted in the widespread acidification of many terrestrial
and aquatic ecosystems (Reuss and Johnson, 1985). International
agreements to reduce emissions of sulphur and nitrogen com-
pounds (Bull et al., 2001) have resulted in an 80% reduction in SOx,
30% reduction in NH3 and 39% reduction in NOx emissions from EU
countries since 1985. As the deposition of S has fallen dramatically
across Europe, many surface waters have begun to recover from
acidification, as manifested by trends of sulphate (SO4), increased
∗ Corresponding author. Tel.: +420 251085433.
E-mail address: jakub.hruska@geology.cz (J. Hruˇska).
pH and increased acid neutralising capacity (ANC) (Evans et al.,
2001; Jenkins et al., 2003; Wright et al., 2005). Despite the fact
that base cation leaching from soils has decreased, recovery of
soils is a much slower process, limited by low weathering rates,
declines in base cation deposition (Hedin et al., 1994) and by net
biomass uptake at some sites (Hruˇska et al., 2002). As S deposition
has decreased, N deposition has become increasingly important
because of its contribution to soil acidification and eutrophication.
It has been shown by studies examining the effects of atmo-
spheric deposition on vegetation; acidification and eutrophication
are among the most important factors responsible for diversity
losses in the long term (Thimonier et al., 1994; Bobbink et al., 2010;
Dupre et al., 2010; Maskell et al., 2010; van den Berg et al., 2011).
Over the past few decades, significant changes in the species diver-
sity and equitability of plant communities, increases of nitrophilous
and acidophilous species and herb generalists in forests
have revealed a common causative connection across various
0304-3800/$ – see front matter © 2012 Elsevier B.V. All rights reserved.
http://dx.doi.org/10.1016/j.ecolmodel.2012.06.025

2. J. Hruˇska et al. / Ecological Modelling 244 (2012) 28–37 29
Fig. 1. Location of research sites Pop Ivan (Ukraine) and Naˇcetín (Czech Republic).
ecosystems (Nygaard and Odegaard, 1999; Hedl, 2004; ˇSamonil
and Vrˇska, 2007, 2008; Durak, 2010; ˇSebesta et al., 2011). In
Eastern Europe, homogenization of the Carpathian beech forests
together with the acidification of topsoil was observed in Poland by
Durak (2010). A similar extent of soil acidification and vegetation
changes were described in beech- and spruce-dominated forests
further east in the Ukraine (ˇSebesta et al., 2011), an area which
includes the site described in this study. Therefore, deposition of
acid pollutants should be considered as a potential threat to the soil
and consequent vegetation development in those ecosystems. It
should also be pointed out that at the same time scale, concurrent
changes in forest management, game pressure, cattle grazing and
climate change have also been documented, all of which are likely
to alter soils and vegetation composition.
Repeated historical soil surveys, originally established in
the 1930s in primeval fir-beech-spruce forests in the Ukraine
Carpathian Mts. (ˇSebesta et al., 2011; Kuˇcera et al., submitted for
publication), and recent intensive measurements of soil solution
chemistry and deposition in the spruce dominated site of Pop Ivan
(Oulehle et al., 2010), allowed us to reconstruct and predict the
soil and soil solution chemistry using the biogeochemical model
MAGIC. The specific objectives of this study are: (1) model the
extent of soil acidification over the last 130 years, (2) predict
the future soil chemistry based on expected trends in S and N
deposition until 2050, and (3) discuss the possible effects of soil
acidification and N cycle alterations on observed changes in the
vegetation composition.
2. Materials and methods
2.1. Site description
The Pop Ivan site is a primeval coniferous forest situated near
the border between Ukraine and Romania at 1480 m a.s.l. (24◦31 E;
47◦57 N) (Fig. 1). Study site corresponds to research plot 11b
(1.50 ha) established by Zlatnik in the 1930s (Zlatník et al., 1938).
The forest cover is dominated by Norway spruce (Picea abies (L.)
Karsten). Soils were classified as Haplic Cambisols (Dystric) or Entic
Podzols; Albic Podzols also rarely occur in the vicinity of study
site at the highest elevations (classified according to Michéli et al.,
2006). The bedrock consists of acid sensitive crystalline schist and
gneiss on study site as well as in surroundings. The Pop Ivan site is
situated on a steep slope oriented to the west; with a mean annual
temperature of 2 ◦C and annual precipitation of 1.9 m. Because
direct human impacts have been minimal in this area, Pop Ivan is
probably among the most natural forests of such extent in Eastern
Central Europe (Elbakidze and Angelstam, 2007).
2.2. Sampling and chemical analyses of precipitation, soil and soil
water
Precipitation (5 collectors for throughfall and 2 collectors for
bulk precipitation) was collected monthly by polyethylene funnels
(surface area of 122 cm2). During the winter season (October–April)
high volume samplers (surface area of 990 cm2) for bulk and
throughfall (surface area of 179 cm2) were used. The contents of
the throughfall samplers were combined to create one sample for
chemical analysis; bulk precipitation collectors were analyzed sep-
arately.
Soil water has been collected since September 2007 using suc-
tion lysimeters at depths of 30 and 90 cm in the mineral soil (3
lysimeters in each depth). Zero-tension lysimeters were installed
under the forest floor (3 replications). Thickness of the forest floor
was 8.7 ± 2.5 cm. All lysimeter samples were collected monthly and
combined to create one sample from each depth for each month.
Ca, Mg, Na, K, Si and Al in soil water were determined by
flame atomic absorption spectrometry (FAAS). pH was measured
using a pH meter with a combination electrode (Radiometer model

3. 30 J. Hruˇska et al. / Ecological Modelling 244 (2012) 28–37
Fig. 2. Estimated trends of (a) sulphur, (b) nitrogen compounds and (c) sum of the base cations (SBC) at Pop Ivan for the period 1880–2050.
GK-2401C). Cl, SO4 and NO3 were measured by exchange ion chro-
matography. Samples processing and analysis were made in an
Accredited Testing Laboratory according to criteria of the ISO/IEC
17025:2005.
Soils were sampled by excavating four 0.5 m2 quantitative pits
using the method described by Huntington et al. (1988). The Ol plus
Of (litter plus fermented) horizons were sampled together as a sin-
gle sample, and then the Oh (humification) horizon (e.g. Zanella
et al., 2009) was sampled separately. Mineral soil was collected for
the depths of: 0–10, 10–20, 20–40 and 40–80 cm. The soil samples
were weighed, and then sieved after air-drying (mesh size of 5 mm
for organic horizons and 2 mm for mineral horizons). Soil moisture
was determined gravimetrically by drying at 105 ◦C. Exchange-
able cations were analyzed in 0.1 M BaCl2 extracts by FAAS. Cation
exchange capacity (CEC) was calculated as the sum of exchange-
able Ca, Mg, Na, K and TEA. Base saturation (BS) was determined as
the fraction of CEC associated with base cations. Total carbon (C)
and total nitrogen were determined using a Carlo-Erba Fisons 1108
analyser.
2.3. Historical soil analysis
To evaluate the historical analytical methods used to deter-
mine the soil chemistry, the original methods of Gedroiz (1926)
were used. Soils samples were extracted in 1 M NH4Cl at the soil
pH (exchange of soil cations with NH4
+). From the eluate, cal-
cium was determined as calcium oxalate (CaC2O4) by titration with
0.1 N KMnO4, Mg concentration was determined by weighing as
the Mg(C9H6ON)2·2H2O precipitate; K concentration by weighing
as potassium chloride (KClO4); and Na concentration by weigh-
ing as sodium chloride (NaCl). For more details see Kuˇcera et al.
(submitted for publication).
2.4. Water fluxes of the soil solution
To assess the water and element fluxes through the soil profiles
we used a measurement of the chloride (Cl) mass budget. Chlo-
rine compounds tend to be highly soluble in water and mobile in
soils, so atmospheric deposition and transport through terrestrial
ecosystems is rapid if there is active hydrologic flow. In addition,
small-watershed studies assume that weathering of Cl is negligible
compared to atmospheric deposition (Juang and Johnson, 1967).
Chlorine is a minor constituent of most rocks and has been assumed
to be relatively unreactive in forest ecosystems (Schlesinger, 1997).
Its content in metamorphic rock underlined our study site may
be considered as minor based on litogeochemical database of the
Czech Geological Survey. The main source of Cl in rivers is glob-
ally sandstones and shales (Graedel and Keene, 1996). Estimated
weathering of Cl from bedrock in Hubbard Brook experimental for-
est (mica schist and quartzite) has been estimated only to 1–2% of
average export in stream water (Lovett et al., 2005). On the other
hand, internal cycling of Cl (vegetation uptake and release from
soil organic matter) may have influence on Cl budget especially
in areas where deposition of Cl is lower than ca. 6 kg ha−1 year−1
(Svensson et al., 2012). The changes in internal Cl cycling assume
changes in biotic control over Cl for which we have no evidence.
However, natural forest at Pop Ivan site may be considered as rela-
tively sTab. with limited anthropogenic influence. Thus, Cl budget
has been used to calculate water flux trough different horizon as
follows:
X =
Clthf
Clss
where X is the water flux (mm) in the respective soil horizon, Clthf
is the chlorine flux in throughfall (mg m−2) and Clss is the respec-
tive soil solution Cl concentration. Solute fluxes were calculated by
multiplying the annual average of each solute by the water flux.
2.5. Trends in deposition of sulphur, nitrogen and base cations
Historical S and N depositions for Pop Ivan (Fig. 2a and b) were
estimated using Czech emissions of SO2, NOx and NH3 for the period
1880–2006. Despite a lack of relevant emission inventories from
Ukraine, the Czech emissions were tightly correlated with total
emissions from Poland, Slovakia and Romania (Berge, 1997) in
the period 1980–2006 (R2 = 0.98; p < 0.001 for SO2 and R2 = 0.93;
p < 0.001 for NOx). For more details of calculation see Oulehle et al.
(2010).
For future S and N deposition up to the year 2050 we used
the current models (Current Legislation scenarios – CLE) of the
Co-operative Programme for Monitoring and Evaluation of the
Long-range Transmission of Air Pollutants in Europe (EMEP,
http://www.ceip.at/emission-data-webdab), which project lower
emissions of S and N compounds in 2010 and 2020 than the values
originally based on the Gothenburg Protocol (Schöpp et al., 2003).
The trends in base cation deposition (Fig. 2c) were similar (but
substantially less steep) to the trend in S deposition, with maxima
in the 1980s, reflecting the maximum dust emissions associated
with coal production and consumption (Hedin et al., 1994). Future
deposition of base cations was held constant at 2008–2010 levels.
2.6. MAGIC model description and calibration
MAGIC (Model of Acidification of Groundwater in Catchments)
was designed to reconstruct past and predict future drainage water
and soil chemistry (Cosby et al., 2001). The MAGIC model uses a
lumped representation of soil properties (Table 1) because under-
standing the catchment runoff chemistry or soil water chemistry
does not require detailed knowledge of the spatial distribution of
the parameters within a catchment or forest plot. Water fluxes,
atmospheric deposition, net vegetation uptake, weathering, and
a description of organic acids are required as external inputs to
MAGIC. We used the new version, MAGIC 7ext developed on the
basis of MAGIC 7 (Cosby et al., 2001; Oulehle et al., 2012), which
includes a new formulation of N retention and loss in soils based
directly on the microbial processes which determine the balance of
N mineralization and immobilization. MAGIC 7ext was described
and successfully tested on three Czech sites (Oulehle et al., 2012).

4. J. Hruˇska et al. / Ecological Modelling 244 (2012) 28–37 31
Table 1
Values of soil fixed parameters and resulting calibrated parameters for the calibration year.
Units Value for soil
+5–30 cm 30–80 cm
Fixed parameters
Discharge, annual m 0.41 0.43
Precipitation, annual m 1.18 1.18
Soil depth m 0.35 0.5
Bulk density (fraction < 2 mm) kg m−3
463 595
CEC mequiv kg−1
89.4 24.7
Al(OH)3 solubility constant log 10 8.45 8.65
SO4 adsorption half saturation mequiv m−3
100 100
SO4 maximum adsorption capacity mequiv kg−1
10 1
pCO2 atm 0.64 0.8
Temperature ◦
C 2.5 2.5
pK1 of organic acids −log 10 2.5 2.5
pK2 of organic acids −log 10 4.1 4.1
pK3 of organic acids −log 10 6.7 6.7
Dissolved organic acid mmol+ m−3
16.5 10
Optimised parameters
Weathering Ca mequiv m−2
0 6
Weathering Mg mequiv m−2
40 5
Weathering Na mequiv m−2
13 7
Weathering K mequiv m−2
0 0
Weathering of (Ca + Mg + K + Na) mequiv m−2
53 18
Selectivity coeff. Al–Ca log −0.13 −0.26
Selectivity coeff. Al–Mg log 0.59 1.34
Selectivity coeff. Al–Na log −3.22 −2.78
Selectivity coeff. Al–K log −5.09 −4.89
Ca initial condition % of CEC 5.1 18.3
Mg initial condition % of CEC 9.0 13
Na initial condition % of CEC 1.7 2.5
K initial condition % of CEC 2.3 3
Initial base saturation (Ca + Mg + Na + K) % of CEC 18.1 36.8
Nitrogen parameters
Initial C pool in soil mol m−2
837
Initial N pool in soil mol m−2
29
Initial C/N in soil mol mol−1
28.9
Plant uptake NH4 mmol m−2
year−1
150
Plant uptake NO3 mmol m−2
year−1
75
Nitrification % of inputs 60
Denitrification mmol m−2
year−1
7
Litter C flux mmol m−2
year−1
11,250
Litter C/N mol mol−1
50
C frac % 21
N frac % 60
Decomp frac % 3.9
The key parameters used in the MAGIC application at Pop Ivan
are summarized in Table 1. Soil concentrations of exchangeable
cations were weighted depending on the measured thickness of
soil horizons and on the measured fine earth (<2 mm) bulk density.
Partial pressure of CO2 in the soil water was assumed to be oversat-
urated with respect to the atmospheric partial pressure of CO2 by
almost a factor of three, a value that is within the observed range
for soils (e.g. Cosby et al., 2001; Oulehle et al., 2007). We used a two
soil layer version of MAGIC to model soil water solution at 30 cm
and 90 cm separately.
MAGIC was calibrated to the average soil water chemistry for
the period 2008–2010 and the soil chemistry in 1935 and 2008. The
calibration proceeded by sequential steps. The first steps involved
calibration of the strong-acid anions; Cl and SO4 were calibrated by
adjusting the deposition inputs such that for both ions the correct
mass balance was obtained.
Next, we calibrated the NO3 concentrations in water. The val-
ues Cfrac and Nfrac, and the initial C and N content of the SOM pool
were specified (Table 1). These values were jointly selected in an
optimization procedure that minimized the differences between
simulated and observed values of a number of criteria, given the
input time-series of litter, runoff, and inorganic N deposition at
each site. The values to be matched included the C content, the
C/N ratio of the soil organic matter, and the concentrations of
inorganic N, DON and DOC in soil water. The simulated values for
the pools of C and N and the C/N ratio in the soil were forced to
match the observed values for the years for which soil measure-
ments were available. N dynamics were simulated only for the top
soil (0–30 cm), assuming little N cycle alteration in the deep mineral
soil.
This optimization procedure resulted in a modelled sum of
strong acid anions (SAA) in soil water equal to that observed. The
next steps involved calibration of the base cations Ca, Mg, Na, and K.
A trial and error process was used to adjust the weathering rates of
Ca, Mg, Na, and K and initial soil exchange pools of these four cations
until modelled concentrations of base cations in the soil water and
modelled pools of base cations in the soil matched those observed
for the calibration period. This step calculated the soil–soil solution
selectivity coefficients for base cations and Al exchange (Table 1).
At this point the modelled sum of base cations (SBC) equalled those
observed for the calibration period, and thus the modelled acid neu-
tralising capacity (ANC) also equalled the observed ANC (ANC was
defined as SBC-SAA).
The final step entailed calibration of the weak acids and bases
such that the simulated concentrations of H+, Ali (inorganic alu-
minium) and organic anions (A−) matched observations. This was
achieved by adjusting the dissociation constants for organic acids,
aluminium hydroxide, fluoride, and sulphate species, and organic

5. 32 J. Hruˇska et al. / Ecological Modelling 244 (2012) 28–37
Fig. 3. Measured (2008–2010, dots) and simulated (1880–2050, lines) changes in soil water chemistry at 30 cm (upper panel) and 90 cm (lower panel) at Pop Ivan.
aluminium complexes. We used a triprotic model for organic acids
with dissociation constants given by Hruˇska et al. (2003).
2.7. Vegetation dynamics
Relevés were performed in the 1930s (Zlatník et al., 1938)
and repeated in the period 1997–2006 on 141 square plots
sized 10 m × 10 m in the massif of Maramuresh Pop-Ivan. Vege-
tation changes after 59–68 years were evaluated by ˇSebesta et al.
(2011). We focused on the herb layer of 22 repeatedly evaluated
phytocoenological plots which were nearest the lysimeters and
precipitation collectors. While a 6-member scale of abundance and
dominance (Braun-Blanquet, 1934) and nomenclature of vascu-
lar plants according to Polívka et al. (1928) were used during the
original survey, repeated relevés were done using an 11-member
abundance and dominance scale (Zlatník, 1953), and nomenclature
was adapted according to Kubát (2002) and Dobrochayeva et al.
(1999). The sets of data were converted to identical units before
processing (see also ˇSebesta et al., 2011).
Changes in the presence of plant species between the periods
were evaluated by means of synoptic sTab., with calculations of
frequency and fidelity for individual taxa applying the phi coef-
ficient according to Chytr´y et al. (2002). Changes in the herb layer
structure were evaluated with the use of an internal diversity index
(Shannon–Wiener’s index) (Hill, 1974). In order to evaluate the
herb layer response to environmental changes, we used phytoindi-
cation with Ellenberg indicator values (EIV according to Ellenberg,
1996) in the software Juice 7.0 (Tich´y, 2002). Differences were
tested by a paired T-test with the critical level of significance = 0.05.
3. Results and discussion
3.1. Soil water chemistry
3.1.1. Sulphate
Mean annual sulphate concentrations (2008–2010) in soil water
varied between 59 and 79 ␮equiv L−1 at 30 cm and between 68 and
87 ␮equiv L−1 at 90 cm, respectively (Fig. 3). Changes in sulphur
deposition were the main factor driving factors in the soil biogeo-
chemistry at Pop Ivan. The amount of historical sulphur deposition
was estimated as proportional to the coal mining in Central Europe
(Fig. 2 and Oulehle et al., 2010) for the period 1880–1990 and the
historical background was estimated to be ca. 50 ␮equiv L−1 for
both depths. The highest concentrations for both depths were esti-
mated to occur during the mid the 1980s (230 ␮equiv L−1), when S
deposition in the region peaked. Due to the low modelled soil sul-
phur adsorption capacity (Table 1), soil water responded quickly to
changes in atmospheric deposition. Thus, the decline of SO4 con-
centrations proportionally followed the decrease of S deposition.
If future S deposition follows CLE expectations (Fig. 2), predicted
soil water SO4 concentrations will decrease to ca 50 ␮equiv L−1
around 2020 and will remain sTab. until 2050. This value is similar
to the SO4 concentration estimated for the historical background
(Fig. 3). Despite this pronounced decline, SO4 will still be the dom-
inant soil water anion in the future.
The SO4 concentrations at Pop Ivan were markedly lower
compared to the Norway spruce research plot Naˇcetín (under-
lined by gneiss and dystric cambisol) in the Ore Mts., Czech
Republic (Oulehle et al., 2006, 2007), where soil water SO4
concentration at the 30 cm depth varied between ca. 400 and
600 ␮equiv L−1 between 2003–2006 and 500–600 ␮equiv L−1 at
90 cm. At Naˇcetín, modelled soil water SO4 peaked in the mid 1980s
(ca. 1200 ␮equiv L−1) as a result of extremely high S deposition
rates of 560 mequiv m−2 year−1 caused by high SOx emissions from
coal-burning power plants. Compared to Naˇcetín, Pop Ivan received
ca. 50% less S deposition. Due to the lower amount of draining soil
water (370 mm at Naˇcetín) and lower soil adsorption capacity at
Pop Ivan, observed and modelled SO4 concentrations at Pop Ivan
were significantly lower compared to heavily acidified soils at the
Czech Republic.
3.1.2. Nitrogen
Mean annual NO3 concentrations in the soil solution at 30 cm
and 90 cm were 32 and 36 ␮equiv L−1, respectively between 2008
and 2010 (Fig. 3). The estimated low NO3 deposition prior to
the 1950s (<3 kg ha−1 year−1 of N-NO3) together with effective
Table 2
Long-term (1880–2050) cumulative N budget simulated by MAGIC for mentioned intervals. Percentage represented distribution of deposited nitrogen between fluxes and
pools.
1880–1950 1950–2010 2010–2050
mmol m−2
% mmol m−2
% mmol m−2
%
Deposition 3477 5585 2615
Soil accumulation 1862 54 1589 28 252 10
DIN leaching 0 0 2528 45 1390 53
DON leaching 1092 31 1005 18 684 26
Denitrification 497 14 420 8 280 11

6. J. Hruˇska et al. / Ecological Modelling 244 (2012) 28–37 33
Fig. 4. Measured (dot) and simulated (1880–2050, line) changes in soil C/N ratio of
the layer 0–30 cm at Pop Ivan.
immobilization of N in soil resulted in low leaching of nitrogen,
mainly in the form of DON (dissolved organic nitrogen). As NO3 and
NH4 deposition have increased since the late 1950s (Fig. 2) leach-
ing of NO3 has dominated the losses of N from the system. Leaching
of NO3 will be the major future pathway of N loss according to the
model prediction (Table 2). Results of isotopic studies have revealed
that NO3 in the stream/soil solution appears to be predominantly
cycled through the microbial pool (Curtis et al., 2011); thus, control
over the soil N retention rate could be attribuTab. to the efficiency
of the microbial community to immobilize nitrogen. However,
labelling experiments with different 15N forms undertaken at Pop
Ivan have revealed that microbial community assimilates exclu-
sively organic N (Tahovská et al., in preparation). The microbial
preference of organic over inorganic N together with relatively high
gross N mineralization processes (ammonification and nitrifica-
tion) suggests that high N-NH4 and mainly N-NO3 content found in
Pop Ivan soils can originate from the organic N pool, but not directly
from inorganic N deposition (Tahovská et al., in preparation). The
limited capacity of the studied ecosystem to retain nitrogen was
successfully demonstrated by the model, which showed only lim-
ited N soil accumulation since the 1950s when the major rise in
deposition occurred (Fig. 2). Consequently, the modelled C/N ratio
(mol mol−1) does not change very much over the last 50 years
(decreasing from 26.7 to 25.5, Fig. 4). We assumed that the soil
C pool has been sTab. over the whole modelled period and the DOC
flux has been set up as a percentage of decomposition to fit the
observed DOC in soil water. DON flux is than calculated according
to the development of the soil C/N ratio. The initial C/N ratio in 1880
was estimated to be 28.9 and it will not change significantly in the
future (Fig. 4). We assumed a “background” denitrification rate of
7 mmol m−2 year−1 as a plausible additional pathway of N loss from
the ecosystem.
3.1.3. Base cations
Mean annual averages of the sum of base cations
(Ca + Mg + Na + K) were between 58–61 ␮equiv L−1 (−30 cm)
and 71–82 ␮equiv L−1 (−80 cm) for the period 2008–2010 (Fig. 3).
Both modelled base cations and SO4 increased over the histor-
ical period 1880–1991 (Fig. 3). The base cations then declined and
decreased further following the decline of SO4. During peak acidi-
fication in the 1980s, concentrations of 152 ␮equiv L−1 (30 cm) and
245 ␮equiv L−1 (90 cm) were estimated, approximately twice the
measured data from the 2000s. The low supply of base cations from
the soil exchange-complex and weathering was a major contribu-
tor to the acidification of soils at Pop Ivan. Measured concentrations
are lower than the historical estimate, as a result of depletion of
the soil cation-exchange complex and also declining base cation
deposition (Fig. 2). A slight increase of SBC concentration is pre-
dicted in the future for both depths as a result of declining strong
anion concentrations. Thus SBCs for both depths will again reach
their historical levels (Fig. 3). SBC concentrations were significantly
lower compared to the heavily acidified Naˇcetín, where concentra-
tions between 180 and 190 ␮equiv L−1 were observed at 90 cm and
around 130 ␮equiv L−1 at 30 cm in the mid-2000s (Oulehle et al.,
2006, 2007).
Weathering rates (Table 1) were estimated within the MAGIC
calibration procedure along with historic soil base saturation. As a
result, high Mg weathering needed to be estimated for the upper
soil to obtain measured concentrations in the soil water (Table 1,
Fig. 3). One possible explanation is that there is intensive exchange
of Mg within the internal flux of the tree canopy with subsequent
cation-exchange between the soil matrix and soil water. Such rapid
short-term litter decomposition/cation exchange is not part of the
MAGIC model and thus must be estimated as weathering rate, even
though it is not actual chemical weathering of the soil particles or
parent material. The weathering rate of Mg for deeper mineral soil
was estimated to be lower and thus more realistic for the crystalline
schist and gneiss underlying Pop Ivan (18 mequiv m−2 annually).
Langan et al. (2001) published 15 weathering rates estimated for
granitic areas in northern Europe calculated in different ways (min-
eral index, PROFILE model, Sr isotopes, depletion) ranging between
10 mequiv m−2 and 85 mequiv m−2 annually, and the high altitude
Pop Ivan fits within this range.
3.1.4. pH, aluminium and Bc/Al ratio
Chronic soil acidification caused a decline in pH and enhanced Al
mobilization at both depths (Fig. 5). Measured pH was 4.49 (30 cm)
and between 4.57 and 4.66 (90 cm) between 2008 and 2010.
As a result of sulphur and nitrogen deposition and consequent
base cation depletion, modelled pH declined to a minimum of 4.3
at 30 cm and 4.4 at 90 cm during the 1980s. These decreases were
about 0.5 pH (30 cm) and 0.7 pH (90 cm) compared to the val-
ues estimated for 1880. After a decade of declining deposition, the
pH increased only slightly (Fig. 5). Future predictions show a slow
increase to pH 4.6 at 30 cm and 4.8 at 90 cm. Soil water in both hori-
zons will stay more acidic in comparison to historical estimates.
Fig. 5. Measured (2008–2010, dots) and simulated (1880–2050, lines) changes in soil water chemistry at 30 cm (upper panel) and 90 cm (lower panel) at Pop Ivan.

7. 34 J. Hruˇska et al. / Ecological Modelling 244 (2012) 28–37
Fig. 6. Measured (dots) and simulated (1880–2050 full lines) changes in soil chemistry of the layer 0–30 cm (upper panel) and 30–90 cm (lower panel) at Pop Ivan.
Similar trends were predicted for Naˇcetín in the Ore Mts., despite
the fact that soil water at 90 cm was more acidified (minima of 4.25
in the 1980s) compared to the estimated pre-industrial pH of 4.7
(Oulehle et al., 2006, 2007).
Dissolved inorganic Al was measured as between
6.6–10.3 ␮mol L−1 (30 cm) and 7.8–8.3 ␮mol L−1 (90 cm) between
2008 and 2010. Al concentrations in the MAGIC model are con-
trolled by Al(OH)3 solubility. The Al(OH)3 solubility constant
for mineral soil was optimized using soil water measurements
between 2008 and 2010 to −pKAl(OH)3
= 8.45 (30 cm) and 8.65
(90 cm). These values were in good agreement with the results
of Oulehle and Hruˇska (2005) at Naˇcetín, who found an equi-
librium of Al3+ concentration in soil water between gibbsite
(−pK = 8.04) and amorphous Al(OH)3 (−pK = 9.66). The highest
modelled Al concentrations were observed during the 1980s
(Fig. 5). The concentration of Al peaked with 50 ␮mol L−1 (30 cm)
and 40 ␮mol L−1 (90 cm). Higher acidic deposition, lower pH and
reduced ability of base cations to neutralize the acidity were the
main factors affecting the mobilization of aluminium in the spruce
stand. MAGIC predictions will lead to further decreases of Al at
both depths as a result of increasing pH (Fig. 5). Nevertheless, Al
concentrations at Pop Ivan are very low in comparison with more
polluted areas in Central Europe. Al concentrations between 40
and 160 ␮mol L−1 were found at Lysina (underlined by granite
and spodosol) in the western Czech Republic (Hruˇska and Krám,
1994) for upper mineral soil, and 110 ␮mol L−1 (Krám et al., 1995)
for the mineral soil solution at 80 cm. Similarly, Oulehle et al.
(2007) observed Al concentrations between 130 and 260 ␮mol L−1
(90 cm) for Naˇcetín, and modelled maximum for the 1980s
exceeded 400 ␮mol L−1.
The depletion of nutrient cations and enhanced concentration
of potentially toxic Al in the soil solution may deteriorate forest
health. The molar Bc/Al ratio ((Ca + Mg + K)/Al) has been widely used
as criterion for the risk of tree damage (Sverdrup et al., 1992; Cronan
and Grigal, 1995; Hruˇska et al., 2001). Based on experiments with
seedlings Sverdrup et al. (1992) showed enhanced mortality if Bc/Al
ratio was lower than 1. Field data from the Czech Republic (Hruˇska
et al., 2001) suggested increasing tree damage with decreasing
Bc/Al in soil solution of the rooting zone in Norway spruce stands.
On the other hand limitation of this concept was show by De Witt
et al. (2010). Only reduction of Mg uptake was observed after long-
term experimental addition of AlCl3 to the rooting zone of Norway
spruce in southern Norway.
One of the results of increasing inorganic aluminium in the soil
solution observed at Pop Ivan was a decline in the Bc/Al ratio (Fig. 5).
Present ratios of 2.4 (30 cm) and 3.1 (90 cm) are much lower than
estimated historical values (12 and 31 respectively), but still above
critical value of 1 that can have a harmful effect on tree roots. Future
estimates show an increase of Bc/Al until 2050. The lowest values
modelled for the 1980s (1.2 and 2.3 respectively) show relatively
favourable soil solution conditions. Thus, soil acidification may not
have created harmful conditions for tree roots.
3.2. Soil chemistry
The soil profile at Pop Ivan reflects chemistry typical for acidified
forest soils in Central Europe. A soil mass-weighted base satura-
tion of 6% was calculated for topsoil (0–30 cm) and 5.5% for deeper
mineral soil (30–80 cm), respectively in 2008 (Fig. 6).
The modelled soil base saturation in the topsoil decreased from
18.1% in 1880 to the 6.0% measured in 2008, and from 36.8% to
5.5% in the deeper mineral soil. The loss of base cations from
cation-exchange sites was an important mechanism for neutral-
izing incoming acidity, and during the peak of acidic deposition
(from the 1950s to the 1990s) this loss was the dominant source of
base cations in the soil water (Fig. 3).
We evaluated our estimates of soil base saturation in the past by
examining historical data. During the 1930s, a soil and vegetation
survey was organized in the Ukraine Carpathians by the Czechoslo-
vak government, and Pop Ivan was one of the investigated sites
(Zlatník et al., 1938). Soil exchangeable Ca and Mg were deter-
mined in 1935 by the method of Gedroiz (1926 – for more details
see Section 2). There was concern over how historical methods can
be compared with modern ones (AAS after 0.1 M BaCl2 extraction).
Thus both methods were applied to soil samples taken at Pop Ivan
and Javornik (a beech forest further west) in 2008 (for more details
see Kuˇcera et al., submitted for publication). A statistically signif-
icant linear relationship (p < 0.001) between the methods for both
Ca and Mg was observed (Fig. 7).
Generally, modern methods slightly underestimated Ca and
overestimated Mg in comparison with historical ones. Using the
obtained equations (Fig. 7), historical exchangeable Ca and Mg were
recalculated to modern equivalents and then used for MAGIC mod-
elling.
Soil saturation of Ca (0–30 cm) was measured as 7.6% in 1935
and decreased to 2.1% in 2008 (Fig. 5). Ca saturation in the deeper
mineral soil (30–80 cm) has declined from 18.3% to 2.5%. Similarly,
soil saturation of Mg declined from 5.1% to 2.0% (0–30 cm) and from
12.4% to 1.4% (30–80 cm) between 1935 and 2008. Relatively high
past deposition of S together with acid sensitive bedrock and high
water fluxes depletes base cations from soils, resulting in such low
base saturation. The lowest base saturation (SBC or Ca2+ and Mg2+)
was estimated for the mid-1980s when S deposition peaked (Fig. 2).
The predicted soil base saturation will increase steadily until 2050,
but this increase will be very slight and base saturation will still be
lower than the 1880 estimate or values measured in 1935 (Fig. 6).
Better recovery was predicted for upper soils (0–30 cm) compared
to mineral soils.
Similar results were obtained for the Lysina catchment in the
western Czech Republic (Hruˇska and Krám, 2003), where estimated

8. J. Hruˇska et al. / Ecological Modelling 244 (2012) 28–37 35
Fig. 7. Comparison between historical (Gedroiz) and modern (AAS) methods for determination of exchangeable Ca (n = 32) and Mg (n = 33) based on samples taken from a
spruce site at Pop Ivan and a beech site at Javornik.
soil base saturation (estimated for a soil mass-weighted base pro-
file 0–90 cm) decreased from 25% in the middle of the 19th century
to 7% measured in 1994 and 5.6% in 2004 respectively. Future pre-
diction shows no recovery until 2050, but base saturation will not
decline. Navrátil et al. (2007) used the SAFE model to estimate
changes in soil base saturation in different soil horizons at Lysina
and concluded there would be significant recovery of organic soil,
but a continuation of acidification (decline of base saturation) for
the B horizon. A lower magnitude of soil base saturation decline
was modelled for Naˇcetín (Oulehle et al., 2007) where it declined
from 14.1% (1850) to 7.5% measured and modelled for 1994. Recov-
ery to ca. 10% was modelled for 2050. Naˇcetín received significant
dry deposition of base cations in the past (as dust from nearby coal
mining open-pit areas), and higher weathering rates also resulted
in a partial mitigation of incoming atmospheric acidity. A decrease
of soil base saturation as a result of acidic deposition has also been
modelled for forest soils elsewhere in Europe (e.g. Malek et al.,
2005; Belyazid et al., 2006).
3.3. Interactions with vegetation development
The number of herb species had decreased after 59 years
(Table 3). A total of 47 herb-layer taxa were recorded in 1938,
of which 16 were not found during the repeat survey. On the
other hand, only 6 plant taxa were new findings in the herb layer
in 1997 (Table 4). Local development – relevant to the investi-
gated lysimetric plot – well corresponded to the general changes
in spruce-dominated forests in whole Maramuresh Pop-Ivan massif
as described by ˇSebesta et al. (2011). Simultaneously, the observed
statistically significant decrease in the Shannon–Wiener diver-
sity index was the result of not only the decreasing number of
taxa but also the decreasing equitability of plant species cover.
In particular, nitrophilous and poorly shade-tolerant generalists
(e.g. Calamagrostis villosa, Rubus idaeus, Athyrium distentifolium)
with the highest fidelity (Table 4) dominated the herb layer more
often during the repeat survey in 1997. The higher frequency of
nitrophilous taxa was also reflected in the significant increase of
EIV value for nitrogen (Table 3). This supports the hypothesis that
enhanced nitrogen load resulting in a higher availability of inor-
ganic nitrogen forms (Fig. 3) can cause changes in the vegetation
composition.
On the other hand, we did not detect an increase of acidophilous
plant species, or a decrease of EIV for soil reaction, despite the
fact that we observed and modelled long-term soil acidification
(Figs. 5 and 6). We also assume that soil acidification was more
likely a result of acidic deposition caused by long distance trans-
port of acidifying pollutants (especially S), rather than on-going
changes caused by vegetation succession or changes within the
“forest cycle” (Korpel’, 1995; Král et al., 2010). A significant spread
of acidophilous plant taxa – probably as a result of anthropogenic
acidification – has been described by ˇSamonil and Vrˇska (2007),
Durak (2010) and ˇSebesta et al. (2011) for naturally base-richer
beech-dominated forests in the Eastern and Western Carpathians.
The originally poor and acid spruce-dominated plant community
at Pop Ivan probably did not respond to the enhanced soil acidifi-
cation through a decrease soil reaction EIV because there was not a
source of more acidophilous taxa. The plant taxa present have been
shown to be resistant to soil acidification (see Falkengren-Grerup,
1990), and therefore this plant community does not seem to have a
significant potential to detect the man-made acidification process
through phytoindication. Insignificant changes in the EIV for soil
reaction can be also associated with the observed gradual move-
ment of broadleaf woody species to higher elevations at Pop Ivan
(ˇSebesta et al., 2011), or with changes to landscape management
in the vicinity of upland pastures, including a cessation of grazing
pressure that in turn leads to nutrient accumulation in the biomass,
and changes in forest management, often manifested by canopy
closure and an alteration of light availability (e.g. Vrˇska et al., 2009).
Those factors also have profound impacts on vegetation dynamics
and may overwhelm the effects induced by soil acidification and
alteration of the N cycle through enhanced N deposition. Never-
theless, the preferential uptake of organic N over inorganic N by
Table 3
Differences in the herb layer vegetation between the years of measurement. EIV – Ellenberg indication values, SD – standard deviation, p-value – level of significance, T-value
– test of statistics, 22 phytocoenological plots.
Variable 1938 year 1997 year T-Value p-Value
Mean SD Mean SD
Number of species 15.27 5.57 13.18 2.86 1.854 0.0779
Shannon–Wiener index 2.12 0.31 1.87 0.24 4.085 0.0005
EIV-light 4.71 0.49 4.72 0.29 −0.115 0.9099
EIV-temperature 4.11 0.25 3.97 0.16 2.522 0.0198
EIV-moisture 5.49 0.18 5.50 0.17 0.616 0.5446
EIV-soil reaction 3.85 0.45 3.97 0.30 −1.279 0.2149
EIV-soil N 4.55 0.40 4.84 0.39 −4.869 <0.0001

9. 36 J. Hruˇska et al. / Ecological Modelling 244 (2012) 28–37
Table 4
Synoptic Tab. of herb species found, sorted by the year of data collection; species
are ordered according to their frequency (%); calculation of Fisher’s exact test –
species with significance p < 0.05 have zero fidelity (/). Fidelity is demonstrated as
an exponent of frequency.
Year of measurement 1997 1938
No. of relevés 22 22
Luzula sylvatica 100 95
Vaccinium myrtillus 100 86
Oxalis acetosella 100 82
Dryopteris dilatata 91 91
Polygonatum verticillatum 86 73
Calamagrostis arundinacea 9132.5 64
Homogyne alpina 77 77
Athyrium distentifolium 9156.7 36
Soldanella montana 55 68
Hieracium transsylvanicum 50 64
Calamagrostis villosa 41 45
Campanula patula ssp. abietina 23 6441.3
Senecio nemorensis agg. 41 45
Luzula luzuloides ssp. rubella 41 41
Rubus idaeus 6446.2 18
Luzula luzuloides 36 45
Solidago virgaurea ssp. minuta 9 6456.7
Anemone nemorosa 14 5039.0
Stellaria nemorum 18 41
Gentiana asclepiadea 18 36
Rumex arifolius 27 23
Hypericum species 5 4547.2
Phegopteris connectilis 14 36
Rubus hirtus 23 23
Festuca picta 9 18
Athyrium filix-femina 9 18
Blechnum spicant 9 14
Prenanthes purpurea 9 14
Lastrea limbosperma . 2335.8
Streptopus amplexifolius 18 .
Euphorbia carniolica . 18
Gymnocarpium dryopteris 9 5
Doronicum austriacum 5 9
Laserpitium krapfii 9 .
Adenostyles alliariae 9 .
Epilobium angustifolium 5 5
Poa chaixii . 9
Hypericum maculatum . 9
Anthoxanthum odoratum . 9
Potentilla aurea . 9
Huperzia selago . 9
Phyteuma orbiculare . 9
Dryopteris filix-mas 5 .
Dryopteris carthusiana 5 .
Petasites albus 5 .
Rumex alpinus . 5
Melampyrum sylvaticum . 5
Crocus heuffelianus . 5
Lycopodium annotinum . 5
Cicerbita alpina . 5
Festuca altissima . 5
Galeobdolon luteum . 5
Leucanthemum rotundifolium . 5
microbes may lead to an alteration in plant–microbe competition,
leaving more inorganic N available for plant uptake and/or leaching.
4. Conclusions
Soils at Pop Ivan have been significantly acidified despite the
long distance from major sources of sulphur deposition in Central
Europe. The observed chemistry of soil water and soil was success-
fully reproduced by the MAGIC model. Moreover, this calibration
was able to satisfactorily reproduce historical soil chemistry data
from the 1930s. At Pop Ivan, the depletion of exchangeable base
cations from the soil played a crucial role in the acidification pro-
cess, which resulted in a decrease of soil base saturation. The higher
soil water anion concentration and lower pH caused by enhanced
acidic deposition resulted in mobilization of dissolved aluminium.
According to the modelled Bc/Al ratio, aluminium stress was not
crucial for spruce stands at Pop Ivan, contrary to conditions in the
more polluted Central Europe.
The model simulation indicates that the spruce ecosystem at
Pop Ivan is susceptible to N leaching, whereby N availability from
mineralization and deposition exceeds the rate of microbial immo-
bilization. The timing of enhanced leaching of inorganic N falls
into the timescale of increased N deposition during the 2nd half
of the 20th century. Thus, higher plant availability of inorganic N
might play an additional role in the on-going vegetation changes
in Eastern Carpathian forests.
Acknowledgements
Fieldwork, equipment, chemical analyses and travelling were
supported by Czech Science Foundation (Project No. 526/07/1187),
salary for Jakub Hruˇska and Filip Oulehle were funded by the
research plan of the Czech Geological Survey (MZP 0002579801),
salary for Pavel ˇSamonil was funded by the research plan of The
Silva Tarouca Research Institute for Landscape and Ornamental
Gardening (Project No. MSM 6293359101).
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