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Ecological Modelling 244 (2012) 28– 37

Contents lists available at SciVerse ScienceDirect

Ecological Modelling

jo u r n al hom ep age : www.elsev ier .com/ locate /eco lmodel

ong-term forest soil acidification, nutrient leaching and vegetationevelopment: Linking modelling and surveys of a primeval spruce forest

n the Ukrainian Transcarpathian Mts.

. Hruskaa,∗, F. Oulehlea,b, P. Samonil c, J. Sebestad, K. Tahovskáe, R. Hlebd, J. Houskad, J. Sikl a

Czech Geological Survey, Klárov 3, 118 21 Prague 1, Czech RepublicCentre for Ecology and Hydrology, Deiniol Road, Bangor LL57 2UW, United KingdomDepartment of Forest Ecology, The Silva Tarouca Research Institute for Landscape and Ornamental Gardening, Lidická 25/27, 657 20 Brno, Czech RepublicFaculty of Forestry and Wood Technology, Mendel University in Brno, Zemedelská 1, 613 00 Brno, Czech RepublicDepartment of Ecosystem Biology, Faculty of Science, University of South Bohemia, Branisovská 31, 370 05 Ceské Budejovice, Czech Republic

r t i c l e i n f o

rticle history:eceived 6 March 2012eceived in revised form 19 June 2012ccepted 22 June 2012vailable online 28 July 2012

eywords:oil acidificationAGIC modelitrogen leachingcid depositionranscarpathian Mts.orway spruceegetation changes

a b s t r a c t

The biogeochemical model MAGIC was applied to simulate long-term (1880–2050) soil and stratified soilsolution (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 ofEurope.

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 valuesestimated 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 magnesiumsaturation 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 incomingacidity, 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 thetop 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 evidencesuggest that N availability from mineralization and deposition exceeds the rate of microbial and plantimmobilization. Thus, soil N accumulation since the 1960s has been limited. A significant increase innitrophilous species as well as a decrease of herb layer diversity was observed between 1936 and 1997.

. Introduction

Emissions of potentially acidifying compounds have acceleratedue to anthropogenic activities since the mid 19th century. Conse-uently, the transport and deposition of sulphur (S) and nitrogenN) have resulted in the widespread acidification of many terrestrialnd aquatic ecosystems (Reuss and Johnson, 1985). Internationalgreements to reduce emissions of sulphur and nitrogen com-ounds (Bull et al., 2001) have resulted in an 80% reduction in SOx,0% reduction in NH3 and 39% reduction in NOx emissions from EU

ountries since 1985. As the deposition of S has fallen dramaticallycross Europe, many surface waters have begun to recover fromcidification, as manifested by trends of sulphate (SO4), increased

∗ Corresponding author. Tel.: +420 251085433.E-mail address: jakub.hruska@geology.cz (J. Hruska).

304-3800/$ – see front matter © 2012 Elsevier B.V. All rights reserved.ttp://dx.doi.org/10.1016/j.ecolmodel.2012.06.025

© 2012 Elsevier B.V. All rights reserved.

pH and increased acid neutralising capacity (ANC) (Evans et al.,2001; Jenkins et al., 2003; Wright et al., 2005). Despite the factthat base cation leaching from soils has decreased, recovery ofsoils is a much slower process, limited by low weathering rates,declines in base cation deposition (Hedin et al., 1994) and by netbiomass uptake at some sites (Hruska et al., 2002). As S depositionhas decreased, N deposition has become increasingly importantbecause 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 eutrophicationare among the most important factors responsible for diversitylosses 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 nitrophilousand acidophilous species and herb generalists in forestshave revealed a common causative connection across various

J. Hruska et al. / Ecological Modelling 244 (2012) 28– 37 29

an (U

eaEtDcfiaascct

tCpc(sMetdav

2

2

t

Fig. 1. Location of research sites Pop Iv

cosystems (Nygaard and Odegaard, 1999; Hedl, 2004; Samonilnd Vrska, 2007, 2008; Durak, 2010; Sebesta et al., 2011). Inastern Europe, homogenization of the Carpathian beech forestsogether with the acidification of topsoil was observed in Poland byurak (2010). A similar extent of soil acidification and vegetationhanges were described in beech- and spruce-dominated forestsurther east in the Ukraine (Sebesta et al., 2011), an area whichncludes the site described in this study. Therefore, deposition ofcid pollutants should be considered as a potential threat to the soilnd consequent vegetation development in those ecosystems. Ithould also be pointed out that at the same time scale, concurrenthanges in forest management, game pressure, cattle grazing andlimate change have also been documented, all of which are likelyo alter soils and vegetation composition.

Repeated historical soil surveys, originally established inhe 1930s in primeval fir-beech-spruce forests in the Ukrainearpathian Mts. (Sebesta et al., 2011; Kucera et al., submitted forublication), and recent intensive measurements of soil solutionhemistry and deposition in the spruce dominated site of Pop IvanOulehle et al., 2010), allowed us to reconstruct and predict theoil and soil solution chemistry using the biogeochemical modelAGIC. The specific objectives of this study are: (1) model the

xtent of soil acidification over the last 130 years, (2) predicthe future soil chemistry based on expected trends in S and Neposition until 2050, and (3) discuss the possible effects of soilcidification and N cycle alterations on observed changes in theegetation composition.

. Materials and methods

.1. Site description

The Pop Ivan site is a primeval coniferous forest situated nearhe border between Ukraine and Romania at 1480 m a.s.l. (24◦31′E;

kraine) and Nacetín (Czech Republic).

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 EnticPodzols; Albic Podzols also rarely occur in the vicinity of studysite at the highest elevations (classified according to Michéli et al.,2006). The bedrock consists of acid sensitive crystalline schist andgneiss on study site as well as in surroundings. The Pop Ivan site issituated on a steep slope oriented to the west; with a mean annualtemperature of 2 ◦C and annual precipitation of 1.9 m. Becausedirect human impacts have been minimal in this area, Pop Ivan isprobably among the most natural forests of such extent in EasternCentral Europe (Elbakidze and Angelstam, 2007).

2.2. Sampling and chemical analyses of precipitation, soil and soilwater

Precipitation (5 collectors for throughfall and 2 collectors forbulk 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 andthroughfall (surface area of 179 cm2) were used. The contents ofthe throughfall samplers were combined to create one sample forchemical 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 (3lysimeters in each depth). Zero-tension lysimeters were installedunder the forest floor (3 replications). Thickness of the forest floorwas 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 byflame atomic absorption spectrometry (FAAS). pH was measuredusing a pH meter with a combination electrode (Radiometer model

30 J. Hruska et al. / Ecological Modelling 244 (2012) 28– 37

d (c) s

GmA1

uOgetwfwaeataa

2

mwpc0tai(

2

wrsescCtIbCaweahsi

Fig. 2. Estimated trends of (a) sulphur, (b) nitrogen compounds an

K-2401C). Cl, SO4 and NO3 were measured by exchange ion chro-atography. Samples processing and analysis were made in anccredited Testing Laboratory according to criteria of the ISO/IEC7025:2005.

Soils were sampled by excavating four 0.5 m2 quantitative pitssing the method described by Huntington et al. (1988). The Ol plusf (litter plus fermented) horizons were sampled together as a sin-le sample, and then the Oh (humification) horizon (e.g. Zanellat al., 2009) was sampled separately. Mineral soil was collected forhe depths of: 0–10, 10–20, 20–40 and 40–80 cm. The soil samplesere weighed, and then sieved after air-drying (mesh size of 5 mm

or organic horizons and 2 mm for mineral horizons). Soil moistureas determined gravimetrically by drying at 105 ◦C. Exchange-

ble cations were analyzed in 0.1 M BaCl2 extracts by FAAS. Cationxchange capacity (CEC) was calculated as the sum of exchange-ble Ca, Mg, Na, K and TEA. Base saturation (BS) was determined ashe fraction of CEC associated with base cations. Total carbon (C)nd total nitrogen were determined using a Carlo-Erba Fisons 1108nalyser.

.3. Historical soil analysis

To evaluate the historical analytical methods used to deter-ine the soil chemistry, the original methods of Gedroiz (1926)ere used. Soils samples were extracted in 1 M NH4Cl at the soilH (exchange of soil cations with NH4

+). From the eluate, cal-ium was determined as calcium oxalate (CaC2O4) by titration with.1 N KMnO4, Mg concentration was determined by weighing ashe Mg(C9H6ON)2·2H2O precipitate; K concentration by weighings potassium chloride (KClO4); and Na concentration by weigh-ng as sodium chloride (NaCl). For more details see Kucera et al.submitted for publication).

.4. Water fluxes of the soil solution

To assess the water and element fluxes through the soil profilese used a measurement of the chloride (Cl) mass budget. Chlo-

ine compounds tend to be highly soluble in water and mobile inoils, so atmospheric deposition and transport through terrestrialcosystems is rapid if there is active hydrologic flow. In addition,mall-watershed studies assume that weathering of Cl is negligibleompared to atmospheric deposition (Juang and Johnson, 1967).hlorine is a minor constituent of most rocks and has been assumedo be relatively unreactive in forest ecosystems (Schlesinger, 1997).ts content in metamorphic rock underlined our study site maye considered as minor based on litogeochemical database of thezech Geological Survey. The main source of Cl in rivers is glob-lly sandstones and shales (Graedel and Keene, 1996). Estimatedeathering of Cl from bedrock in Hubbard Brook experimental for-

st (mica schist and quartzite) has been estimated only to 1–2% of

verage export in stream water (Lovett et al., 2005). On the otherand, internal cycling of Cl (vegetation uptake and release fromoil organic matter) may have influence on Cl budget especiallyn areas where deposition of Cl is lower than ca. 6 kg ha−1 year−1

um of the base cations (SBC) at Pop Ivan for the period 1880–2050.

(Svensson et al., 2012). The changes in internal Cl cycling assumechanges 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 budgethas been used to calculate water flux trough different horizon asfollows:

X = Clthf

Clss

where X is the water flux (mm) in the respective soil horizon, Clthfis the chlorine flux in throughfall (mg m−2) and Clss is the respec-tive soil solution Cl concentration. Solute fluxes were calculated bymultiplying 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) wereestimated using Czech emissions of SO2, NOx and NH3 for the period1880–2006. Despite a lack of relevant emission inventories fromUkraine, the Czech emissions were tightly correlated with totalemissions from Poland, Slovakia and Romania (Berge, 1997) inthe 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 usedthe current models (Current Legislation scenarios – CLE) of theCo-operative Programme for Monitoring and Evaluation of theLong-range Transmission of Air Pollutants in Europe (EMEP,http://www.ceip.at/emission-data-webdab), which project loweremissions of S and N compounds in 2010 and 2020 than the valuesoriginally based on the Gothenburg Protocol (Schöpp et al., 2003).

The trends in base cation deposition (Fig. 2c) were similar (butsubstantially less steep) to the trend in S deposition, with maximain the 1980s, reflecting the maximum dust emissions associatedwith coal production and consumption (Hedin et al., 1994). Futuredeposition 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 waterand soil chemistry (Cosby et al., 2001). The MAGIC model uses alumped representation of soil properties (Table 1) because under-standing the catchment runoff chemistry or soil water chemistrydoes not require detailed knowledge of the spatial distribution ofthe parameters within a catchment or forest plot. Water fluxes,atmospheric deposition, net vegetation uptake, weathering, anda description of organic acids are required as external inputs toMAGIC. We used the new version, MAGIC 7ext developed on thebasis of MAGIC 7 (Cosby et al., 2001; Oulehle et al., 2012), which

includes a new formulation of N retention and loss in soils baseddirectly on the microbial processes which determine the balance ofN mineralization and immobilization. MAGIC 7ext was describedand successfully tested on three Czech sites (Oulehle et al., 2012).

J. Hruska et al. / Ecological Modelling 244 (2012) 28– 37 31

Table 1Values of soil fixed parameters and resulting calibrated parameters for the calibration year.

Units Value for soil

+5–30 cm 30–80 cm

Fixed parametersDischarge, annual m 0.41 0.43Precipitation, annual m 1.18 1.18Soil depth m 0.35 0.5Bulk density (fraction < 2 mm) kg m−3 463 595CEC mequiv kg−1 89.4 24.7Al(OH)3 solubility constant log 10 8.45 8.65SO4 adsorption half saturation mequiv m−3 100 100SO4 maximum adsorption capacity mequiv kg−1 10 1pCO2 atm 0.64 0.8Temperature ◦C 2.5 2.5pK1 of organic acids −log 10 2.5 2.5pK2 of organic acids −log 10 4.1 4.1pK3 of organic acids −log 10 6.7 6.7Dissolved organic acid mmol+ m−3 16.5 10

Optimised parametersWeathering Ca mequiv m−2 0 6Weathering Mg mequiv m−2 40 5Weathering Na mequiv m−2 13 7Weathering K mequiv m−2 0 0Weathering of

∑(Ca + Mg + K + Na) mequiv m−2 53 18

Selectivity coeff. Al–Ca log −0.13 −0.26Selectivity coeff. Al–Mg log 0.59 1.34Selectivity coeff. Al–Na log −3.22 −2.78Selectivity coeff. Al–K log −5.09 −4.89Ca initial condition % of CEC 5.1 18.3Mg initial condition % of CEC 9.0 13Na initial condition % of CEC 1.7 2.5K initial condition % of CEC 2.3 3Initial base saturation

∑(Ca + Mg + Na + K) % of CEC 18.1 36.8

Nitrogen parametersInitial C pool in soil mol m−2 837Initial N pool in soil mol m−2 29Initial C/N in soil mol mol−1 28.9Plant uptake NH4 mmol m−2 year−1 150Plant uptake NO3 mmol m−2 year−1 75Nitrification % of inputs 60Denitrification mmol m−2 year−1 7Litter C flux mmol m−2 year−1 11,250Litter C/N mol mol−1 50

acsPuafsa

tccam

uwosieC

C frac %

N frac %

Decomp frac %

The key parameters used in the MAGIC application at Pop Ivanre summarized in Table 1. Soil concentrations of exchangeableations were weighted depending on the measured thickness ofoil horizons and on the measured fine earth (<2 mm) bulk density.artial pressure of CO2 in the soil water was assumed to be oversat-rated with respect to the atmospheric partial pressure of CO2 bylmost a factor of three, a value that is within the observed rangeor soils (e.g. Cosby et al., 2001; Oulehle et al., 2007). We used a twooil layer version of MAGIC to model soil water solution at 30 cmnd 90 cm separately.

MAGIC was calibrated to the average soil water chemistry forhe period 2008–2010 and the soil chemistry in 1935 and 2008. Thealibration proceeded by sequential steps. The first steps involvedalibration of the strong-acid anions; Cl and SO4 were calibrated bydjusting the deposition inputs such that for both ions the correctass balance was obtained.Next, we calibrated the NO3 concentrations in water. The val-

es Cfrac and Nfrac, and the initial C and N content of the SOM poolere specified (Table 1). These values were jointly selected in an

ptimization procedure that minimized the differences between

imulated and observed values of a number of criteria, given thenput time-series of litter, runoff, and inorganic N deposition atach site. The values to be matched included the C content, the/N ratio of the soil organic matter, and the concentrations of

2160

3.9

inorganic N, DON and DOC in soil water. The simulated values forthe pools of C and N and the C/N ratio in the soil were forced tomatch the observed values for the years for which soil measure-ments were available. N dynamics were simulated only for the topsoil (0–30 cm), assuming little N cycle alteration in the deep mineralsoil.

This optimization procedure resulted in a modelled sum ofstrong acid anions (SAA) in soil water equal to that observed. Thenext steps involved calibration of the base cations Ca, Mg, Na, and K.A trial and error process was used to adjust the weathering rates ofCa, Mg, Na, and K and initial soil exchange pools of these four cationsuntil modelled concentrations of base cations in the soil water andmodelled pools of base cations in the soil matched those observedfor the calibration period. This step calculated the soil–soil solutionselectivity coefficients for base cations and Al exchange (Table 1).At this point the modelled sum of base cations (SBC) equalled thoseobserved for the calibration period, and thus the modelled acid neu-tralising capacity (ANC) also equalled the observed ANC (ANC wasdefined 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 wasachieved by adjusting the dissociation constants for organic acids,aluminium hydroxide, fluoride, and sulphate species, and organic

32 J. Hruska et al. / Ecological Modelling 244 (2012) 28– 37

in soi

aw

2

ast(ppdloaw(p

wffis(hc1t

3

3

3

v8

TLp

Fig. 3. Measured (2008–2010, dots) and simulated (1880–2050, lines) changes

luminium complexes. We used a triprotic model for organic acidsith dissociation constants given by Hruska et al. (2003).

.7. Vegetation dynamics

Relevés were performed in the 1930s (Zlatník et al., 1938)nd repeated in the period 1997–2006 on 141 square plotsized 10 m × 10 m in the massif of Maramuresh Pop-Ivan. Vege-ation changes after 59–68 years were evaluated by Sebesta et al.2011). We focused on the herb layer of 22 repeatedly evaluatedhytocoenological plots which were nearest the lysimeters andrecipitation collectors. While a 6-member scale of abundance andominance (Braun-Blanquet, 1934) and nomenclature of vascu-

ar plants according to Polívka et al. (1928) were used during theriginal survey, repeated relevés were done using an 11-memberbundance and dominance scale (Zlatník, 1953), and nomenclatureas adapted according to Kubát (2002) and Dobrochayeva et al.

1999). The sets of data were converted to identical units beforerocessing (see also Sebesta et al., 2011).

Changes in the presence of plant species between the periodsere evaluated by means of synoptic sTab., with calculations of

requency and fidelity for individual taxa applying the phi coef-cient according to Chytry et al. (2002). Changes in the herb layertructure were evaluated with the use of an internal diversity indexShannon–Wiener’s index) (Hill, 1974). In order to evaluate theerb layer response to environmental changes, we used phytoindi-ation with Ellenberg indicator values (EIV according to Ellenberg,996) in the software Juice 7.0 (Tichy, 2002). Differences wereested by a paired T-test with the critical level of significance = 0.05.

. Results and discussion

.1. Soil water chemistry

.1.1. SulphateMean annual sulphate concentrations (2008–2010) in soil water

aried between 59 and 79 �equiv L−1 at 30 cm and between 68 and7 �equiv L−1 at 90 cm, respectively (Fig. 3). Changes in sulphur

able 2ong-term (1880–2050) cumulative N budget simulated by MAGIC for mentioned intervools.

1880–1950 1

mmol m−2 % m

Deposition 3477 5Soil accumulation 1862 54 1DIN leaching 0 0 2DON leaching 1092 31 1Denitrification 497 14

l water chemistry at 30 cm (upper panel) and 90 cm (lower panel) at Pop Ivan.

deposition were the main factor driving factors in the soil biogeo-chemistry at Pop Ivan. The amount of historical sulphur depositionwas estimated as proportional to the coal mining in Central Europe(Fig. 2 and Oulehle et al., 2010) for the period 1880–1990 and thehistorical background was estimated to be ca. 50 �equiv L−1 forboth depths. The highest concentrations for both depths were esti-mated to occur during the mid the 1980s (230 �equiv L−1), when Sdeposition in the region peaked. Due to the low modelled soil sul-phur adsorption capacity (Table 1), soil water responded quickly tochanges 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), predictedsoil water SO4 concentrations will decrease to ca 50 �equiv L−1

around 2020 and will remain sTab. until 2050. This value is similarto 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 lowercompared to the Norway spruce research plot Nacetín (under-lined by gneiss and dystric cambisol) in the Ore Mts., CzechRepublic (Oulehle et al., 2006, 2007), where soil water SO4concentration at the 30 cm depth varied between ca. 400 and600 �equiv L−1 between 2003–2006 and 500–600 �equiv L−1 at90 cm. At Nacetín, modelled soil water SO4 peaked in the mid 1980s(ca. 1200 �equiv L−1) as a result of extremely high S depositionrates of 560 mequiv m−2 year−1 caused by high SOx emissions fromcoal-burning power plants. Compared to Nacetín, Pop Ivan receivedca. 50% less S deposition. Due to the lower amount of draining soilwater (370 mm at Nacetín) and lower soil adsorption capacity atPop Ivan, observed and modelled SO4 concentrations at Pop Ivanwere significantly lower compared to heavily acidified soils at theCzech 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 2008and 2010 (Fig. 3). The estimated low NO3 deposition prior tothe 1950s (<3 kg ha−1 year−1 of N-NO3) together with effective

als. Percentage represented distribution of deposited nitrogen between fluxes and

950–2010 2010–2050

mol m−2 % mmol m−2 %

585 2615589 28 252 10528 45 1390 53005 18 684 26420 8 280 11

J. Hruska et al. / Ecological Mod

Ft

imNiomtcoolIspgtPflsid((Cflotwf7t

3

(a

ig. 4. Measured (dot) and simulated (1880–2050, line) changes in soil C/N ratio ofhe layer 0–30 cm at Pop Ivan.

mmobilization of N in soil resulted in low leaching of nitrogen,ainly in the form of DON (dissolved organic nitrogen). As NO3 andH4 deposition have increased since the late 1950s (Fig. 2) leach-

ng of NO3 has dominated the losses of N from the system. Leachingf NO3 will be the major future pathway of N loss according to theodel prediction (Table 2). Results of isotopic studies have revealed

hat NO3 in the stream/soil solution appears to be predominantlyycled through the microbial pool (Curtis et al., 2011); thus, controlver the soil N retention rate could be attribuTab. to the efficiencyf the microbial community to immobilize nitrogen. However,abelling experiments with different 15N forms undertaken at Popvan have revealed that microbial community assimilates exclu-ively organic N (Tahovská et al., in preparation). The microbialreference of organic over inorganic N together with relatively highross N mineralization processes (ammonification and nitrifica-ion) suggests that high N-NH4 and mainly N-NO3 content found inop Ivan soils can originate from the organic N pool, but not directlyrom inorganic N deposition (Tahovská et al., in preparation). Theimited capacity of the studied ecosystem to retain nitrogen wasuccessfully demonstrated by the model, which showed only lim-ted N soil accumulation since the 1950s when the major rise ineposition occurred (Fig. 2). Consequently, the modelled C/N ratiomol mol−1) does not change very much over the last 50 yearsdecreasing from 26.7 to 25.5, Fig. 4). We assumed that the soil

pool has been sTab. over the whole modelled period and the DOCux has been set up as a percentage of decomposition to fit thebserved DOC in soil water. DON flux is than calculated accordingo the development of the soil C/N ratio. The initial C/N ratio in 1880as estimated to be 28.9 and it will not change significantly in the

uture (Fig. 4). We assumed a “background” denitrification rate of mmol m−2 year−1 as a plausible additional pathway of N loss fromhe ecosystem.

.1.3. Base cationsMean annual averages of the sum of base cations

Ca + Mg + Na + K) were between 58–61 �equiv L−1 (−30 cm)nd 71–82 �equiv L−1 (−80 cm) for the period 2008–2010 (Fig. 3).

Fig. 5. Measured (2008–2010, dots) and simulated (1880–2050, lines) changes in soi

elling 244 (2012) 28– 37 33

Both modelled base cations and SO4 increased over the histor-ical period 1880–1991 (Fig. 3). The base cations then declined anddecreased further following the decline of SO4. During peak acidi-fication in the 1980s, concentrations of 152 �equiv L−1 (30 cm) and245 �equiv L−1 (90 cm) were estimated, approximately twice themeasured data from the 2000s. The low supply of base cations fromthe soil exchange-complex and weathering was a major contribu-tor to the acidification of soils at Pop Ivan. Measured concentrationsare lower than the historical estimate, as a result of depletion ofthe soil cation-exchange complex and also declining base cationdeposition (Fig. 2). A slight increase of SBC concentration is pre-dicted in the future for both depths as a result of declining stronganion concentrations. Thus SBCs for both depths will again reachtheir historical levels (Fig. 3). SBC concentrations were significantlylower compared to the heavily acidified Nacetín, where concentra-tions between 180 and 190 �equiv L−1 were observed at 90 cm andaround 130 �equiv L−1 at 30 cm in the mid-2000s (Oulehle et al.,2006, 2007).

Weathering rates (Table 1) were estimated within the MAGICcalibration procedure along with historic soil base saturation. As aresult, high Mg weathering needed to be estimated for the uppersoil to obtain measured concentrations in the soil water (Table 1,Fig. 3). One possible explanation is that there is intensive exchangeof Mg within the internal flux of the tree canopy with subsequentcation-exchange between the soil matrix and soil water. Such rapidshort-term litter decomposition/cation exchange is not part of theMAGIC model and thus must be estimated as weathering rate, eventhough it is not actual chemical weathering of the soil particles orparent material. The weathering rate of Mg for deeper mineral soilwas estimated to be lower and thus more realistic for the crystallineschist and gneiss underlying Pop Ivan (18 mequiv m−2 annually).Langan et al. (2001) published 15 weathering rates estimated forgranitic areas in northern Europe calculated in different ways (min-eral index, PROFILE model, Sr isotopes, depletion) ranging between10 mequiv m−2 and 85 mequiv m−2 annually, and the high altitudePop Ivan fits within this range.

3.1.4. pH, aluminium and Bc/Al ratioChronic 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 consequentbase cation depletion, modelled pH declined to a minimum of 4.3at 30 cm and 4.4 at 90 cm during the 1980s. These decreases wereabout 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, thepH increased only slightly (Fig. 5). Future predictions show a slowincrease 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.

l water chemistry at 30 cm (upper panel) and 90 cm (lower panel) at Pop Ivan.

34 J. Hruska et al. / Ecological Modelling 244 (2012) 28– 37

emist

Sti(

62tfb(ol(m(armsbcpaa1f((e

ohaasreBOets

sPecem

Fig. 6. Measured (dots) and simulated (1880–2050 full lines) changes in soil ch

imilar trends were predicted for Nacetín in the Ore Mts., despitehe fact that soil water at 90 cm was more acidified (minima of 4.25n the 1980s) compared to the estimated pre-industrial pH of 4.7Oulehle et al., 2006, 2007).

Dissolved inorganic Al was measured as between.6–10.3 �mol L−1 (30 cm) and 7.8–8.3 �mol L−1 (90 cm) between008 and 2010. Al concentrations in the MAGIC model are con-rolled by Al(OH)3 solubility. The Al(OH)3 solubility constantor mineral soil was optimized using soil water measurementsetween 2008 and 2010 to −pKAl(OH)3

= 8.45 (30 cm) and 8.6590 cm). These values were in good agreement with the resultsf Oulehle and Hruska (2005) at Nacetín, who found an equi-ibrium of Al3+ concentration in soil water between gibbsite−pK = 8.04) and amorphous Al(OH)3 (−pK = 9.66). The highest

odelled Al concentrations were observed during the 1980sFig. 5). The concentration of Al peaked with 50 �mol L−1 (30 cm)nd 40 �mol L−1 (90 cm). Higher acidic deposition, lower pH andeduced ability of base cations to neutralize the acidity were theain factors affecting the mobilization of aluminium in the spruce

tand. MAGIC predictions will lead to further decreases of Al atoth depths as a result of increasing pH (Fig. 5). Nevertheless, Aloncentrations at Pop Ivan are very low in comparison with moreolluted areas in Central Europe. Al concentrations between 40nd 160 �mol L−1 were found at Lysina (underlined by granitend spodosol) in the western Czech Republic (Hruska and Krám,994) for upper mineral soil, and 110 �mol L−1 (Krám et al., 1995)or 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 Nacetín, and modelled maximum for the 1980sxceeded 400 �mol L−1.

The depletion of nutrient cations and enhanced concentrationf potentially toxic Al in the soil solution may deteriorate forestealth. The molar Bc/Al ratio ((Ca + Mg + K)/Al) has been widely useds criterion for the risk of tree damage (Sverdrup et al., 1992; Cronannd Grigal, 1995; Hruska et al., 2001). Based on experiments witheedlings Sverdrup et al. (1992) showed enhanced mortality if Bc/Alatio was lower than 1. Field data from the Czech Republic (Hruskat al., 2001) suggested increasing tree damage with decreasingc/Al in soil solution of the rooting zone in Norway spruce stands.n the other hand limitation of this concept was show by De Wittt al. (2010). Only reduction of Mg uptake was observed after long-erm experimental addition of AlCl3 to the rooting zone of Norwaypruce in southern Norway.

One of the results of increasing inorganic aluminium in the soilolution observed at Pop Ivan was a decline in the Bc/Al ratio (Fig. 5).resent ratios of 2.4 (30 cm) and 3.1 (90 cm) are much lower than

stimated historical values (12 and 31 respectively), but still aboveritical value of 1 that can have a harmful effect on tree roots. Futurestimates show an increase of Bc/Al until 2050. The lowest valuesodelled for the 1980s (1.2 and 2.3 respectively) show relatively

ry of the layer 0–30 cm (upper panel) and 30–90 cm (lower panel) at Pop Ivan.

favourable soil solution conditions. Thus, soil acidification may nothave created harmful conditions for tree roots.

3.2. Soil chemistry

The soil profile at Pop Ivan reflects chemistry typical for acidifiedforest soils in Central Europe. A soil mass-weighted base satura-tion of 6% was calculated for topsoil (0–30 cm) and 5.5% for deepermineral soil (30–80 cm), respectively in 2008 (Fig. 6).

The modelled soil base saturation in the topsoil decreased from18.1% in 1880 to the 6.0% measured in 2008, and from 36.8% to5.5% in the deeper mineral soil. The loss of base cations fromcation-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 ofbase cations in the soil water (Fig. 3).

We evaluated our estimates of soil base saturation in the past byexamining historical data. During the 1930s, a soil and vegetationsurvey 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 detailssee Section 2). There was concern over how historical methods canbe compared with modern ones (AAS after 0.1 M BaCl2 extraction).Thus both methods were applied to soil samples taken at Pop Ivanand Javornik (a beech forest further west) in 2008 (for more detailssee Kucera et al., submitted for publication). A statistically signif-icant linear relationship (p < 0.001) between the methods for bothCa and Mg was observed (Fig. 7).

Generally, modern methods slightly underestimated Ca andoverestimated Mg in comparison with historical ones. Using theobtained equations (Fig. 7), historical exchangeable Ca and Mg wererecalculated to modern equivalents and then used for MAGIC mod-elling.

Soil saturation of Ca (0–30 cm) was measured as 7.6% in 1935and decreased to 2.1% in 2008 (Fig. 5). Ca saturation in the deepermineral 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 from12.4% to 1.4% (30–80 cm) between 1935 and 2008. Relatively highpast deposition of S together with acid sensitive bedrock and highwater fluxes depletes base cations from soils, resulting in such lowbase 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 belower than the 1880 estimate or values measured in 1935 (Fig. 6).

Better recovery was predicted for upper soils (0–30 cm) comparedto mineral soils.

Similar results were obtained for the Lysina catchment in thewestern Czech Republic (Hruska and Krám, 2003), where estimated

J. Hruska et al. / Ecological Modelling 244 (2012) 28– 37 35

F termis

sfitddcabtwfedmiom2

3

(ooigiasstI(wo

TD–

ig. 7. Comparison between historical (Gedroiz) and modern (AAS) methods for depruce site at Pop Ivan and a beech site at Javornik.

oil base saturation (estimated for a soil mass-weighted base pro-le 0–90 cm) decreased from 25% in the middle of the 19th centuryo 7% measured in 1994 and 5.6% in 2004 respectively. Future pre-iction shows no recovery until 2050, but base saturation will notecline. Navrátil et al. (2007) used the SAFE model to estimatehanges in soil base saturation in different soil horizons at Lysinand concluded there would be significant recovery of organic soil,ut a continuation of acidification (decline of base saturation) forhe B horizon. A lower magnitude of soil base saturation declineas modelled for Nacetín (Oulehle et al., 2007) where it declined

rom 14.1% (1850) to 7.5% measured and modelled for 1994. Recov-ry to ca. 10% was modelled for 2050. Nacetín received significantry deposition of base cations in the past (as dust from nearby coalining open-pit areas), and higher weathering rates also resulted

n a partial mitigation of incoming atmospheric acidity. A decreasef soil base saturation as a result of acidic deposition has also beenodelled for forest soils elsewhere in Europe (e.g. Malek et al.,

005; Belyazid et al., 2006).

.3. Interactions with vegetation development

The number of herb species had decreased after 59 yearsTable 3). A total of 47 herb-layer taxa were recorded in 1938,f which 16 were not found during the repeat survey. On thether hand, only 6 plant taxa were new findings in the herb layern 1997 (Table 4). Local development – relevant to the investi-ated lysimetric plot – well corresponded to the general changesn spruce-dominated forests in whole Maramuresh Pop-Ivan massifs described by Sebesta et al. (2011). Simultaneously, the observedtatistically significant decrease in the Shannon–Wiener diver-ity index was the result of not only the decreasing number ofaxa but also the decreasing equitability of plant species cover.

n particular, nitrophilous and poorly shade-tolerant generalistse.g. Calamagrostis villosa, Rubus idaeus, Athyrium distentifolium)ith the highest fidelity (Table 4) dominated the herb layer more

ften during the repeat survey in 1997. The higher frequency of

able 3ifferences in the herb layer vegetation between the years of measurement. EIV – Ellenbe

test of statistics, 22 phytocoenological plots.

Variable 1938 year

Mean SD

Number of species 15.27 5.57

Shannon–Wiener index 2.12 0.31

EIV-light 4.71 0.49

EIV-temperature 4.11 0.25

EIV-moisture 5.49 0.18

EIV-soil reaction 3.85 0.45

EIV-soil N 4.55 0.40

nation of exchangeable Ca (n = 32) and Mg (n = 33) based on samples taken from a

nitrophilous taxa was also reflected in the significant increase ofEIV value for nitrogen (Table 3). This supports the hypothesis thatenhanced nitrogen load resulting in a higher availability of inor-ganic nitrogen forms (Fig. 3) can cause changes in the vegetationcomposition.

On the other hand, we did not detect an increase of acidophilousplant species, or a decrease of EIV for soil reaction, despite thefact that we observed and modelled long-term soil acidification(Figs. 5 and 6). We also assume that soil acidification was morelikely a result of acidic deposition caused by long distance trans-port of acidifying pollutants (especially S), rather than on-goingchanges caused by vegetation succession or changes within the“forest cycle” (Korpel’, 1995; Král et al., 2010). A significant spreadof acidophilous plant taxa – probably as a result of anthropogenicacidification – has been described by Samonil and Vrska (2007),Durak (2010) and Sebesta et al. (2011) for naturally base-richerbeech-dominated forests in the Eastern and Western Carpathians.The originally poor and acid spruce-dominated plant communityat Pop Ivan probably did not respond to the enhanced soil acidifi-cation through a decrease soil reaction EIV because there was not asource of more acidophilous taxa. The plant taxa present have beenshown to be resistant to soil acidification (see Falkengren-Grerup,1990), and therefore this plant community does not seem to have asignificant potential to detect the man-made acidification processthrough phytoindication. Insignificant changes in the EIV for soilreaction can be also associated with the observed gradual move-ment of broadleaf woody species to higher elevations at Pop Ivan(Sebesta et al., 2011), or with changes to landscape managementin the vicinity of upland pastures, including a cessation of grazingpressure that in turn leads to nutrient accumulation in the biomass,and changes in forest management, often manifested by canopyclosure and an alteration of light availability (e.g. Vrska et al., 2009).

Those factors also have profound impacts on vegetation dynamicsand may overwhelm the effects induced by soil acidification andalteration of the N cycle through enhanced N deposition. Never-theless, the preferential uptake of organic N over inorganic N by

rg indication values, SD – standard deviation, p-value – level of significance, T-value

1997 year T-Value p-Value

Mean SD

13.18 2.86 1.854 0.07791.87 0.24 4.085 0.00054.72 0.29 −0.115 0.90993.97 0.16 2.522 0.01985.50 0.17 0.616 0.54463.97 0.30 −1.279 0.21494.84 0.39 −4.869 <0.0001

36 J. Hruska et al. / Ecological Mod

Table 4Synoptic Tab. of herb species found, sorted by the year of data collection; speciesare ordered according to their frequency (%); calculation of Fisher’s exact test –species with significance p < 0.05 have zero fidelity (/). Fidelity is demonstrated asan exponent of frequency.

Year of measurement 1997 1938No. of relevés 22 22

Luzula sylvatica 100 95Vaccinium myrtillus 100 86Oxalis acetosella 100 82Dryopteris dilatata 91 91Polygonatum verticillatum 86 73Calamagrostis arundinacea 9132.5 64Homogyne alpina 77 77Athyrium distentifolium 9156.7 36Soldanella montana 55 68Hieracium transsylvanicum 50 64Calamagrostis villosa 41 45Campanula patula ssp. abietina 23 6441.3

Senecio nemorensis agg. 41 45Luzula luzuloides ssp. rubella 41 41Rubus idaeus 6446.2 18Luzula luzuloides 36 45Solidago virgaurea ssp. minuta 9 6456.7

Anemone nemorosa 14 5039.0

Stellaria nemorum 18 41Gentiana asclepiadea 18 36Rumex arifolius 27 23Hypericum species 5 4547.2

Phegopteris connectilis 14 36Rubus hirtus 23 23Festuca picta 9 18Athyrium filix-femina 9 18Blechnum spicant 9 14Prenanthes purpurea 9 14Lastrea limbosperma . 2335.8

Streptopus amplexifolius 18 .Euphorbia carniolica . 18Gymnocarpium dryopteris 9 5Doronicum austriacum 5 9Laserpitium krapfii 9 .Adenostyles alliariae 9 .Epilobium angustifolium 5 5Poa chaixii . 9Hypericum maculatum . 9Anthoxanthum odoratum . 9Potentilla aurea . 9Huperzia selago . 9Phyteuma orbiculare . 9Dryopteris filix-mas 5 .Dryopteris carthusiana 5 .Petasites albus 5 .Rumex alpinus . 5Melampyrum sylvaticum . 5Crocus heuffelianus . 5Lycopodium annotinum . 5Cicerbita alpina . 5Festuca altissima . 5

ml

4

lEfwfcc

Galeobdolon luteum . 5Leucanthemum rotundifolium . 5

icrobes may lead to an alteration in plant–microbe competition,eaving more inorganic N available for plant uptake and/or leaching.

. Conclusions

Soils at Pop Ivan have been significantly acidified despite theong distance from major sources of sulphur deposition in Centralurope. The observed chemistry of soil water and soil was success-ully reproduced by the MAGIC model. Moreover, this calibration

as able to satisfactorily reproduce historical soil chemistry data

rom the 1930s. At Pop Ivan, the depletion of exchangeable baseations from the soil played a crucial role in the acidification pro-ess, which resulted in a decrease of soil base saturation. The higher

elling 244 (2012) 28– 37

soil water anion concentration and lower pH caused by enhancedacidic deposition resulted in mobilization of dissolved aluminium.According to the modelled Bc/Al ratio, aluminium stress was notcrucial for spruce stands at Pop Ivan, contrary to conditions in themore polluted Central Europe.

The model simulation indicates that the spruce ecosystem atPop Ivan is susceptible to N leaching, whereby N availability frommineralization and deposition exceeds the rate of microbial immo-bilization. The timing of enhanced leaching of inorganic N fallsinto the timescale of increased N deposition during the 2nd halfof the 20th century. Thus, higher plant availability of inorganic Nmight play an additional role in the on-going vegetation changesin Eastern Carpathian forests.

Acknowledgements

Fieldwork, equipment, chemical analyses and travelling weresupported by Czech Science Foundation (Project No. 526/07/1187),salary for Jakub Hruska and Filip Oulehle were funded by theresearch plan of the Czech Geological Survey (MZP 0002579801),salary for Pavel Samonil was funded by the research plan of TheSilva Tarouca Research Institute for Landscape and OrnamentalGardening (Project No. MSM 6293359101).

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