Tree species traits cause divergence in soil acidification during four decades of postagricultural...

Tree species traits cause divergence in soil acidificationduring four decades of postagricultural forestdevelopmentAN D E SCHR I JVER * , P I ETER D E FRENNE * , J EROEN STAELENS * , GOR IK VERSTRAETEN * ,

BART MUYS † , LARS VESTERDAL ‡ , KAREN WUYTS * , LOTTE V AN NEVEL * , S TEPHAN IE

SCHELFHOUT * , S TEFAAN D E NEVE § and KRIS VERHEYEN*

*Laboratory of Forestry, Faculty of Bioscience Engineering, Ghent University, Geraardsbergse Steenweg 267, 9090 Gontrode

(Melle), Belgium, †Division Forest, Nature and Landscape, Department of Earth and Environmental Sciences, K.U.Leuven,

Celestijnenlaan 200E Box 2411, 3001 Leuven, Belgium, ‡Forest & Landscape Denmark, University of Copenhagen, Rolighedsvej

23, DK-1958 Frederiksberg C, Denmark, §Department of Soil Management and Soil Care, Faculty of Bioscience Engineering,

Ghent University, Coupure Links 653, 9000 Ghent, Belgium

Abstract

A change in land use from agriculture to forest generally increases soil acidity. However, it remains unclear to what

extent plant traits can enhance or mitigate soil acidification caused by atmospheric deposition. Soil acidification is

detrimental for the survival of many species. An in-depth understanding of tree species-specific effects on soil acidifi-

cation is therefore crucial, particularly in view of the predicted global increases in acidifying nitrogen (N) deposition.

Here, we report soil acidification rates in a chronosequence of broadleaved deciduous forests planted on former ara-

ble land in Belgium. This region receives one of the highest loads of potentially acidifying atmospheric deposition in

Europe, which allowed us to study a ‘worst case scenario’. We show that less than four decades of forest develop-

ment caused significant soil acidification. Atmospheric deposition undoubtedly and unequivocally drives postagri-

cultural forests towards more acidic conditions, but the rate of soil acidification is also determined by the tree

species-specific leaf litter quality and litter decomposition rates. We propose that the intrinsic differences in leaf litter

quality among tree species create fundamentally different nutrient cycles within the ecosystem, both directly through

the chemical composition of the litter and indirectly through its effects on the size and composition of earthworm

communities. Poor leaf litter quality contributes to the absence of a burrowing earthworm community, which retards

leaf litter decomposition and, consequently, results in forest-floor build-up and soil acidification. Also nutrient uptake

and N2 fixation are causing soil acidification, but were found to be less important. Our results highlight the fact that

tree species-specific traits significantly influence the magnitude of human pollution-induced soil acidification.

Keywords: atmospheric deposition, burrowing earthworms, dinitrogen-fixing species, exchangeable aluminium, exchangeable

calcium, litter quality, nutrient uptake, soil pH

Received 30 June 2011 and accepted 3 August 2011

Introduction

Forest is the dominant natural ecosystem in most

parts of Europe and eastern North America, but centu-

ries of agricultural expansion and urbanization have

transformed large parts of natural forest into cultural

landscapes (Flinn & Vellend, 2005). In recent years,

however, the reverse trend has been observed, as farm-

lands have been reforested (Heil et al., 2007; Cramer

et al., 2008). The fertilization and liming associated with

periods of agricultural management cause nutrient

accumulation in the topsoil and (partially) neutralize

the acidity of the original soil. Reforestation of agricul-

tural land affects soil acidity through a variety of pro-

cesses such as nutrient uptake, litter decomposition,

and symbiosis with dinitrogen (N2)-fixing bacteria. To

balance the higher root uptake of cations compared to

anions, roots excrete more protons (H+) than hydroxyls

(OH�), which acidifies the soil (Nilsson et al., 1982).

This acidification becomes irreversible when plant

material is removed, for example, by harvesting wood,

but is neutralized when cations are released by the

decomposition of organic matter (Nilsson et al., 1982).

Furthermore, slowly decomposing litter acidifies

because a forest floor (FF) is built up, which causes a

time delay in the cycling of buffering cations (Nilsson

et al., 1982). In addition, slowly decomposing litter pro-

duces more organic acids, which can play an important

role in the natural production of H+ in the soil (de VriesCorrespondence: An De Schrijver, tel. + 329 264 9031,

fax + 329 264 9092, e-mail: [email protected]

© 2011 Blackwell Publishing Ltd 1127

Global Change Biology (2012) 18, 1127–1140, doi: 10.1111/j.1365-2486.2011.02572.x

& Breeuwsma, 1986). Actinorhizal tree species, which

are symbiotic with N2-fixing bacteria, increase the con-

centration of reactive N in the ecosystem, which may

result in soil acidification (Van Miegroet & Cole, 1984;

Compton et al., 2003).

In addition to these plant traits related to soil acidifi-

cation, trees promote the dry deposition of acidifying

(SOx, NOy) and potentially acidifying (NHx) compounds

(Van Breemen et al., 1982) by efficiently capturing atmo-

spheric pollutants (Augusto et al., 2002). Despite large

emission reductions during the past decades, potentially

acidifying emissions of nitrogen (N) and sulphur (S) in

many European countries are still higher than the med-

ium- and long-term targets formulated in the Gothen-

burg Protocol and the European NEC directive (http://

ec.europa.eu/environment/air/pollutants/ceilings.htm,

accessed 25 October 2011). Given the particularly high

rates of acidifying dry deposition in industrialized

regions and the discontinuation of lime application after

agricultural abandonment, rapid soil acidification can

be expected in recently established postagricultural for-

ests in Western Europe. Indeed, several studies have

reported significant soil acidification rates in postagri-

cultural forests (e.g. Markewitz et al., 1998; Ritter et al.,

2003; Berthrong et al., 2009).

Soil acidification is associated with decreases in

exchangeable base cations and increases in potentially

toxic aluminium (Bowman et al., 2008). Increased

exchangeable Al3+ concentrations in the soil solution

can lead to increased Al3+ toxicity and nutrient imbal-

ances (Cronan & Grigal, 1995). Several studies have

reported shifts towards this Al buffer range in forest

and grassland soils in Europe (e.g. Blake et al., 1999; De

Schrijver et al., 2006; Bowman et al., 2008). Seedlings of

Fraxinus excelsior L. and Acer pseudoplatanus L., for

instance, suffer direct damage to root structure and

function or reduced Ca2+ and Mg2+ uptake due to high

Al3+ concentrations in the soil (>100 lmol Al3+ L�1)

(Weber-Blaschke et al., 2002). Soil acidification, lowered

Ca2+ concentrations and elevated Al3+ concentrations

are also found to be detrimental for other ecological

groups in forests, such as soil microorganisms (Illmer

et al., 1995; Joner et al., 2005) including arbuscular

mycorrhizal fungi (Goransson et al., 2008), snails (Gar-

denfors, 1992), earthworms (Edwards, 2009). For

instance, high exchangeable Al3+ concentrations and

low pH values inhibit earthworm growth and cocoon

production (Van Gestel & Hoogerwerf, 2001). Burrow-

ing earthworm species are especially negatively

affected by soil acidity (Muys & Granval, 1997).

Because earthworms are important organisms for litter

decomposition and soil bioturbation (Edwards, 2009),

their disappearance will influence nutrient cycling

(Reich et al., 2005). The herbal vegetation of the forest

understory also suffers from soil acidity (Falkengren-

Grerup & Tyler, 1993; Koptsik et al., 2001). Houle et al.

(2007) further suggested that increased soil acidity has

caused the decline of sugar maple (Acer saccharum

Marsch.) populations in eastern North America by

reducing Ca2+ availability and elevating Mn2+ concen-

trations.

An in-depth understanding of tree species-induced

soil acidification is crucial for soil protection, particu-

larly in view of the predicted global increases in acidify-

ing N deposition (Galloway et al., 2004). However, it

remains unclear to what extent plant traits can enhance

or mitigate soil acidification due to atmospheric deposi-

tion. The present study reports soil acidification rates in

a chronosequence of broadleaved deciduous forests

planted on former arable land in Belgium, in

northwestern Europe. Despite the reduction in S+Ndeposition from 4054 molc ha

�1 yr�1 in 1990 to

2109 molc ha�1 yr�1 in 2007, this region still receives

one of the highest loads of potentially acidifying S and

N deposition in Europe (UNECE-EC, 2008; Van Steerte-

gem, 2010). This high deposition load allowed us to

study a ‘worst case scenario’ of acidifying atmospheric

input. Furthermore, we examined the plant traits that

drive soil acidification by studying five broadleaved

tree species that differ in (i) net H+ excretion through

nutrient uptake by tree roots, (ii) FF build-up as a conse-

quence of slow litter decomposition, (iii) leaf litter qual-

ity, and (iv) N2-fixing capacity. We focused on changes

in soil pH and exchangeable soil Ca2+ and Al3+ concen-

trations because of their positive (Ca2+) or negative (Al3+)

influences on forest ecosystem function (Cronan &

Grigal, 1995; McLaughlin & Wimmer, 1999). In the

oldest forest in our chronosequence, we also studied the

size and composition of the earthworm population.

Materials and methods

Study sites

The study sites were isolated forests surrounded by agricul-

tural land and situated in the western part of Flanders in

northern Belgium. Before reforestation, the soils had been

under continuous agricultural use (farmland) for at least

50 years. All forests were located in a 20 9 20 km2 area and

were replicated monoculture plot experiments, that is, com-

mon gardens with ages of 11 years (Groenenburg, 50°49′N, 2°57′E), 14 years (Eversam, 50°56′N, 2°41′E), 17 years (Eeuwenh-

out, 50°46′N, 2°46′E), 25 and 35 years (Mortagne forest, 50°46′N, 3°21′E) at the sampling dates. Each common garden

contained at least five tree species, of which we studied Alnus

glutinosa (L.) Gaertn., Fraxinus excelsior L. or Fraxinus americana

L., Prunus avium L., Quercus robur L. and Tilia cordata Mill. The

tree species were planted in adjacent blocks, which allowed

their direct comparison, given the similar initial soil

© 2011 Blackwell Publishing Ltd, Global Change Biology, 18, 1127–1140

1128 A. DE SCHRIJVER et al.

characteristics, topography, climate, and atmospheric wet

deposition. In the neighbourhood of each studied forest, two

agricultural fields with the same soil type were sampled as

control plots.

The Groenenburg, Eversam and Eeuwenhout forests are

plantations in which the five studied species were planted in

at least three randomly distributed replicate blocks of about

30 9 30 m2. The Mortagne forest consists of monoculture

stands of A. glutinosa, Q. robur, T. cordata, P. avium, and

F. americana with a surface area of at least 0.55 ha. In this for-

est, stands were divided into three blocks of at least 0.18 ha to

generate three replicates. For A. glutinosa and F. americana,

one extra stand was available, allowing replicability tests; soil

pH and plant available Ca2+ and Al3+ concentrations were not

significantly different between these replicates. Trees were

originally planted on a 2 m 9 2.5 m grid in all sites. Blocks

were only selected when they were at a sufficient distance

from the forest edge (at least 20 m) to avoid elevated deposi-

tion due to edge effects (Wuyts et al., 2008). All plots were

homogeneous, dense stands without a shrub layer, except for

the Fraxinus stand in the Mortagne forest, in which a shrub

layer of Corylus avellana L. and A. pseudoplatanus L. and

F. americana seedlings were present.

All soils are well-drained, quaternary niveo-aeolian loamy

luvisols with an average composition of 40–60% silt, 30–50%

sand and ca. 10% clay. The quaternary topsoil lies on tertiary

alternating layers of clayey sand and clay. The long-term

mean (1901–2000) annual temperature and precipitation of

the region are 9.6 °C and 805 mm, respectively (Royal

Meteorological Institute of Belgium).

Soil sampling

In each block, soil was sampled at six random locations and

four depths within the plough layer: 0–5, 5–10, 10–20, and 20–

30 cm. We avoided sampling close to stems to avoid areas

with stemflow-induced soil acidification (Falkengren-Grerup

& Bjork, 1991). Soil sampling was performed in September

2008 in the Groenenburg, Eversam and Eeuwenhout forests,

and in September 1997 and September 2007 in the Mortagne

forest. Samples taken in each block were pooled and dried for

48 h at 40 °C before sieving through a 2 mm mesh. The soil

pH-KCl of each sample was analysed using a glass electrode

(Orion, Orion Europe, Cambridge, England, model 920A) after

extracting 14 mL soil in a 70 mL KCl (1 M) solution. NH4+-ace-

tate-EDTA extractable Ca2+ and Al3+ were analysed by atomic

absorption spectrophotometry (AA220, Agilent Technologies

Belgium, Diegem, Belgium) after shaking 10 g of dry soil in

50 mL NH4+-acetate-EDTA solution (192.5 g NH4

+-acetate,

50 mL acetic acid, and 29.225 g EDTA, diluted to 2 L) for

30 min.

Calculation of predictor variables

1. To explain the observed soil pH and exchangeable Ca2+ and

Al3+ concentrations, five potential predictor variables were

evaluated: (i) the cumulative H+ input through acidifying N

and S deposition, (ii) the net excretion of H+ associated with

nutrient uptake by tree roots, (iii) the net excretion of H+ asso-

ciated with FF build-up, (iv) the leaf litter quality index, and

(v) whether a tree species is N2 fixing or not.

H+ input through cumulative wet and dry deposition of

acidifying components (NHx, NOy, and SOx) was calculated

based on modelled EMEP deposition data (http://webdab.

emep.int/Unified_Model_Results/, accessed 25 October 2011)

(Fig. 1). Forests are rough surface structures that promote the

dry deposition of acidifying pollutants (Van Breemen et al.,

1982). A forest’s capacity for capturing dry deposition

depends on its canopy height, leaf area index, foliage longev-

ity, canopy structure, the form or shape of leaves, and the dis-

tance to the forest edge (Augusto et al., 2002). Because the

studied plantations are very dense blocks of broadleaved tree

species within an otherwise open landscape, we corrected the

dry deposition EMEP data, which are valid for short (0.3 m)

vegetation types (D. Simpson, personal communication), by

multiplying these values by dry deposition factors deduced

from the relationship between tree height and net throughfall

fluxes of SO42� and Na+ published by Erisman & Draaijers

(2003). The amount of dry-deposited nonacid cations was sub-

tracted from the calculated potentially acidifying deposition.

Alnus

0

100

200

300

Cum

ulat

ive

H+

inpu

t (ke

q ha

–1)

Atmospheric depositionNutrient uptake and storage in woody biomassForest floor build-up

Quercus Tilia Prunus0 20 40 0 20 40 0 20 40 0 20 40 0 20 40

Fraxinus

Forest age (years)

Fig. 1 Calculated cumulative H+ input through potentially acidifying atmospheric deposition, nutrient uptake and storage into woody

biomass and forest floor accumulation (± standard error) in kmolc ha�1.


TREE SPECIES TRAITS AND SOIL ACIDIFICATION 1129

Modelled EMEP wet and dry deposition data were avail-

able for the period from 1980 to 1997. We used the data from

1980 for the period from 1972 to 1979. Dry deposition values

were adjusted using a block-specific dry deposition factor. We

used the following relationship between tree height and net

throughfall values of SO42�:

NTF ¼ 0:9123 � e0:0884�h ð1Þð1Þ

[deduced from Fig. 4 in Erisman & Draaijers (2003)] with NTF

is the net throughfall deposition (eq.ha�1 yr�1) of SO42� and

h is forest height (m). Because canopy leaching or uptake of

SO42� is often assumed to be small or negligible compared to

the deposition fluxes (Staelens et al., 2008), we assumed NTF

to equal dry deposition. Furthermore, because dry deposition

of H+ obtained from the EMEP model was calculated for a low

vegetation height of 0.3 m, Eqn (1) can be used to deduce that

dry deposition values equal ca. 1 eq.ha�1 yr�1. Consequently,

we calculated the relative increase of dry deposition based on

the average tree height for each block and this for each year

and used this factor to correct the EMEP dry deposition val-

ues. For each tree species, average tree heights for each block

and year were deduced from exponential regressions fit to the

available data of forest age vs. tree height. The increase of dry

deposition relative to shorter vegetation ranged from 1 for the

youngest forests to 4.4 (Tilia) and 7.7 (Fraxinus) for the oldest.

Subsequently, for each block, we summed wet and dry

deposition values for each year and calculated their

cumulative values.

We also corrected the potentially acidifying deposition for

the neutralizing effect of base cations (K+, Ca2+, and Mg2+)

(UBA, 2004) by subtracting dry deposition values of base

cations based on NTF measurements in a nearby deciduous

forest of 30 m in height. The following relationship between

tree height and net throughfall values of Na was used to esti-

mate stand-specific deposition values:

NTF ¼ 1:8532 � e0:0688�h ð2Þð2Þ

[deduced from fig. 4 in Erisman & Draaijers (2003)] with NTF

is the net throughfall deposition (eq.ha�1 yr�1) of Na+ and h is

the forest height (m). Because canopy leaching or uptake of

Na+ is reported to be negligible (Staelens et al., 2008), we

assumed NTF to equal dry deposition. For a low vegetation

height of 0.3 m, Eqn (2) predicts dry deposition values of ca. 2

for Na+. We calculated the relative increase of dry deposition

of base cations in each block with the same method used for

H+ deposition values above. Factors ranged from 1 for the

youngest forests to 3.0 (Tilia) and 4.7 (Fraxinus) for the oldest.

Then, we calculated cumulative values for dry deposition of

base cations, which were subtracted from the calculated

cumulative proton input for each block.

2. To get a representative estimate of the acidification asso-

ciated with nutrient uptake by tree roots, we calculated the

excess cation accumulation in both stems and branches for all

studied tree species according to the method described by

Nilsson et al. (1982) (Fig. 1). Excess cation accumulation was

obtained by subtracting the total anion amount from the total

cation amount stored in the woody biomass. The aboveground

volume of both stems and branches (m3 ha�1) (Table S1, Sup-

porting Information) was calculated using the formulas pub-

lished by Dagnelie et al. (1985). In each block, stem cores were

taken at about 40 cm above the root collar from three ran-

domly selected trees. Furthermore, five representative

branches were sampled in each block. Stem cores and

branches were dried and milled, and their total cation concen-

trations (K, Ca, Mg, Na) were analysed by atomic absorption

spectrophotometry (Varian AA220). Total P concentration was

determined according to the Scheel (1936) colorimetric

method, with molybdenum vanadate as the reagent, after acid

wet digestion. Nitrogen concentrations were determined using

a modified Kjeldahl method. We used the values reported by

Hagen-Thorn et al. (2004) as average S concentrations. Iron

(Fe), aluminium (Al) and chloride (Cl) were not measured

because these elements have very low concentrations in wood

(Nilsson et al., 1982). The nutrient concentrations measured in

stems and branches are presented in Tables S2 and S3. We

used the wood density values reported by de Vries et al.

(2003) to calculate the aboveground biomass. The above-

ground nutrient stock was calculated by multiplying the bio-

mass fractions by their respective nutrient concentrations.

Root biomass is considerably lower than aboveground bio-

mass, and several studies have described a strong relationship

between aboveground and belowground tree biomass (e.g.

Santantonio, 1990; Finer et al., 2011). Because no tree species-

specific root/shoot (R/S) values were available in literature,

we calculated the belowground biomass using a fixed R/S

ratio of 0.25, based on the studies of Wang et al. (2008), Cairns

et al. (1997), Cannell (1982) and Pallardy (2008). Yang & Luo

(2011) showed that R/S ratio is not significantly correlated

with stand age. We calculated belowground nutrient values

using the values of wood density and nutrient concentrations

calculated for the stems.

Nitrogen, which may enter roots as NH4+ or NO3

� and is

generally the nutrient taken up in the greatest amount by

plants, plays a critical role in determining cation/anion bal-

ance (Nilsson et al., 1982). It is not clear whether broadleaved

trees prefer NH4+ or NO3

� as an inorganic N source (e.g. Min

et al., 1998; Choi et al., 2005; Gessler et al., 2005). While NO3�

is readily available because of its high solubility in soil water,

plants must use energy to reduce NO3� before it can be incor-

porated into amino acids, which is not the case for NH4+ (Tem-

pler & Dawson, 2004). It is thus more likely that most

inorganic N is taken up as NH4+. We calculated the excess cat-

ion accumulation for two scenarios of N uptake: (a) N is taken

up solely as NH4+ and is consequently 100% acidifying, and

(b) N uptake is partly neutralizing when taken up as NO3–N

that originates from atmospheric deposition. NH4+ uptake is

directly acidifying because it is compensated for by H+ release.

In contrast, the uptake of NO3� originating from atmospheric

NO3� deposition is a neutralizing process because OH� is

released. However, the uptake of NO3� released from decom-

posing material through nitrification is also acidifying, not as a

consequence of the uptake itself (which is neutralizing because

of the release of OH�), but as a consequence of the double H+

production during the nitrification process.



In scenario (b), we calculated the NH4+/total N ratio of root

uptake for each tree species under the assumption that all

atmospherically deposited NO3� (on average 10 kg NO3

�–Nha�1 yr�1 over the past 20 years) has been taken up by the

trees. We calculated yearly N uptake rates by summing the

yearly N storage in woody biomass with the yearly N flux

through leaf litterfall. We then calculated the proportion of N

taken up as NH4+ by subtracting 10 kg NO3

�–N from the total

yearly N uptake, resulting in 89% for Alnus, 81% for Quercus,

74% for Tilia, 76% for Prunus and 85% for Fraxinus. We believe

case (a) to be most probable, but we also calculated option (b)

as a minimum acidification scenario (maximal possible neu-

tralization).

3. To obtain a representative estimate of the acidification

associated FF build-up as a consequence of slow litter decom-

position, we calculated the excess cation accumulation in the

FF according to the method described by Nilsson et al. (1982).

Excess cation accumulation was obtained by subtracting the

total anion amount from the total cation amount stored in the

FF biomass. In each plot, FF biomass was measured by care-

fully collecting the complete organic layer (litter layer + frag-

mentation layer + humus layer) in three squares of

0.3 9 0.3 m2. FF sampling and soil sampling were conducted

on the same dates. Total FF biomass was dried at 70 °C and

milled, and total N, total P and total cation (K, Ca, Mg, Na,

and Al) concentrations were analysed using the methods

described above. We used the values reported in Van Ranst

et al. (2002) as average S concentrations. We calculated excess

cation accumulation by assuming N uptake to be solely acidi-

fying (100% of the N uptake as NH4+ and 0% as NO3

�).4. Litter quality was determined by collecting leaf litterfall

during autumn 2009 and analysing the total concentrations of

C, N, lignin, phosphorus (P), K, Ca, and Mg. We placed four

circular littertraps of 0.4 m diameter in each block in the

Groenenburg, Eversam and Eeuwenhout forests and five cir-

cular littertraps of 0.6 m diameter in each block in the Morta-

gne forest. Leaf litterfall was collected fortnightly during the

complete leaf fall period (August 1 through December 31). At

each sampling, leaf litterfall mass was weighed. After weigh-

ing, we pooled litter samples for the blocks by each fortnightly

sampling period and ground and analysed the total cation

concentrations (K, Ca, and Mg by atomic absorption spectro-

photometry; Varian AA220) in the combined samples. Total P

concentration was determined according to the Scheel (1936)

colorimetric method, with molybdenum vanadate as the

reagent, after acid wet digestion with a 1 : 5 HClO4 : HNO3

solution. Carbon and N concentrations were determined by

elemental analysis (Variomax CNS, Germany). A single lignin

analysis was performed on the pooled fortnightly samples

from the oldest forest (35 years, Mortagne) according to the

method of Van Soest et al. (1991). To determine whether it was

appropriate to use the values obtained in the Mortagne forest

for the younger forest stands, we analysed the lignin concen-

trations in a subsample (N = 1 or N = 2 per forest age and per

tree species) of leaf litter. Because we only had one or two rep-

licates per tree species and age class, we could not perform

statistical analyses on this dataset, but the calculated coeffi-

cients of variation were <5% for all tree species. We therefore

decided to use the Mortagne values for all age classes.

For each block, we then calculated a time-weighted average

nutrient concentration to obtain the average concentration of

each nutrient during the leaf litterfall period. For some nutri-

ents, the leaf litter nutrient concentrations changed signifi-

cantly with forest age (Table S4), which is likely to be related

to changes in soil conditions. To account for this, we calcu-

lated age-weighted average nutrient concentrations (Xav)

according to the following formula:

where Xn yr is the leaf litter nutrient concentration during

autumn in a forest of n years old. Based on these data, a ‘leaf

litter quality index’ was calculated (see Statistics).

5. Symbiosis with the N2-fixing bacteria Alnus glutinosa was

accounted for by including a dummy variable with a value of

one for Alnus symbionts (N2-fixing) and zero for other species

(not N2-fixing).

Earthworm collection

Earthworms are key representatives of the soil fauna

(Edwards, 2009) and are very important organisms for litter

decomposition and soil bioturbation. We sampled earthworms

in the oldest forest of our chronosequence (Mortagne forest,

35 years), using three methods to obtain a maximum sampling

of three ecological earthworm categories (epigeic, endogeic,

and anecic species): litter sorting, extraction with a mustard

solution, and hand sorting of the soil. At three randomly

selected sites within each block, a 0.5 m 9 0.5 m frame was

placed on the forest soil, in which the vegetation was carefully

removed and checked for earthworms. Leaves and small

branches were collected in plastic bags, together with the

earthworms found in the previous step. These samples were

hand sorted for earthworms, which were collected in plastic

pots with ethanol (95%). Then, additional earthworms were

extracted from the soil within the same frame by mustard

powder. Mustard expellants (6 g mustard powder L�1) were

prepared in 10 L containers at least 12 h before use (Valckx

et al., 2011). Thirty litres of this emulsion was applied per

frame in several steps; the frequency of pouring depended on

the percolation rate of the solution. When the earthworms

emerged, they were collected with pincers and immediately

placed in another ethanol-filled pot. Finally, a soil subsample

(0.3 m 9 0.3 m with a 20 cm depth) in the centre of the frame

was dug out with a spade and hand sorted over a brightly

Xav ¼ 11 � X11yr þ ð14� 11Þ � X14yr þ ð17� 14Þ � X17yr þ ð25� 17Þ � X25yr þ ð35� 25Þ � X35yr

35;



coloured plastic bag, and the earthworms collected using this

method were placed in a third ethanol-containing pot. All

worms were preserved in ethanol (95%) for a few hours and

then transferred to a 5% formalin solution for fixation

(3 days), after which they were transferred to a 95% ethanol

solution again for further preservation and identification. All

earthworms were identified according to the key published by

Sims & Gerard (1999) and were subsequently categorized into

specific ecological groups (epigeic, endogeic or anecic species).

All worms were individually weighed, including gut contents,

after drying on filter paper at room temperature for 1 min

until their weight became stable. The number of earthworms

sampled by digging in the 0.09 m2 plot was extrapolated to

the 0.25 m2 plot. We then calculated the total biomass and

number of worms per ecological group and per m2. The bio-

mass of the burrowing earthworm species was calculated by

summing the biomasses of the endogeic and anecic species.

Statistics

It is important to note that we did not intend to produce a

mechanistic model of the causes of soil acidification. Rather,

our main aim was to explore whether the observed soil acidifi-

cation is related to the five studied predictor variables. To

determine whether these predictor variables were statistically

related to soil pH and exchangeable Ca2+ and Al3+ concentra-

tions, multilevel models were generated in R 2.12.0 (R Devel-

opment Core Team, 2010). Based on Zuur et al. (2009), we first

selected the optimal random-effects structure based on a like-

lihood ratio test of models with similar fixed components, but

different random components using the lme function of the

nlme library. The resulting optimal model included both ‘for-

est site’ and ‘block’ nested within ‘forest site’ as random

effects. At the same time, we addressed the likelihood that

cases obtained from the same forest site or block share auto-

correlated characteristics. Modelling this hierarchical nature of

the data using random effect terms in a multilevel model leads

to partial optimal pooling across levels (Qian et al., 2010). First,

we constructed a null model with randomly varied intercepts.

To avoid overfitting, the four predictors were first evaluated

as stand-alone predictors based on the �2 log likelihood infor-

mation criterion [i.e. deviance (Hox, 2002)]. Subsequently,

extra predictors were added one-by-one to the first model

with the lowest deviance containing only one predictor. If the

deviance decreased significantly (chi-square P < 0.05; Hox,

2002), this procedure was repeated. To avoid overfitting and

to simplify the model, only variables with P-values < 0.1 were

considered for the final multilevel models. To detect possible

multicollinearity (Graham, 2003) between the different predic-

tor variables, variance inflation factors (VIF) were calculated

according to Quinn & Keough (2002). These VIF values were

low (<3), indicating low collinearity. The output of the final

multilevel models is given in Table 2, and the model output

based on single predictor variables is shown in Table S5.

To account for correlations among litter variables (C, N, P,

K, Ca, Mg, and lignin) and to obtain a general characterization

of the leaf litter quality, we performed a principal component

analysis with varimax rotation in SPSS 15.0. This analysis

extracted only one factor, thus reducing the set of litter vari-

ables to one component. The obtained sample scores for the

first principal component, which explained 75% of the

observed variation, were used for multilevel analysis as a sin-

gle ‘leaf litter quality index’. The calculated leaf litter quality

index was significantly (n = 75, P < 0.01) positively correlated

with total Ca (r = 0.94), K (r = 0.88), Mg (r = 0.87), P (r = 0.85)

and N (r = 0.83) concentrations in leaf litter, and negatively

correlated with C (r = �0.91) and lignin (r = �0.83). Hence, a

high leaf litter quality index was associated with high nutrient

concentrations and low C and lignin values.

Comparability of initial soil conditions

To confirm that initial soil acidity conditions prior to reforesta-

tion were comparable, pH and ln-transformed Ca2+ and Al3+

concentrations of the deepest layer (20–30 cm) of forests with

different ages (11, 14, 17, 25 and 35 years) were compared

using one-way ANOVA in SPSS 15.0. We used the soil data of all

tree species for this comparison (N = 15 per forest). We con-

sider the deepest soil layer to be less influenced by atmospheric

deposition, nutrient uptake, N2 fixation and leaf litter quality.

Despite the observed differences (Table 1), we consider the af-

forested soils to be initially comparable given the comparable

soil type and agricultural history (rotation of wheat, beets, and

potatoes) and the observed influence of atmospheric deposi-

tion and nutrient uptake (see Tables 2 and S5).

Results

After 35 years of forest development, the pH of the

mineral topsoil (0–5 cm) in the postagricultural forests

was significantly decreased from ca. 6.5 in farmland

soil to values ranging from 3.2 under Alnus and 3.5

under Quercus to 4.5 under Fraxinus (Fig. 2a). The

strong decline in topsoil pH was reflected in changes in

exchangeable soil Ca2+ and Al3+ concentrations

(Figs 2b,c, 3 and 4). The Ca2+ concentrations decreased

from ca. 2500 mg kg�1 to values ranging from ca.

1500 mg kg�1 under Fraxinus to 400 mg kg�1 under

Quercus. The Al3+ concentrations increased steeply from

ca. 50 to ca. 240 mg kg�1 under Fraxinus and to

530 mg kg�1 under Alnus.

Table 1 Mean ± standard deviation of pH-KCl and

exchangeable Ca2+ and Al3+ concentrations (mg kg�1) at a soil

depth of 20–30 cm

Age (years) pH-KCl Ca2+ Al3+

11 5.0 ± 0.2b 2087 ± 490b 63 ± 17a

14 4.4 ± 0.5a 1726 ± 368ab 91 ± 34a

17 4.5 ± 0.3a 1376 ± 201a 107 ± 39ab

25 4.3 ± 0.2a 1608 ± 271ab 65 ± 17a

35 4.1 ± 0.2a 1416 ± 270a 145 ± 78b

Means with different superscript letters within the same

column are significantly different (P < 0.05).



In areas that had been reforested for 17 years or

more, the topsoil (0–5 cm) pH and Ca2+ and Al3+ con-

centrations differed significantly (P < 0.001) among the

tree species, with lower pH-KCl and Ca2+ concentra-

tions but higher Al3+ concentrations under Alnus and

Quercus than under Tilia, Prunus, and Fraxinus. The dif-

ferences in soil pH and Ca2+ and Al3+ concentrations

between species progressed to deeper soil layers with

increasing forest age. At 25 years, the tree species effect

became significant (P < 0.001) in the 5–10 cm soil layer,

and at 35 years, it was significant at the 10–20 cm depth

(P < 0.001). In the deepest sampled soil layer (20–30 cm), no significant differences between species were

found at any of the forest ages studied (11–35 years).

Based on this multispecies chronosequence, we eval-

uated the driving factors of soil acidification under the

five studied deciduous broadleaved tree species. The

analyses showed that atmospheric deposition was sig-

nificantly related to soil pH and Ca2+ and Al3+ concen-

trations across the whole soil profile (Table 2). Nutrient

uptake was not retained as a source of soil acidification

in the final models (Table 2), although multilevel mod-

els with nutrient uptake as single predictor variable

showed its negative relationship with soil acidity

(Table S5) throughout the upper 30 cm of the soil pro-

file. The multilevel model analyses gave comparable

results for both scenarios of root N uptake (scenario

1: N uptake solely as NH4+; scenario 2: N uptake partly

as NO3� deposited as atmospheric NO3

� deposition).

The final models did not retain the N2-fixing symbiosis

of Alnus trees as significant factor, but this factor was

significantly related to topsoil pH (0–10 cm) and to soil

Al3+ concentrations when analysed as a single predic-

tive factor in the multilevel models (Table S5). The final

models retained FF build-up and leaf litter quality as

significant factors for changes in the topsoil (0–10 cm)

pH, exchangeable Ca2+ and Al3+ concentrations, and

this effect declined with soil depth (Table 2). The FF is

built up under Quercus and Alnus over time (Fig. 1),

whereas under Tilia, Prunus and Fraxinus, only minor

FF biomasses were recorded over the complete chrono-

sequence. Leaf litter index values were significantly

(P < 0.001) higher for Prunus and Fraxinus than for

other species. The litter quality index for Alnus was

found to be significantly higher than that for Tilia,

which again was significantly higher than that for Quer-

cus. High values of the leaf litter quality index reflect

high nutrient concentrations, particularly Ca2+, low C

and lignin concentrations, and consequently low C : N

and lignin : N ratios, in the leaf litter. High litter qual-

ity indices after 35 years of forest development were

associated with high soil pH (Fig. 5) and exchangeable

Ca2+ concentrations (Fig. S1) but low exchangeable

Al3+ concentrations (Fig. S2), except for Alnus.

From the age of 17 years, Al3+ concentrations

decreased with increasing soil depth (Table 3, Fig. 2c)

for all studied tree species, indicating the diminishing

impact of acidifying processes with soil depth. Under

Alnus and Quercus, the soil pH and Ca2+ concentrations

increased with soil depth (Fig. 2a and b), and the soil

layers were sharply and significantly differentiated

(Table 3). Under Alnus, Ca2+ concentrations were sig-

nificantly higher in the top 0–5 cm than in the underly-

ing 5–10 cm soil layer (Table 3). Under Tilia, Prunus,

and Fraxinus, the soil layers showed no sharp distinc-

tion in pH or Ca2+ concentration (Fig. 2a and b;

Table 3).

After less than four decades of forest development

on former agricultural land, clear effects on the

number, mass, and composition of the earthworm

Table 2 Output of the final multilevel model analysis show-

ing the effects of (i) cumulative potentially acidifying atmo-

spheric deposition, (ii) leaf litter quality, and (iii) H+ excretion

associated with forest floor build-up on pH and exchangeable

Ca2+ and Al3+ concentrations at different soil depths within

the plough layer (0–5, 5–10, 10–20 and 20–30 cm). The H+

excretion associated with nutrient uptake and N2 fixation were

not retained in the final models

Response

variable Predictor variable df t-Value P value

0–5 cm

pH Atmospheric deposition 47 �4.2 <0.001Leaf litter quality 47 3.8 <0.001Forest floor build-up 47 �4.2 <0.001

Ca Atmospheric deposition 47 �2.9 <0.005Leaf litter quality 47 3.1 0.004

Forest floor build-up 47 2.4 0.022

Al Atmospheric deposition 47 6.3 <0.001Leaf litter quality 47 �1.8 0.082

Forest floor build-up 47 5.8 <0.001

5–10 cm

pH Atmospheric deposition 48 �4.2 <0.001Forest floor build-up 48 �2.8 <0.001

Ca Atmospheric deposition 48 �5.8 <0.001Leaf litter quality 48 2.5 0.015

Al Atmospheric deposition 48 5.1 <0.001Forest floor build-up 48 3.7 <0.001

10–20 cm

pH Atmospheric deposition 49 �5.2 <0.001Ca Atmospheric deposition 49 �6.0 <0.001Al Atmospheric deposition 48 4.5 <0.001

Forest floor build-up 48 4.5 <0.001

20–30 cm

pH Atmospheric deposition 49 �3.8 <0.001Ca Atmospheric deposition 49 �3.5 <0.001Al – – – –



community (Table 4) were found. We found a signifi-

cantly higher biomass and number of epigeic earth-

worms under Alnus than under the four other tree

species. Endogeic species were more prevalent (signifi-

cantly higher biomass and numbers) under Fraxinus

and Prunus, whereas the highest biomasses (and num-

bers for Fraxinus) of anecic species were found under

Tilia and Fraxinus.

0

1000

2000

3000

4000

Exc

hang

eabl

e C

a2+

(m

g kg

–1)

Exc

hang

eabl

e A

l3+

(m

g kg

–1)

Alnus

0

200

400

600

Quercus Tilia Prunus0 20 40 0 20 40 0 20 40 0 20 40 0 20 40

Fraxinus

Soil depth (cm)0–510–20 20–30

3.0

4.0

5.0

6.0

7.0

Soi

l pH

(K

Cl)

Forest age (years)

(a)

(b)

(c)

Fig. 2 The properties of three soil layers (0–5, 10–20, and 20–30 cm) in agricultural fields (0 years) and postagricultural forests contain-

ing five broadleaved deciduous tree species (11–35 years). (a) Soil pH, (b) exchangeable Ca2+, (c) exchangeable Al3+. The x-axis gives

the forest age in years, and the y-axis gives the soil pH-KCl and exchangeable Ca2+ and Al3+ concentrations in mg kg�1. Error bars

denote standard errors.



Discussion

Since the 1980s, scientists have debated the relative

importance of acidifying deposition vs. natural pro-

cesses as the major cause of forest soil acidification. The

controversy began with Krug & Frink (1983), who pos-

tulated that acidification by atmospheric pollutants is

superimposed on long-term acidification induced by

changes in land use and consequent vegetation succes-

sion. Others refuted this theory (Johnson et al., 1984;

Seip & Dillon, 1984; Wright, 1984), and the discussion

continued in the 1990s (Frink, 1996; Kirchner, 1996; Lik-

ens et al., 1996). Different calculation methods and dif-

ferences in the ecosystems studied underlay this debate

(de Vries & Breeuwsma, 1986). Here, we show that

atmospheric deposition contributes significantly to the

observed soil acidification in the former plough layer

(0–30 cm) of postagricultural forests in a region that

received approximately the greatest deposition in

Europe. In addition, our final multilevel model analysis

strongly suggests that natural processes are significant

and important in influencing acidification of the topsoil

(0–10 cm). Atmospheric deposition undoubtedly and

unequivocally drives postagricultural forests towards

more acidic conditions, but the rate of soil acidification

is also determined by the tree species-specific litter

quality and litter decomposition rates. Our discussion

is built around the conceptual model presented in

Fig 6.

After less than four decades of forest development,

initially limed, loamy soils with high pH and high

exchangeable Ca2+ concentrations were significantly

acidified. The extent of acidification depended on the

planted tree species; the topsoil below Quercus and

Alnus trees was ca. 10 and 20 times, respectively, more

acid (in terms of H+ concentration) than under Fraxinus.

Furthermore, the acidification of deeper soil layers

occurred after a shorter reforestation period under

Alnus and Quercus than under Tilia, Prunus, or Fraxinus.

Atmospheric deposition was not significantly different

among the tree species studied (Fig. 1), indicating that

other factors underlie this divergence in soil acidity.

We hypothesize that intrinsic differences in leaf litter

quality among tree species produce fundamentally dif-

ferent nutrient cycling across these ecosystems, both

directly, through the chemical composition of the litter,

and indirectly, through its effects on the size and com-

position of earthworm communities (Reich et al., 2005).

Poor leaf litter quality contributes to the absence of a

burrowing earthworm community, which retards leaf

litter decomposition and consequently results in FF

build-up (Fig. 6). This results in soil acidification

because (i) cations are held back from active circulation

and soil buffering when a FF is built up (Nilsson et al.,

1982), and (ii) the organic acids produced when litter

decomposes slowly play an important role in the natu-

ral H+ production of the soil (de Vries & Breeuwsma,

1986). The bioturbation activities of earthworms can

also counteract soil acidification (Fig. 6).

Leaf litter that is rich in N or poor in lignin (Cotrufo

et al., 2000; Prescott, 2010) has been shown to be

pH (KCl)

Ca

(mg

kg–1

)

03 4 5 6 7 8

1000

2000

3000

4000

Farmland11 years14 years17 years25 years35 years

Fig. 3 pH-KCl values vs. Ca2+ concentrations (mg kg�1) in the

topsoil (0–5 cm) for the studied farmlands and postagricultural

forests.

pH (KCl)

Al (

mg

kg–1

)

0

200

400

600

Farmland11 years14 years17 years25 years35 years

3 4 5 6 7 8

Fig. 4 pH-KCl values vs. Al3+ concentrations (mg kg�1) in the

topsoil (0–5 cm) for the studied farmlands and postagricultural

forests.



positively correlated to the rate of leaf litter decomposi-

tion, at least in ecosystems where microbial activity

dominates litter processing and/or Ca-demanding

earthworms are not abundant (Hobbie et al., 2006). In

the presence of earthworms, however, species differ-

ences in litter Ca have been found to determine litter

decomposition rates, apparently because of the positive

influence of litter Ca on earthworm populations (Reich

et al., 2005; Hobbie et al., 2006). We calculated leaf litter

quality indices as a global indication of leaf litter qual-

ity; high values of this index represent high concentra-

tions of Ca, Mg, K, N and P and low values of C and

lignin, and consequently low C : N and lignin : N

ratios. The tree species studied differ significantly in

leaf litter quality, with significantly higher values for

Fraxinus and Prunus, and significantly lower values for

Quercus. Tilia and Alnus have intermediate leaf litter

quality.

Many studies have linked litter decomposition rates

to earthworm activity (e.g. Staaf, 1987; Muys & Lust,

3.0–2 0 2

–2 0 2

–2 0 2

–2 0 2 –2 0 2

3.5

4.0

4.5

5.0

5.5

Soil

pH

AlnusQuercusTiliaPrunusFraxinus

11 years

3.0

3.5

4.0

4.5

5.0

5.5 14 years

3.0

3.5

4.0

4.5

5.0

5.5

Soil

pH

17 years

Leaf litter quality index

3.0

3.5

4.0

4.5

5.0

5.5 25 years

3.0

3.5

4.0

4.5

5.0

5.5 35 years

Fig. 5 Principal component analysis-extracted leaf litter quality indices vs. soil pH-KCl values for the studied tree species in each

forest.

Table 3 Mean values ± standard deviation of pH-KCl and bioavailable Ca2+ and Al3+ concentrations (mg kg�1) for forests

� 17 years old [Eeuwenhout (17 years), Mortagne forest (25 and 35 years)] at different depths within the plough layer (N = 9)

Depth (cm) Alnus Quercus Tilia Prunus Fraxinus

pH-KCl

0–5 3.3 ± 0.2a 3.6 ± 0.2a 3.9 ± 0.3a 4.0 ± 0.2a 4.3 ± 2.6b

5–10 3.5 ± 0.2a 3.7 ± 0.3a 3.8 ± 0.2a 3.7 ± 0.2a 4.0 ± 0.1a

10–20 3.8 ± 0.3b 3.9 ± 0.4b 4.0 ± 0.3a 3.9 ± 0.2a 4.1 ± 0.2ab

20–30 4.4 ± 0.3c 4.2 ± 0.3c 4.3 ± 0.3b 4.3 ± 0.3b 4.2 ± 0.1b

Ca2+

0–5 860 ± 283b 582 ± 263a 1235 ± 415ab 1413 ± 285ab 1503 ± 369b

5–10 571 ± 298a 473 ± 361a 886 ± 296a 1002 ± 173a 1089 ± 213a

10–20 905 ± 377b 898 ± 378b 1087 ± 303ab 1121 ± 309ab 1249 ± 247ab

20–30 1559 ± 286c 1370 ± 208c 1381 ± 224b 1641 ± 347b 1472 ± 227b

Al3+

0–5 394 ± 109c 379 ± 97c 225 ± 68ab 222 ± 40b 196 ± 39b

5–10 381 ± 138c 375 ± 128c 231 ± 78b 279 ± 73b 134 ± 51a

10–20 246 ± 143b 209 ± 83b 152 ± 67ab 220 ± 91b 107 ± 48a

20–30 90 ± 52a 103 ± 23a 101 ± 48a 124 ± 88a 85 ± 51a

Means with different superscript letters within the same column are significantly different (P < 0.05).



1992; Neirynck et al., 2000; Bohlen et al., 2004; Hobbie

et al., 2006). Under Quercus and Alnus, a build-up of the

litter and fragmentation layers was noted (Fig. 1), caus-

ing the humus type to change from eumull to oligomull

(Jabiol et al., 1995) over 35 years. In contrast, under

Fraxinus, Prunus and Tilia, a eumull humus type was

still present in the oldest forests, and scarce litter

remained at the end of the summer. Earthworm sam-

pling in the 35-year-old Mortagne forest showed a sig-

nificantly higher biomass and number of burrowing

earthworm species under Tilia, Prunus and Fraxinus

than under Alnus and Quercus, consistent with Muys

et al. (1992), Neirynck et al. (2000), and Reich et al.

(2005).

Under Alnus and Quercus, soils became less acidic

with increasing soil depth and sharp distinctions in soil

pH and exchangeable Ca and Al concentrations were

observed between the different soil layers (Fig. 2). In

contrast, in the soil under Tilia, Prunus, and especially

Fraxinus, the different soil layers lack differentiation,

and there is no clear relationship between soil depth

and soil pH or Ca concentrations. The absence of clear

soil layer differentiation under Tilia, Prunus, and

Fraxinus is likely caused by bioturbation of the topsoil

as a result of earthworm burrowing activities. In tem-

perate climates, the topsoil (0–15 cm) may be com-

pletely turned over every 10–20 years by earthworms

(Edwards, 2009). The consumption of litter along with

soil mineral particles, as performed by anecic earth-

worm species, affects litter decomposition (Daniel,

1992) and enhances physical, chemical and biological

soil transformations (Bardgett, 2005), which ensures the

conservation of soil fertility and accounts for buffering

processes against acidification (Muys et al., 1992; Reich

et al., 2005) under Tilia and Fraxinus. Under Prunus and

Fraxinus, the highest abundances of endogeic earth-

worms were found; these species are critical for biotur-

bation (Marhan & Scheu, 2006; Felten & Emmerling,

2009) because they create an extensive and coherent

subhorizontal burrow network to a depth of 50 cm. As

such, the topsoil layer, which is most vulnerable to

atmospheric deposition, is mixed with less acidic and

more base-rich soil from the deeper layers. Burrowing

earthworms thus probably have a decisive influence on

the rate of soil acidification; the permanent mixing of

topsoil with litter, earthworm excrements and soil from

deeper soil layers that are richer in base cations, is cru-

cial for mitigation of the soil acidification that inevita-

bly results from atmospheric deposition (Fig. 6). The

release of base cations towards the topsoil as a conse-

quence of the rapid decomposition of leaf litter (also

stimulated by epigeic earthworms) is also important in

this matter. Burrowing earthworms thus counteract soil

acidification because they cause an acceleration of base

cation circulation through the forest ecosystem. The

beneficial effect of leaching base cations from the litter

layer can be seen in the significantly higher exchange-

able Ca concentrations in the topsoil compared to the

underlying soil, and this under all tree species, except

for Quercus.

Intensively managed farmlands generally exhibit

small earthworm populations, while old-growth decid-

uous forest ecosystems may harbour a sevenfold

increased earthworm density relative to managed

farmland (Smith et al., 2008). The intensive tillage and

Table 4 Mass (g m�2) and number (m�2) of earthworms under the five tree species in the Mortagne forest (35 years)

Epigeic Mass Number Endogeic Mass Number Anecic Mass Number

Alnus glutinosa 22.8 ± 7.0b 223.3 ± 122.3b 0.1 ± 0.1a 3.7 ± 6.4a 0.1 ± 0.2a 1.3 ± 2.3a

Quercus robur 7.8 ± 7.1a 20.4 ± 19.3a 0a 0a 2.6 ± 4.5a 1.3 ± 2.3a

Tilia cordata 2.5 ± 2.4a 5.3 ± 2.3a 0a 0a 22.9 ± 14.9ab 7.7 ± 3.6a

Prunus avium 8.5 ± 8.2a 73.6 ± 61.1a 3.9 ± 1.1c 28.6 ± 18.8b 0.2 ± 0.3a 1.3 ± 2.3a

Fraxinus excelsior 9.3 ± 3.1a 34.1 ± 15.7a 2.0 ± 1.9b 12.4 ± 18.2ab 39.4 ± 26.8b 10.4 ± 11.0a

The numbers presented are the average values (±standard deviation) of three sampling points in each block. Means with different

superscript letters within the same column are significantly different (P < 0.05).

Atmospheric deposi on of (poten ally) acidifyingpollutants

Atmosphere

Leaf li erquality

Forestgrowth

Leaf li erFilteringcapacity

Tree N2 fixa onq y

Forest floorbuild-up

Leaf li erdecomposi on

Microbialcommunity

Forest floor

Burrowing earthwormcommunityMineral soil

Soil acidifica on

Fig. 6 Conceptual model of tree species-related causes of soil

acidification in temperate forests.



chemical applications carried out in farmlands are per-

nicious for earthworms (Edwards et al., 1995; Callaham

et al., 2003). The reforestation of former farmland with

tree species with high leaf litter quality most likely pro-

motes a significant increase in earthworm population.

Tree species with a lower leaf litter quality, such as

Quercus, provide less palatable food for earthworms

and thus prevent the establishment of large earthworm

populations (Reich et al., 2005).

Despite the higher leaf litter quality index, soils

under Alnus acidified faster than under Quercus and

Tilia. The ability of Alnus to symbiotically fix atmo-

spheric N2 can enhance gross soil nitrification and acid-

ification rates through the formation of HNO3, which is

a strong acid (Van Miegroet & Cole, 1984). N2 fixation

can also enhance soil acidification by H+ production as

well as by the leaching of cations associated with

mobile NO3� (Van Miegroet & Cole, 1984). In the 35-

year-old forest, significantly higher epigeic earthworm

masses and numbers were sampled under Alnus com-

pared to the other tree species studied. Epigeic species

use the litter layer both as habitat and food source (Fel-

ten & Emmerling, 2009) and are more tolerant of low

pH values (Edwards, 2009). These epigeic worms,

which rarely burrow (Edwards, 2009) and consequently

cause no bioturbation, are often the only earthworms

found in forests characterized as or evolving to a moder

or mor humus type (Haimi & Boucelham, 1991; Muys

& Granval, 1997).

Nutrient uptake was not retained as a significant fac-

tor causing soil acidification in our final statistical

model. However, when considered as a single predic-

tive factor, nutrient uptake was significantly correlated

with soil acidification (Table S5), although it was less

significant than atmospheric deposition and leaf litter

related parameters. Indeed, several studies report that

nutrient uptake and the subsequent harvesting of

aboveground biomass causes soil acidification (e.g. Ak-

selsson et al., 2007; Vanguelova et al., 2010). By remov-

ing trees through conventional stem-only or whole tree

harvesting, nutrients are irreversibly removed from the

ecosystem, and the latter harvesting type is obviously

the most conducive to soil acidification (Vanguelova

et al., 2010). Nevertheless, in regions with high atmo-

spheric pollution pressure, nutrient uptake appears to

be of secondary importance (Fig. 1). After harvesting,

the decomposition of soil roots, branches, and twigs

will release cations, which may partly neutralize the

acidification by nutrient uptake (Nilsson et al., 1982).

Conclusion

Our results highlight the significant influence of spe-

cies-specific plant traits on the magnitude of human

pollution-induced soil acidification. Postagricultural

forest soils are definitely acidifying as a consequence of

atmospheric depositions, but litter quality and the here-

with related build-up of the FF are important drivers

too. Also nutrient uptake and N2 fixation are causing

soil acidification, but were found to be less important.

For long-term sustainable protection of forest soils and

ecosystems, source-oriented measures aimed at reduc-

ing potentially acidifying N and S emissions should be

complemented with forest management that avoids

homogeneous stands of tree species with a low litter

quality.

Acknowledgements

We are grateful to Luc Willems, Greet De bruyn, Elke Declerckand Kris Ceunen for assisting with the fieldwork, earthwormspecies identification and the chemical analysis of litter, woodand soil samples. We thank Arno Thomaes and Luc De Keers-maeker for providing information on the forests. We also thankMargot Vanhellemont and Andreas Demey for their valuablecomments on the article. We furthermore thank the Forest andNature Agency of the Flemish Government for their permissionto conduct research in their forests. This article was writtenwhile A. D. S. and J. S. held a postdoctoral fellowship and P. D.F. held a PhD fellowship from the Research Foundation – Flan-ders (FWO). K. W. and L. V. N. were paid by Ghent Universitywith a postdoctoral fellowship (BOF) and as a teaching assis-tant, respectively. G. V. was paid by the Flanders Institute forthe Promotion of Innovation by Science and Technology (IWT).

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