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Applied Soil Ecology 83 (2014) 101–108

Contents lists available at ScienceDirect

Applied Soil Ecology

jo ur nal home page: www.elsev ier .com/ locate /apsoi l

onnectivity of litter islands remaining after a fire and unburnt forestetermines the recovery of soil fauna

ndrey S. Zaitseva,b,∗, Konstantin B. Gongalskya,c, Tryggve Perssond, Jan Bengtssond

A.N. Severtsov Institute of Ecology and Evolution, Russian Academy of Sciences, Leninsky prospect, 33, Moscow 119071, Russian FederationInstitute of Animal Ecology, Justus-Liebig-University, Heinrich-Buff-Ring 26, 35392 Giessen, GermanyBiogeography Department, Faculty of Geography, M.V. Lomonosov Moscow State University, Vorobyevy Gory 1, Moscow 119899, Russian FederationDepartment of Ecology, Swedish University of Agricultural Sciences (SLU), PO Box 7044, Uppsala SE-750 07, Sweden

r t i c l e i n f o

rticle history:eceived 20 December 2012eceived in revised form 15 January 2014ccepted 20 January 2014vailable online 16 February 2014

eywords:orest wildfireorridorsoil macrofaunaoil mitesribatidacosystem recovery

a b s t r a c t

Forest wildfires can dramatically affect soil communities and reduce abundance and diversity of soilfauna. The recovery of soil animals after a fire depends both on immigration from the unburnt forest andon local survival in less-burnt spots, but the relative importance of these mechanisms is poorly known.Therefore, these factors were studied with regard to soil macrofauna and soil mites seven years aftera wildfire occurring in a pine forest area with shallow soil in 2001 in central Sweden. Three replicatetransects, each consisting of four plots were studied. The plots were located in (i) the unburnt forestclose to the fire edge; (ii) slightly burnt patches directly attached to the unburnt forest; (iii) slightly burntpatches surrounded by bare rock but connected to the unburnt forest edge by a corridor with mostlyunburnt litter and vegetation; and (iv) island patches not connected with a corridor to the unburnt forestedge. The hypothesis was that that soil animals would particularly disperse from the unburnt forest tomoderately burnt plots inside a burnt area via the network of less-burnt corridors. Poor dispersers wouldbe especially few in the island patches lacking connection to the “mainland”, whereas good disperserswould be independent of gaps in connectivity. As expected, the highest abundance of both macrofaunaand oribatid mites was found in the unburnt forest. Resident soil macro- and mesofauna representativeshad half the abundance in the edge and corridor plots as compared to the control, but their abundancewas not lower in the island plots than in the corridor plots indicating on-site survival and recovery.Mobile mesostigmatid mites did not show any significant reduction of abundance in any of the plots. Theabundance of soil-dwelling oribatid mites did not differ between islands and unburnt forest, but mobileaboveground oribatids had significantly lower abundance on the islands than in the unburnt forest. Theopposite was observed for aboveground and belowground oribatid mite species richness. In conclusion,

belowground animals showed mainly local survival and seemed to be independent of corridors presence,whereas most aboveground and mobile macro- and mesofauna seemed to be more responsive to isolationinduced by forest fires. Soil and litter corridors connecting unburnt patches inside the burnt forests withthe unburnt edges were important mainly for less mobile groups of soil macrofauna. This supports theidea that there is a relatively slow process of soil ecosystem recovery and that external colonization ofthe burnt areas dominates over the local survival and recovery from refuges.

. Introduction

Forest wildfires occur in the boreal zone of Europe with an

nterval of 50–200 years (Fauria and Johnson, 2008), maintain-ng pyrogenic ecosystem cycles in the forests (Niklasson andranström, 2000). Human activity and global climatic change

∗ Corresponding author at: A.N. Severtsov Institute of Ecology and Evolution, Rus-ian Academy of Sciences, Leninsky pr. 33, 119071 Moscow, Russian Federation.el.: +7 963 711 9154; fax: +7 495 954 5534.

E-mail address: andrey.zaytsev@hotmail.com (A.S. Zaitsev).

929-1393/$ – see front matter © 2014 Elsevier B.V. All rights reserved.ttp://dx.doi.org/10.1016/j.apsoil.2014.01.007

© 2014 Elsevier B.V. All rights reserved.

normally increase fire frequency (Beck et al., 2011). Forest firesreturn biogenic elements bound in the tree biomass to the soil asash. However fires also result in a significant loss of organic matterand macro-elements like carbon, nitrogen, phosphorus and others(Bormann et al., 2008). Because most ground vegetation and theupper organic layer are combusted during fire, the hydrothermalconditions in soil will change. The organic matter remaining aftera fire is mostly well-decomposed and offers a poor substrate for

decomposer organisms. These factors determine the developmentof post-fire soil biota communities, which form a soil food webstructure drastically different from the pre-fire one (Huhta, 1971;Holliday, 1991; Gongalsky and Persson, 2013).

1 Soil Ecology 83 (2014) 101–108

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02 A.S. Zaitsev et al. / Applied

There are two major factors affecting the recovery of soil animalopulations after forest fires: (1) immigration from the unburntorest into the burnt area, and (2) survival in less-burnt patches oreep soil layers including hatching of eggs that may withstand fireGongalsky et al., 2012). The relative importance of these two mech-nisms depends on fire severity, season of fire, fire heterogeneitynd weather conditions. For small soil animals like mites or collem-olans the size of the burnt area seems to be important (Malmströmt al., 2008). Another important set of factors is related to the ini-ial microhabitat differentiation within the burnt forest and theirneven destruction by fire. Such heterogeneity is often observed inoreal forests growing in moraine landscapes (Wang et al., 2012).or example, our earlier study showed that boggy depressions suf-ered much less from fires than the surrounding hills, where soilas burnt down to the bedrock and left no habitat for recovery

f the soil community (Gongalsky et al., 2012). We also noticedhat such depressions and unburnt areas may sometimes be con-ected to each other and to the unburnt forest by a network ofeakly burnt corridors of forest floor and create a complex system

f islands and passages available for colonization similar to thosenown for the lake islands as was assumed by Jonsson et al. (2009).resence of such corridors might be catalytic for recolonization ofefaunated areas by soil organisms (Rantalainen et al., 2005, 2008).o better understand spatial patterns of soil fauna recovery after are in boreal forests we aimed at combining this knowledge withhe set of driving forces related to the dispersal ability of differentoil animal taxa (Berthet, 1964; Ojala and Huhta, 2001; Åström andengtsson, 2011; Lehmitz et al., 2012).

The aim of this study was to evaluate the role of corridorsormed by relatively unburnt litter and soil in the recovery of soilauna communities after a fire as compared to isolated patchesithout any connection to unburnt areas in a boreal forest. Weypothesized that soil animals would particularly disperse fromhe unburnt forest to moderately burnt plots inside a burnt areaia the network of less-burnt corridors. Animals with low mobil-ty should show slower recovery than mobile animals under suchonditions.

. Material and methods

.1. Study area

The study area is situated 20 km south of Stockholm in centralweden (59◦10′N, 18◦20′E) at the border of the Tyresta nationalark and nature reserve, which together make up 4700 ha of mostlyoniferous forest. Its climate and geography are described in detailn the paper of Gongalsky and Persson (2013). Before the fire whichamaged 22 ha in August 2001 a sparse forest of Scots pine (Pinusylvestris) and common silver birch (Betula pendula) was growingn the rocky outcrops, whereas the mire elements were to a largextent covered by Scots pine and hairy birch (Betula pubescens). Theine forest is dominated by 150–300 year-old trees and has nevereen affected by large-scale forestry. It has a well-documentedistory of post-fire recovery of both vegetation (Schimmel andranström, 1996; de Chantal and Granström, 2007) and soil biota

Malmström et al., 2008; Gongalsky and Persson, 2013).The soil depth varied between a few cm on the rocky outcrops

o 20–30 cm in the wet depressions in the unburnt forest. The fireas not very intense, but was strong enough to burn away the

hin soil layer on rocky outcrops covered by lichens and mosses.

fter the fire, the burnt area had a large proportion of bare rocks

60%) intermixed with moist/wet depressions (40%) filled with–10 cm deep organic layer. Initial pH in the forest soil was 4.3nd increased to 4.5–4.7 in the burnt area (Gongalsky and Persson,013).

degrees of isolation/distance from the unburnt forest (control forest, edge plot,corridor plot and island plot). Plots within the burnt area are surrounded by barerock.

2.2. Sampling design

Three replicate transects were established at a distance of about200 m from each other. Each transect consisted of four plots, includ-ing one forest control about 30 m outside the burnt area and threeplots inside the burnt area representing different isolation levelsfrom the control (Fig. 1). The “edge” plot was located at a burntpatch directly attached to the unburnt forest. The “corridor” plotwas located inside the burnt area at a distance of ca. 20 m to theforest edge, consisting of a less-burnt island surrounded by barerock but connected to the forest by a weakly burnt corridor. Theconnecting corridor was never broader than the plot itself and had∼0.5–1.0 m in width. The “island” plot was selected at the samedistance as the previous one, but not connected to other litterpatches or forest matrix. Thus, the plots represented an increasingdegree of isolation from the unburnt forest through the gradientcontrol > edge > corridor > island. Each plot had an area of 2–10 m2

(Fig. 1), and the plots were selected at places with similar soil depthand soil moisture.

Samples were collected in early November 2008. Five intactsoil monoliths per plot, including the litter layer, were randomlytaken with the help of a steel 10 cm × 10 cm frame to a depthof 8–12 cm (depending on the humus layer depth). The distancebetween the samples varied between 50 cm and 1 m (see data treat-ment below). The samples were placed in plastic bags in the fieldand stored at 4 ◦C for a maximum of 3 days before animal extractionin a set of 60 simultaneously operated modified Tullgren funnels.The samples were evenly spread out to a thickness of 1–2 cm on20 cm × 20 cm nets with 5 mm mesh in the funnels. The soil sampleswere always divided into two subsamples (about 0–5 and 5–10 cmdepth) before extraction to ensure high extraction efficiency. Thesamples were then heated and dried for 4 days by means of con-ventional light bulbs, and extracted animals were collected andpreserved in 70–80% ethanol.

2.3. Soil animal identification and ecological classification

All soil animals not belonging to microfauna (mainly Nematoda),or mesofauna (mainly Collembola, Acari, and Enchytraeidae) andbeing larger than 2 mm in any of its dimensions were ascribedto macrofauna. Macrofaunal groups were sorted, counted and

Soil Ec

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A.S. Zaitsev et al. / Applied

dentified under a dissecting microscope. Due to problematic iden-ification of the juvenile and larval stages of some groups, mostdentifications were done down to the family level with the help ofpecial literature (Mamaev, 1972; Dindal, 1990; Klausnitzer, 1994,996). In total, 60 samples were processed.

Among the mesofauna, oribatid and mesostigmatid mites wereaken into account. Oribatid mites were counted and identified tohe species level, using the keys of Balogh and Balogh (1992a,b),hilarov and Krivolutsky (1975) and Weigmann (2006). Due to the

arge number of mites, enumeration and species identification wereade only for two random samples from each plot. In total, 24

amples were processed. Higher abundance of mites compared toacrofauna and the averaging procedure for getting a single data

oint per plot in our study allowed us to count mites only in soilamples with the total area of 200 cm2. This seems sufficient forescription of the abundance and diversity of mites and approxi-ately equals to 10 cylindrical soil cores with the diameter of 5 cmhich is a standard sampling effort for this type of studies (see e.g.

aitsev et al., 2002).The soil invertebrates were classified into three groups accord-

ng to their dispersal ability and following the classificationsuggested by Ghilarov (1965), Mamaev (1972) and Dindal (1990),amely (1) winged and mobile animals; (2) surface-litter animals,oderately mobile; and (3) endogeic animals, resident or very littleobile. All groups possessing wings (e.g., insects) and spiders were

onsidered as highly mobile macroinvertebrates. Other arachnids,olluscs and myriapods (except for Geophilidae) were considered

s moderately mobile, and earthworms, geophilids, and larvae oflateridae and Cantharidae were ascribed to resident macroin-ertebrates. Oribatid mites were classified as resident (Berthet,964), and mesostigmatid mites as mobile animals (Krantz andalter, 2009). Soil macrofaunal groups and oribatid mites were

urther subdivided into above- (surface and litter-dwelling) orelowground (soil-dwelling) ones (Krivolutsky, 1995; Gongalskynd Persson, 2013).

.4. Data treatment

Faunal abundance was expressed as mean number of individu-ls m−2 per plot. Samples within a plot (n = 5 for macrofauna, and

= 2 for Oribatida and Mesostigmata) were considered as pseudo-eplicates, and were only used to form a single mean value perlot. Under such treatment the given sampling effort seems suf-cient for adequate description of even macrofauna community

or the purposes of our research. To test if ‘treatment’, i.e. degreef plot isolation (DPI) (control, edge, corridor and island) influ-nced soil animal abundance and ecological structure, the dataere analysed with one-way ANOVA. The analysis was done onon-transformed data for all mite groups and overall macrofauna,ecause the Kolmogorov–Smirnov test did not reject normalityp > 0.05) in the non-transformed state. Abundance of abovegroundnd belowground macrofauna and of different mobility groups ofacrofauna required log-transformation to achieve normality. If

he ANOVA showed significant overall differences between treat-ents, the individual differences were tested with a Tukey HSD test.

tatistical calculations were performed using Statistica 8 software.We also calculated taxa gain and loss between island, corridor,

nd edge, relative to control plots. Taxa gain was the number of taxataxa richness) found in the respective plot and not detected in theontrol plot. Taxa loss was the opposite, i.e. absence of the taxon inhe respective plots in comparison with control. Taxa gain and lossere calculated at the resolution of identified taxonomic groups for

acrofauna and at species level for oribatid mites. This was done

eparately for aboveground and belowground-living macrofaunaroups and oribatid mite species. Since the groups studied con-ained different number of taxa, gain and loss were standardized

ology 83 (2014) 101–108 103

as the percentage of the total recorded amount of taxa in controlfor above- and belowground oribatid species and soil macrofaunagroups.

3. Results

3.1. Abundance

There was a significant impact of DPI on the total abundanceof macrofauna (ANOVA: F = 5.91, p < 0.01) (Table 1). The edgeplots hosted 3,600 ind. m−2 (Table 1) with inter-transect variationbetween 1,700 and 6,000. At the corridor plots, there were about2,570 ind. m−2 (Table 1). The island plots had the lowest abundanceof soil macrofauna, 2,040 ind. m−2 (Table 1). Total macrofaunalabundance in the control plots reached the highest values, 5,780ind. m−2 (Table 1).

Oribatid abundance was significantly lower in the island plotsthan in the control plots (ANOVA: F = 4.34, p < 0.05), and oribatidabundance in the edge and corridor plots demonstrated inter-mediate values (Table 1). The abundance of oribatid mites was,on average, about 58,000 ind. m−2 in the unburnt control anddid not exceed 120,000 in the different transects. Mesostigmatidmite abundance also seemed to decrease from the control (about5700 ind. m−2) to the island plots (about 2,100 ind. m−2), butthe decline was not statistically significant (ANOVA: n.s., p = 0.12)(Table 1).

3.2. Diversity patterns and taxonomic composition

In total 39 soil macrofaunal taxonomic groups (mainly fam-ily level) were identified, of which 25 were found in the control,29 in the edge, 25 in the corridor and 23 in the island treat-ments. The total number of aboveground taxa ranged from 15 inthe corridor and island to 19 in the edge. Belowground taxa werefewer, ranging from 8 in the island plots to 10 in the edge plots.The mean number of macrofaunal taxonomic groups was highest(n = 18) at the corridor plots and lowest in the island plots (n = 13)(Table 1). The most abundant group at the edge plot were thripses(about 1650 ind. m−2), and other groups had lower abundance, e.g.diplopods with 300 and Cecidomyiidae larvae with 400 ind. m−2.One of the edge plots was unique with several Raphidiidae lar-vae emerging from the samples. In the corridor plot, the mostabundant groups were thripses (500 ind. m−2), and Cecidomyiidaelarvae (520). The third most abundant group was Cantharidae lar-vae (300 ind. m−2). In the island plots diplopods (450 ind. m−2)dominated followed by chironomid larvae (390) and spiders (250).All diplopods found here belonged to the same species, Proteroiu-lus fuscus, and were only found in one of the transects, wherethey reached an abundance of 1340 ind. m−2 and largely affectedoverall macrofauna abundance (Table 1). The three most abundantgroups in control plots were thripses (1800 ind. m−2), diplopods(900 ind. m−2), and cecidomyiid larvae (700 ind. m−2). Molluscs,earthworms, and Stratiomyiidae larvae were found exclusively inthe control plots.

Sixty-six oribatid mite species were identified in the soil sam-ples. Of these, 57 species were found in the control, 45 in the edgeand 40–41 in the corridor and island plots. On average, the controlplots hosted high mite diversity, 36 species (Table 1). Nine orib-atid mite species present in edge, corridor or island plots were notrecorded in the control. In contrast nine other species were presentexclusively in the control plots (Table 1). The number of species

found exclusively in each of the burnt treatments was lower thanin the control, ranging from 0 (edge) to 4 species (island) (Table 1).

The analysis of macrofaunal and oribatid taxonomic composi-tion indicated both gains and losses in taxa richness in the partially

104 A.S. Zaitsev et al. / Applied Soil Ecology 83 (2014) 101–108

Table 1Mean abundance (ind. m−2 ± SE, n = 3) and mean number of taxonomic groups (±SE) of the soil fauna taxa studied in three transects, each with four plots (control, edge,corridor and island) with different degree of isolation after burning. L: larvae, I: imagines, M: mobile, W: moderately mobile (wandering), R: resident, A: aboveground, B:belowground. Estimates with different letters within the same line are significantly different at p < 0.05 according to Tukey HSD tests.

Ecologicalgroups

Control Edge Corridor Island

Macrofauna mean abundance

OligochaetaLumbricidae R B 7 ± 7 0 0 0

ArachnidaAraneae M A 433 ± 70 227 ± 66 213 ± 27 247 ± 24Pseudoscorpiones W A 7 ± 7 7 ± 7 0 7 ± 7

MyriapodaGeophilidae R B 60 ± 60 7 ± 7 73 ± 37 113 ± 74Lithobiidae W A 40 ± 11 20 ± 20 20 ± 12 20 ± 12Julidae W A 900 ± 253 307 ± 229 153 ± 113 447 ± 447

InsectaThysanopteraAll Thysanoptera families M A 1787 ± 743 1653 ± 983 500 ± 31 207 ± 37

HemipteraLygaeidae M A 0 13 ± 13 0 7 ± 7Tingidae M A 0 13 ± 13 40 ± 20 20 ± 12

HomopteraAphidoidea M A 0 73 ± 73 33 ± 33 27 ± 13Ortheziidae M A 480 ± 253 187 ± 123 0 0Cicadellidae M A 33 ± 7 7 ± 7 60 ± 41 7 ± 7

DipteraAsilidae, L M B 20 ± 20 7 ± 7 0 0Cecidomyidae, L M B 707 ± 397 393 ± 225 520 ± 245 213 ± 173Ceratopogonidae, L M B 87 ± 87 0 53 ± 35 0Chironomidae, L M B 153 ± 115 733 ± 53 273 ± 148 387 ± 242Empididae, L M B 0 7 ± 7 0 0Muscidae, L M B 7 ± 7 0 7 ± 7 0Stratiomyidae, L M A 33 ± 33 0 0 0Psychodidae, L M B 0 0 7 ± 7 0Sciaridae, L M B 0 7 ± 7 0 0Tipulidae, L M B 0 0 33 ± 24 7 ± 7

ColeopteraByrrhidae, L W B 0 7 ± 7 0 13 ± 7Carabidae, L M A 47 ± 24 27 ± 7 33 ± 13 27 ± 17Coleoptera, L M A 0 7 ± 7 0 0Curculionidae, L M A 0 0 7 ± 7 0Elateridae, L R B 480 ± 208 100 ± 30 733 ± 47 147 ± 109Pselaphidae, I M A 20 ± 20 40 ± 40 13 ± 13 0Coleoptera, unidentified, I M A 7 ± 7 0 0 0Staphylinidae, L M A 7 ± 7 20 ± 20 27 ± 7 7 ± 7Cantharidae, L R B 427 ± 118 280 ± 110 307 ± 64 733 ± 73Carabidae, I M A 7 ± 7 13 ± 13 53 ± 13 47 ± 37Chrysomelidae, I M A 0 7 ± 7 20 ± 20 0Curculionidae, I M B 0 13 ± 7 13 ± 7 7 ± 7Staphylinidae, I M A 20 ± 12 53 ± 35 13 ± 7 7 ± 7

Raphidiidae M A 0 20 ± 20 0 0Noctuidae, L M A 13 ± 13 13 ± 7 27 ± 13 7 ± 7Arionidae W A 0 0 0 7 ± 7Cochlicopidae W A 7 ± 7 0 0 0Total macrofauna abundance 5780 ± 1235a 3600 ± 1236ab 2573 ± 276b 2040 ± 745b

Mean number of macrofauna taxonomic groups 15.0 ± 1.5c 16.3 ± 1.2b 18.0 ± 0.6a 13.3 ± 0.9d

Total number of macrofauna taxonomic groups 27 29 25 23Total number of macrofauna aboveground taxonomic groups 16 19 15 15Total number of macrofauna belowground taxonomic groups 9 10 10 8

Oribatida species mean abundanceAchipteria nitens (Nicolet, 1855) R A 467 ± 467 0 33 ± 33 0Adoristes ovatus (C.L.Koch, 1839) R B 300 ± 176 67 ± 44 0 0Atropacarus striculus (C.L.Koch, 1835) R A 1050 ± 721 2417 ± 1743 2033 ± 1459 850 ± 437Brachychthonius berlesei Willmann, 1928 R B 2667 ± 1453 133 ± 88 1783 ± 906 333 ± 209Neobrachychthonius marginatus (Forsslund, 1942) R B 3533 ± 895 1000 ± 1000 0 167 ± 167Caleremaeus sp. R B 17 ± 17 0 0 17 ± 17Camisia biurus (C.L.Koch, 1839) R B 133 ± 44 100 ± 76 33 ± 17 17 ± 17C. lapponica (Trägårdh, 1910) R B 0 0 17 ± 17 0C. spinifer (C.L.Koch, 1835) R B 150 ± 126 0 0 0Carabodes areolatus Berlese, 1916 R A 33 ± 33 83 ± 83 0 17 ± 17C. femoralis (Nicolet, 1855) R A 333 ± 240 83 ± 44 17 ± 17 17 ± 17C. labyrinthicus (Michael,1879) R A 150 ± 76 0 0 17 ± 17C. minusculus Berlese, 1923 R A 83 ± 44 117 ± 73 250 ± 250 33 ± 17Cepheus cepheiformis (Nicolet, 1855) R A 17 ± 17 17 ± 17 0 0Ceratoppia bipilis (Hermann, 1804) R A 0 0 17 ± 17 17 ± 17Ceratozetes minimus (Sellnick, 1928) R A 483 ± 252 4217 ± 4167 117 ± 117 0

A.S. Zaitsev et al. / Applied Soil Ecology 83 (2014) 101–108 105

Table 1 (Continued)

Chamobates borealis (Trägårdh, 1902) R B 867 ± 719 333 ± 285 33 ± 33 350 ± 301Conchogneta traegardhi (Forsslund, 1947) R B 83 ± 83 1600 ± 1266 50 ± 50 0Cultroribula dentata Willmann, 1950 R B 167 ± 120 50 ± 50 50 ± 29 100 ± 50Damaeus onustus C.L.Koch, 1841 R A 50 ± 50 0 0 0Diapterobates humeralis (Hermann, 1804) R A 17 ± 17 0 0 0Dissorhina ornata (Oudemans, 1900) R B 167 ± 167 367 ± 273 883 ± 693 1183 ± 491Eremaeus sp. R B 150 ± 126 17 ± 17 0 0Eupelops torulosus (C.L.Koch, 1836) R A 717 ± 174 117 ± 67 283 ± 259 100 ± 100Euphthiracarus monodactylus (Willmann, 1919) R B 50 ± 50 0 0 0Euzetes globulus (Nicolet, 1855) R A 17 ± 17 0 0 0Galumna lanceata Oudemans, 1900 R A 300 ± 252 517 ± 252 2750 ± 1247 1500 ± 473Galumna sp. R A 33 ± 33 717 ± 419 267 ± 167 33 ± 33Hemileius initialis (Berlese, 1908) R B 1533 ± 819 117 ± 73 33 ± 33 0Heminothrus targionii (Berlese, 1885) R B 150 ± 126 33 ± 33 117 ± 60 17 ± 17Hypochthonius rufulus C.L.Koch, 1835 R B 83 ± 83 0 0 0Liacarus coracinus (C.L.Koch, 1840) R A 317 ± 67 83 ± 44 83 ± 60 100 ± 58L. subterraneus (C.L.Koch, 1844) R A 117 ± 60 50 ± 50 0 0Liochthonius sellnicki (Thor, 1930) R B 83 ± 83 500 ± 289 0 0Malaconothrus monodactylus (Michael, 1888) R B 83 ± 60 483 ± 388 2650 ± 2070 1550 ± 889Micreremus brevipes (Michael, 1888) R B 0 0 0 17 ± 17Microtritia minima (Berlese, 1904) R B 900 ± 900 1500 ± 1475 0 1867 ± 1792Nanhermannia coronata Berlese, 1913 R B 1250 ± 757 950 ± 759 533 ± 484 0N. nana (Nicolet, 1855) R B 350 ± 350 33 ± 33 17 ± 17 0Nothrus anauniensis Canestrini et Fanzago, 1876 R B 2233 ± 1684 383 ± 186 717 ± 470 17 ± 17N. pratensis Sellnick, 1928 R B 433 ± 433 0 33 ± 33 0Medioppia falcata (Paoli, 1908) R B 0 50 ± 50 0 150 ± 150M. fallax (Paoli, 1908) R B 5217 ± 3389 917 ± 667 1333 ± 882 1167 ± 392Oppia sp. R B 0 33 ± 33 33 ± 33 17 ± 17M. subpectinata (Oudemans, 1901) R B 5017 ± 5017 67 ± 67 0 67 ± 33Oppiella nova (Oudemans, 1902) R B 9850 ± 3623 15400 ± 11797 7267 ± 3370 2017 ± 1169Oribatula tibialis (Nicolet, 1855) R B 117 ± 73 17 ± 17 17 ± 17 50 ± 29Phthiracarus laevigatus (C.L.Koch, 1841) R A 233 ± 44 500 ± 176 133 ± 73 33 ± 33Platynothrus peltifer (C.L.Koch, 1839) R B 67 ± 67 0 0 0Porobelba spinosa (Sellnick, 1920) R A 117 ± 117 17 ± 17 0 50 ± 50Protoribates variabilis Rajski, 1958 R B 0 0 0 100 ± 58Quadroppia quadricarinata (Michael, 1885) R B 0 0 0 250 ± 87Rhysotritia duplicata (Grandjean, 1953) R B 750 ± 520 333 ± 240 1350 ± 153 1883 ± 977Scheloribates laevigatus (C.L.Koch, 1835) R B 1167 ± 627 1317 ± 417 1300 ± 841 1450 ± 1097Sellnickochthonius zelawaiensis (Sellnick, 1928) R B 5383 ± 2919 217 ± 217 100 ± 100 383 ± 217Steganacarus carinatus (C.L.Koch, 1841) R A 683 ± 438 467 ± 467 100 ± 100 67 ± 33Suctobelbella acutidens (Forsslund, 1941) R B 167 ± 167 0 500 ± 500 0S. falcata (Forsslund, 1941) R B 1867 ± 1206 1333 ± 667 2150 ± 1775 350 ± 176S. hamata Moritz, 1970 R B 467 ± 467 0 0 0S. perforata (Strenzke, 1950) R B 0 0 0 167 ± 167S. similis (Forsslund, 1941) R B 167 ± 167 0 0 0S. subcornigera (Forsslund, 1941) R B 0 67 ± 67 167 ± 167 0S. vera (Moritz, 1964) R B 367 ± 367 550 ± 333 100 ± 100 467 ± 467Tectocepheus velatus (Michael, 1880) R B 6567 ± 3767 1667 ± 867 2400 ± 1652 1717 ± 749Trichoribates trimaculatus (C.L.Koch, 1835) R A 150 ± 126 0 17 ± 17 0Trimalaconothrus foveolatus Willmann, 1931 R B 33 ± 33 83 ± 44 483 ± 459 833 ± 463

Total Oribatida abundance 57933 ± 21602a 39117 ± 12258ab 30250 ± 4617ab 19533 ± 3628b

Mean number of Oribatida species 35.0 ± 1.5a 27.3 ± 0.3b 23.7 ± 3.9b 25.0 ± 2.3b

Total number of Oribatida species 57 45 40 41Total number of Oribatida aboveground species 21 14 13 13Total number of Oribatida belowground species 36 31 27 28Total number of Oribatida “unique” species 8 0 2 4Mesostigmata mean abundance

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Total Mesostigmata abundance W unknown

urnt as compared to the control plots (Fig. 2). For the above-round macrofauna, gains and losses were fairly similar in islandnd corridor plots relative to the control, whereas the gains werewice as large as the losses in the edge plots (Fig. 2A). Among theelowground macrofauna, losses were higher than gains in the

sland plots and lower in corridor and edge plots (Fig. 2B). Thepecies losses among oribatid mites were higher than the gains foroth aboveground and belowground species. For aboveground mitepecies, the loss was especially high (up to 62% of the total num-er of species) in comparison with the control treatment (Fig. 2C)

nd was not compensated by the species gain. Species losses foroth island and corridor plots were similar as well as species gain.he relative species loss was smaller for belowground (about 30%)han for aboveground oribatids (about 60%) (Fig. 2D), and the gain

667 ± 905 4283 ± 1184 2633 ± 784 2067 ± 533

of species was somewhat higher for the belowground oribatidmites.

3.3. Ecological structure of soil fauna communities

Most (76%) of the macrofaunal groups could be classified asmobile (flying) macroinvertebrates, 15% as wandering (moderatelymobile), and 10% as resident (Table 1, Fig. 3). The abundance ofmobile macrofauna tended to be lower (p = 0.055) in the island

treatment (ca. 1,200 ind. m−2) than in the control treatment (ca.3,900 ind. m−2) (Fig. 3). The abundance of moderately mobile wan-dering macrofauna did not differ significantly between treatments(ANOVA: n.s., p > 0.05). Resident macrofauna were less abundant in

106 A.S. Zaitsev et al. / Applied Soil Ecology 83 (2014) 101–108

Ffs

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Fig. 4. Mean abundance (ind. m−2 ± SE) of above- and below-ground invertebratesin the four plots with different degrees of isolation after burning of (A) soil macro-

ig. 2. Relative gains and losses of aboveground (A) and belowground (B) soil macro-auna taxonomic groups and aboveground (C) and belowground (D) oribatid mitepecies in edge, corridor and island plots in comparison with control plots.

he edge, corridor and island plots than in the control plots (Fig. 3),ANOVA: F = 3.77, p < 0.05).

Mobile (Mesostigmata) and resident (Oribatida) mitesesponded differently to increasing degree of isolation. While themount of slowly moving oribatid mites was gradually decreasingrom the control to the island treatment, mobile mesostigmatid

ites did not demonstrate any significant reduction of theirumbers across all plots (Table 1).

Another contrast was found for the abundance of above- andelowground living macrofauna and oribatids. The abundance ofboveground macrofauna was significantly lower in the corridornd island treatments (ca. 1200 and 1100 ind. m−2, respectively)han in the control treatment (about 3900 ind. m−2) (ANOVA:

= 11.7, p < 0.001) (Fig. 4A). The abundance of belowground macro-auna was only nearly significantly different between treatmentsnd was in the range of ca. 950–1,950 ind. m−2 (ANOVA: F = 2.75,

> 0.05) (Fig. 4A). Nor the abundance of aboveground oribatid mites

as not significantly different between treatments ranging from

,400 ind. m−2 in the edge and 2,800 ind. m−2 in the island plotsANOVA: n.s., p > 0.05) (Fig. 4A). The abundance of belowground

ig. 3. Mean abundance of mobile, moderately mobile and resident groups of soilacrofauna (ind. m−2 ± SE) in the four plots with different degrees of isolation after

urning. Estimates with different letters for the same category of macrofauna are sig-ificantly different at p < 0.05 according to Tukey HSD tests. For mobile macrofaunaifferences were nearly significant (p ∼ 0.055).

fauna and (B) oribatid mites. Estimates with different letters for the same categoryof fauna are significantly different at p < 0.05 according to Tukey HSD tests.

oribatids declined from the control to the island treatment(ANOVA: F = 6.19, p < 0.05) (Fig. 4B). The edge and corridor treat-ments had intermediate abundances and were not significantlydifferent from the other two treatments.

4. Discussion

Moderately burnt soil patches in the fire area that showedvarious degrees of isolation from the unburnt forest hosted dif-ferent numbers of soil animals seven years after the wildfire. Thelitter “islands”, completely surrounded by bare rock, had the low-est abundance of both macrofauna and mites, while similar patchesconnected by corridors to the unburnt forest “mainland” had highersoil animal abundance. This demonstrates an important but proba-bly group-specific role of corridors which connect the interior ofthe burnt areas with the unburnt surroundings for maintainingrecolonization (Kounda-Kiki et al., 2009). In our study corridorsmoderately affected colonization of the least mobile soil animalgroup–oribatid mites, but had no effect on the recovery of mobilemacrofauna and mesostigmatid mites (Lehmitz et al., 2011, 2012).Edge plots (treatments) attached to the unburnt forest did not dif-fer significantly in faunal abundance from the other treatments forany of the faunal groups with the exception of resident macrofauna.This supports the idea that there is a relatively slow process ofsoil ecosystem recovery and that external colonization of the burntareas dominates over the local survival and recovery from refuges.

The rates of soil fauna recovery of about 35% for both macro-and mesofauna observed seven years after the wildfire in our studycorresponded well with the degrees of recovery of about 30–40%observed 5–10 years after fire in boreal forests in Scandinavia

(Malmström et al., 2009; Gongalsky and Persson, 2013). It has beenshown that burnt areas in the boreal zone are initially colonized bymobile taxonomic groups, mostly flies, which possibly experience

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ack of predation during the first stages of colonization (Gongalskynd Persson, 2013).

Soil animal survival and recovery after fire is largely dependentn the quantity and quality of the remaining organic matter. Inur study, the fire was fairly moderate at the study plots, wherehe main difference between the unburnt and burnt plots washe lack of all ground vegetation and litter layer (L) in the burntlots initially after the fire, whereas the fragmented litter layerF) was partly affected and the humus layer proper (H) was morer less intact. The litter and fragmented litter layers have higherarbon availability and microbial biomass than deeper soil layersPersson et al., 2000; Berg et al., 2001). Therefore, we expectedhat microbi-detritivorous soil fauna, for example, litter-dwellingribatid mites, millipedes and lumbricids would initially decreaseramatically in comparison with the unburnt forest. Also fungiv-rous oribatid mites were expected to decrease in abundance,ecause burning reduces fungal biomass (Pietikäinen and Fritze,995). Pietikäinen and Fritze (1995) found that prescribed burningeduced microbial biomass in the humus layer to 45 and 60% of thatn unburnt clear-cut according to the fumigation–extraction (FE)nd substrate-induced respiration (SIR) methods, respectively, oneear after burning. Fritze et al. (1993) found that there was a recov-ry in the ratio of microbial carbon to total carbon during the first 12ears after fire, which could be explained by an increasing supplyf energy-rich carbon sources derived from the post-fire plant suc-ession. This indicates that microbial-feeding animals would have

smaller food supply in burnt than in unburnt forests during aelatively long period.

There are also other indirect effects of post-fire organic matteruality described in the literature (Certini, 2005). The amount ofio available nutrients from ash, which moves down the soil pro-le, may directly influence microbial communities and the biomassCairney and Bastias, 2007), which in turn creates potential for a

ore rapid recovery of microbi-detritivorous micro- and meso-auna. Another pathway associated with post-pyrogenic organic

atter turnover is instant regrowth of post-fire vegetation (likeosses) which are an important food source for plant-feedingacrofauna like aphids, thripses and scale insects. When the

round vegetation has recovered, phytophagous groups becomeain components of the soil macrofaunal community. After 5–6

ears, mobile soil-living predators can be found in the burnt areaSaint-Germain et al., 2004). In contrast, resident soil mesofauna, inur study represented by oribatid mites, mainly relies on local sur-ival due to very low mobility and shortly after burning is mainlyepresented by soil-dwelling oribatid species (Gongalsky et al.,012). However, direct comparison of this study and our previousesearch on this issue is not possible due to considerable geographicifferences between the study plots. Many mesostigmatid mitesre mobile and restored their abundance seven years after firendependent on the degree of plot isolation. The availability of cor-idors did not seem to have any effect on the mesostigmatid miteommunity recovery. This might be potentially explained by bet-er survival or recovery of their prey (nematodes, collembolans andther mites) in the unburnt islands in comparison with macrofaunaKoehler, 1997). Slow moving oribatids, which probably are sensi-ive to drought on bare rocks, require more time for recolonizationf isolated unburnt patches. The presence of even small corridorsignificantly improves the process of recovery of the oribatid com-unity abundance in burnt forests (Starzomski and Srivastava,

007).In contrast to our initial assumption, resident macrofaunal

roups were able to survive moderate forest fires in deeper soil

orizons, while mobile macrofauna were more vulnerable to theildfire event (Gongalsky et al., 2012). It was also reflected in the

verall number of taxonomic groups found in the four treatments.nother explanation may be related to the dynamics of mobile

ology 83 (2014) 101–108 107

aboveground taxonomic groups. Lower abundance and diversityof suitable prey (aboveground resident groups) in the islands maydeter quickly moving hunting animals from staying long on theislands (Moretti et al., 2006).

Overall the comparison of above- and belowground abundancesof macro- and mesofauna in different treatments revealed com-pletely different reactions to burning and degrees of patch isolation.The recolonization of aboveground macrofauna is more deter-mined by habitat suitability than by isolation (Wool and Inbar,1998). This was also demonstrated by a balance between gainsand losses of aboveground taxa in both island and corridor plots.Belowground macrofauna had similar abundance in isolated andconnected patches suggesting that local survival was important.However the high gain of taxa belonging to the belowgroundmacrofauna in corridor plots in comparison with the loss of taxademonstrates the impact of colonization through corridors. The lossof taxa in the island treatment for this group exceeded the gain oftaxa, which indicated restrictions in recolonization.

Both the island and corridor plots/treatments were character-ized by massive species loss of aboveground oribatid mites. Herehowever the reason seems to be different. Despite better resistanceof aboveground oribatids against desiccation (Krivolutsky, 1995)they are not able to penetrate and recolonize partially burnt areasdue to habitat destruction. This is supported by recent findings ofLehmitz et al. (2012) who evaluated the speed of active substratecolonization by different oribatid species and the known waysof passive oribatid distribution (Krivolutsky et al., 2004; Lehmitzet al., 2011). During the fire event, belowground-dwelling oribatidshave more chances to survive in the partially burned soil patchesbut surface-living oribatids cannot quickly leave upper soil hori-zons, where they often occur in high numbers in boreal forest soils(Krivolutsky, 1995). Hence recolonization of the unburnt islandsfrom the outside would also be important for them. This is substan-tiated as well by their relatively smaller species loss rates than forthe aboveground oribatids. Presence of corridors to some extentincreases recolonization rate of the burnt area by belowground-dwelling oribatid mites but does not allow active recovery of thespecies composition.

It seems that corridors are not as important in recovering abun-dance of aboveground macrofauna as niche differentiation andmicrohabitats suitability in the partially burnt organic layer of soilin the islands (Starzomski and Srivastava, 2007). The same is appli-cable for the recovery of the taxonomic richness of this group whichis successful without respect to the presence of corridors. Corridorsare important for the recovery of the species richness of below-ground macrofauna but not overall abundance which is probablylimited by the food availability. Aboveground oribatids do not bene-fit from the corridors neither for the recovery of species richness norfor abundance. Especially for this group there was no relationshipbetween abundance and species richness patterns, which suggeststhat recovery patterns are rather induced by niche availabilityincrease than colonization possibility (Setälä et al., 2005). Existingsoil and litter corridors connecting less burnt patches inside theburnt forests with the unburnt edges are important especially forthe recovery of the abundance of belowground oribatid mites butnot of species richness. The loss of belowground oribatid speciesrichness is rather determined by habitat availability. Belowgroundanimals have a strategy of local survival in the less-burnt patches,and also for belowground mites, corridors are extremely importantin abundance recovery (Gandhi et al., 2001).

We conclude that aboveground and mobile macro- and meso-fauna groups of soil animals are much more vulnerable to forest

fires than belowground taxa. On the other hand, the former twogroups have more possibilities for recovery of their abundances butnot of taxonomic richness which takes a much longer time (Zaitsevet al., 2006). It will be essential in the future to evaluate the effect

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f isolation on the overall soil food-web diversity, performance andunctioning and elaborate strategies on mitigating negative effectsf forest fires (Wardle et al., 2003).

cknowledgements

We thank Dr. Anna Malmström (SLU) for her help during samp-ing and fruitful discussions. The authors are grateful to Dr. Klemenskschmitt (Justus-Liebig-University, Giessen, Germany) for somealuable suggestions during the manuscript preparation. Funds forhe investigation were received from the Swedish Research Coun-il Formas and the Swedish Institute, the Russian Foundation forasic Research grant no. 11-04-00245-ɑ and “Living Nature: Cur-ent State and Prospects of Development” Programme (Zhivayariroda: Sovremennoe Sostoyanie i Problemy Razvitia) operated byhe Russian Academy of Sciences.

eferences

ström, J., Bengtsson, J., 2011. Patch size matters more than dispersal distance in amainland–island meta community. Oecologia 167, 747–757.

alogh, J., Balogh, P., 1992a. The Oribatid Mites Genera of the World, 1. NaturalHistory Museum, Budapest.

alogh, J., Balogh, P., 1992b. The Oribatid Mites Genera of the World, 2. NaturalHistory Museum, Budapest.

eck, P.S.A., Goetz, S.J., Mack, M.C., Alexander, H.D., Jin, Y., Randerson, J.T., Loranty,M.M., 2011. The impacts and implications of an intensifying fire regimeon Alaskan boreal forest composition and albedo. Glob. Change Biol. 17,2853–2866.

erg, M., de Ruiter, P., Didden, W., Janssen, M., Schouten, T., Verhoef, H., 2001. Com-munity food web, decomposition and nitrogen mineralisation in a stratifiedScots pine forest soil. Oikos 94, 130–142.

erthet, P.L., 1964. Field study of the mobility of Oribatei (Acari), using radioactivetagging. J. Anim. Ecol. 33, 443–449.

ormann, B.T., Homann, P.S., Darbyshire, R.L., Morrissette, B.A., 2008. Intense forestwildfire sharply reduces mineral soil C and N: the first direct evidence. Can. J.For. Res. 38, 2771–2783.

airney, J.W.G., Bastias, B.A., 2007. Influence of fire on forest soil fungal communities.Can. J. For. Res. 37, 207–215.

ertini, G., 2005. Effects of fire on properties of forest soils: a review. Oecologia 143,1–10.

e Chantal, M., Granström, A., 2007. Aggregations of dead wood after wildfire actas browsing refugia for seedlings of Populus tremula and Salix caprea. For. Ecol.Manag. 250, 3–8.

indal, D.L. (Ed.), 1990. Soil Biology Guide. John Wiley & Sons, New York.auria, M.M., Johnson, E.A., 2008. Climate and wild fires in the North American boreal

forest. Philos. Trans. R. Soc. B 363, 2317–2329.ritze, H., Pennanen, T., Pietikäinen, J., 1993. Recovery of soil microbial biomass and

activity from prescribed burning. Can. J. For. Res. 23, 1286–1290.andhi, K.J.K., Spence, J.R., Langor, D.W., Morgantini, L.E., 2001. Fire residuals as

habitat reserves for epigaeic beetles (Coleoptera: Carabidae and Staphylinidae).Biol. Conserv. 102, 131–141.

hilarov, M.S., 1965. Zoological Methods in Soil Diagnostics. Nauka, Moscow (inRussian, English summary).

hilarov, M.S., Krivolutsky, D.A. (Eds.), 1975. Keys for Soil-Inhabiting Mites. Nauka,Moscow (in Russian).

ongalsky, K.B., Malmström, A., Zaitsev, A.S., Shakhab, S.V., Bengtsson, J., Persson, T.,2012. Do burned areas recover from inside? An experiment with soil fauna in aheterogeneous landscape. Appl. Soil Ecol. 59, 73–86.

ongalsky, K.B., Persson, T., 2013. Recovery of soil macrofauna after wildfires inboreal forests. Soil Biol. Biochem. 57, 182–191.

olliday, N.J., 1991. Species responses of carabid beetles (Coleoptera: Carabidae)during post-fire regeneration on boreal forest. Can. Entomol. 123, 1369–1389.

uhta, V., 1971. Succession in the spider communities of the forest floor after clear-cutting and prescribed burning. Ann. Zool. Fenn. 8, 483–542.

onsson, M., Yeates, G.W., Wardle, D.A., 2009. Patterns of invertebrate density andtaxonomic richness across gradients of area, isolation, and vegetation diversityin a lake-island system. Ecography 32, 963–972.

ology 83 (2014) 101–108

Klausnitzer, B., 1994. Die Larven der Käfer Mitteleuropas. 2 Band. G. Fischer Verlag,Jena.

Klausnitzer, B., 1996. Die Larven der Käfer Mitteleuropas. 3 Band. G. Fischer Verlag,Jena.

Koehler, H.H., 1997. Mesostigmata (Gamasina Uropodina), efficient predators inagroecosystems. Agric. Ecosyst. Environ. 62, 105–117.

Kounda-Kiki, C., Celini, L., Ponge, J.F., Mora, P., Sarthou, C., 2009. Nested variation ofsoil arthropod communities in isolated patches of vegetation on a rocky outcrop.Soil Biol. Biochem. 41, 323–329.

Krantz, G.W., Walter, D.E. (Eds.), 2009. A Manual of Acarology. , 3rd ed. Texas TechUniversity Press, Lubbock, Texas.

Krivolutsky, D.A. (Ed.), 1995. Oribatid Mites. Nauka, Moscow (in Russian).Krivolutsky, D.A., Lebedeva, N.V., Gavrilo, M.V., 2004. Soil microarthropods in the

feathers of Antarctic birds. Dokl. Biol. Sci. 397, 342–345.Lehmitz, R., Russell, D., Hohberg, K., Christian, A., Xylander, W.E.R., 2011. Wind

dispersal of oribatid mites as a mode of migration. Pedobiologia 54, 201–207.Lehmitz, R., Russell, D., Hohberg, K., Christian, A., Xylander, W.E.R., 2012. Active

dispersal of oribatid mites into young soils. Appl. Soil Ecol. 55, 10–19.Malmström, A., Persson, T., Ahlström, K., 2008. Effects of fire intensity on survival

and recovery of soil microarthropods after a clear-cut burning. Can. J. For. Res.38, 2465–2475.

Malmström, A., Persson, T., Ahlström, K., Gongalsky, K.B., Bengtsson, J., 2009. Dynam-ics of soil meso- and macrofauna during a 5-year period after clear-cut burningin a boreal forest. Appl. Soil Ecol. 43, 61–74.

Mamaev, B.M., 1972. Identification Keys for Insect Larvae. Prosveshenie, Moscow (inRussian).

Moretti, M., Duelli, P., Obrist, M.K., 2006. Biodiversity and resilience of arthro-pod communities after fire disturbance in temperate forests. Oecologia 149,312–327.

Niklasson, M., Granström, A., 2000. Numbers and sizes of fires: long-term spatiallyexplicit fire history in a Swedish boreal landscape. Ecology 81, 1484–1499.

Ojala, R., Huhta, V., 2001. Dispersal of microarthropods in forest soil. Pedobiologia45, 443–450.

Persson, T., Karlsson, P.S., Seyferth, U., Sjöberg, R.M., Rudebeck, A., 2000. Carbon min-eralisation in European forest soils. In: Schulze, E.D. (Ed.), Carbon and NitrogenCycling in European Forest Ecosystems (Ecological Studies), 142. Springer-Verlag, pp. 257–275.

Pietikäinen, J., Fritze, H., 1995. Clear-cutting and prescribed burning in conifer-ous forests: comparison of effects on soil fungal and total microbial biomass,respiration activity and nitrification. Soil Biol. Biochem. 27, 101–109.

Rantalainen, M.-L., Fritze, H., Haimi, J., Pennanen, T., Setälä, H., 2005. Colonisation ofnewly established habitats by soil decomposer organisms: the effect of habitatcorridors in relation to colonisation distance and habitat size. Appl. Soil Ecol. 28,67–77.

Rantalainen, M.-L., Haimi, J., Fritze, H., Pennanen, T., Setälä, H., 2008. Soil decomposercommunity as a model system in studying the effects of habitat fragmentationand habitat corridors. Soil Biol. Biochem. 40, 853–863.

Saint-Germain, M., Drapeau, P., Hébert, C., 2004. Comparison of Coleoptera assem-blages from a recently burnt and unburnt black spruce forests of northeasternNorth America. Biol. Conserv. 118, 583–592.

Schimmel, J., Granström, A., 1996. Fire severity and vegetation response in the borealSwedish forest. Ecology 77, 1436–1450.

Setälä, H., Berg, M.P., Jones, T.H., 2005. Trophic structure and functional redundancyin soil communities. In: Bardgett, R.D., Usher, M.B., Hopkins, D.W. (Eds.), Biolog-ical Diversity and Function in Soils. Cambridge University Press, Cambridge, pp.236–249.

Starzomski, B.M., Srivastava, D.S., 2007. Landscape geometry determines commu-nity response to disturbance. Oikos 116, 690–699.

Wang, X., Hamann, A., Cumming, S.G., 2012. Measuring boreal forest fragmentationafter fire: which configuration metrics are best? Ecol. Indic. 13, 189–195.

Wardle, D.A., Hornberg, G., Zackrisson, O., Brundin, M.K., Coomes, D.A., 2003. Long-term effects of wildfire on ecosystem properties across an island area gradient.Science 300, 972–975.

Weigmann, G., 2006. Hornmilben (Oribatida). In: Dahl, F. (Ed.), Tierwelt Deutsch-lands, 76. Goecke & Evers, Keltern.

Wool, D., Inbar, M., 1998. Colonization of ecological islands: galling aphid popula-tions (Sternorrhyncha: Aphidoidea: Pemphigidae) on recovering Pistacia treesafter destruction by fire. Eur. J. Entomol. 95, 41–53.

Zaitsev, A.S., Chauvat, M., Pflug, A., Wolters, V., 2002. Oribatid mite diversity and

community dynamics in a spruce chronosequence. Soil Biol. Biochem. 34,1919–1927.

Zaitsev, A.S., Wolters, V., Waldhardt, R., Dauber, J., 2006. Long-term succession oforibatid mites after conversion of croplands to grasslands. Appl. Soil Ecol. 34,230–239.