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Applied Soil Ecology 83 (2014) 159–169

Contents lists available at ScienceDirect

Applied Soil Ecology

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

arthworm invasion alters enchytraeid community composition andndividual biomass in northern hardwood forests of North America

irí Schlaghamerskya,∗, Nico Eisenhauerb, Lee E. Frelichc

Department of Botany and Zoology, Faculty of Science, Masaryk University, Kotlárská 2, 611 37 Brno, Czech RepublicInstitute of Ecology, Friedrich Schiller University Jena, Dornburger Str. 159, 07743 Jena, GermanyDepartment of Forest Resources, University of Minnesota, 1530 Cleveland Avenue N., St. Paul, MN 55108, USA

r t i c l e i n f o

rticle history:eceived 29 January 2013eceived in revised form 4 September 2013ccepted 11 September 2013vailable online 13 October 2013

eywords:arthworm invasionumbricidaenchytraeidaeorth Americaoilesofauna

a b s t r a c t

European earthworms are invading many ecosystems worldwide and fundamentally transform habitatsby acting as dominant ecosystem engineers. However, there is little knowledge of the consequences ofearthworm invasion on the composition and diversity of native soil organisms. Particularly functionallysimilar groups, such as enchytraeids (Annelida: Enchytraeidae), may be affected through changes in thechemical and physical properties of the soil, but also due to competition for resources. In 2010–2011, westudied the impact of earthworm invasion on enchytraeids at two sites in the northern hardwood forestsof North America: one site within the Chippewa National Forest in northern Minnesota and one site inthe Chequamegon-Nicolet National Forest in northern Wisconsin, USA. At each site, three plots weresampled along a transect, representing (1) a non-invaded or very slightly invaded area, (2) the leadingedge of earthworm invasion and (3) a heavily invaded area with an established population of the anecicearthworm Lumbricus terrestris (among other species). In total, 29 enchytraeid (morpho)species wereidentified (some yet to be formally described, several first or second records for the continent); of those24 occurred at the Minnesota site and 17 at the Wisconsin site. The structure of enchytraeid assemblagesdiffered significantly among the three invasion stages, although this was not equally pronounced at thetwo sites. Each stage was characterized by one or several indicator species. Mean enchytraeid densities(10,700–30,400 individuals/m2) did not differ significantly among the invasion stages, but were lowestat the leading edge of earthworm invasion at both sites. In the heavily invaded plot at the Minnesota site,the mean enchytraeid density and biomass in L. terrestris middens were significantly higher than in soilin-between the middens. This was due to a pronounced effect of L. terrestris middens in the uppermost

3 cm of soil. Differences in biomass among earthworm invasion stages were most apparent for meanindividual biomass. This was significantly higher in the heavily invaded area than at the leading edgeor in the non-invaded area at the Minnesota site. Compositional changes of the enchytraeid assemblageare likely to result in changes in the functioning of soil foods webs. Our results suggest that earthworminvasions can cause a loss of native species in soil, including heretofore unknown ones, that might gounnoticed.

© 2013 Elsevier B.V. All rights reserved.

. Introduction

Earthworms (Crassiclitellata) and enchytraeids or potwormsEnchytraeidae) are the two dominant groups of annelid “worms”

Annelida) in soil. Although operating on different spatial scales,elonging to soil macrofauna and mesofauna, respectively, they areoth considered of high ecological importance. The much larger

∗ Corresponding author.E-mail addresses: jiris@sci.muni.cz,

chlaghamersky@mail.muni.cz (J. Schlaghamersky).

929-1393/$ – see front matter © 2013 Elsevier B.V. All rights reserved.ttp://dx.doi.org/10.1016/j.apsoil.2013.09.005

and more popular earthworms act as “ecosystem engineers”, hav-ing the capacity to transform terrestrial ecosystems substantially(Jones et al., 1994; Lavelle et al., 1997; Bohlen et al., 2004). At leastsince Darwin’s seminal work (Darwin, 1881), people have beenaware of their important role in soil formation and fertility. Sincethen, numerous studies have increased our understanding of theirbiology and function in terrestrial ecosystems. Much less is knownabout their small relatives, the enchytraeids. Together with spring-

tails (Collembola) and oribatid mites (Oribatida), they are one ofthe three most abundant groups of soil mesofauna (Coleman et al.,2004). In temperate climate soils of low pH, high humidity and withthick organic horizons they were found to be represented by few

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60 J. Schlaghamersky et al. / Appl

pecies (Schlaghamersky, 2002; Standen, 1980), however, reach-ng total mean annual densities up to 145,000 ind./m2 (Didden,993). Under such conditions the densities of many other groups,

ncluding earthworms and springtails, are rather low or in someases absent, and enchytraeids play a key role in decompositionnd nutrient cycling (Petersen and Luxton, 1982). For this reason,hey have also been the focus of studies looking at the fate of car-on storage in mires or tundra soils under global warming scenariose.g. Briones et al., 1998; Carrera et al., 2011).

Under cold or acidic conditions enchytraeids might be the mostffective group in mixing and aerating the soil, provided that con-itions do not get too dry (Petersen and Luxton, 1982; Swift et al.,998), a role otherwise ascribed mostly to earthworms (and inhe tropics also to termites and ants). Even in soils of higher pHith thriving earthworm populations and mull humus, however,

nchytraeids compose a considerable portion of soil fauna; theirssemblages are species-rich and mean annual densities rangerom several thousands to tens of thousands of individuals perquare meter (Didden, 1993; Schlaghamersky, 1998). Despite anncreasing number of studies on potential interactions betweenarthworms and enchytraeids, this topic requires further investi-ation. According to Beylich and Graefe (2012) the existence andutcome of such interactions depend on the combination of speciesnd site conditions, being often overridden by other environmen-al factors under field conditions. Most observations, often basedn mesocosm experiments, indicate detrimental effects of earth-orms on enchytraeids (Górny, 1984; Huhta and Viberg, 1999;

agerlöf and Lofs-Holmin, 1987; Makulec and Pilipiuk, 2000; Räty,004; Räty and Huhta, 2003; Schlaghamersky, 1998; Tao et al.,011), while a few studies have shown detrimental effects ofnchytraeids on earthworms (Haukka, 1987; Sandor and Schrader,012). Positive effects have also been shown: middens and gal-

eries of anecic earthworms host enchytraeid aggregations due tohe accumulation of organic matter (Dózsa-Farkas, 1978; Schradernd Seibel, 2001), and earthworm presence in experimental ves-els can increase enchytraeid abundance (Sandor and Schrader,012).

Large areas of North America had no earthworms prior to Euro-ean colonization at least since the last ice age (Addison, 2009;ohlen et al., 2004; Gates, 1929, 1976a,b; Reynolds, 1977; Reynoldst al., 2002; Tiunov et al., 2006; but compare Schwert, 1979). Sincearthworms (mostly European species of the family Lumbricidae)ere introduced both accidentally (ship ballast, gardening, farm-

ng) and deliberately (fish bait, composting) to North America, theyave been spreading across the temperate hardwood forest regionBohlen et al., 2004; Tiunov et al., 2006). Although the presencef European earthworms in previously earthworm-free forests inhe Midwest and Northeast USA was noted and their effects on soiltructure have been under study for almost 50 years (Langmaid,964; Gates, 1972; Alban and Berry, 1994), only recently haveetailed studies of earthworm impacts on ecosystem processes andative biota been published (e.g. Bohlen et al., 2004; Hale et al.,005; Holdsworth et al., 2007a). Especially in areas with suitablelimate and soils devoid of native earthworms, exotic earthwormsause substantial changes in the soil environment with conse-uences for entire ecosystems (Bohlen et al., 2004; Hendrix et al.,008; Langmaid, 1964; Nielsen and Hole, 1964). Some of the speciesere shown to remove the entire organic soil layer (mixing it into

he mineral soil) within a short time (Hale et al., 2005; Eisenhauert al., 2007), with detrimental impacts on reproduction of plantpecies dependent on these layers (Gundale, 2002; Hale et al., 2006;oldsworth et al., 2007b). When dominant tree species, such as

ugar maple, are affected, this might lead to fundamental changesn the forest stand composition (Frelich et al., 2012). Such changesan facilitate the invasion of other exotic plant species (Nuzzo et al.,009). Furthermore, invasive earthworms alter nutrient cycling and

il Ecology 83 (2014) 159–169

availability as well as soil microbial communities (Hale et al., 2008;Eisenhauer et al., 2011). For the sake of simplicity, the foregoingcompilation of invasive earthworm effects focused on earthwormsin general; however, individual species or at least their groupingsbased on major ecological traits, are known to differ substantiallyin their effects on other soil organisms (Eisenhauer, 2010).

Very few studies have looked at the impact of earthworm inva-sions on other soil fauna (Burke et al., 2011; Cameron et al., 2013;Eisenhauer et al., 2007; McLean and Parkinson, 1998; Snyder et al.,2009; Straube et al., 2009). As pointed out in two review paperson this topic, not a single one looked at the effects on enchy-traeids in North America (Migge-Kleian et al., 2006; Parkinsonet al., 2004). The invasion of exotic earthworms into completelyearthworm-free soils, as found in the Great Lakes Area of NorthAmerica, provides a unique opportunity to find out more aboutthe effect of earthworm-induced changes on other soil fauna ingeneral and on enchytraeids in particular. It is also an opportunityto study the structure of enchytraeid assemblages in soils with athick organic layer, as present in non-invaded forest areas, and howthis changes with the arrival and establishment of earthworms.Where both groups have evolved together (Europe), soils with athick duff layer unmixed with mineral soil (mor or raw humus)are only found under acidic conditions, whereas earthworm-freeforests of North America have similar humus also on soils of mod-erately acid to neutral pH, allowing us to uncouple effects of soilacidity and earthworms. The only comparable data on interactionsof earthworms and enchytraeids under similar conditions are fromearthworm-free deciduous (birch) forest stands interspersed in theboreal zone in Finland, mostly based on mesocosm experimentswith introduced earthworms (Huhta and Viberg, 1999; Räty andHuhta, 2003; Räty, 2004). Earthworms were shown to increase soilpH and to reduce enchytraeid populations in lab experiments, withone accidentally introduced enchytraeid species thriving in Lum-bricus terrestris middens.

Similar to earthworms, enchytraeids are not a homogenousgroup. For instance, different species prefer soils of differing acid-ity or moisture range, they differ in their tolerance to disturbancesor stress such as drought, and they differ in their preference formicrohabitats, such as different layers of the upper soil (Graefeand Schmelz, 1999). Terrestrial enchytraeids have been very lit-tle studied in North America and our knowledge of their diversityand distribution is poor (Schlaghamersky, 2013a). Combined withthe generally difficult enchytraeid taxonomy, this makes the studyof effects on enchytraeid assemblages a challenging task. However,without looking at effects on the species level, many importantresponses might go unnoticed, including changes in enchytraeidcommunity composition and potentially significant loss in biodi-versity.

Therefore, we studied two sites in northern hardwood forests ofMinnesota and Wisconsin (USA) where an on-going invasion by asuite of European earthworm species (including epigeic, endogeicand anecic species) had been observed and well documented in pre-vious years (Hale et al., 2005; Holdsworth et al., 2007a), comparingenchytraeid density, biomass and community structure along twoearthworm invasion gradients. In soils with well-separated organichorizons, enchytraeids are expected to concentrate in these layersoverlying mineral soil. Would the activity of earthworms, workingthe organic material into the mineral soil, reduce enchytraeid num-bers and biomass, or would the enchytraeids just shift into deeperhorizons now enriched with organic material? And if so, wouldthis be accompanied by a shift in community structure (speciescomposition and dominance)? Would the species pool available

provide species able to thrive under the changed conditions? Whatwould happen with specialized dwellers of the organic horizons?Would there be a new equilibrium in areas with well-establishedearthworm populations?

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ig. 1. Location of the two study sites: Minnesota site (M) – Ottertail Peninsula on Lhe Rainbow Lake Wilderness, Chequemegon section of Chequamegon-Nicolet Nati

. Materials and methods

.1. Study sites

The two sites were located 250 km apart in northern MinnesotaChippewa National Forest) and Wisconsin (Chequamegon sectionf Chequamegon-Nicolet National Forest), USA, a region devoid ofative earthworms (Fig. 1). The site within the Chippewa Nationalorest is situated on the Ottertail Peninsula at Leech Lake, beingdentical to the “Section 19” site in Hale et al. (2005), geographicoordinates: 47◦16′0.00′′ N, 94◦23′48.60′′ W (corresponding to theurrent position of the leading edge of earthworm invasion alonghe studied transect with a length of 390 m at 443–449 m a.s.l.). Theite within the Chequamegon-Nicolet National Forest is situated atower Lake in the Rainbow Lake Wilderness north of the town ofrummond, geographic coordinates: 46◦26′3.06′′ N, 91◦19′36.00′′

(position of leading edge as above; length of transect 1700 mrom 380 m to 405 m a.s.l.). Both study sites were covered with

esic forests approximately 80–100 years old after logging in thearly 1900s. At the Minnesota site, sugar maple (Acer saccharum)as dominant, with yellow birch (Betula alleghaniensis) and bass-ood (Tilia americana) as secondary species in the tree layer. Soilas a deep, well-drained and light-colored silty clay loam Eutrob-

ralf (Warba series) associated with the Guthrie Till Plain (USDA,997). It had a mean soil pH (H2O) of 6.2 and silt loam texture withean percentages of 31%, 9% and 60% sand, silt and clay, respec-

ively. The climate is humid, continental, and cold temperate with

mean annual temperature of 3.9 ◦C and mean annual precipita-ion of 672 mm; mean monthly temperatures range from −5 ◦C inanuary to 19.8 ◦C in July (30-year averages for 1971–2000; PRISMlimate Group; see also Daly et al., 2008).

ake, Chippewa National Forest; Wisconsin site (W) – vicinity of Tower Lake withinorest.

At the Wisconsin site, sugar maple was also dominant, followedby aspen (Populus tremuloides and P. grandidentata), basswood, redmaple (Acer rubrum) and ash (Fraxinus spp.). The mean soil pH (H2O)was 4.4 and soil texture was loam to silt loam with mean percent-ages of 39%, 48% and 13% sand, silt and clay, respectively. The soils inthe wider area are Fragiorthods and Haplorthods (Albert, 1995), butno information was available for the exact study area. The climateis milder than at the above site, with a mean annual temperatureof 5.2 ◦C and mean annual precipitation of 843 mm; mean monthlytemperatures range from −0.3 ◦C in January to 25.4 ◦C in July (30-year averages for 1971–2000; source as above).

The two study sites were used in previously published studiesof European earthworm invasion impacts on native plant com-munity structure, and were chosen to include the leading edgeof earthworm invasion (Hale et al., 2005, 2006; Holdsworth et al.,2007a,b). At the Minnesota site three parallel transects 10 m apartand 150 m in length, had been placed perpendicular to the lead-ing edge of earthworm invasion in 1999 (Hale et al., 2005). For thepresent study, one of these transects was used. As the visible lead-ing edge of earthworm invasion had further advanced, the transectwas extended to a total length of 390 m to reach ground with a stateof the litter and vegetation indicating the absence of earthworms(Eisenhauer et al., 2011). At the Wisconsin site one transect of 400 mlength had been placed in 2001 (Holdsworth et al., 2007a). As thestate of the litter layer indicated that at least the epigeic earthwormspecies Dendrobaena octaedra had penetrated substantially fartherinto the forest stand, the transect was extended to a total length of

1700 m for the present study to reach uninvaded ground. Accordingto Hale et al. (2005) and Holdsworth et al. (2007a), areas with thickorganic horizons and a light brown A horizon were characterized byno earthworms or a small biomass (ca. 1 g m−2 ash-free dry mass,

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FDM) of D. octaedra, whereas the middle areas of the transects,ear the visible front of the leading edge of invasion, with thinrganic horizons, had a moderate biomass (ca. 1–3 g m−2 AFDM)f Lumbricus rubellus, Aporrectodea spp., and Octolasion tyrtaeum,hile heavily invaded areas (with organic horizons absent at mid-

ummer) and black A horizons were characterized by the presencef high earthworm biomass (ca. 3–5 g m−2 AFDM) of L. terrestris,. rubellus and Aporrectodea spp. The latter have been invaded ateast ten years previously. At the Minnesota site the heavily invadedreas were characterized by a high density of L. terrestris middensith rather regular distances among them (analysis of photographs

howed that about 45% of the soil surface was covered by mid-ens), with almost bare soil between them (covered by fresh leaf

itter in fall). At the Wisconsin site middens were much scarcer inhe well invaded areas and hardly discernible among leaf litter andrass/sedge ground-layer vegetation.

.2. Sampling and sample processing

The sites were sampled in autumn 2010 and again in spring toarly summer 2011 (Minnesota site: Sept. 24, 2010, June 2, 2011;isconsin site: Oct. 3, 2010, July 4, 2011). Along the transects, three

lots (10 m × 4 m oriented perpendicular to the transect) were sam-led, representing three different situations: non-invaded (or with

low density of the epigeic earthworm D. octaedra only), lead-ng edge (with established populations of epigeic and endogeicarthworm species) and invaded (with established populations ofll three ecological groups of earthworms, including adults of thenecic L. terrestris). These sampling plots were situated at 0 m,45 m and 385 m of the transect at the Minnesota site and at 0 m,00 m and 1700 m of the transect at the Wisconsin site. The struc-ure of the upper soil horizons and earthworm presence in the soilores confirmed our expectations concerning the stage of earth-orm invasion. Soil cores of 4.8 cm in diameter were taken down

o a depth of 12 cm (rarely 9 or 15 cm, depending on the presencef stones and roots in the soil; including the litter layer) at randomositions within these three plots along the transects. Stratifiedandom sampling was applied in the “invaded” plot at the Min-esota site, where the density of L. terrestris middens was very highsee above): one half of the soil cores was taken from these middensnd one half from the almost bare ground in-between middens atoth sampling dates. In 2010 six soil cores from each position of theransect were extracted (3 × 6 = 18 soil cores per site). In 2011 fouroil cores per transect position (3 × 4 = 12 soil cores in total) werextracted from the Minnesota site and six soil cores per transectosition (3 × 6 = 18 soil cores in total) from the Wisconsin site. Alloil cores were subdivided into 3 cm layers and stored separately.

Microannelids were extracted from these 3-cm layers by aodified O’Connor wet funnel extraction (24 h without heat-

ng with subsequent heating of the soil surface up to 44 ◦Cithin 4 h) as developed for the quantitative extraction of enchy-

raeids (for a comparison of extraction efficiency of the original’Connor method and some modifications see Kobeticová andchlaghamersky, 2003). Prior to extraction the soil was kept cooluring transport and stored in a refrigerator at ca. 4 ◦C. Extractednchytraeids were kept in water-filled Petri dishes at the same tem-erature. Water used in the extraction and keeping of enchytraeidsas adjusted to the mean soil pH of the corresponding site. Tem-erature control and pH adjustments were adopted to minimizenchytraeid mortality. Each extraction included one soil core (sub-ivided into 3-cm layers) from each of the three transect positionsrom a given site (to rule out that potential effects of storage time

ay influence effects of transect positions). The extraction fromnother three soil cores was started when it could be expected thathe identification of the preceding batch would be almost com-leted by the time new extracted specimens were available. In our

il Ecology 83 (2014) 159–169

experience (confirmed by Didden (2003) for individual species)enchytraeids might have rather high mortality, increasing overtime, when kept in water, whereas they survive well in soil keptat low temperatures. Due to the time required for live identifica-tion of samples, storage time of the soil samples prior to extractionvaried between one and a maximum of 80 days, whereas storagetime of extracted enchytraeids prior to identification did not exceedfive days. Abrahamsen (1972) did not find any significant effect ofstorage of soil samples for up to three months. Didden (2003) wasonly able to show an effect on enchytraeid abundance significantat the 10% level for storage prior to extraction up to 168 days (andnone for longer storage times). However, he warned that shifts inspecies composition might be more substantial than in abundanceand thus recommended to process field samples within 6 weeks.

Enchytraeids and other microannelids were identified under alight microscope, partially using Nomarski differential contrast, tospecies or – in cases of juveniles, injured or dead specimens – togenus (whenever possible). The approximate body length of allspecimens (stretched live individuals) was recorded. Based on thismeasurement, the species identity (each species or genus havinga characteristic length: volume ratio) and an average density fac-tor, the body mass (fresh weight) of each specimen was calculatedaccording to Abrahamsen (1973), resulting in individual massesthat were added up to total biomass values for soil core × soil layercombinations and entire soil cores. As biomass data are more oftengiven as dry weight, fresh weights were also transformed to dryweights by division by 5.56 (Edwards, 1967).

2.3. Statistical analysis

Arithmetic means, standard deviation and standard error werecomputed for density and biomass data. However, all statisti-cal testing was done under consideration of limits given by thenon-normal or unknown (due to small sample size) distribution.Data were log10-transformed to meet the assumptions of paramet-ric tests (unpaired t-test; Analysis of Variance with subsequentTukey–Kramer Multiple Comparisons Test) and if not achievedwe applied non-parametric alternatives (Mann–Whitney U test;Kruskal–Wallis Test with subsequent Dunn’s Multiple ComparisonTest).

Potential differences in enchytraeid abundance and biomassamong plots (earthworm invasion stages) were tested using Gen-eral Linear Modeling (GLM) with data for individual soil layers(3 cm thick, four layers down to 12 cm depth) considered asrepeated measures (Eisenhauer et al., 2007). Individual soil cores(spaced > 1 m) from a given plot were considered replicates, inde-pendently of the sampling date (but the sampling date was includedas one factor in the analyses). Repeated measures analysis wasnot considered appropriate because spatial variability in soil andenchytraeid populations are high – even at distances of 5–10 cm(e.g. Kobeticová and Schlaghamersky, 2003) – and to avoid anypotential effects caused by disturbance associated with taking thefirst set of cores during fall 2010, the second set of soil cores takenduring spring-summer 2011 was not spatially paired with the firstset.

In order to investigate earthworm invasion effects on the com-position of the enchytraeid assemblage (including the only othermicroannelid, Parergodrilus heideri), we used two complimentarynon-parametric multivariate techniques: non-metric multidimen-sional scaling (NMDS) and multi-response permutation procedures(MRPP). NMDS, an unconstrained ordination method robust forecological data, is based on ranked distances among sample units

and uses an iterative search to depict these data in as few dimen-sions as possible while minimizing stress. Analyses were carriedout using the program PCORD version 5. In consideration of thelarge difference in the species composition at both sites, data from

ied Soil Ecology 83 (2014) 159–169 163

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Fig. 2. NMDS ordination plot for enchytraeid assemblages at the Minnesota site(Ottertail Peninsula at Leach Lake, Chippewa National Forest); the three invasionstages sampled are distinguished by different symbols, each point represents theassemblage within one soil core (extracted in fall 2010 or spring 2011); MRPP:

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he Minnesota and Wisconsin sites were analyzed separately. Thehree invasion stages represented the “treatments”. Species withess than three occurrences in soil cores taken at a given site

ere excluded, following standard practice for multivariate analy-es in which rare species do not have enough co-occurrences withther species to evaluate their contribution to community structurePeck, 2010). Maximum standardization by species was appliedeach species value was divided by the maximum among all soilores, expressing values among soil cores relative to the maximumumber attained by a given species, equalizing input of specieso the analysis). For the Minnesota site 23 species and 30 samp-ing units (soil cores) were used, for the Wisconsin site 13 speciesnd 36 sampling units. From the three-dimensional output, two-imensional plots showing the first two axes were plotted. MRPPas then run to test for pairwise differences in treatment group

omposition by comparing dissimilarities within and among treat-ent groups with random samples of soil cores. MRPP provides two

ieces of information, the agreement statistic ‘A’, which is a mea-ure of within group agreement, and a P-value. MRPP and NMDSperate on dissimilarity matrices and we utilized the asymmet-ical Sørensen distance measure. Indicator species analyses wereonducted using the method of Dufrene and Legendre (1997) forhe calculation of indicator values. The Monte Carlo test of signif-cance of observed maximum indicator value for species was run

ith 4999 permutations.

. Results

.1. Species composition

In total, 2218 enchytraeid specimens were extracted andxamined (1349 and 869 at the Minnesota and Wisconsin site,espectively) and 29 enchytraeid (morpho)species were identifiedsome yet to be formally described), including 24 at the Minnesotaite and 17 at the Wisconsin site (Table 1). At the Minnesota sitepecies richness increased from the non-invaded (16) to the heavilynvaded area (19), but the higher number of species was com-ined with a 1.5× higher number of identified specimens. At theisconsin site, species richness dropped from the non-invaded

nd leading edge area (both 15) to the heavily invaded area (12),lthough the number of identified specimens from the latter waslso higher (1.2× and 1.6×, respectively). Enchytronia parva, Frid-ricia cylindrica, Hemifridericia parva, and Stercutus niveus wereecorded for the first time in North America (cf. Coates et al.,008). Enchytraeus dichaetus was found a few months earlier inebraska prairie soil (Schlaghamersky, 2013b), this is thus its sec-nd record from North America. Hemifridericia bivesiculata wasound for the first time since its description from a single local-ty in the Canadian Arctic Archipelago; Henlea conchifera had beennown so far only from Siberia and several localities in the Cana-ian Arctic Archipelago (Christensen and Dózsa-Farkas, 2006). Atoth sites one other microannelid, P. heideri (Parergodrilidae), waslso found, presenting first records of this terrestrial polychaete fororth America (Schlaghamersky and Frelich, 2012). Four Frideri-

ia spp., three Oconnorella spp., one Marionina sp. and one Achaetap. could not be identified to species. Where material is sufficient,hese will be described as new species to science. These and furtheraxonomic and faunistic details will be presented elsewhere.

.2. Similarity of enchytraeid assemblages

The percentage representation of individual species changedubstantially along the transects (Table 1). At the Minnesota site,enlea perpusilla reached 29% in the uninvaded plot, but decreasedith increasing earthworm presence, taken over by Enchytraeus

chance-corrected within-group agreement – A = 0.095; probability of a smaller orequal delta – P < 0.001.

buchholzi s.l. (21%) in the leading edge plot and Buchholzia appendic-ulata in the heavily invaded plot (22%), where all Fridericia speciestogether exceeded 39%. At the Wisconsin site, Marionina sp. 1 wasmost abundant, presenting 52%, 27% and 19% of all enchytraeidsin the uninvaded, leading edge and heavily invaded plots, respec-tively. However, Cognettia glandulosa increased toward the heavilyinvaded plot (from 8% to 19% and 38%), where it overtook Mario-nina sp. 1. Also, E. parva reached high percentages, in particular inthe less invaded plots, whereas Fridericia spp. (mostly F. cylindrica)had a peak at the leading edge (17%). Mean values of percent-age representation (dominance) can be misleading as variabilityamong soil cores can be high. Nevertheless, the overall picture wascorroborated by the results of the ordination (NMDS) of enchy-traeid assemblage composition in soil cores. For the Minnesotasite it showed a rather good separation of the assemblages fromthe three invasion stages: soil cores from the non-invaded plotand from the heavily invaded plot were well separated, whereassoil cores from the leading-edge plot had an intermediate posi-tion with some overlap with the soil cores of the heavily invadedplot (Fig. 2). Separation of the three invasion stages was less appar-ent for the Wisconsin site, however, most of the soil cores fromthe heavily invaded plot formed a cluster separated from the otherinvasion stages (Fig. 3). Thus the structure of enchytraeid assem-blages differed significantly among the three earthworm invasionstages studied, although this was not equally pronounced at thetwo sites.

Changes in enchytraeid community composition in responseto earthworm invasion were further corroborated by indicatorspecies analysis. For the Minnesota site, it showed four indicatorspecies for the area not invaded by earthworms (Bryodrilus librus:marginally significant, P = 0.0501; Bryodrilus sp. 1: P = 0.0064, H.parva: P = 0.0028; and Henlea ventriculosa: P = 0.0116), one indica-tor for the front of invasion (C. glandulosa: P = 0.0004) and threefor the highly invaded area (B. appendiculata: P = 0.0004, Frideri-cia galba: P = 0.0004, Fridericia paroniana: P = 0.0152). Also Fridericiaspp. (mostly juvenile representatives of the genus; treated as a

“species” in the analysis), were an indicator of the highly invadedarea (P = 0.0498). For the Wisconsin site the indicator speciesanalysis yielded one indicator species for the non-invaded area

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159–169Table 1List of microannelids (Enchytraeidae and Parergodrilus heideri) with their mean individual body mass (based on measurements in the present study) and percentage representation in the sampling plots along the earthworminvasion gradient at both sites: not invaded (no earthworms or only low density of the epigeic Dendrobaena octaedra), leading edge of invasion (epigeic and endogeic species present: D. octaedra, Apporectodea spp., Lumbricusrubellus), highly invaded (suite of earthworm species including established Lumbricus terrestris population). Percentage representations (rounded to the first decimal number) within the assemblage of a given plot are based ontotal numbers of specimens (given in the last row) identified at least to genus; injured or dead specimens not included if genus not identified; juveniles could usually be assigned to genus but not to species. Species names inbold mark first records for North America (only P. heideri was already reported in a separate paper – Schlaghamersky and Frelich, 2012; Bryodrilus librus was hitherto reported from North America under the synonym B. parvus),underlined species are presumably endemic to North America.

Species Mean individualbiomass (�g)

Minnesota: Chippewa National Forest Wisconsin:Chequamegon-NicoletNational Forest

Notinvaded

Leadingedge

Highly invaded– Total

Highly invaded– Bare ground

Highly invaded– Midden

Notinvaded

Leadingedge

Highlyinvaded

Achaeta sp. 1 23 – – – – – 4.6 4.0 0.3Bryodrilus librus (Nielsen & Christensen, 1959) 173 1.7 – 0.4 1.5 – 0.7 – –Bryodrilus cf. diverticulatus Cernosvitov, 1929 395 1.3 – – – – 0.4 0.5 –Bryodrilus cf. ehlersi Ude, 1892 403 – – – – – 0.4 1.0 2.1Bryodrilus spp. 78 1.3 – – – – – – –Buchholzia appendiculata (Buchholz, 1862) 139 0.3 9.9 22.3 26.3 20.7 – – –Buchholzia spp. (139) – – 2.4 3.8 1.8 – – –Cognettia glandulosa (Michaelsen, 1888) 104 – 8.5 – – – 8.2 19.8 38.1Cognettia spp. (104) – – – – – 0.7 – 0.3Enchytraeus buchholzi s. l. Vejdovsky, 1879 38 8.1 20.8 12.1 15.8 10.6 0.4 1.5 5.1Enchytraeus dichaetus Schmelz & Collado, 2010 69 – 2.1 1.3 2.3 0.9 – – –Enchytraeus spp. (38) – – 0.2 – 0.3 – – –Enchytronia parva Nielsen & Christensen, 1959 20 – – – – – 23.2 19.3 8.0Fridericia bulboides Nielsen & Christensen, 1959 174 – 1.1 – – – – 0.5 0.3Fridericia cylindrica Springett, 1971 704 – – – – – – 12.4 –Fridericia galba (Hoffmeister, 1843) 2728 – – 10.4 4.5 12.8 – – –Fridericia paroniana Issel, 1904 158 – – 8.2 6.8 8.8 – – –Fridericia ratzeli (Eisen, 1872) 3147 – 0.4 0.9 – 1.2 – – –Fridericia sp. 1 (cf.agilis/agricola) 270 – – 0.2 0.8 – – – –Fridericia sp. 2 1410 0.3 0.7 0.4 – 0.6 – – –Fridericia sp. 3 422 – – – – – 0.3 – –Fridericia sp. 4 422 – – – – – 0.3 – –Fridericia spp. 274 5.7 6.3 19.3 16.5 20.4 1.8 5.0 11.6Hemifridericia bivesiculata Christensen & Dózsa-Farkas,

2006(11) 1.3 0.4 0.9 0.8 0.9 – – –

Hemifridericia parva Nielsen & Christensen, 1959 (11) 12.4 6.7 0.2 – 0.3 0.4 1.5 4.8Hemifridericia spp. 11 0.3 1.8 0.2 0.8 – – 4.0 0.6Henlea conchifera Christensen & Dózsa-Farkas, 1999 796 2.0 1.1 0.4 0.8 0.3 – – –Henlea nasuta (Eisen, 1878) 552 – – 1.9 – 2.7 – – –Henlea perpusilla Friend, 1911 130 29.2 13.0 6.3 5.3 6.7 0.4 1.0 2.7Henlea ventriculosa (Udekem, 1854) 267 3.0 1.1 – – – 0.7 1.0 –Henlea spp. 294 2.7 0.4 0.6 – 0.9 0.4 – –Marionina sp. 1 25 5.7 11.3 5.2 6.8 4.6 51.8 26.7 18.8Oconnorella sp. 1 83 17.1 3.2 2.6 2.3 3.3 – – –Oconnorella sp. 2 130 5.7 2.5 3.0 4.5 1.8 – – –Oconnorella sp. 3 190 1.0 – 0.2 0.8 – 0.4 1.0 2.7Oconnorella spp. 74 – – – – – – – 1.9Stercutus niveus Michaelsen, 1888 50 1.7 1.1 0.2 – 0.3 5.7 0.5 2.7Parergodrilus heideri Reisinger, 1925 4 0.3 0.7 – – – – 0.5 0.3

Total number of identified specimens 298 284 462 133 329 282 207 336Total number of (morpho)species 16 17 19 14 16 15 15 12Mean density ± SE (103 ind./m2) 25.0 ± 4.6 19.2 ± 3.5 30.4 ± 7.6 15.5 ± 5.3 45.3 ± 11.1 13.3 ± 4.7 10.7 ± 3.5 16.0 ± 6.3

J. Schlaghamersky et al. / Applied Soil Ecology 83 (2014) 159–169 165

Fig. 3. NMDS ordination plot for enchytraeid assemblages at the Wisconsin site(near Tower Lake, Chequamegon section of Chequamegon-Nicolet National Forest);the three invasion stages sampled are distinguished by different symbols, each pointr2a

(ip

3

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fcc

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Fig. 4. Mean enchytraeid abundance per soil core (±standard error of the mean) inthe highly invaded plot in the Chippewa NF (Minnesota) inside and outside Lum-bricus terrestris middens (sampling in autumn 2010 and spring 2011) in four soillayers (0–3 cm, 3–6 cm, 6–9 cm, and 9–12 cm); Repeated Measures Analysis of Vari-ance, effective hypothesis decomposition; Soil depth × L. terrestris middens: F = 8.21,P = 0.006.

epresents the assemblage within one soil core (extracted in fall 2010 or spring011); MRPP: chance-corrected within-group agreement – A = 0.046; probability of

smaller or equal delta – P < 0.001.

Marionina sp. 1: P = 0.0118), none for the invasion front, and twondicators for the heavily invaded area (E. buchholzi: P = 0.0002, H.erpusilla: P = 0.0194).

.3. Density, biomass and vertical distribution

Mean enchytraeid densities per plot (both samplings pooled,ounded off to whole hundreds) ranged from 19,200 ± 3500 to0,400 ± 7600 individuals/m2 (±SE) at the Minnesota site and from0,700 ± 3500 to 16,000 ± 6300 individuals/m2 (±SE) at the Wis-onsin site (Table 1). At both sites the plot representing the leadingdge of earthworm invasion had the lowest enchytraeid density,hereas the heavily invaded plot had a somewhat higher density

han the (almost) non-invaded plot. However, none of these appar-nt differences was statistically significant (Kruskal–Wallis Test: all

> 0.05).At the Minnesota site 76–86% of all specimens were extracted

rom the upper 6 cm of the soil profile, at the Wisconsin site theorresponding value was 69–74% (range of means based on all soilores taken per plot; data not shown).

In the heavily invaded plot at the Minnesota site, the meannchytraeid density in L. terrestris middens and in soil in-etween the middens was 45,300 ± 11,100 and 15,500 ± 5300

ndividuals/m2 (±SE), respectively. The difference was significantsing the unpaired t-test (P = 0.041), which was justified by theesults of the Kolmogorov–Smirnov test for normal distribution andhe F-test for equality of variance. The GLM approach revealed thathis result was due to a pronounced effect of L. terrestris middensn the uppermost soil layer (0–3 cm), while enchytraeid densitiesid not differ significantly in the lower layers (midden presence xoil layer interaction, P = 0.006; Fig. 4).

The mean enchytraeid biomass in L. terrestris middens and inoil between middens was 27.16 ± 8.14 g and 4.87 ± 2.81 g fw/m2

±SE), respectively (or 4.89 ± 1.46 and 0.88 ± 0.50 g dw/m2). Theifference in biomass was significant with the Mann–Whitney Uest (P = 0.016; data did not meet the assumptions of paramet-

ic tests). Mean enchytraeid biomass (based on both samplings)er plot ranged from 2.31 ± 0.57 to 16.02 ± 5.50 g fw/m2 (±SE)or 0.41 ± 0.10 to 2.88 ± 0.99 g dw/m2) at the Minnesota site androm 0.53 ± 0.12 to 1.55 ± 0.90 g fw/m2 (±SE) (or 0.10 ± 0.02 to

Fig. 5. Mean enchytraeid biomass (fresh weight; ±standard error of the mean) alongearthworm invasion gradients in two forests (see Table 1) based on samplings inautumn 2010 and spring/summer 2011; different letters mark statistically differentvalues, see text.

166 J. Schlaghamersky et al. / Applied So

Fig. 6. Mean individual biomass (±standard error of the mean) of enchytraeids alongeTs

0yunptppd(udbtPs

wlPiWtwt

4

gav

arthworm invasion gradients in two forests (for explanation of invasion stages seeable 1); sampling in autumn 2010 and spring/summer 2011; different letters marktatistically different values (post hoc test: Tukey’s HSD), see text.

.28 ± 0.16 g dw/m2) at the Wisconsin site (Fig. 5). The GLM anal-sis, taking into account sampling date and site effects (andsing data from individual soil layers as repeated measures), didot yield significant differences in enchytraeid biomass amonglots. However, for the Minnesota site, one-way ANOVA of log-ransformed biomass data yielded significant differences amonglots (P = 0.0353). The subsequent Tukey–Kramer Multiple Com-arisons Test identified the biomass in the leading edge plot toiffer significantly (P < 0.05) from that in the heavily invaded plotfor the ANOVA, data were log-transformed to avoid bias due tonequal variance; after transformation SDs were not significantlyifferent: Bartlett statistic = 5.304; P = 0.0705). The enchytraeidiomass in the heavily invaded plot was also significantly higherhan in the non-invaded plot in spring 2011 (Kruskal–Wallis Test:

= 0.0066; Dunn’s Multiple Comparison Test: P < 0.05; data nothown).

At the Minnesota site, mean enchytraeid individual biomassas significantly higher in the heavily invaded area than at the

eading edge or in the non-invaded area (Kruskal–Wallis Test: = 0.0031; post hoc Tukey’s HSD test: non-invaded vs highlynvaded – P = 0.0041; leading edge vs highly invaded – P = 0.0287).

e found no significant differences in mean individual biomass athe Wisconsin site, although the value for the non-invaded areaas substantially lower than those for the other two plots along

he studied transect (Fig. 6).

. Discussion

Our study showed a modest impact of earthworm invasion onross enchytraeid community properties. Total density was lessffected than biomass, in particular, when calculated as mean indi-idual biomass. At both sites a drop in enchytraeid densities at

il Ecology 83 (2014) 159–169

the leading edge was apparent but not statistically significant. Itis worth mentioning that a similar, but statistically significant dropwas observed at both sites for microbial biomass and basal respi-ration, with some recovery in the well invaded areas (Eisenhaueret al., 2011). Results on the invasion’s impact on enchytraeid speciesrichness were ambiguous. Its apparent increase in the heavilyinvaded area of the Minnesota site, however, might have been pri-marily an effect of the higher enchytraeid abundance connectedto the high density of middens. Working on the species levelallowed us, however, to show significant changes in the communitystructure, i.e. the representation of individual species forming theenchytraeid assemblages at the different stages of earthworm inva-sion. We identified indicator species for non-invaded and invadedsites (one also for the leading edge of earthworm invasion), indicat-ing significant interactions between earthworms and enchytraeids.Indicator species for the non-invaded areas were either typical ofsoils with a high organic content (Bryodilus spp.) or of very smallbody size (H. parva and Marionina sp. 1). At this stage it is difficultto assess the risk of some of these species (or others so far goingunnoticed in North American soils) being extirpated in earthworm-invaded forests, but the results of the present study show thatthis risk might be real. At the Wisconsin site, the representationof Marionina sp. 1, E. parva, and the smallest Henlea species, H. per-pusilla, decreased from the non-invaded to the highly invaded area,the same was observed for H. parva and the presumably nativeOconnorella sp. 1 at the Minnesota site (Table 1). In the case ofsmall-bodied species, however, one can hope for survival at lowdensities thanks to the preservation of suitable microsites withinthe invaded forests. C. glandulosa, indicating the leading edge at theMinnesota site, inhabits the organic horizons of water-logged soils(Graefe and Schmelz, 1999; Jänsch and Römbke, 2003) and mighthave profited from soil compaction and a still sufficient duff layerat the leading edge of earthworm invasion. At the Wisconsin siteits representation increased toward the highly invaded area.

Enchytraeid species reaching larger body size became moredominant with increasing earthworm abundance and this alsobecame manifest in the above-mentioned enchytraeid biomassdata. Increasing individual biomass might indicate functionalchanges in response to earthworm invasion. Larger enchytraeids,such as many representatives of the genera Fridericia and Henlea,should be better able to move through compacted mineral soil.Furthermore, L. terrestris, in its capacity as an ecosystem engineer,created burrows with middens at their openings that presented thepreferred habitat of larger-bodied enchytraeid species, in particu-lar representatives of the genus Fridericia (Table 1). This is in goodagreement with the observations of Dózsa-Farkas (1978) from dryhornbeam-oak forests in Hungary.

The effect observed by us at the Minnesota site when compar-ing the three invasion stages might be somewhat biased by thestratified random sampling of earthworm middens and surround-ing, almost bare soil, with equal representation of both habitatsin the heavily invaded area. We believe completely random samp-ling would have led to higher bias, given the limited number ofsoil cores that could be processed. The trends found at the Wis-consin site, where middens were much rarer and not specificallysampled, were overall similar, although the two forests and theirenchytraeid assemblages were different. Here, E. buchholzi and H.perpusilla were indicators of the invaded area – these are rathersmall, ubiquitous and stress-tolerant species but avoiding stronglyacid soils and often found in mineral soils, including arable ones(Graefe and Schmelz, 1999; Jänsch and Römbke, 2003). The gravi-metric water content was lowest at the leading edge of earthworm

invasion (Eisenhauer et al., 2011). Decreasing soil water contentshould expose small-bodied enchytraeid species to a higher riskof desiccation than large ones. Soil water content is also associ-ated with soil density and pore size distribution and large-bodied

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J. Schlaghamersky et al. / Appl

pecies should be better able to dig through soil compacted byarthworm activity than small ones. Assumed changes in the com-osition of the microbial community (Eisenhauer et al., 2011) mightave affected the assemblages of enchytraeids due to specific feed-

ng preferences. Loss of the duff layer, being the preferred habitat ofany enchytraeid species, was probably the most important single

actor leading to the observed changes in enchytraeid communitytructure, particularly the decrease of species of small body size.herefore the middens, representing the only “duff” remaining onhe soil surface, were hotspots of enchytraeid abundance. Soil com-action and direct disturbance by earthworm movements mightave selected for species able to withstand this type of stress, eitherhanks to larger size or other traits leading to high stress tolerance,uch as in the case of E. buchholzi, regarded as an r-strategist ofigh tolerance to various types of stress or disturbance, includingrought (Graefe and Schmelz, 1999; Jänsch and Römbke, 2003). Aotential increase of soil pH due to earthworm presence (see for

nstance Räty, 2004) would probably not have led to the observedhanges, at least not at the Minnesota site, where the initial pH ofon-invaded soil was not very low.

We do not present data on the vertical distribution of enchy-raeids in detail, because these are less informative than expected:here a thick organic layer occurred, most of the sampled soilrofile corresponded to this layer, whereas in the heavily invadedreas almost the entire soil core was mineral soil. A direct com-arison of distribution according to soil depth would thereforee misleading and a comparison of densities reached in differ-nt (sub)horizons was not possible as these layers were often notomparable between soil cores. A rather sharp decrease of enchy-raeid numbers with increasing soil depth is typical for most soils,lthough less pronounced in arable ones (Didden, 1993). The highercentage of individuals extracted from the two upper layers0–6 cm depth) however indicates that sampling double this depthovered the vast majority of enchytraeids present in the soil profile.

The leading edge of earthworm invasion is by definition a bor-er zone and thus it is not surprising that earthworm impact andoil conditions were rather heterogenous among the soil coresxtracted in the corresponding plots at either site. This is reflectedy the overlap of the assemblages found in individual soil cores ashown by the NMDS ordination. Earthworms have been spreadingather fast at the study sites and the leading edge was thereforeot only fuzzy but also shifting. This was particularly true for theisconsin site, where D. octaedra had invaded further ground very

uickly in the few years between the studies of Holdsworth et al.2007a,b) and the present one. The sharpness of the leading edgeas much lower than at the Minnesota site. Thus the exact posi-

ion of the sampled plots representing the leading edge at the twoites might not have represented exactly the same degree of inva-ion. Therefore it is not surprising that the results for the two sitesere particularly different for the two leading edges. Also L. ter-

estris had not reached the same high density in the most heavilynvaded area identified (and sampled) at the Wisconsin site as athe Minnesota site. Obviously, data from more situations with ann-going earthworm invasion under otherwise similar conditionsould further enhance our understanding of earthworm invasion

mpacts on enchytraeids. Alternatively, experiments are neededirectly manipulating earthworms densities in the field and inves-igating the responses of other soil organisms. For instance, it haseen shown that earthworm extraction using the octet method canuccessfully decrease earthworm densities and thereby dampenarthworm effects on ecosystem processes (Eisenhauer et al., 2008;zlavecz et al., 2013). From the larger number of studies on the

mpact of earthworm invasions on microarthopods, comprising allhe other dominant members of mesofauna, we know that theltimate outcome depends on the species or ecological groups ofarthworms involved, their densities, season, and the time since

il Ecology 83 (2014) 159–169 167

invasion or the time scale of observation (Migge-Kleian et al., 2006;Straube et al., 2009; Eisenhauer, 2010). Beylich and Graefe (2012)came to the same conclusion analysing co-occurrence of earth-worms and enchytraeids in North German long-term monitoringsites. Therefore, one has to be careful when generalizing the find-ings of the present study. Nevertheless, our results are in goodagreement with previous knowledge on both negative and pos-itive effects of earthworms on enchytraeids, including the role ofearthworm middens as enchytraeid habitat (see Introduction). Fur-thermore, this is, together with Schlaghamersky (2013a,b), oneof the first descriptions of enchytraeid assemblages within NorthAmerica, providing quantitative data on the individual speciespresent at clearly defined sites. A discussion on the proportion ofnative and introduced, potentially invasive, enchytraeids in NorthAmerica has hardly begun and would still suffer from of a lack ofdata. The presence of many species known from Europe, Greenlandand Siberia even in remote Northern temperate forests, includingold growth not yet or just recently reached by invasive earthworms(Schlaghamersky, 2013a) is supportive of the idea of a natural Hol-arctic distribution of many enchytraeids. Nurminen (1973) found astrong similarity of enchytraeid faunas in the vicinity of Montreal,Canada, and northern Europe. He also suggested that many speciesdescribed from North America might also have been described fromEurope under other names. It is obvious that enchytraeids have alsobeen brought to North America by human activity (Gates, 1976b).The enchytraeid faunas of the sites of the present study includespecies of Holarctic distribution (possibly in part due to introduc-tion), one species hitherto known only from the North AmericanArctic, and several species not yet known to science and thus prob-ably endemic to the continent (Table 1).

5. Conclusions

The present study shows that detailed investigation of com-munities in soil is necessary to appreciate the effects of invasiveearthworms on native ecosystems: while gross data on total enchy-traeid density, diversity and biomass did not indicate pronouncedearthworm effects, we found distinct changes in the compositionand shifts in mean individual biomass of the enchytraeid assem-blage. These compositional changes are likely to result in changesin the functioning of soil food webs. Moreover, our results suggestthat often underappreciated belowground invasions, such as byEuropean earthworms, can cause the loss of soil-dwelling species –even heretofore unknown ones – and thus deserve more scientificattention.

Acknowledgements

Jirí Schlaghamersky gratefully acknowledges funding by the Ful-bright Program (stating that neither the Government of the UnitedStates nor any agency representing it has endorsed the conclusionsor approved the contents of this publication) and by the Ministryof Education, Youth and Sports of the Czech Republic (ResearchPlan MSM0021622416). Nico Eisenhauer gratefully acknowledgesfunding by the Deutsche Forschungsgemeinschaft (DFG, GermanScience Foundation; Ei 862/1). Both above-mentioned authors werekindly hosted by the Department of Forest Resources and Center forForest Ecology at the University of Minnesota. Cindy Buschena andSusan Barrott from this department provided invaluable logisticsupport. We thank Linda Parker of Chequamegon-Nicolet NationalForest for help with logistical arrangements for field work.

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