Earthworms as colonizers of natural and cultivated soil environments

Applied Soil Ecology 50 (2011) 1–13

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Review

Earthworms as colonizers of natural and cultivated soil environments夽 H. Eijsackers a,b,c,∗ a

Wageningen University and Research Centre, Wageningen, The Netherlands Institute of Ecological Sciences, Vrije Universiteit Amsterdam, Amsterdam, The Netherlands c Department of Zoology, University of Stellenbosch, Stellenbosch, South Africa b

a r t i c l e

i n f o

Article history: Received 17 December 2010 Received in revised form 6 July 2011 Accepted 12 July 2011 Keywords: Earthworms Colonization Succession Cultivated soils Dispersal

a b s t r a c t For cultivated soils, the important function of earthworms as ecosystem engineers and their major contribution to the composition and functioning of soil ecosystems with a varying species diversity has been extensively addressed. However, the role of earthworms as colonizers of virgin, uncultivated soil in the process of soil formation has been little researched and long underrated. To better understand this role, the following questions need to be considered: (1) what makes an early colonizer successful, what are its characteristics, and which species are the most successful and under what circumstances are they successful?; (2) what are the limiting factors in these colonization processes with respect to environmental conditions and also to interspecific interactions?; (3) what do earthworms contribute to the further colonization by other soil animals?; and (4) how do they impact the soil itself and what could therefore be the consequences for soil management and restoration? These questions have recently been addressed from the perspective of new (or ‘alien’) earthworm species invading ecosystems, suggesting a massive influx of species, competitive to the originally present fauna. This idea is, however, contrary to colonization, which suggests a gradual exploration of a previously uninhabited area. Unlike recent research, this review approaches colonization primarily as a spatial dispersal process and part of natural succession processes, and is mainly illustrated with examples of Palearctic species, either in Europe or introduced elsewhere. To begin, the various stages of colonization: dispersal, establishment, population growth and interspecies relations are analysed. Next, the colonization processes, the possible limiting environmental factors and the sequence of the appearance and establishment of species are described. Dispersal rates and sequences of colonization by different earthworm species are given for different soil ecosystems. For colonization, limiting environmental factors such as pH, soil type and heavy metal contents as well as the presence of organic matter seem to play a more important role than inherent ecological characteristics like r/K selection. Finally, the role of earthworms in the early colonization of soils that are earthworm-free because of non-cosmopolitan distribution, drained former sea bottom, permanently water-logged soils or anaerobic, acid peaty soils are reviewed. If we understand the role of earthworms in succession, we will be able to improve their role in soil restoration and soil management. © 2011 Elsevier B.V. All rights reserved.

Contents 1. 2. 3.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Colonization as a succession process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Dispersal, establishment, and population growth characteristics of earthworms as colonizers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1. Dispersal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.1. Passive dispersal of cocoons . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.2. Passive dispersal of earthworms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.3. Active dispersal through the soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.4. Active dispersal over the soil surface . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

夽 This paper was prepared for a large part during study leaves at the Department of Zoology of the University of Stellenbosch, South Africa, and the Institute for Environment and Sustainability of the EU-Joint Research Centre, Ispra, Italy. ∗ Correspondence address: POB 9101, 6700 HB Wageningen, The Netherlands. Tel.: +31 317 481563; fax: +31 317 484449. E-mail address: [email protected] 0929-1393/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.apsoil.2011.07.008

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3.2. Establishment of earthworms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3. Population growth . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4. Competition, facilitation and succession . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Colonization in structured soils . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1. Colonization process and limiting factors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2. Earthworms and organic matter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Putting the role of earthworms as colonizers into perspective . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1. Various scales of earthworm colonization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2. Successful early colonizers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3. Enabling and limiting environmental factors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.4. Field validation and application in soil management and restoration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1. Introduction The role of colonizers in soil formation of virgin, uncultivated soil has long been underrated. This holds true both for virgin soil that has never been inhabited and for anthropogenically influenced soil that has become ‘sterile’ and then inhabited again, such as sanitized soils (Tamis and Udo de Haes, 1995) and dredged sediments (Eijsackers et al., 2009). Several laboratory experiments have shown the intrinsic relation between soil community structure (species diversity) and system functioning (Thompson et al., 1993; Naeem et al., 1994; Tilman and Downing, 1994; Heemsbergen et al., 2004; Hedde et al., 2010). Although Darwin (1882) had already described these underlying processes, it was not until Jones et al. (1994) coined the term ecosystem engineer that the role of the earthworm in soil formation was more accurately described. The term was then further defined by Lawton (1994), Anderson (1995) and Lavelle (1997), especially the major contribution of earthworms to the composition and functioning of soil ecosystems with varying species diversity. In all of these experiments, however, earthworms were added to cultivated, mixed soil. For this reason, two important questions still remain on the role of earthworms as ecosystem engineers, namely what role do earthworms play and how do they impact the soil as invaders or early colonizers under natural soil conditions? This topic has recently received a good deal of attention from the perspective of the ecological consequences of new (‘exotic’ or ‘alien’) species invading ecosystems, especially in North America and elsewhere in the Nearctic, and has been impressively described in a special issue of Biological Invasions (Hendrix, 2006; Hendrix et al., 2006, 2008). In this issue also meta-distribution over time (decades to centuries) is discussed for different areas in the world. However, the idea of invasion suggests a massive influx of species competing with the local biota. In contrast to the gradual process of colonization, this influx can cause major changes in the structure and functioning of ecosystems especially for agronomic and forestry systems (Ljungström, 1972). Unlike recent research this paper approaches colonization as part of a step-wise succession process and deals mainly with spatial dispersal processes and rates of Palearctic species, either within Europe or introduced in other continents to cultivated soils. A complementary review on the colonization of contaminated land and soil and sediment waste deposits has recently been published (Eijsackers, 2010). Moreover, this review only briefly treats the meta-distribution patterns and processes of earthworm species, as well as the consequences of earthworm colonization for soil ecosystems, as this has already been described in relation to the North American situation by Hendrix et al. (2006). The paper starts with an analysis of Bradshaw’s (1993) various stages of colonization: getting there, establishing, growth (including interspecies relations). In focusing on the

6 6 7 8 8 9 9 9 9 10 10 10 10

role of earthworms as colonizers, the following questions are addressed: (1) What makes an early colonizer successful, and what kind of characteristics play a role in their success? Are these population-biological characteristics such as reproductive system, reproductive capacity or growth rate, or are these ecological characteristics such as resilience against adverse conditions (freezing, desiccation) and pH-distribution range? (2) Which species are the most successful and under what circumstances are they successful? Is success related to soil or humus type? (3) What are the limiting factors in these colonization processes with respect to environmental conditions and interspecific interactions? Assuming that dispersal, more specifically primary introduction, is mostly human-driven, soil treatments and land use, in general, could be main steering factors. However, both interspecific competition and facilitation of the soil conditions could also play a role. (4) What do earthworms contribute to the soil’s further colonization by other soil animals? (this issue is not further treated in this paper because of excellent reviews by Hendrix (2006) and Hendrix et al. (2006)). How do earthworms impact the soil with respect to bulk density, organic matter content and distribution, and what are the consequences of this impact for soil management and restoration? 2. Colonization as a succession process Begon et al. (1996) define succession as ‘the non-seasonal, directional and continuous pattern of colonization and extinction on a site by specific populations’. Primary succession deals with soils that have not been previously inhabited by soil life due to adverse conditions: permanently water-logged, recently drained sediments (although benthic fauna may have had a distinct impact on sediment structure); soils with extreme acidity or alkalinity; or deeper soil layers becoming available through landslides or open cast mining. Globally, primary succession in soils is the case where earthworms have been absent from an area such as an island. Secondary succession, in contrast, deals with soils where soil life has been present and active but has become nearly or totally eradicated and has to be restored. This is the case after a natural disaster, such as inundation, frost or fire, or due to changed land use, such as forestation of natural meadows, reclamation of peat land or abandonment of arable land. Facilitation may play a role in primary succession of virgin soil material that is devoid of biological life. Materials such as lavaashes, newly deposited silts or drained lake and sea bottoms are weathered, but lack structure and internal coherence. To prepare these materials for colonization, organic matter is needed. Organic

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matter reduces erosion and regulates soil moisture content while providing rooting substrate. Organic matter and nutrients as well as cover are provided by the first plant species. Then soil animals can colonize (Begon et al., 1996). Approaching succession as a process, Bradshaw (1993) distinguished three states: getting there, establishing, and growth. Den Boer (1961) mentioned three possible driving forces for the start of ‘getting there’ (=dispersal): escape from current adverse conditions, overflow from overcrowding and increased chance to (re)found populations, and natural diffusive dispersal due to random individual movement. 3. Dispersal, establishment, and population growth characteristics of earthworms as colonizers 3.1. Dispersal Dispersal can be sub-divided into passive dispersal, either by anthropogenic or natural processes, and active dispersal over the soil surface or through the soil. In contrast to migration, dispersal is undirected, although environmental conditions influence the direction and rate of dispersal. 3.1.1. Passive dispersal of cocoons The likelihood of medium- to long-range dispersal is attributable to earthworms escaping to the soil surface, e.g. after heavy rains, followed by wash-off of cocoons and earthworms, and eventual further transport by streams. Surprisingly, few observations exist for long-distance transport of cocoons. Transport by sea currents seems to be limited for earthworms, as only a few species have euryhaline cocoons and can withstand seawater (Lee, 1985). Haeck (1969) observed that the dispersal of freshwater-tolerant Allolobophora chlorotica into newly drained, polder soils is closely related to the distribution of the ditches and waterways in the polders; in these polders water is continuously circulated and, as such, transport of earthworm cocoons by streaming water may be the main dispersal mechanism. Lumbricus rubellus, Aporrectodea caliginosa and A. chlorotica were regular first colonizers along dikes and in the polder itself, with semi-aquatic Eiseniella tetraeda at a few places along canals. Terhivuo and Saura (2006) demonstrated transport along streams by analysing genetic clone variability of earthworms for a catchment in a NorthEuropean mountain side. As part of a study on earthworm dispersal by surface run-off, Schwert and Dance (1979) and Schwert (1980) collected cocoons from a stream in Ontario, Canada, and they found most of them to be viable. Furthermore, an exceptionally high dispersal rate of earthworms in Canadian hardwood forests, mentioned by Dymond et al. (1997), may be attributed to cocoon dispersal during excessive rain and temporary inundation of the area. Meijer (1972) found a small confined earthworm colony in the heart of a newly drained polder on a former sea bottom some kilometres from the surrounding mainland. Before being drained, this area was home to a colony of gulls. Apparently, these birds imported earthworm cocoons to the area through the mud on their feet. Lee (1985) and Schwert (1980) also attributed cocoon dispersal partly to avian phoresy. The presence of earthworm species introduced relatively recently to the South Sea islands, such as Gough and Marion, could also be attributed to dispersal by birds, although human transport seems to have had the most quantitative impact (Lee, 1985; James, 2004). Dispersal by cattle along their tracks has also regularly been observed (Hoogerkamp et al., 1983). Marinissen and van den Bosch (1992) observed and experimentally proved that tractor-wheels provide excellent phoresy. From additional modelling they concluded that passive transport of cocoons has a dominant influence on

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the rate of population expansion, at least in regularly cultivated soils. To survive transport to and first establishment in primary soils, cocoons must be able to withstand adverse conditions. The cocoons of L. rubellus are resistant to cold and drought (Jensen and Holmstrup, 1997; Marinissen, 1992) and can withstand inundation and flooding (Roots, 1956; Pizl, 1999; Plum, 2005; Zorn et al., 2005). In general, earthworms survive frost mainly by their frost-resistant cocoons (Holmstrup, 2003). L. rubellus and Lumbricus castaneus overcome quiescence and anhydrobiosis by their resistant cocoons (Bouché, 1987). On the other hand, Baker and Whitby (2003) suggest that soil moisture content is a good indicator of cocoon survival and hence establishment of Aporrectodea longa. Bengtsson et al. (1986) showed that cocoons of Dendrodrilus rubidus are quite sensitive to lower soil pH (<4.5). These observations illustrate the important role of cocoon survival, but the idea needs further and more thorough investigation. 3.1.2. Passive dispersal of earthworms Few reports deal with the question of why earthworms start to disperse. Gates (1972) mentions active mass migration of Perionyx spp. in hills in Burma without providing a specific driving factor. Atlavinyté and Payarskaite (1962) and Schwert (1980) report massive dispersal after and in response to heavy rain. Mathieu et al. (2010) observed in a rather space-limited experimental set-up that dispersal of Aporrectodea icterica could be triggered by increased earthworm density, while low habitat quality and pre-use by conspecific individuals also trigger dispersal. In a field experiment Grigoropoulou and Butt (2010) observed that artificially increased densities of Lumbricus terrestris lead to higher dispersal. Furthermore, species that live close to the surface (epigeics) are more prone to physical dispersal forces such as flooding, wind, or phoresy than deep burrowing anecic species (Terhivuo, 1988). Locally, humans play a dominant role in earthworm introduction and redistribution by transporting soil and plant material. In addition to earthworm dispersal through plant material and adhering soil, Plisko (2001) observed that the distribution of introduced Pontoscolex corethrurus was related to relatively high rainfall and proximity to urban and agricultural areas. Proulx (2003) and Hale and Host (2005) found a relationship between dispersal in temperate hardwood forests in the USA and an anthropogenic index consisting of density of cabins/resorts, boat landings, roads, and campsites. Holdsworth et al. (2007) found a relationship between earthworm distribution and distance to roads and cabins, and Cameron and Bayne (2009) between the distribution of alien earthworm species and road age. Cameron and Bayne (2009) mention truck transport and bait abandonment as the most important distribution factors. Genetic analysis showed that human-mediated dispersal ‘jumps’ due to multiple introductions are more common than diffuse dispersal via road networks or earthworm mobility itself (Cameron et al., 2008). Passive dispersal by rain events, inundation over several weeks and floods or water streams has also been reported by Haeck (1969), Schwert (1980), Terhivuo and Saura (2006) and Bodt (unpublished data). However, these studies focus on relatively short distances (a few kilometres up to tens of kilometres) and without distinguishing cocoons from earthworms. This passive dispersal may also be triggered by overcrowding (Grigoropoulou and Butt, 2010) or to escape adverse conditions such as heavy rain or flooding. Zorn et al. (2008) observed experimentally that A. chlorotica tends to escape flooded soil, while L. rubellus clearly avoids and escapes from wet and flooded soil. In the laboratory Perreault and Wahlen, 2006 measured less burrowing activity of A. caliginosa and L. terrestris in wet (−5 kPa) than dry (−11pKa ) soils. Surface applications of irritating fluids such as manure slurries (Edwards, 1988a, 2004) with high ammonia content or dissolved salts or flooding with salt

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seawater (Piearce and Piearce, 1979) cause surfacing. According to Edwards (1988b), total dissolved salt levels above 1.4% generally cause avoidance, and ammonia levels above 0.5 ␮g g−1 dry soil and salt levels above 2.9% in solution are lethal. Certain pesticides also force earthworms to the surface (Mather and Christensen, 1998; Christensen and Mather, 2004). Because of this combination of avoidance/escape and toxicity, it is difficult to specify how many specimens can disperse after surfacing before dying or being predated. On a much larger spatial scale, Tiunov et al. (2006) reviewed meta-distribution and meta-dispersal processes in north-eastern Europe and Russia and in the USA Western Lakes District, with temperature as a main limiting factor for dispersal and establishment. All intercontinental introductions of European earthworm species to Africa, Asia, Australia or North America can be attributed to human transport of materials containing earthworms or their cocoons, such as potted plants (Lee, 1985). In South Africa, European and Asian earthworm species can be found in gardens all over the country (Ljungström, 1972; Reinecke, 1983) and seem to have moved or been moved along the routes of the early European settlers. In Brazil, 20 of 54 newly found species are introduced (Baker et al., 2006) and in Argentina, the invasion and further dispersal of the Palearctic as well as the Australasian species is partly explained by human immigration waves (Mischis and Herrera, 2006). In Iceland, the distribution of the five most common earthworm species currently found is mostly associated with urban and, to some extent, agricultural areas (Rundgren, 2007). Rundgren (2007) further suggests that earthworms may have been transported from Norway to Iceland in drifting ice blocks. For an overview of these global distribution processes, see Hendrix et al. (2008). 3.1.3. Active dispersal through the soil The active dispersal mechanism of earthworms – crawling – differs between various groups of earthworms: epigeics, endogeics and acenics (Bouché, 1972). Epigeic earthworm species (e.g. L. rubellus, Dendrobaena octaedra and Eisenia fetida) crawl through the upper soil-litter layer, endogeics (A. caliginosa and A. chlorotica) steadily burrow through the humic and mineral soil layer, and anecics (L. terrestris and A. longa) dig deep vertical, more or less permanent burrows. According to these descriptions, epigeics seem to disperse most actively through the soil and anecics the least. However, differences in dispersal rate and distance are less distinct when considering movement over the soil surface (see Section 3.1.4). Dispersal of earthworms has been measured by introducing earthworms into different types of earthworm-free soils, such as polder soils of former sea bottoms that are totally devoid of earthworms (Van Rhee, 1969a,b) and soils in New Zealand and Australia where European species were absent and where the population densities of indigenous species were negligible (Stockdill, 1966, 1982; Springett, 1972). Other introductions relate to the recovery of open-cast mining soil or spoil heaps (Eijsackers, 2010). In The Netherlands earthworms were introduced at a specific point, and dispersal was directly measured by observing the progress of the earthworm-front, i.e. by determining the maximum dispersal distance from the point source. Hoogerkamp et al. (1983) observed an almost immediate dispersal of the introduced A. caliginosa, which progressed linearly over time to 72 m in 8 years. After 8 years, the population level was steady up to 36 m after which the earthworm numbers tailed off. Differences in population development were considerable between the different transects measured: at 36 m, numbers varied between 140 and 410 earthworms m−2 . However, the maximum dispersal was the same for the different transects: 72 m. The latter observation suggests that dispersal and population development are two separate processes: dispersal is not directly related to population size and, hence, to crowding as a dispersal

factor. Inoculated A. longa in the UK (Butt et al., 1999) showed a maximum dispersal of 2.5 m in a heavy compacted clay after 10 months, with a mean dispersal of 0.6 m and 80% of the observed surface casts within this distance from the introduction point. For the same species Baker (2004) measured a maximum of 7 m and a mean of 2.5 m after 7 years in cultivated grassland in Australia. For L. rubellus, Eijsackers (unpublished data) measured a dispersal between 5 and 11 m per year (mean 7 m yr−1 ). Table 1 summarizes dispersal rates, all of which were measured over several years in the same area in the Netherlands, thereby excluding the impacts of different soil types and minimizing differences in climate conditions. Earthworms were introduced at one confined source in known numbers, known species composition and age structure. In this set-up, no other dispersal processes except active burrowing occurred. The mean rates in Table 1 were measured some years after introduction. According to Hoogerkamp et al. (1983), earthworm population development started following an adaptation period of 3–4 years after introduction. Lee (1985; based on data of Stockdill, 1966, 1982) and Van der Werff et al. (1998) found a similar lagphase in population development. Eijsackers (unpublished data in Table 1) observed dispersal by A. longa and L. rubellus in the first year after introduction, although Butt et al. (1997) recorded much slower dispersal after the first 10 months than the mean rates given in Table 1. Ligthart and Peek (1997) also observed that A. longa immediately started to disperse and reached its highest dispersal rate after 4 years, followed by a gradual decline. A. caliginosa showed a very slow increase, with the highest dispersal rate after 6 years while L. rubellus started to disperse only after 5 years. To verify whether dispersal is influenced by different field conditions, field studies in which the total dispersal period (a series of years) was specified with the maximum distance reached, were selected for analysis. These are plotted for managed agricultural land (pastures, meadows, arable land) and for landfills combined with sediment deposits (Fig. 1A and B, respectively). In agricultural land the dispersal distances showed limited variation between the various studies and earthworm species and proceeded more or less linearly with an adaptation period of 2–6 years. Variation was greater in landfill sites and sediment deposits, possibly due to the greater variability in substrate, but there was a similar limited lagphase. Table 2 gives regression equations, correlation coefficients and number of observations of dispersal rates of various earthworm species in grassland, waste deposits, and fly ash and mine waste deposits (the latter derived from Eijsackers, 2010). Because of the limited number of suitable long-term studies, only a few general trends can be indicated. The highest dispersal rates were recorded for the endogeic species A. chlorotica and A. caliginosa which appeared to flourish in grassland and fly ash + mine colliery deposits, but not in municipal waste deposits (possibly because the substrate provides less suitable environmental conditions without mineral soil, and possible because some waste deposits are capped) .Of the two anecic species, A. longa fared well in all three types of situation, while L. terrestris fared best in grassland on fertile mineral soil. Among the epigeic species, L. rubellus was ubiquitous while the other epigeic species encountered (L. castaneus, D. rubidus, Eisenia tetralix) were associated with the presence of a surface organic layer. But it has to be realized that rates highly depend on local conditions, as can be deduced from the variability in dispersal rates in Table 1, which have been measured in similar grasslands in one area in The Netherlands (Oostelijk Flevoland). In several instances dispersal started only after some years. Soil substrate and management seem to influence dispersal to some extent. Lower dispersal rates can be related to bulk density as observed by Eijsackers et al. (2005). Capowicz et al. (2009) observed

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Table 1 Annual dispersal rate (m yr−1 ) of earthworm species introduced in worm-free clayey polder soils in the Netherlands. Species

Land use

Dispersal rate (m yr−1 )

Reference

Lumbricus rubellus

Grazed grassland Arable land Grazed grassland

7 14 8

Eijsackers (unpublished data) Marinissen and van den Bosch (1992) Stein et al. (1992)

Aporrectodea caliginosa

Grass strips orchards Grazed grassland Arable land Grazed grassland Grassland

6 9 7 11 6

Van Rhee (1969a,b) Hoogerkamp et al. (1983) Marinissen and van den Bosch (1992) Stein et al. (1992) Ligthart and Peek (1997)

Allolobophora chlorotica

Grass strips orchards

4

Van Rhee (1969a,b)

Aporrectodea longa

Grazed grassland Grazed grassland Grassland

5 8 6

Eijsackers (unpublished data) Stein et al. (1992) Ligthart and Peek (1997)

Lumbricus terrestris

Grazed grassland Grassland

4 1.5

Hoogerkamp et al. (1983) Ligthart and Peek (1997)

Adapted from Eijsackers (2010).

in experimental field-cages that anecics, mainly L. terrestris but also Aporrectodea giardi, A. caliginosa and Aporrectodea rosea, are able to colonize compacted soil zones. However, they also noticed indications of avoidance of these zones. Chemical factors may also play a role in dispersal. Amendment of soil with fly ash resulted in 30–40% reduced burrow volume with fewer and smaller burrows by native megascolecid and introduced Aporrectodea trapezoides earthworms (Yunusa et al., 2009), although earthworm densities in this study were high (800 m−2 ) and fly ash was added in limited amounts (500 g m−2 ). Ma and Eijsackers (1989), however, did not observe reduced burrowing activity of L. rubellus in clay loam soil supplied with fly ash up to 30% by weight. Facilitative interaction between species during the dispersal may also occur. Butt et al. (1999) observed a far better dispersal by A. chlorotica when introduced with A. longa (see further Section 3.4). 3.1.4. Active dispersal over the soil surface Nightly surface crawls, as described for anecics, generally occur for all types of earthworm species. Anecic earthworms leave burrows at night and show homing ability, as recorded for L. terrestris, up to a distance of 0.7 m away from their burrow (Nuutinen and Butt, 2004). When using soil surface pitfalls, Mather and Christensen (1998) collected at least seven species of earthworms (A. longa, L. terrestris, A. rosea, A. caliginosa, A. chlorotica, Octolasion cyaneum and Lumbricus festivus). Anecics were, by far, the most frequently present, endogeics to a slightly lesser degree and epigeics to

a far lesser extent. Pitfall trapping by Bouché (1976), Martins (1976) and Callaham et al. (2003) gave similar results. Bouché (1976) captured substantial numbers of endogeics (A. caliginosa, A. icterica, A. rosea) and recognized the importance of surface movements, but he did not elaborate on it. Following the individual mucous tracks of earthworms, Mather and Christensen (1988, 1992) studied surface movements and observed that L. terrestris can crawl 19 m in one night and A. longa 23 m. The mean track length for L. terrestris measured 9.0 ± 3.8 m (range 3.8–19.3 m), and adults, sub-adults and some large juveniles were observed crawling over the soil surface. Van Breemen (personal communication) found earthworms in collectors of tree stem flow water and assumed that the earthworms must have crawled into the trees. Pizl (1999) observed that earthworms climb tree trunks to escape flood waters. According to Hendrix et al. (2006), P. corethrurus was found both in the soil and in the trees of montane cloud forest in the Lesser Antilles. Considering both the observations on passive and active dispersal of earthworms and the heterogeneous character of most soils, it seems reasonable to assume that diffusive processes, as mentioned by Den Boer (1961), drive active dispersal through the soil, as shown for polder soils (Addendum by Eijsackers and Oude Voshaar, in Hoogerkamp et al., 1983), and for an experimental sediment deposit (Eijsackers et al., 2009). With respect to active dispersal by earthworms over the soil surface, Mather and Christensen (1988, 1992, 1998) observed a considerable reach but no directionality. Earthworms can, however, detect and avoid adverse conditions, such as acid or highly alkaline soils (Satchell, 1955; Satchell and Stone, 1977), remediated versus contaminated

Table 2 Linear regressions (y = qx − s), correlation coefficients (r2 ) as derived from Table 1 and Fig. 1A and B and number of observations (n) of earthworm dispersal rates for various earthworm species in grassland, fly ash + mine colliery depots, municipal waste deposits (low correlation coefficients or low numbers of observations in italics). Grassland

Aporrectodea longa Aporrectodea rosea Aporrectodea caliginosa Allolobophora chlorotica Lumbricus terrestris Lumbricus festivus Lumbricus rubellus

Fly ash + mine colliery 2

Regression

r

n

Regression

y = 6.72x + 17.40

0.18

5

y = 9.19x − 29.04 y = 13.73x − 29.36 y = 7.39x − 15.22

0.62 0.43 0.96

8 3 5

y = 5.73x + 1.70

0.98

4

y = 6.38x − 30.86 y = 5.98x − 34.96 y = 7.03x − 39.98 y = 5.39x − 32.12 y = 2.11 + 10.65 y = 4.07x + 27.30 y = 2.65x + 48.05 Scattered distribution y = 9.50x + 19.50 Scattered distribution

Lumbricus castaneus Dendrodrilus rubidus Eisenia tetralix Data on fly ash and mine colliery depots are adapted from Eijsackers (2010).

Waste deposits 2

r

n

Regression

r2

n

0.69 0.59 0.75 0.44 0.19 0.47 0.06

5 4 6 4 6 5 7

y = 15.05x − 40.26

0.79

5

y = 2.84x + 5.20

0.44

5

0.45 0.12

5 5

y = 6.34x − 12.81 y = 5.30x − 12.50

0.94 0.81

4 5

6

H. Eijsackers / Applied Soil Ecology 50 (2011) 1–13 80

A

Along

70

Dispersal distance (meters)

rosea (Wilcke, 1952) are not known as early colonizers, however. Another factor that influences establishment is the survival capacity of cocoons. According to Pizl (1999), recovery from surviving cocoons is the main mechanism for survival, especially for epigeic species. Cocoons can survive drought spells (Bouché, 1987; Doube and Auhl, 1998), inundation and flooding (Roots, 1956; Pizl, 1999; Plum, 2005; Zorn et al., 2005), and frost (Holmstrup et al., 1998; Holmstrup and Zachariassen, 1996; Marinissen, 1992). However, no clear relation between cocoon characteristics of specific earthworm species and success of establishment has been observed.

Acal Lrub Lterr

60

Linear (Acal)

50

Linear (Along) Linear (Lrub)

40

Linear (Lterr)

30 20 10 0 0

2

4

6

8

10

12

3.3. Population growth

Dispersal period (years)

Dispersal distance (meters)

B

140

Along Etetr

120

Lcast Lrub

100

Linear (Along) Linear (Etetr)

80

Linear (Lcast) Linear (Lrub)

60 40 20 0 0

2

4

6

8

10

12

14

16

Dispersal period (years) Fig. 1. Linear dispersal in metres during different periods of years after inoculating various earthworm species in: (A) grasslands or cultured grassland soils in Australia (Lee, 1981), Switzerland (Daniel et al., 1996) and the Netherlands (Van Rhee, 1969b; Ligthart and Peek, 1997; Hoogerkamp et al., 1983). Acal: Aporrectodea caliginosa, Along: Aporrectodea longa, Lrub: Lumbricus rubellus, Lter: Lumbricus terrestris. (B) Municipal waste and harbour sludge deposits in the UK (Butt et al., 1997, 1999, 2004), Germany (Brockmann et al., 1980) and the Netherlands (Ma and van den Ham, unpublished data). Along: Aporrectodea longa, Etret: Eiseniella tetralix, Lcast: Lumbricus castaneus, Lrub: Lumbricus rubellus.

soils (Stephenson et al., 1998), or specific contaminants such as copper (Eijsackers, 1983; Ma, 1984, 1988; Eijsackers et al., 2005; Lukkari and Haimi, 2005). This avoidance behaviour may hamper colonization by earthworms for many years, especially in cases of persistent contaminants such as heavy metals, so that soil restoration is retarded or even prevented. 3.2. Establishment of earthworms Biological characteristics influence the establishment of earthworm populations. Reproduction strategies by parthenogenesis or self-fertilization may certainly be advantageous for the initial establishment of individual specimens (James and Hendrix, 2004), but these strategies have hardly been studied in relation to colonization. For further dispersal, small cocoons that can endure adverse environmental conditions should be effective, whereas for establishment, large cocoons should be potentially the most successful. Unfortunately, data on cocoon characteristics of various earthworm species is limited. Recent knowledge from culturing is mostly related to manure and compost species like Eisenia species, although Lowe and Butt (2002) provide an overview on culturing of soil-dwelling earthworm species. According to Lavelle (1981), cocoon size varies from 1 to 25 mm and relates positively to the size of the adult earthworm. Earthworm species with large cocoons such as A. longa and A.

After the initial establishment by hatching cocoons or invading adults, populations have to grow. Here environmental conditions also play a role. In addition to the observations on cold hardiness of cocoons by Holmstrup et al. (1998), survival has been observed in juvenile and mature earthworms at extremely low temperatures for many months. Nuutinen and Butt (2009) found A. caliginosa actively burrowing below a deep frozen surface layer and viable cocoons of A. caliginosa and L. terrestris in the frozen layer. D. octaedra was observed in frozen soil of −5/−20 ◦ C (Parkinson et al., 2004), and −12/−14 ◦ C (Leirikh et al., 2004). However, the latter authors commented that the low rates of embryonic and postembryonic development prevent the life cycle from being completed in one season. Population modelling shows that maturation time, in particular, determines whether a species will survive (Klok, 2007), as shown for L. rubellus and L. terrestris under chemical stress (Klok and De Roos, 1998), under inundation stress (Klok et al., 2005) and combined chemical and inundation stress (Klok et al., 2007). Most studies on chemical stress focus on survival and reproduction, presumably because, as stated by Klok and De Roos (1998), studying maturation time requires lengthy experimental observations. Bengtson et al. (1979) introduced different earthworm species in Icelandic grasslands and suggested that the better colonization of A. caliginosa compared to L. rubellus is due to A. caliginosa’s relatively longer total and reproductive life span. Lavelle (1979) suggested a combination of these factors, which he called a demographic factor. It consists of the number of cocoons produced per adult per year, divided by the time needed for maturation multiplied by the total life expectancy. However, he did not apply this factor any further. To explore what factors may play a role in establishment success, in Table 3 information on population aspects obtained from Wilcke (1952), Edwards and Bohlen (1996) and Holmstrup et al. (1990) is combined with data on environmental preferences: moisture by Zorn et al. (2008), pH and C/N by Bouché (1972) and cold hardiness by Holmstrup et al. (1990). This information is compared with an r–K classification based on population-biological, functional-morphological and physiological characteristics by Satchell (1980e), assuming that opportunistic r-oriented earthworm species would be better colonizers. From the table no apparent relations between r–K classification and population-biological characteristics can be deduced, although it should be noted that these data are from different sources and provide only a general picture of these relations. The r–K classification of Satchell (1980e) can be compared with ecological classifications (epigeic, endogeic and anecic) by Bouché (1977) for European earthworm species and by Lee (1985) for Australasian earthworm species. After further extension with an Adversity factor (Greenslade, 1983), Lee (1987) found more precise classification of the ecological positions of the species in the

−1 (−5) −1 (−5) 35–65 7.30–24.38 7.30–47.81 3.9–8.3 3.7–8.5 55 55 (5) 3

8 27

17.5 19

48 ± 8 40 ± 5

72.5 74

−1 (−5) (45) 55–65

37 50 36 53 ± 10 39 ± 5

53 60 48.5

3.5–8.5 4.3–8.1 4.1–8.5

8.34–30.05

35–45 (55) (2.5)

3 6 2.5

106 8 27

16 10 12.5

7.64–39.73 3.9–8.5 8 24 14 65

8.63–19.65 4.4–7.5 3.9–8.1 66 38.5 55 30 37 ± 7 11 8.5 11 42

55–25 50–0

85–70 50–40 50–35

100–85

100–90 100–85

Eisenia. fetida Dendrobaena subrubicunda Lumbricus castaneus Lumbricus rubellus Aporrectodea longa Allolobophora chlorotica Aporrectodea rosea Aporrectodea caliginosa

4 2.5

7.96–47.81

0

Cold hardiness (◦ C) Moisture preference loamy soil (%) C/N ratio soil pH-range Total period (weeks) Period of growth of worm (weeks) Incubation time of cocoons (weeks/days ± SD) Number cocoons (worm−1 yr−1 ) r to K range (r = 100,K = 0)

Size of cocoons (Ø in mm)

7

r–K–A framework, but still showed limited correspondence with the population-biological data collected by Wilcke (1952). Bouché (1987) commented that only some hygrophilic epigeic species are really r strategists, and inhospitable, hostile environments are occupied by some r-selected individuals or are devoid of earthworms. 3.4. Competition, facilitation and succession

Species

Table 3 r/K range of ecologically different earthworm species (adapted from Satchell, 1980e), combined with cocoon sizes (Edwards and Bohlen, 1996), total time of development (incubation time and growth period) and reproduction capacity (numbers of cocoons) of various earthworm species (Wilcke, 1952), experimental cocoon incubation data in days by Holmstrup et al. (1990), pH and C/N preferences according to Bouché (1972), moisture preferences with incidental observations between brackets according to Zorn et al. (2008) and cold hardiness as observed by Holmstrup et al. (1990).

H. Eijsackers / Applied Soil Ecology 50 (2011) 1–13

Over the last decade the number of studies on the interactions between earthworm species has increased. From a review, Uvarov (2009) concluded that there are contrasting interaction patterns between anecic (with mostly positive interactions) and epi/endogenic species (mostly competitive). In a series of studies, Decaëns et al. (2008, 2009, 2011) analysed competitive interactions as drivers of species assemblage dynamics. According to Uvarov (2009), these interactions strongly affect structure and functioning of earthworm communities. However, this area needs further study to conclude how these interactions are steered by genetical and physiological characteristics as data on these characteristics are scarce. Competition can result from direct interactions like consumption of cocoons, as Dalby et al. (1998) described for Microscolex dubius in the presence of A. longa. Most literature on competition refers to competition for food sources (organic matter), which results in a longer maturation time and a lower mass gain of other species. This is explained by the steeper growth curves of D. veneta, L. rubellus and A. chlorotica as compared to other species (Butt, 1998). Similar observations concern A. longa reducing the abundance of M. dubius (Baker et al., 1999; Hendrix et al., 2006), and the abundance and biomass of A. caliginosa (Baker et al., 2002), as well as megascolecid species (Baker et al., 1999). Lowe and Butt (2002) observed a negative impact of L. rubellus on L. terrestris when co-cultured with a limited amount of surface organic matter, while lower competitive abilities of A. trapezoides versus Argilophilus marmoratus became smaller with decreasing habitat quality (Winsome et al., 2006). Eriksen-Hamel and Whalen (2007) observed slower growth rates for L. terrestris, A. chlorotica, A. caliginosa, and A. longa in multispecies cultures, especially when larger (more competitive?) earthworms were present. Sánchez-deLéon and Zou (2004) showed that introduced A. caliginosa increase the number and biomass of endogeic P. corethrurus. These observations could be interpreted as examples of out competing by introduced (alien) species in some cases, but there could also be spatial exclusion between species or mutually excluding environmental preferences. In a study by Baker (1998), introduced species showed different distributions than endemic species, due to other environmental preferences. Valkx et al. (2009) concluded from a kriging analysis that A. caliginosa and A. rosea have similar sized and overlapping patches, so they are apparently not competitive. However, these patches are not related to the areas where L. terrestris and A. longa occur. Jiménez and Rossi (2006) observed spatial competitive exclusion in the Colombian ‘Llanos’ for two endogeic Andiodrilus and Glossodrilus species. Thus, competition does occur specifically for food and organic matter. From a colonization perspective, Dunger (1969) and Bernier et al. (1993) described a succession of species whereby epigeic species comminuted collected raw humus. In doing so, they seemed to facilitate the establishment of endogeic species, which subsequently consumed the amount of humus present. This action, in turn, retarded the appearance of anecic species, which could be labelled inhibition. Facilitation was also observed in relation to dispersal of A. chlorotica when inoculated with A. longa (Butt et al., 1999), while Ligthart and Peek (1997) observed that A. longa is

8

H. Eijsackers / Applied Soil Ecology 50 (2011) 1–13

always found at the dispersion front, followed by A. caliginosa. Comparing the dispersal rates of earthworm species inoculated either combined (Hoogerkamp et al., 1983) or individually (Eijsackers, data in Table 1), the dispersal rates of individual L. rubellus or A. longa were lower than rates when combined. All these data refer to vegetated areas, mostly with humus top layers. There are only a few studies in bare soil substrate or substrate with a very thin layer of litter. Eijsackers et al. (2009) studied how freshly deposited clayey sediment was colonized in the sequence L. castaneus > L. rubellus > A. caliginosa > E. tetralix > A. chlorotica > A. rosea, from the 5th month onwards, with periods of several months to years in between during which no colonization occurred. On freshly deposited flyash, L. rubellus and A. caliginosa appeared first, followed by A. chlorotica and E. tetralix after some more years (Eijsackers et al., 1983). The observation that colonization occurs in ‘waves’ with intermittent periods without new appearances, as also observed by Dunger (1968, 1969) and Pizl (1992), suggests that the process is a combination of competitive interaction between species and environmental changes and facilitation by which a species changes the soil substrate to favour a subsequently appearing species. Much of the success of early colonisers can be attributed to ecological flexibility or endurance, with respect to an available thin layer of organic matter or raw humus and low pH (Dunger, 1968, 1969; Pizl, 1992) and possibly combined with a high incidence of desiccating or freezing conditions (Bouché, 1987; Holmstrup et al., 1998). L. rubellus, a successful colonizer, combines a broad ecological tolerance with respect to humus quality and pH with the ability to change its behaviour from surface-active to more deep-burrowing according to the conditions, even creating surface casts (Butt et al., 1999). Reproduction strategy (polyploidy, parthenogenesis) might be advantageous for first establishment, but it is only incidentally mentioned in the literature and seems to be ancillary to ecological factors. Species with a high reproductive potential, such as E. fetida, are hardly found under early succession conditions, as they require and inhabit organic-rich (raw humus, manure) substrates and prefer higher temperatures. Growth rates may also play an important role in establishing earthworm populations, but they are seldom mentioned in the literature (Butt, 1998) and deserve more research. Curry and Boyle (1987) and Curry and Schmidt (2006) investigated establishment in cutover peat, and observed that initial growth and survival are not always good predictors of long-term establishment success. In the sequence of appearing species during succession, a certain pattern can be observed with endogeics (Allolobophora and Aporrectodea) and epigeics (Lumbricus) as first colonizers, depending on the soil type. Therefore, the ecological grouping in epigeics, endogeics and anecics is helpful in interpreting ecological development and succession (see also Addison, 2009). From the European papers reviewed, anecic earthworm species come last in succession because they need the deepest topsoil layer. In North America and Nearctic in well-developed top soils, anecics establish at an early stage of colonization. Classification based on r–K selection (Satchell, 1980e; Lee, 1985) does not seem to help to structure colonization succession. The sequence of appearance may be explained by a difference in mobility capacity, in addition to exposure to physical dispersal forces at the soil surface (Terhivuo, 1988) and ecological flexibility. Hoogerkamp et al. (1983) and Eijsackers et al. (2009) observed that adult specimens arrive first and are at the dispersal front, followed by sub-adults. In laboratory experiments, Butt (1998) observed that some species are more successful in active (crawling or burrowing) dispersal than other species. More specific behavioural data are needed on dispersal activities of early colonizing species such as D. rubidus and L. rubellus (Ma and Eijsackers, 1989).

4. Colonization in structured soils 4.1. Colonization process and limiting factors Most authors who have investigated colonization of primary soil substrates stress the importance of topsoil, i.e. a layer of organic matter should be available (Dunger, 1969). Surprisingly, few reports on earthworm succession after natural catastrophes such as fires, landslides or floods have been published. Lee (1981) reported earthworm recovery after a scrubland fire: after a rapid, temporary recovery of the topsoil species in the first few months after the incident, no earthworm colonization occurred for the next 4 years. A. caliginosa and L. rubellus then began to colonize the area. Dunger (1968) observed a similar sequence in a mining area where the top soil had been removed: colonization started 5 years after reclamation and was fully established after 10 years. The impacts of flooding were summarized by Plum (2005). From her meta-analysis she concluded that flooding reduces species diversity and earthworm abundance in winter as well as in spring and summer. Zorn et al. (2004) studied flooding in riverbanks combined with contamination by heavy metals and organic contaminants. The impact of flooding was much higher than the impact of the contamination, but populations recovered well after each inundation event of a few weeks. Klok et al. (1997, 2005) investigated earthworm growth rate under stress and concluded from modelling studies that the length of the dry period between the inundation episodes determined the growth and the number of mature earthworms and, hence, reproductive capacity and viability of the earthworm populations. Piearce and Piearce (1979) and Piearce (1982) observed that 1 year after flooding with seawater, earthworms still avoided the previously flooded soils, and that after 3 years, A. caliginosa, A. chlorotica and L. rubellus were the dominant re-colonizers. Ivask et al. (2007) studied earthworm communities of coastal meadows and floodplain grasslands in the Matsalu bay in Estonia. They concluded that A. caliginosa and L. rubellus are highly tolerant of flooding with low saline waters. Fly ash deposits are a special case of saline impacts as these are characterized by high pH and salt levels. Both Satchell and Stone (1977) and Ma and Eijsackers (1989) concluded that only after weathering, which gradually reduces the salt concentrations, are conditions suitable for earthworm colonization. Given the type of soil and land use, the sequence in which species colonize may differ. Eijsackers (1983) studied arable fields on very poor, poorly structured, bare, sandy soils for the first 8 years after abandonment. L. rubellus was the first colonizer, 1 year after abandonment, in a few cases accompanied by A. caliginosa, L. castaneus, and D. octaedra established a few years later in low numbers and very few plots. Only sampling plots with 3.2–3.6% OM were colonized in the first year with 10–30 earthworms m−2 , whereas two plots with lower OM contents (1.9 and 2.5%) had zero and 5 earthworms m−2 , respectively. Chamberlain and Butt (2008) found no earthworms in sand dune transects with OM contents <1%; the first observation of D. octaedra and L. rubellus was at OM contents >4%. Boyd (1957) observed a similar succession with D. rubidus in dunes on the Hebrides. Pizl (1992) compared succession in an arable field, a primary uncultivated fallow, a 10-year-old fallow and a 60-to-70-year-old fallow. In the third year, L. rubellus and D. octaedra colonized first, followed by A. rosea 10 years later. A. caliginosa, a common first invader in arable and pasture land, was found in the cultivated field and in other sites in negligible numbers. In naturally reforested mountain meadows (spruce, 5–15 years old), endogeic A. icterica and A. caliginosa were present in large numbers, especially at a higher altitude (Bernier et al., 1993), sometimes with L. cas-

H. Eijsackers / Applied Soil Ecology 50 (2011) 1–13

taneus, L. terrestris and some D. octaedra. After further succession these high numbers of endogeics disappeared. Sagot et al. (1999) observed a correlation between earthworm distribution and the age of the trees in semi-natural alpine forests. In newly reclaimed peat sites, Curry and Cotton (1983) found D. octaedra, D. rubidus and L. rubellus, which is not surprising, given the C/N-ratio of the raw humus of these freshly reclaimed sites, with A. caliginosa, A. chlorotica and L. festivus in low numbers. A. longa, A. rosea and L. terrestris appeared after 10–15 years, indicating significant profile development at these older sites then. Eijsackers (2010) compared the sequences of colonization for pulverised ash deposits, colliery soil heaps, abandoned agricultural fields and meadows and reclaimed peat sites, but he did not observe consistently similar patterns.

4.2. Earthworms and organic matter As discussed in the previous section, an important factor for succession in these various soil habitats is the formation/presence of an organic matter layer to enable initial colonization by earthworms. In open-cast brown-coal-lignite mining planted with alder and poplar after cessation of mining activities, the rate of litter decomposition rose from 5 to 27% after earthworms established themselves in year 7 (Dunger, 1968). After 10 years, earthworms had decomposed 70% of the litter-fall. From experiments in mesocosms, Frouz et al. (2007) concluded that only treatments of post-brown coal-mining substrate with earthworms, especially L. rubellus, result in improved litter breakdown. Billings (1938) observed similar phenomena in the natural succession of old field soils, although these were cultivated and succession occurred over many decades. Burtelow et al. (1998) studied an earthworm-free forest soil that had been abandoned some 100 years ago and invaded by earthworms. They observed that earthworm activities reduced organic matter content in the upper A soil layer from 48.1 to 30.8%, thereby increasing the pH from 4.1 to 5.5. Microbial biomass and microbial activity had changed considerably, increasing both C and N flux. The disappearance of the organic top-layer, however, did not coincide with an increase in the organic matter content of the underlying soil layer, so earthworm activity had decreased the Cstore. After endogeic earthworm species nearly totally consumed the available organic matter in a mountainous coniferous forest (Bernier et al., 1993), spruce growth and new accumulation of raw humus were stimulated. Satchell (1980a,b,c,d) tried to improve the quality of a Calluna podzol by planting birch and introducing L. terrestris. Introduction was not successful because the pH was too low (3.3–3.5). Only D. octaeadra and Eisenia eiseni, typical raw humus species, established, but they did not change the humus layer. Hoogerkamp et al. (1983) studied the impacts of introduced earthworms in a soil with a well-developed mat of grass, roots and raw humus. After 3.5 years, the mat still partially existed, after 5–7 years there was a heterogeneous A1 -horizon and after 10 years a well-worked homogenous A1 -horizon. The C/N-ratio, infiltration capacity, conductivity for water and air, and penetration resistance had also altered considerably. As a consequence, root contents had changed and cumulative dry matter yields had increased. Further development of soil structure (Rogaar and Boswinkel, 1978; Ligthart and Peek, 1997) and burying of plant seeds (Van der Reest and Rogaar, 1988) have also been reported. These secondary observations are, however, less relevant for early succession during which the disappearance of the top layer of organic raw humus and the formation of a mixed humus-rich top soil layer are of primary importance.

9

5. Putting the role of earthworms as colonizers into perspective 5.1. Various scales of earthworm colonization From the literature on colonization by earthworms, different scales can be distinguished: colonization of continents and isles, colonization of catchments, and colonization of habitats such as grasslands and forest stands. Each colonization has different spatial scales, time scales, mechanisms and success factors. For example, the colonization of Australia, New Zealand, and southern Africa by European earthworm species has occurred over centuries, is mainly anthropogenic and has been successful in relation to manmade land use patterns (Lee, 1985). Colonization of catchments occurs over years to decades, is mainly driven by water transport and human disposal but also incidentally by avian phoresy, and has environmental site characteristics as success factors (Schwert, 1980; Terhivuo and Saura, 2006). Colonization of habitats such as grasslands, forest stands, arable fields, but also former mining pits and deposits occurs over years to decades, is mainly driven by earthworm burrowing in combination with their survival capacities, and can be supported by the addition of top soil or organic matter and a vegetation cover (Hoogerkamp et al., 1983; Bernier et al., 1993; Butt, 2008). 5.2. Successful early colonizers A selection of successful colonizers can be based on: (1) the number of times a particular species arrives first, (2) the range of habitats in which a species can establish itself, (3) the range of environmental conditions a species can endure, and (4) the dominance in numbers or biomass of a species in early successional species communities. According to these characteristics, L. rubellus, L. castaneus and D. octaedra are mentioned mostly as primarily colonizing earthworm species in acid sandy soils, provided that there is some vegetative cover. In these conditions ecological plasticity is expressed in survival under drought and frost conditions but also in a broad food spectrum (Póspieck and Skalski, 2006). As can be concluded from the monitoring data collected by Bouché (1972) in Table 3, L. rubellus has the widest pH-range of 4.9 units (3.5–8.4), followed by L. castaneus with 4.5 units (3.9–8.4) and D. octaedra 4.4 units (3.3–7.7). L. rubellus has a reasonably broad moisture range (Zorn et al., 2008), while Plum (2005) found L. rubellus in all locations examined in flooded grassland, and A. caliginosa, E. tetralix, O. lacteum, D. octaedra, A. chlorotica and L. castaneus to a consecutively lesser extent. In North American forests, the epi-endogeic L. rubellus causes the most rapid removal of litter during primary colonization (Hale et al., 2005). For acidic, drier and metal polluted soils Curry and Cotton (1983) mention L. rubellus, D. rubidus, D. octaedra and L. eiseni as the most successful colonizers. For succession in metal polluted areas, physiological similarities influence colonization (Spurgeon, 1997). Tolerant species found close to the smelter in the Avonmouth area (L. terrestris, L. rubellus and L. castaneus) have higher calciferous gland activities than sensitive species (A. chlorotica, A. caliginosa and A. rosea). These calciferous glands play an important role in metal sequestration (Vijver et al., 2004). In clayey soils, A. caliginosa is mentioned mostly as a successful colonizer. Various authors (e.g. Bengtson et al., 1979) mentioned A. caliginosa’s capability to aestivate during adverse drought conditions as a success factor for colonization. Additionally, the species has a broad soil moisture range (35–65%) in loamy clay soil (Zorn et al., 2008) and a slightly smaller pH-range (3.7–8.5) than L. rubellus. For clayey soils with a better buffering and cation exchange capacity and a more neutral pH than sandy soils, the pH-range is less important.

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Lee (1985), Dunger (1968) and Bernier et al. (1993) showed that it takes a longer period than the first colonization phase in which the epigeic earthworm species arrive before endogeic species can successfully establish themselves. 5.3. Enabling and limiting environmental factors The presence and subsequent disappearance of a layer of organic matter is typical in primary succession. D. rubidus is a raw humus species and, according to Kobel-Lamparski and Lamparski (1987), L. rubellus also feeds on raw humus. In changing the raw humus layer into a more decomposed humus type, the pH also changes. This change in pH improves the conditions for species with a narrower pH-range such as endogeics (A. caliginosa), as studied in natural alpine forests by Bernier et al. (1993) and in recovering mine deposits by Dunger (1969). According to these authors, comminution of raw humus is the first major change in the disappearance of the litter layer, followed by a change in C/N-ratio (Bernier et al., 1993; Billings, 1938). Soil depth, the mixing and reworking of the organic top-layer with the underlying mineral soil layer, is primarily related to the appearance and burrowing activities of endogeics. When a deeper well-worked soil becomes available, anecic species such as L. terrestris establish themselves (Dunger, 1969). Reworking organic matter is linked with a suite of derived changes from physicochemical characteristics to crop production increases (Hoogerkamp et al., 1983; Baker et al., 2006) and forest ecosystem changes (Parkinson et al., 2004; Hendrix et al., 2006; McLean et al., 2006; Migge-Kleian et al., 2006; Addison, 2009). 5.4. Field validation and application in soil management and restoration A field validation under natural primary succession conditions of the experimental studies on the ecosystem engineering activities of earthworms (Jones et al., 1994) is lacking. However, many studies have been done on how earthworms impact soil structure, drainage capacity, aeration, seed burial, nutrient availability and release and primary plant production (for an overview, see Edwards, 2004). That said, these aspects have been studied at one location on cultivated fields (Hoogerkamp et al., 1983) – so do not describe primary succession as such – while recent studies in North American hardwood forests have investigated the broad impacts of these natural processes (Hendrix et al., 2006). Eisenhauer et al. (2008, 2009a,b) studied the impact of soil activities on aboveground plant growth and observed clear positive relations. Belote and Jones (2009) state that earthworm invasion can suppress and facilitate invasion of plant species other than the ones originally present. A point of discussion for future research is how this invasion literally proceeds. Observations of extremely fast dispersal of several tens of metres per year (Dymond et al., 1997) are not supported by experimental field data as presented in this review. The invasion of European earthworms in Australia has been geographically extensive, but it remains patchy at regional and field scales (Baker et al., 2006). Given the ecological consequences of changed forest floor dynamics, the discussion on rates and patterns of dispersal deserves more attention, including regulations to minimize introduction of new exotic species (Callaham et al., 2006). Further, to gain a complete understanding of the consequences of earthworm comminuting, burrowing and bioturbating activities, indirect impacts also have to be studied and assessed for their ecological implications. These include: (1) drainage in combination with stimulated leaching of nutrients or contaminants, (2) aeration in combination with changes in aerobic breakdown as well as anaerobic sorption processes, and (3) bioturbation in combination with the vertical distribution of nutrients and contaminants in the

soil profile. These impacts have to be related to the environmental conditions in the various soil layers (O2 -content, pH) with consequences for breakdown and sorption capacity for soil contaminants. This is particularly relevant when earthworm colonization, possibly by artificial inoculation (Butt et al., 1999; Butt, 2008), is used to restore degraded land. To successfully colonize and improve soil conditions, earthworm species have to be applied in the most appropriate way in relation to the purpose of a particular management practice. In doing so, three main questions have to be considered (Bengtsson, 2002): (1) what soil characteristics are to be improved?; (2) is it primary colonization of bare ground or soils under harsh conditions, revitalisation of cleaned soil, or secondary colonization after land use change?; and (3) can disturbed areas be restored by regeneration from within or by sources outside the disturbed area? Environmental conditions such as pH and OM-content can be improved by adding lime and organic wastes, respectively. Wanner and Dunger (2002) further suggested that by creating a specific ‘landscape’ with crests and troughs, colonization conditions can be improved by the accumulation of organic matter and moisture in the troughs. The selection of species also has to agree with the purpose of the introduction. For instance, the choice of E. fetida for soil improvement is not relevant and neither is the introduction of L. terrestris in podzolic soils under heath land (Satchell, 1980a,b,c,d). In most disposal situations (mine waste, dredged sediments, municipal waste), no internal sources exist and earthworms have to be introduced from outside by inoculating earthworms, top soiling or adding plant material. When internal regeneration is possible, resilience processes also have a strong spatial component. In this case, soil heterogeneity is important, too and has to be maintained or restored. In this context, soil improvement is a classical topic of debate in soil science, including the positive contributions of earthworms to soils (see Lavelle et al., 1998; Parmelee et al., 1998; Brown et al., 2004). Equally important in this debate is the role that earthworms can play in restoring contaminated or deteriorated soils. Understanding the role of earthworms in succession by recognizing the similarity between colonization processes and natural succession processes will enable us to improve their role in restoration; introducing the wrong group of earthworm species for restoration purposes works contra-productively. In employing earthworms for soil restoration, various papers (Ma and Eijsackers, 1989; Butt, 2008; Snyder and Hendrix, 2008) describe how to use earthworms in natural attenuation, an ecological alternative for the land-useimprovement-based term ‘reclamation’. Acknowledgements The author gratefully acknowledges the support that he received from the University of Stellenbosch, especially Professors Sophie and Adriaan Reinecke, and from the JRC/IES, especially former director Professor Manfred Grasserbauer and Dr Jan Marco Muller. The comments of Sten Rundgren, Kees van Gestel and James Curry during the preparation of the paper, as well as from five anonymous reviewers and the linguistic editing by Peter Griffith greatly improved the paper. References Addison, J.A., 2009. Distribution and impacts of invasive earthworms in Canadian forest ecosystems. Biol. Invasions 11, 59–79. Anderson, J.M., 1995. Soil organisms as engineers: microsite modulation of macroscale processes. In: Jones, C.G., Lawton, J.H. (Eds.), Linking Species and Ecosystems. Chapman and Hall, London, pp. 94–106. Atlavinyté, O., Payarskaite, A.I., 1962. The effect of erosion on earthworms (Lumbricidae) during the growing season. Zool. Zh. 41, 1631–1636.

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