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Title: Deciduous woodland exposed to elevated atmospheric CO2 has species-specific impacts
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on anecic earthworms.
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John Scullion*1, Andrew R. Smith2, Dylan Gwynn-Jones1, David L. Jones2, Douglas L. Godbold3
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Aberystwyth, SY23 3DA, UK
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2 School
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UK
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* Corresponding author Tel 0044 1970 622304 Fax 0044 1970 622350 email [email protected]
Institute of Biological, Environmental and Rural Sciences, Aberystwyth University, Penglais,
of Environment, Natural Resources and Geography, Bangor University, Bangor, LL57 2UW,
Institute of Forest Ecology, University of Natural Resources and Life Sciences, 1190 Vienna, Austria.
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Key words: carbon, litter, respiration, elevated CO2, deciduous woodland
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The authors have no conflicts of interest regarding this research.
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Research was supported by CIRRE under the Bangor-Aberystwyth University initiative and funders
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had no input into this research.
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Abstract
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Elevated atmospheric CO2 induced reductions in litter quality can adversely affect earthworms.
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However, this understanding is based on laboratory rather than field research and relates to
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single earthworm and tree species. Here earthworm populations were investigated under Alnus
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glutinosa, Betula pendula, and Fagus sylvatica in a Free Air Carbon dioxide Enrichment field
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experiment. Litters from this experiment were also fed to Lumbricus terrestris L. at two rates with
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live weight change and cast properties assessed. Elevated CO2 (580 ppmv) reduced litter N (-
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12%) with a corresponding increase in C:N ratio, especially for A. glutinosa. In the field, elevated
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CO2 caused a shift in overall population composition, mainly characterised by reduced anecic
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biomass (–25%); endogeic and epigeic species were less affected. CO2 effects on total biomass
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were most pronounced for A. glutinosa (e.g. field total biomass -47% vs. -11% overall). Growth of
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L. terrestris was lower when fed elevated CO2 litter (-18%), although increased inputs of A.
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glutinosa litter mitigated this effect. In mesocosms, fresh cast respiration was lower (-14%) for
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elevated CO2 litter, an effect more pronounced for A. glutinosa (-24%). When normalised for C
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content, elevated CO2 effects on cast respiration were again negative and most marked for A.
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glutinosa litter. Litter N concentration, and possibly ease of litter mineralisation were factors
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affecting litter resource quality Litter N and P concentrations varied with A. glutinosa > B.
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pendula > F. sylvatica; F. sylvatica had the highest cellulose content. Field earthworm
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biomass was higher under A. glutinosa compared with B. pendula and F. sylvatica (+17 and
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+70% respectively); live weight increased with A. glutinosa litter in the feeding trial almost
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three times more than for B. pendula, whereas it decreased for F. sylvatica. Cast respiration
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was highest for A. glutinosa, intermediate for B. pendula (ca. -36%) and lowest for F.
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sylvatica (ca. -78%). Earthworm responses to elevated CO2 were complex, being
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characteristic of individual tree and earthworm species; responses were more adverse for
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trees with higher quality litter and for anecic earthworms.
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Key-words: carbon, litter quality, respiration, earthworms, elevated CO2 , deciduous woodland
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1. Introduction
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Earthworms play a major role in temperate woodland function by recycling leaf litter, and promoting
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the release of its nutrients. Earthworm casts are organic rich ‘hot spots’ of microbial and mesofaunal
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activity (Gorres et al., 2001) which influence broader soil ecology (Scheu and Parkinson, 1994). In
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woodlands, earthworms influence root development and depth distribution (Fisk et al., 2004), and
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alter understory herb communities (Frelich et al., 2006) on a scale comparable with those of deer
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grazing (Fisichelli et al., 2013). Earthworms increase litter incorporation into soil, although the extent
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is dependent on species composition (Suarez et al., 2006). These activities alter soil carbon, nitrogen
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and phosphorus dynamics (e.g. Madritch and Lindroth, 2009; Sackett et al., 2013). Earthworms have
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wider environmental and ecological roles; they promote infiltration and recharge of groundwater
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(Shutz et al., 2008); particular species are an important food resource for birds and mammals (Kruuk
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and Parish, 1981; Peach et al., 2004). Therefore, the size and composition of earthworm communities
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may significantly impact on woodland food webs and ecosystem function.
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Numerous studies (e.g. Neirynck et al., 2000) have shown that earthworm communities are affected
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by litter characteristics of different tree species. Factors considered to affect earthworm dietary
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preference and nutrition include litter C:N ratio and concentrations of substances such as
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polyphenols, tannins and lignin (Neilson and Boag, 2003). Litter colonisation by microorganisms,
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particularly fungi (Tiunov and Scheu, 2000), may also be important in determining food selection.
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Atmospheric carbon dioxide (CO2) concentrations for the end of this century are predicted to reach
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550-900 ppm depending on emissions scenarios (Karl et al., 2009). Growth of trees under elevated
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CO2 changes both the chemical and physical properties of their leaves but the effect varies between
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species. Thus, elevated CO2 increased leaf starch (Lindroth et al., 1993) in Populus tremuloides and
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Quercus rubra but had no effect in Acer saccharum; sugars were decreased in P. tremuloides,
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increased in Q. rubra and unchanged in A. saccharum (Lindroth et al., 1993). Increased lignin content
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and lignin:N ratios have been reported in leaf litter of Liriodendron tulipifera, Betula pendula, A.
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saccharum and Picea sitchensis (e.g. Cotrufo et al., 1994) but lignin decreased in leaves of Fagus
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sylvatica and B. pendula (Blaschke et al., 2002; Oksanen et al., 2005). Elevated CO2 effects on litter
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decomposition have also been found to vary with species (e.g. Rouifed et al., 2010).
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Most studies of soil faunal responses to elevated CO2 have been in grassland. Changes have been
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found in nematode communities (Yeats et al., 2003), but not in earthworm biomass (Niklaus, 2001;
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Chevallier et al., 2006). Marked effects of elevated CO2 on soil biota may be confounded with
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changes in plant community composition (Allard et al., 2004). Indeed Arnone et al. (2013) reported
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that reductions in plant biodiversity eliminated otherwise stimulatory effects of elevated CO2 on
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earthworm activity. Numerous studies have been undertaken into the consequences of elevated CO 2
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for insect herbivory in woodlands (e.g. Lindroth et al., 1993). More recently, there has been interest in
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the potential effects of elevated CO2 on earthworms (e.g. Meehan et al., 2010) relating to woodland
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ecosystems. Kasurinen et al. (2007) found that growth of Lumbricus terrestris was lower when fed B.
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pendula litter from elevated as compared with ambient CO2 treatments and attributed this effect to
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changes in litter chemistry. Similar responses were obtained by Meehan et al. (2010) in P.
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tremuloides litter feeding trials, with lower growth of the same earthworm species even though litter
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consumption rates were unaffected. In the field, elevated CO2 may affect earthworm populations
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through mechanisms other than litter quality, for example due to altered soil moisture regimes (e.g.
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Niklaus et al., 2007). In grassland, surface casting increased under elevated CO2 (Zaller and Arnone,
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1997).
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The aims of this study were to investigate whether responses to elevated CO2 in three diverse (in
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terms of litter quality) tree species affected field populations of earthworms and whether their litter
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varied as a food resource for the anecic earthworm L. terrestris L.. Previous research (Kasurinen et
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al., 2007; Meehan et al., 2010) reported on interactions between a single earthworm and litter species
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only, and did not include integrated field investigations. Litter resource quantity was expected to be
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higher but quality lower under elevated CO 2 (e.g. Meehan et al., 2010); reductions in litter quality were
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expected to have an adverse effect on litter feeding earthworms, but simultaneous increases in litter
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inputs might at least partly mitigate this effect, especially with intrinsically high litter quality. The study
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involved a Free Air CO2 Enrichment (FACE) field experiment and an ex situ feeding trial to investigate
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impacts of litter from this experiment on earthworms. The feeding trial complemented the field survey
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and focussed on how varying litter quality and availability affected growth of L. terrestris. This species
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dominated earthworm biomass at the study site and is a direct consumer of leaf litter in woodland
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(e.g. Suarez et al., 2006).
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2. Materials and methods
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2.1. Free air CO2 exposure field experiment
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The BangorFACE experiment was established in March 2004 on the coastal plain in north Wales, UK
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(53o 14’ N; 4o 01’W). The climate is Hyperoceanic, with annual rainfall of about 1000 mm. The soil is a
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fine loamy brown earth over gravel (Rheidol series) classified as a Dystric Cambisol in the FAO
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system (Teklehaimanot et al., 2002). The topography consists of a shallow slope of approximately 1–
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2° on a deltaic fan. The aspect is towards the north-west, at an altitude of 13 to 18 m above sea level.
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The depth of the water table ranges between 1 and 6 m. Further site details are given in Hoosbeck et
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al. (2011); in particular they found no differences in soil properties between control and FACE plots
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prior to tree planting.
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Trees were planted on two adjacent fields in a block design. The main experimental plots were
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surrounded by a 10 m buffer strip containing the same species planted at the same density and
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pattern. Four ambient CO2 and four elevated CO2 main plots within the plantation formed a complete,
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replicated block design. Carbon dioxide enrichment started at the beginning of April 2005 by injecting
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pure CO2 through laser-drilled holes in tubing mounted on eight masts (Miglietta et al., 2001).
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Elevated CO2 concentrations, measured at 1 minute intervals, were within 30% of the pre-set target
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concentration of 580 ppmv. The main experimental plots were approximately 8 m in diameter, with
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saplings of Alnus glutinosa L. Gaertn., B. pendula Roth and F. sylvatica L. planted in split-plots at 80
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cm spacing (seven single or mixed species plantings) in a hexagonal design. For the purposes of this
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study only single species split-plots (c. 9 m2) were investigated since the composition of litterfall under
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mixed species was too heterogeneous. When the field sampling was undertaken in autumn 2007 A.
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glutinosa and B. pendula were ca. 5.5 m tall, and F. sylvatica ca. 2.5 m tall; with the exception of F.
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sylvatica split-plots the canopy had closed.
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2.2. Leaf litter collection and analysis
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Collection of leaf litter commenced from September 2005 onwards. One litter basket (30 X 30 cm)
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was placed in the centre of each species split-plot (24 for the purposes of this study) and emptied
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weekly until all leaves had abscised. Litter was briefly rinsed with water to remove surface
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contamination, sorted into individual species, and then placed into paper bags for drying at 80 ºC for
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24 hours. The dry weight of each species was determined; litterfall data for the period September to
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November 2007 preceding the earthworm sampling are reported here. At the time of earthworm
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sampling, for A. glutinosa in the ambient treatment 88% and in FACE 79% of the total litterfall for the
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year had occurred. In B. pendula these values were 97% and 94%. For F. sylvatica not all dead
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leaves were abscised (< 50% estimated) in autumn, with some retained until spring or longer.
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Freshly fallen litter was collected from the split plots sampled for earthworms in November 2007 to
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determine treatment effects on its chemical composition and as a substrate for the feeding trial. Leaf
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material was processed as above, but dried at a lower temperature of 30°C with sub-samples taken
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for analyses and the remainder used in the feeding trial. One sample of milled litter was analysed for
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each field split-plot (24 in total). Leaf quality was assessed by measuring lignin, cellulose, nutrients (N
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and P) and C:N ratio using standard methods (Harborne and Harborne, 1998). Briefly, acid detergent
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fibre was determined (Van Soest and Wine, 1967), then lignin in residues from this analysis was
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dissolved using a saturated potassium permanganate/buffer solution; lignin was determined by weight
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loss and cellulose by weight loss from these residues after ignition at 500 oC. C and N were measured
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using a LECO CHN analyser (LECO Corp., MI); P was determined colorimetrically (Murphy and Riley,
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1962).
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2.3. Earthworm survey of field experiment
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Sampling of earthworms, and recording of surface cast and burrow openings (one location near the
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centre of each tree species split-plot and adjacent to litter collection), were carried out under moist soil
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conditions when earthworm activity was high (November 2007). Owing to the potential disruption
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caused to the experiment by sampling, a single survey was timed to coincide with other sampling and
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the volume of soil from which earthworms could be recovered was restricted to that required by this
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survey. Prior to earthworm sampling, casts were cleared from the soil surface of quadrats (30 by 30
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cm) then collected at roughly weekly intervals over the following 18 days, dried at 105oC and weighed;
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numbers of surface burrow openings were recorded at these locations immediately prior to earthworm
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sampling and some larger burrows showed evidence of midden formation. Soil blocks including any
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surface litter (30 by 30 by 30 cm) were then excavated and hand-sorted for earthworms, with deeper
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dwelling earthworms extracted using a chemical expellant (0.5% formalin) applied to the base of the
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resulting pit. Earthworms were stored in moist moss before counting and weighing (after 24h
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depuration); this depuration period resulted in some loss in their condition. After identification to
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species level based on their morphology (Sims and Gerard, 1999), individuals were classified as
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mature (clitellate) or immature; a sub-set of individuals from each species were preserved in formalin
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and re-examined to confirm live identifications. The potential for misidentification of species in this
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process and uncertainty in earthworm taxonomy (e.g. James et al., 2010) is recognised, but protocols
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were followed both rigorously and consistently such that data for different treatments were
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comparable. Some partial samples and smaller immature specimens were included in overall
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population estimates but could not be reliably allocated to individual species.
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2.4. L. terrestris mesocosm feeding trial
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The litter feeding trial was set up in January 2008. Mesocosm conditions were not designed to
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simulate those in the field nor field earthworm densities; rather they aimed to provide a favourable and
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neutral feeding environment for earthworms in which to evaluate short-term, responses to adding litter
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of varying quality. Litter collected from split plots on the field site was fed to L. terrestris at two rates
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based on notional average feeding levels for this species (Edwards and Bohlen, 1996). Whilst these
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rates are referred to as ‘excess’ (approximately 125% of estimated ‘average’ feeding rates) and
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‘deficit’ (approximately 75% of estimated ‘average’) they were not intended to relate to actual rates of
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consumption in the field; rather they aimed to provide an assessment of potential
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compensatory responses to reductions in litter quality.
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For practical reasons, the excess treatments were started two days after the deficit treatments, but
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each treatment used the same supply of earthworms and litters, and was subjected to identical
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environmental conditions. Litter from each field tree species split-plot was added to duplicate pots,
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giving 48 pots for each feeding rate, and averages calculated for subsequent data analysis.
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Litter from each split-plot was milled to <0.2 mm to facilitate direct ingestion by earthworms (Marashi
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and Scullion (2003) who found marked reductions in mineral particles > 0.5 mm in casts and to
earthworm
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minimise the effects of varying physical constraints between tree species on feeding rates (Neilson
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and Boag, 2003). Litter was added at 7-8 day intervals at 5 or 3 mg g-1 earthworm live weight for
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excess and deficit treatments respectively. In the excess treatments, surface cover of litter at each
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subsequent addition was usually < 15% except for F. sylvatica where litter accumulated. In the deficit
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treatments surface litter had been incorporated by earthworms in most A. glutinosa and B. pendula
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mesocosms (F. sylvatica inputs ceased early for deficit treatment) by the time subsequent additions
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were made. These observations confirmed that the input rates chosen bracketed potential feeding
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rates.
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The feeding trial was conducted in darkness under favourable temperature (16°C) and soil moisture
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(60% water-filled pore space) conditions. From a batch of L. terrestris of varying sizes and maturity
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(Neptune Ecology, Ipswich) smaller clitellate individuals (2.4 - 2.6 g live weight) were selected to allow
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for potential weight loss or gain. Species identity was confirmed (Sims and Gerard, 1999) prior to use.
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Four individuals were then added to each 2 litre (13 cm diameter, 15 cm deep with perforated lids)
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mesocosm containing an artificial soil medium (sand, peat and kaolin – ratio 7:3:1 by volume; OECD,
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1984), mixed with soil (1g) from each field plot to provide a microbial inoculum corresponding to the
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field. This high earthworm density aimed to encourage rapid processing of litter and cast production.
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The medium used ensured that the only food resource was added litter and that pots were free of
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other earthworms; use of the field site soil would have introduced confounding effects of treatments
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on soil organic matter. Previous studies (Kasurinen et al., 2007; Meehan et al., 2010) have used
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defaunated or sterilised soils.
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The feeding experiment lasted for 25 and 27 days, with the final litter applications on day 21 and 22,
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for deficit and excess feeding respectively. Assessments included estimates of % surface litter
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removal (indicator of litter palatability – week 1 only as subsequent recordings were confounded by
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remaining residues), surface casting rates and cast C content (index of litter ingestion) and earthworm
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weight change (index of litter feed quality). Weight changes were calculated on a weekly basis from
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differences between initial and final weight of partially depurated (24h on moist filter paper)
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earthworms. Surface casts were removed at weekly intervals and at the end of the trial, weighed after
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oven drying, then used for loss-on-ignition determinations; ‘fresh’ (< 20 h old) casts deposited over a
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single day following weekly cast removal from the ‘excess’ feeding mesocosms (amounts on the
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deficit feeding pots were insufficient) were used for respiration measurements as a proxy for litter
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digestibility after adjusting for C contents (mineralisation constant). The period over which ‘fresh’ casts
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were collected was limited so as to avoid the rapid changes that occur in cast microbial activity
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(Scullion et al., 2003). It was assumed that bulk organic contents in casts would not alter significantly
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over a longer collection period.
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Cast respiration measurements used a colorimetric assay based on CO2 absorption by an alkaline
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gel containing a pH indicator dye (MicrorespTM Macaulay Land Use Research Institute, Aberdeen UK);
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this microplate system enabled the measurement of CO2 production from the small amounts of cast
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materials available and over a 6h incubation (Campbell et al., 2003). Cast organic contents were
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estimated by loss-on-ignition at 400oC (Gallenkamp muffle furnace) and converted to carbon using a
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standard factor of 1.724. Mineralisation constant (C respiration-to-total organic C ratio - Gilsotres et
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al., 1992) was calculated from respiration and C data.
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2.5. Data analysis
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Field responses within earthworm ecotypes (Bouche, 1977) were assessed using two-way ANOVA
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following data transformation where appropriate; population results were analysed as aggregate data
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for anecic (mainly Lumbricus terrestris with some Aporrectodea longa Ude), endogeic (mainly
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Aporrectodea caliginosa Sav. with some Allolobophora chlorotica Sav., Aporrectodea rosea Sav. and
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Octolasion cyaneum Sav) and epigeic (Lumbricus castaneus L. and Lumbricus rubellus Hoffm.)
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species. A. longa might be characterised as endo-anecic (e.g. Felten and Emmerling, 2009). Epigeic
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species were not abundant, so were not analysed in this way.
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The field experiment was treated as a split-plot design with CO2 regime (main) and tree species (split)
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treatments and included a block term. Data from the excess feeding treatments were also analysed in
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this way, since litter collected from the field experiment was fed on a split plot basis and therefore
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potentially retained variations inherited from the field.
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together, a three-way ANOVA was used and the experiment treated as a split-split-plot design with
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feeding rate as an additional factor. Relationships between mesocosm weight change data (excess
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feed treatments only) and litter quality parameters were investigated by linear and partial correlation
When both feeding rates were analysed
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analyses. All data analyses used GenStat 13th Edition (VSN International Ltd, Hemel Hempstead,
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UK).
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For the feeding trial, weight changes were adjusted for time to allow for direct comparison between
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the excess and deficit litter treatments; the analyses included all litter species for the excess
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treatment, but only A. glutinosa and B. pendula for the combined excess-deficit comparisons. F.
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sylvatica litter proved a poor food resource for earthworms, with high mortality rates in the deficit
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treatments. The response variables were change in mean weight per live individual and cast
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parameters averaged over two field split-plot replicate pots. Use of this index favoured the F. sylvatica
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litter treatments (excess feed treatment) where deaths were more frequent (22% average mortality),
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and also discounted occasional random losses from other litter treatments (< 5% average mortality).
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3. Results
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3.1. Field experiment
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Litterfall rates (Table 1) tended to be slightly (+9%) higher under elevated CO2, an effect close to
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statistical significance (P = 0.067). Elevated CO2 reduced litter quality (Table 1) but this effect was
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significant only for N (-11%) concentration (P = 0.038) and C:N (+16%) ratio (P = 0.010). Overall, litter
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quantities were markedly higher (P < 0.001) for A. glutinosa compared with B. pendula and F.
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sylvatica. Annual leaf fall for F. sylvatica was less than half that of B. pendula.
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indices differed between tree species, with a general ranking of A. glutinosa > B. pendula > F.
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sylvatica. Highest N concentrations were found in A. glutinosa litter, almost 3 times those in F.
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sylvatica (P < 0.001). Similar concentrations of phosphorous were found in litter of A. glutinosa and
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B. pendula, both higher (P < 0.01) than for F. sylvatica. Cellulose contents in F. sylvatica litter were
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almost double (P < 0.001) those in A. glutinosa and B. pendula.
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Aporrectodea caliginosa was dominant numerically (> 60% of total abundance), whereas L. terrestris
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was the dominant species in terms of biomass (> 50% of total biomass); these were the only two
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species (A. caliginosa (n = 22) and L. terrestris (n = 24)) recovered from most locations and together
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represented > 75% of biomass. A. chlorotica and A. longa were present at 14 locations. A. rosea, L.
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castaneus, L. rubellus (all n = 11) and, less frequently, O. cyaneum (n = 6) were also recovered in low
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densities.
Most litter quality
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There was no significant effect of elevated CO2 on population biomass (Table 2) but a marked (ca.
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50%) decrease under A. glutinosa with limited changes under F. sylvatica and B. pendula. This
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resulted in a significant (P = 0.032) CO2 X species interaction (Fig 1). Anecic species biomass was
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significantly (ca. 25%) lower (P = 0.046) under elevated compared with ambient CO2; CO2 X species
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interaction trends were similar to those for the population as a whole (P = 0.061). Endogeic species
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did not show any direct significant treatment effects (Table 2). Of the 13 points at which epigeic
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species were recovered 9 were under elevated CO2 (mean biomass 7.2 gm-2) compared with 4 (mean
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biomass 2.1 gm-2) under ambient CO2.
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A. glutinosa supported a higher (ca. - 70%) total biomass than F. sylvatica, with B. pendula
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intermediate. Anecic species biomass was twice as high under A. glutinosa (P = 0.025) compared
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with F. sylvatica. Of the eight split-plots for each tree species, epigeic earthworms were recorded at 4
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(mean biomass 3.3 gm-2), 3 (mean biomass 2.6 gm -2) and 6 (mean biomass 7.8 gm -2) points under A.
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glutinosa, B. pendula and F. sylvatica respectively.
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Mean densities of surface burrows were unaffected by CO2 regime and field casting rates were highly
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variable with no significant treatment responses. Burrow densities differed significantly (P = 0.007)
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between tree species in order of A. glutinosa > B. pendula > F. sylvatica, trends broadly following
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those for anecic species abundance. Although not recorded systematically, it was observed that
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larger burrow openings often showed evidence of litter incorporation.
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3.2. Mesocosm feeding trial
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Trends in litter removal in week 1 of the feeding trials were similar to those for casting. Mean removal
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rates of 89 and 87% (excess litter input excluding F. sylvatica) were estimated for ambient and
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elevated CO2 litter respectively; these high removal rates reflected the high earthworm densities used.
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Overall, cast production was not affected by the CO2 regime of the litter source. For excess only data,
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there were significant (P = 0.002) CO2 X species interactions (Fig. 2); earthworms fed elevated CO2
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A. glutinosa litter increased cast production compared with ambient litter, but those fed elevated CO2
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B. pendula litter had lower casting rates, whilst the generally low rates of casting for F. sylvatica were
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unaffected. Casting rates were markedly (ca. -40%) lower (P < 0.001) for the deficit compared with
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the excess litter input. A further CO2 X species X feed input (P < 0.001) interaction was obtained (Fig.
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3a); whereas elevated CO2 A. glutinosa litter fed in excess caused a marked increase in casting, for
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all other comparisons elevated CO2 decreased casting rates.
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Cast production for A. glutinosa was almost twice that for B. pendula, and more than 10 times that for
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F. sylvatica (excess only). Litter removal was estimated at 95% for A. glutinosa, 80% for B. pendula
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and 35% for F. sylvatica. There was a significant species X litter input rate (P = 0.017) interaction
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effect with the reduction in casting rates under deficit inputs more pronounced for A. glutinosa than for
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B. pendula.
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For the excess feeding treatments (Table 3), there was no significant overall CO2 regime litter effect
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on live weight change, with data for F. sylvatica highly variable and mostly negative. For excess with
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deficit data (excluding F. sylvatica), growth rates (Table 3) were significantly lower (-22%) for elevated
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compared with ambient CO2 litter (P = 0.011), and > 70% higher for excess compared with deficit
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inputs (P < 0.001). Elevated CO2 caused a more pronounced reduction in weight for A. glutinosa
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compared with B. pendula (interaction P = 0.023) and, with both species and input rate (Fig. 3b), with
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the benefits of increased litter input marked only for A. glutinosa (interaction P = 0.037).
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Growth rates were almost three times higher for A. glutinosa compared with B. pendula (P < 0.001).
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At the deficit litter input there was a more pronounced decrease in growth rates for A. glutinosa than
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for B. pendula litter (interaction P = 0.002) compared with the higher input.
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Correlation analyses of data (n = 24) from the excess feeding treatment indicated significant positive
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associations between weight change and litter nitrogen (r = 0.875, P < 0.001) and phosphorus (r =
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0.590, P = 0.002); negative relationships were obtained with C/N ratio (r = - 0.844, P < 0.001) and
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cellulose (r = - 0.688, P < 0.001). Litter nitrogen, C/N ratio, phosphorus and cellulose were
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significantly inter-correlated; with partial correlations, adjusting for all other variables, only litter N
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demonstrated an independent relationship (r = 0.753, P < 0.001) with weight change. Weight changes
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were also closely associated with fresh cast mineralisation constant (r = 0.726, P < 0.001).
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Cast respiration rate and mineralisation constant were markedly (ca. - 42% P = 0.023 and ca. - 29% P
321
= 0.036 respectively) lower for elevated compared with ambient CO2 litter (Table 4). Elevated CO2
322
litter reductions in mineralisation constant were more pronounced (interaction P < 0.001) for A.
323
glutinosa than for the other two litters (Fig. 4). Respiration also differed significantly between litter
324
species (P < 0.001) with A. glutinosa rates ca. 55% higher than those of B. pendula and more than
325
four times those of F. sylvatica. Cast organic C content was unaffected by CO2 regime but lower for F.
326
sylvatica. Litter species differences were also significant and followed the same trend described
327
previously for respiration.
328
4. Discussion
329
Elevated CO2 can modify woodland earthworm populations, with potential consequences for key
330
ecological processes (e.g. Scheu and Parkinson, 1994; Madritch and Lindroth, 2009). Our findings
331
from field and mesocosm studies show that earthworm responses to elevated CO2 are complex, being
332
characteristic of individual tree and earthworm ecotypes. Studies of other invertebrate responses to
333
elevated CO2 (e.g. herbivores – Couture et al., 2012) have found similar variations.
334
Although recent studies have argued the additional role of root and mycorrhizal derived inputs
335
(Godbold et al., 2006; Pollierer et al., 2007), leaf litter is recognised as a major C input to woodland
336
soils whose quality affects belowground food chains. Variations in litter quality, due to CO2 treatments,
337
tree species and their interactions exerted a strong influence on the growth of the anecic species L.
338
terrestris; in the field trends were similar but not always significant suggesting that other factors (e.g.
339
litter quantity) may have partially offset differences in litter quality.
340
Differences in litter quality and in earthworm responses attributable to elevated CO2 in the litter
341
feeding trial were broadly consistent with those in similar studies, at least for A. glutinosa and B.
342
pendula (Kasurinen et al., 2007; Meehan et al., 2010) and in wider studies of elevated CO2 effects on
343
detrivores (e.g. David and Gillon, 2009). Both earthworm studies attributed adverse elevated CO2
344
effects on L. terrestris to reductions in litter N concentrations. Our data support this conclusion as only
345
the litter N-earthworm association was independent of other quality parameters and only N was
346
significantly affected by CO2 regime. Markedly lower mineralisation constants in fresh casts with
347
elevated CO2 litter (Table 4) may indicate that this litter, in addition, was less readily metabolised by
348
micro-organisms than that grown under ambient CO 2; given the correlation between mineralisation
349
constants and earthworm weight changes, it is possible that this was true also for earthworm gut
350
micro-organisms.
351
In the field elevated CO2 adversely affected species identified as anecic (Table 2) dominated by L.
352
terrestris; pronounced CO2 effects on overall biomass were apparent mainly under A. glutinosa. L.
353
terrestris is a relatively long-lived species and most individuals were large sub-adult or mature, so
354
differences are unlikely to represent short-term responses reflecting conditions at sampling. Since
355
litter feeding earthworms are likely to most affected by the changes in litter quality, which might
356
precede other elevated CO2 effects in mature woodland, these findings are consistent with the
357
identification of L. terrestris in the field populations. Field outcomes for A. glutinosa were corroborated
358
by those from the feeding trial, but this was not the case for B. pendula. Indirect, beneficial elevated
359
CO2 impacts in the field, for example increased litterfall and higher soil moisture (Niklaus et al., 2007),
360
may have mitigated the smaller elevated CO2 induced reductions in B. pendula and F. sylvatica litter
361
quality; soil moisture can significantly influence woodland earthworm distribution (Stoscheck et al.,
362
2012). Feeding trial data suggest that increased ingestion (casting) rates with elevated CO2 litter, also
363
found by Zaller and Arnone (1997) in grassland, could compensate for reductions in resource quality
364
at elevated CO2. However, this input compensation occurred only where there were unrealistically
365
large increases in the supply of A. glutinosalitter compared with small increases in the field. With
366
limited litter supply or litter of lower intrinsic quality no such compensatory effect was observed either
367
in mesocosm or field investigations. Indeed earthworms responded to elevated CO2 by reducing
368
casting, possibly indicating a decline below a threshold in palatability. It is also possible changes in
369
leaf thickness or surface chemistry may have influenced feeding rates and this aspect of elevated
370
CO2 impacts on litter characteristics merits further investigation.
371
Any response among endogeic and epigeic species would have been delayed until tree derived
372
inputs, rather than existing soil C, became a more dominant food source (endogeic) and contributed
373
to the development of a litter layer (epigeic). Earthworm community effects beyond the anecic
374
grouping may therefore have become more pronounced had the field trial continued.
375
Marked reductions in fresh cast respiration with elevated CO 2 litter may represent a delayed or
376
reduced potential for organic matter mineralisation, with consequences for earthworm mediated C
377
sequestration (e.g. Don et al., 2008) and nutrient release (e.g. Madritch and Lindroth, 2009). Other
378
studies (e.g. Parsons et al., 2008) have found slower decomposition rates of elevated CO 2 litter and
379
grassland cast mineral N concentrations were 18% lower under high CO 2 conditions (Chevallier et al.,
380
2006).
381
In both field and feeding studies, anecic earthworms were favoured under A. glutinosa, followed by B.
382
pendula and F. sylvatica. This trend broadly reflected variations in litter quality but in the field would
383
have been reinforced by litterfall rates. As found by Hedde et al. (2007), fresh F. sylvatica litter was a
384
poor resource for L. terrestris and may have been unpalatable given the low mesocosm casting rates
385
and cast C contents observed. Since F. sylvatica supported field earthworm populations here and in
386
other studies (e.g. Cesarz et al., 2007), this litter may require partial decomposition before it is a
387
useful food resource for anecic earthworms.
388
Whilst litter quality responses to elevated CO2 are generally negative, especially for N, the extent of
389
these effects does vary between species (e.g. Cotrufo et al., 1994; Norby et al., 2001). For this
390
reason, it is difficult to extrapolate findings to mixed deciduous woodland. Nevertheless, given the
391
important role of L. terrestris in woodland soil hydrology (e.g. Shutz et al., 2008), in broader
392
ecosystem function (Fisk et al., 2004; Suarez et al., 2006) and for understory plant communities
393
(Frelich et al., 2006), any effects on their abundance would have significant hydrological and
394
ecological implications.
395
5. Conclusions
396
In woodlands earthworm populations and ecosystem services mediated by them will alter where
397
elevated atmospheric CO2 induces marked changes in litter characteristics and this impact will occur
398
within several seasons. It is of note that reductions in anecic earthworms occurred even under N-
399
fixing A. glutinosa, although this may be explained by the absence of any increase in N-fixation at
400
elevated CO2 found in the field study site (Hoosbeck et al., 2011). Broader ecosystem responses,
401
such as shifts in tree species composition attributable to climate change (rainfall and temperature)
402
impacts (e.g. Ferreira et al., 2010; Rouifed et al., 2010) may be at least as important as intra-specific
403
CO2 responses in the longer term.
404
Climate variations may also directly affect earthworm populations and interact with litter quality as
405
found by David and Gillon (2009) for millipedes. A fuller understanding of earthworm and other
406
decomposer community responses to CO2 variations would allow wider ecosystem consequences to
407
be predicted with more confidence. For ecosystem assemblages, these consequences may depend
408
on unique combinations of inter- and intra-specific traits (Bradley and Pregitzer 2007). Earthworm
409
responses to higher atmospheric CO2 may affect soil-understory-tree interactions and ultimately
410
atmospheric feedbacks. An improved understanding of these feedbacks would allow the inclusion of
411
soil animal impacts on decomposition in future modelling of global change scenarios (Wall et al.,
412
2008).
413
Acknowledgements
414
This study was undertaken with the support of CIRRE under the Bangor-Aberystwyth University
415
initiative. Technical support from T. Castle and A. Vaughan, and figure preparation from J. Bussell is
416
gratefully acknowledged. Anonymous reviewers for suggested improvements to the MS.
417
References
418
Allard, V., Newton, P.C.D., Lieffering, M., Soussana, J-F., Grieu, P., Matthew, C., 2004. Elevated CO2
419
effects on decomposition processes in a grazed grassland. Glob. Change Biol. 10, 1553–1564.
420
Arnone III, J.A., Zaller, J.G., Hofer, G., Schmid, B., Ko¨rner, C., 2013. Loss of plant biodiversity
421
eliminates stimulatory effect of elevated CO2 on earthworm activity in grasslands. Oecologia 171,
422
613–622.
423
Blaschke, L., Forstreuter, M., Sheppard, L.J., Leith, I.K., Murray, M.B., Polle, A. 2002. Lignification in
424
beech (Fagus sylvatica) grown at elevated CO2 concentrations: interaction with nutrient availability
425
and leaf maturation. Tree Physiol. 22, 469-477.
426
427
428
429
430
431
Boerner, R.E.J., Rebbeck, J. 1995. Decomposition and nitrogen release from leaves of 3 hardwood
species grown under elevated O3 and/or CO2. Plant Soil 170,149-157.
Bouche, M 1977. Strategies lombriciennes. In: Lohm U., Persson, T. (Eds.), Soil Organisms as
Components of Ecosystems, Biol. Bull. (Stockholm) 25, 122-132.
Bradley, K.L., Pregitzer, K.S. 2007. Ecosystem assembly and terrestrial carbon balance under
elevated CO2. Trends Ecol. Evol. 22, 538-547.
432
Campbell, C.D., Chapman, S.J., Cameron, C.M., Davidson, M.S., Potts, J.M., 2003. A rapid microtiter
433
plate method to measure carbon dioxide evolved from carbon substrate amendments so as to
434
determine the physiological profiles of soil microbial communities by using whole soil. Appl.
435
Environ. Microbiol. 69, 3593-3599.
436
437
Cesarz, S., Fahrenholz, N., Migge-Kleian, S., Platner, M., Schaefer, C. 2007. Earthworm communities
in relation to tree diversity in a deciduous forest. Euro. J. Soil Biol. 43, 61-67.
438
Chevallier, A.J.T., Lieffering, M., Carran, R.A., Newton, P.C.D. 2006. Mineral nitrogen cycling through
439
earthworm casts in a grazed pasture under elevated atmospheric CO 2. Glob. Change Biol. 12, 56-
440
60.
441
442
443
444
Cotrufo, M.F., Ineson, P., Rowland, A.P. 1994. Decomposition of tree leaf litters grown under elevated
CO2 – effect of litter quality. Plant Soil 163, 121-130.
Couture, J.J., Meehan, T.D., Lindroth, R.L. 2012. Atmospheric change alters foliar quality of host
trees and performance of two outbreak insect species. Oecologia 168, 863–876.
445
David, J.F., Gillon, D. 2009. Combined effects of elevated temperatures and reduced leaf litter quality
446
on the life-history parameters of a saprophagous macroarthropod. Glob. Change Biol. 15, 156-165.
447
Don, A., Steinberg, B., Scho¨ning, I., Pritsch, K., Joschko, M., Gleixner, G., Schulze, E-D. 2008.
448
449
450
451
452
Organic carbon sequestration in earthworm burrows. Soil Biol. Biochem. 40, 1803–1812.
Edwards, C.A., Bohlen, P.J. (1996). Biology and Ecology of Earthworms, Third ed, Chapman & Hall,
London.
Felten D, Emmerling C (2009) Earthworm burrowing behaviour in 2D terraria with single- and multispecies assemblages. Biol. Fertil. Soils 45, 789-797.
453
Ferreira, V., Goncalves, A.L., Godbold, D.L., Canhoto, C. 2010. Effect of elevated atmospheric CO2
454
on the performance of an aquatic detrivore through changes in water temperature and litter quality.
455
Glob. Change Biol. 16, 3284-3296.
456
Fisichelli, N.A., Frelich, L.E., Reich, P.B., Eisenhauer, N. 2013. Linking direct and indirect pathways
457
mediating earthworms, deer, and understory composition in Great Lakes forests. Biol. Invas. 15,
458
1057-1066.
459
460
Fisk, M.G., Fahey, T.J., Groffman, P.M., Bohlen, P.J. 2004. Earthworm invasion, fine-root
distributions, and soil respiration in North temperate forests. Ecosys. 7, 55-62.
461
Frelich, L.E., Hale, C.M., Scheu, S., Holdsworth, A.R., Heneghan, L., Bohlen, P.J., Reich, P.B. 2006.
462
Earthworm invasion into previously earthworm-free temperate and boreal forests. Biol. Invas. 8,
463
1235-1245.
464
465
Gilsotres, F., Trasarcepeda, M.C., Ciardi, C., Ceccanti, B., Leiros, M.C., 1992. Biochemical
characterisation of biological activity in very young mine soils. Biol. Fertil. Soils 13, 25-30.
466
Godbold, D.L., Hoosbeek, M.R., Lukac, M., Cotrufo, M.F., Janssens, I.A., Ceulemans, R., Polle, A.,
467
Velthorst, E.J., Scarascia-Mugnozza, G., De Angelis, P., Miglietta, F., Peressotti, A. 2006.
468
Mycorrhizal hyphal turnover as a dominant process for carbon input into soil organic matter. Plant
469
Soil 281, 15-24.
470
Gorres, J.H., Savin, M.C., Amador, J.A. 2001. Soil micropore structure and carbon mineralization in
471
burrows and casts of an anecic earthworm (Lumbricus terrestris) Soil Biol. Biochem. 33,1881-
472
1887.
473
474
Harborne, J.B., Harborne, A.J. 1998. Phytochemical Methods: A Guide to Modern Techniques of
Plant Analysis. Taylor and Francis Group, London, p302.
475
Hedde, M., Bureau, F., Akpa-Vinceslas, M., Aubert, M., Decaens, T. 2007. Beech leaf degradation in
476
laboratory experiments: Effects of eight detritivorous invertebrate species. Appl. Soil Ecol. 35, 291-
477
301.
478
Hoosbeck, M.R., Lukac, M., Velthorts, E., Smith, A.R., Godbold, D.L. 2011. Free atmospheric CO2
479
enrichment increased above ground biomass but did not affect symbiotic N 2-fixation and soil
480
carbon dynamics in a mixed deciduous stand in Wales. Biogeosciences 8, 353-364.
481
James S., Porco D., Decaëns T., Richard B., Rougerie R., Erséus C. (2010) DNA Barcoding Reveals
482
Cryptic Diversity in Lumbricus terrestris L., 1758 (Clitellata): Resurrection of L. herculeus (Savigny,
483
1826). PLoS One, 5, e15629.
484
485
Karl, T.R., Melillo, J.M., Peterson, T.C. (eds.) 2009. Global Climate Change Impacts in the United
States. Cambridge University Press, New York.
486
Kasurinen, A., Peltonen, P.A., Julkunen-Tiitto, R., Vapaavuori, E., Nuutinen, V., Holopainen, T.,
487
Holopainen, J.K. 2007. Effects of elevated CO2 and O3 on leaf litter phenolics and subsequent
488
performance of litter-feeding soil macrofauna. Plant Soil 292, 25-43.
489
490
491
492
493
494
Kruuk, H., Parish, T. 1981. Feeding specialization of the European badger Meles meles in Scotland.
J. Anim. Ecol. 50, 773-788.
Lindroth, R.L., Kinney, K.K., Platz, C.L. 1993. Responses of deciduous trees to elevated atmospheric
CO2: productivity, phytochemistry and insect performance. Ecology 74, 763-777.
Madritch, M.D., Lindroth, R.L. 2009. Removal of invasive shrubs reduces exotic earthworm
populations. Biol. Invas. 11, 663-671.
495
496
497
498
Marashi, A.R.A., Scullion, J. 2003. Earthworm casts form stable aggregates in physically degraded
soils. Biol. Fertil. Soils 37, 375–380.
Meehan, T.D., Crossley, M.S., Lindroth, R.L. 2010. Impacts of elevated CO2 and O3 on aspen leaf
litter chemistry and earthworm and springtail productivity. Soil Biol. Biochem. 42, 1132-1137.
499
Miglietta, F., Peressotti, A., Vaccari, F.P., Zaldei, A., deAngelis, P., Scarascia-Mugnozza, G. 2001.
500
Free-air CO2 enrichment (FACE) of a poplar plantation: the POPFACE fumigation system. New
501
Phytol. 150, 465-476.
502
503
504
505
Murphy J, Riley JP (1962) A modified single solution method for the determination of phosphate in
natural waters. Analytica Chimica Acta 27: 31-36.
Neilson, R., Boag, B. 2003. Feeding preferences of some earthworm species common to upland
pastures in Scotland. Pedobiol. 47, 1-8.
506
Neirynck, J., Mirtcheva, S., Sioen, G., Lust, N. 2000. Impact of Tilia platyphyllos Scop., Fraxinus
507
excelsior L., Acer pseudoplatanus L., Quercus robur L. and Fagus sylvatica L. on earthworm
508
biomass and physic-chemical properties of a loamy topsoil. Forest Ecol. Manag. 133, 275-286.
509
Niklaus, P.A., Glockler, E., Siegwolf, R., Korner, C. 2001. Carbon allocation in calcareous grassland
510
under elevated CO2: a combined C-13 pulse-labelling/soil physical fractionation study. Funct. Ecol.
511
15, 43-50.
512
Niklaus, P.A., Alphei, J., Kampichler, C., Kandeler, E., Korner, C., Tscherko, D., Wohlfender, M. 2007.
513
Interactive effects of plant species diversity and elevated CO 2 on soil biota and nutrient cycling.
514
Ecology 88, 3153-3163.
515
516
517
518
Norby, R.J., Cotrufo, M.F., Ineson, P., O'Neill, E.G., Canadell, J.G., 2001. Elevated CO 2, litter
chemistry, and decomposition: a synthesis. Oecologia 127, 153-165.
OECD 1984. Guideline for Testing Chemicals; Earthworm Acute Toxicity. Test No. 207, Organisation
for Economic Co-operation and Development. Paris.
519
Oksanen, E., Oksanen, E., Riikonen, J., Kaakinen, S., Holopainen, T., Vapaavuori, E. 2005. Structural
520
characteristics and chemical composition of birch (Betula pendula) leaves are modified by
521
increasing CO2 and ozone. Glob. Change Biol. 11, 732-748.
522
Parsons, W.F.J., Bockheim, J.G., Lindroth, R.L. 2008. Independent, interactive, and species-specific
523
responses of leaf litter decomposition to elevated CO 2 and O3 in a northern hardwood forest.
524
Ecosys. 11, 505-519.
525
526
527
528
Peach, W.J., Robinson, R.A., Murray, K.A. 2004. Demographic and environmental causes of the
decline of rural Song Thrushes Turdus philomelos in lowland Britain. Ibis 146, 50-59.
Pollierer, M.M., Langel, R., Koerner, C., Maraun, M., Scheu, S. 2007. The underestimated importance
of belowground carbon input for forest soil animal food webs. Ecol. Lett. 10, 729-736.
529
Rouifed, S., Handa, I.T., David, J-F., Hattenschwiler, S, 2010. The importance of biotic factors in
530
predicting global change effects on decomposition of temperate forest leaf litter. Oecologia 163,
531
247-256.
532
533
534
535
536
537
Sackett, T.E., Smith, S.M., Basiliko, N. 2013 Indirect and direct effects of exotic earthworms on soil
nutrient and carbon pools in North American temperate forests. Soil Biol. Biochem. 57, 459-467.
Scheu, S., Parkinson, D. 1994. Effects of invasion of an aspen forest (Canada) by Dendrobaena
octaedra (Lumbricidae) on plant growth. Ecology 75, 2348-2361.
Schutz, K., Nagel, P., Dill, A., Scheu, S. 2008. Structure and functioning of earthworm communities in
woodland flooding systems used for drinking water production. Appl. Soil Ecol. 39, 342-351.
538
Scullion, J., Goodacre, R., Elliott, G., Huang, W., Worgan, H., Gwynn-Jones, D., Griffith, G., Darby,
539
R., Bailey, M., Clegg, C., Draper, J. 2003. Food quality and microbial succession in ageing
540
earthworm casts: standard microbial indices and metabolic fingerprinting. Pedobiol. 47, 440-446.
541
Sims, R.W., Gerard, B.M. 1999. Earthworms.. Barnes R.S.K., Crothers, J.H. (Eds.), Synopsis of the
542
British Fauna (New Series) No. 31 (revised), Linnean Society of London,Field Studies Council,
543
Shrewsbury, UK.
544
Stoschecka, L.M., Sherman, R.E., Suarez, E.R., Fahey, T.J. 2012. Exotic earthworm distributions did
545
not expand over a decade in a hardwood forest in New York state. Appl. Soil Ecol. 62, 124– 130
546
Suarez, E.R., Fahey, T.J., Yavitt, J.B., Groffman, P.M., Bohlen, P.J. 2006. Patterns of litter
547
disappearance in a northern hardwood forest invaded by exotic earthworms. Ecol. Appl. 16, 154-
548
165.
549
Teklehaimanot, Z., Jones, M., Sinclair, F.L. 2002. Tree and livestock productivity in relation to tree
550
planting configuration in a silvopastoral system in North Wales, UK. Agroforestry Systems 56, 47-
551
55.
552
553
Tiunov, A.V., Scheu, S. 2000. Microfungal communities in soil, litter and casts of Lumbricus terrestris
L. (Lumbricidae): a laboratory experiment. Appl. Soil Ecol. 14, 17-26.
554
555
556
557
558
559
560
561
562
563
564
565
566
567
568
569
570
571
572
573
Van Soest, P.J., Wine, R.H. 1967. The use of detergents in analysis of fibrous feeds 4. Determination
of plant cell-wall constituents. J. Assoc Official Analytical Chem. 50, 50-55.
Wall, D.H., Bradford, M.A., St John, M.G., et al., 2008. Global decomposition experiment shows soil
animal impacts on decomposition are climate-dependent. Glob. Change Biol. 14, 2661-2677.
Yeats, G.W., Newton, P.C.D., Ross, D.J. 2003. Significant changes in soil microfauna in grazed
pasture under elevated carbon dioxide. Biol. Fertil. Soils 38, 319-326.
Zaller, J.G., Arnone, J.A. 1997. Activity of surface-casting earthworms in a calcareous grassland
under elevated atmospheric CO2. Oecologia 111, 249-254.