F A C U L T Y O F S C I E N C E U N I V E R S I T Y O F C O P E N H A G E N

PhD thesis Marie Dam

Global change effects on plant-soil interactions

Academic advisor: Søren Christensen

Submitted: 23/12/14

Name of department: Department of Biology Author): Marie Dam Title / Subtitle: Global change effects on plant-soil interactions Academic advisor: Søren Christensen Submitted: 12. måned 2014 Cover photos: Marie Merrild and Marie Dam

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Preface and acknowledgements

This dissertation is the result of a three year Ph.D. project completed at the Terrestrial Ecology Section,

Department of Biology, University of Copenhagen, as part of the Danish Centre of Excellence, CLIMAITE,

funded by the Villum Kann Rasmussen foundation. Work done as a visiting scholar at the Department of

Biology, Colorado State University is also included.

I owe a debt of gratitude to several people who has aided this project. First of all, to my supervisor Søren

Christensen for guidance, support and valuable discussions challenging both of our perspectives. Thank you,

for always having an open door. Secondly, to Mette Vestergård for great mentorship, common love of

nematodes and for being a friend. And to the rest of the Terrestrial Ecology Section, particularly the soil

biology group with whom I have been associated with for a number of years and value highly. I am also

grateful to Diana Wall and the rest of Wall lab at CSU, for inspiring collaboration and for showing me new

methods in nematology.

Being a part of the CLIMAITE project has been fruitful in a great number of ways. I am thankful especially

to Martin Holmstrup for visits to his lab and collaborating on soil fauna data. This is what led me to the new

and exciting world of oribatology. I am thankful also to my cherished Ph.D. colleague, Marie Merrild, and

the rest of the CLIMAITE Ph.D. group, who always offered rewarding meetings both scientifically and

socially. And a special thanks to the people who kept CLIMAITE going, both technically and scientifically,

particularly Claus Beier.

Finally, I absolutely would not have been able to complete this dissertation without my husband Andreas’

eternal support. Thank you, my love, for keeping our family together. And thank you to the rest of my family

for beloved emotional and practical support.

Marie Dam

Copenhagen, December 2014

Table of Contents Summary............................................................................................................................................................ 2

Sammendrag ...................................................................................................................................................... 3

List of papers ..................................................................................................................................................... 4

General introduction .......................................................................................................................................... 5

Soil organisms ............................................................................................................................................... 5

Microorganisms – fungi and bacteria ........................................................................................................ 5

Microfauna – nematodes and protozoa ...................................................................................................... 6

Acari – oribatids ........................................................................................................................................ 6

Soil food web ............................................................................................................................................. 7

Chains and pyramids ................................................................................................................................. 7

Top-down and bottom-up controls ............................................................................................................ 8

Nematodes as indicators ............................................................................................................................ 8

Decomposition and soil fauna ....................................................................................................................... 8

The plant-soil system ..................................................................................................................................... 9

Plant allocation of resources ...................................................................................................................... 9

Root input to soil ..................................................................................................................................... 10

Quality of organic input ........................................................................................................................... 10

Disturbance of the plant-soil system ....................................................................................................... 11

Temperate grassland and shrubland ecosystems ......................................................................................... 11

Preserving grasslands requires some measure of disturbance ................................................................. 11

Growth strategy of grasses ...................................................................................................................... 12

Dry heathland is a distinct type of shrubland. ......................................................................................... 12

Global change .............................................................................................................................................. 13

Fire ........................................................................................................................................................... 13

CO2, warming and precipitation .............................................................................................................. 14

Research objectives and aims .......................................................................................................................... 15 References……………………………………………………………………………………………………16

Paper I…………………..………………………………………………………………..…………….…… 23

Paper II……….………….………………………………………………………………….………...…….. 41

Paper III………..………….……………………………………………………………………….…...…… 59

Paper IV………..………….………..………………………………………………………….……………. 87

General discussion……..….……………………………………………………………………….………... 98

Conclusions and perspectives…..…………………………………………………………..……….…..…. 100

References…………………….……………………………………………………………………..…….. 102

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Summary Global change is expected to increasingly affect composition and functioning of soil communities. In

terrestrial ecosystems, the plant-soil interactions will be of particular importance for the ecosystem response,

including feed-back responses that may further increase climate change. The aim of this dissertation has been

to determine how soil food web structure and function is affected when the quantity and quality of plant

input is altered under global change.

By studying the abundance and composition of soil organisms, particularly those in the rhizosphere, closely

associated with living plants, we are able to determine effects of global change on the plant-soil system. By

extraction and microscopy of nematode communities, we are able to characterize the trophic structure of a

significant part of the rhizosphere community. The work compiled for this dissertation is based on field

experiments in temperate heathland and grassland. This includes characterization of a decomposer system

under global change as defined by the plants present (Paper I), understanding the mechanism shaping the

system response by manipulating input from living plants (Paper II) or by manipulating plant input by

burning (Paper III). Furthermore, by way of meta-analysis, the role of organisms in global change effects on

ecosystem function is modelled (Paper IV).

Among CO2, warming and summer drought, CO2 is the factor most consistently impacting soil organisms.

CO2 increases abundance of microorganisms and nematodes, and increase flow through the fungal

decomposition pathway, most likely due to an increase of plant resources allocated belowground and an

increased C:N ratio of the plant input. In line with this, CO2 increase is found to generally correlate with

plant biomass in the meta-analysis. Hence, plant allocation of resources is of significant impact

belowground, but the allocation is also influenced by other factors: When grasses are defoliated in the

growing season, there is an increased shoot regrowth at the expense of the belowground system. Both

microorganisms and nematodes are reduced, despite positive CO2 effects. Furthermore, the plant functional

type (shrub or grass) is more strongly determining the rhizosphere community structure than any global

change factor. Frequent burning of prairie vegetation changes the soil community to an extent that alters the

decomposition rate.

Together, these results suggest that not only the global change effects on established ecosystems, but also the

global change effects on plant community composition as well as land use management may determine the

composition and function of soil food webs in the future.

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Sammendrag Globale forandringer af klima, miljø og landskabsanvendelse forventes i stigende grad at påvirke

organismesammensætningen og processerne i jordens økosystemer. I de terrestriske økosystemer vil

samspillet mellem planter og organismer i den jord, de vokser i, være af særlig betydning for

økosystemresponsen. Med hensyn til klimaforandringer, kan dette samspil og de processer, der foregår her,

have betydning for feed-back mekanismer, der yderligere forstærker de globale forandringer. Formålet med

denne afhandling har i særlig grad været at undersøge, hvordan jordbundsfødenettet og dets funktion

påvirkes når tilførslen af organisk substrat via planter ændres af miljøforandringer eller -forstyrrelser.

Ved at undersøge de trofiske strukturer og antallet af organismerne i jordbunden, særligt rhizosfæren,

kan vi analysere jordbundssamfundene og deres respons på miljøforandringer og forstyrrelser i den

overjordiske del af økosystemet. Ved uddrivning og mikroskopi af nematoder kan den trofiske

sammensætning af en betydelig del af rhizosfærefødenettet karakteriseres. For at forstå de naturlige systemer

og de mekanismer, der afgør systemernes funktion og respons på miljøændringer, baserer denne afhandling

sig på eksperimenter i felten i hede og græsland. Dette omfatter karakterisering af jordbundssamfund knyttet

til forskellige plantetyper under indflydelse af forhøjet CO2 og forandringer af klimaet (Paper I),

undersøgelse af de mekanismer, der former disse samfund, ved manipulation af materialetilførslen fra

levende planter (Paper II), og ved manipulation af tilførslen af dødt plantemateriale ved afbrænding (Paper

III). Desuden kan metanalyse af data for globale miljøændringer, mikroorganismer og økosystem processer

vise generelle tendenser i mikroorganismernes rolle i økosystemernes respons (Paper IV).

Blandt CO2, opvarmning og sommertørke, er CO2 den faktor, der mest gennemgående påvirker

jordbundsorganismerne. CO2 øger antallet af miroorganismer og nematoder, og driver i højere grad

omsætningen igennem den svampe-definerede del af nedbryderfødenettet, sandsynligvis på grund af

allokering af øgede planteressourcer til rodsystemet og øget C:N forhold i plantematerialet. Således

korrelerer forhøjet CO2 også på tværs af studier i metaanalysen med øget plantebiomasse. Plantens

prioritering af ressourcer over og under jorden er altså afgørende, men påvirkes også af andre faktorer.

Eksempelvis er der betydelig skudgenvækst af græsser efter defoliering i vækstsæsonen og deraf følger et

reduceret antal mikroorganismer og nematoder i rhizosfæren, på trods af positive CO2 effekter. Desuden er

plantetypen, som jordbundsorganismerne lever i interaktion med, mere betydende for

organismesammensætningen af både mikroorganismer og nematoder end CO2. Jævnlig afbrænding af

plantevæksten på prærien ændrer nedbrydersamfundet så meget, at kulstofomsætningen påvirkes.

konklusioner.

Samlet set må vi altså forvente, at ikke blot de atmosfæriske og klimatiske ændringers påvirkninger

af eksisterende økosystemer, men også effekterne på plantesammensætningen i økosystemerne samt

menneskets forvaltning heraf kan påvirke jordbundsorganismernes sammensætning og funktion i fremtidens

klima.

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List of papers

PAPER I

Dam, M., Bergmark, L., Vestergård, M. Elevated CO2 increases fungal decomposer channel in rhizospheres

of contrasting plant species. Manuscript in preparation.

PAPER II

Dam, M. & Christensen, S. Elevated CO2 stimulates soil biota – defoliation modifies effects depending on

season. Manuscript in preparation.

PAPER III

Dam, M., Soong, J., Wall, D.H., Cotrufo, M.F. Fire in the tallgrass prairie: effects on soil communities and

trophic transfers of carbon during litter and pyrogenic organic matter decomposition. Manuscript in

preparation.

PAPER IV

García-Palacios, P., Vandegehuchte, M. L., Ashley Shaw, E., Dam, M., Post, K. H., Ramirez, K. S., Milano

de Tomasel, C., Wall, D. H. (2014). Are there links between responses of soil microbes and ecosystem

functioning to elevated CO2, N deposition and warming? A global perspective. Global Change

Biology, 1–11. doi:10.1111/gcb.12788

Other papers not included in this dissertation: Dam, M., Vestergård, M., & Christensen, S. (2012). Freezing eliminates efficient colonizers from nematode

communities in frost-free temperate soils. Soil Biology and Biochemistry, 48, 167–174.

doi:10.1016/j.soilbio.2012.01.017

Christensen, S., Dam, M., Vestergård, M., Petersen, S. O., Olesen, J. E., & Schjønning, P. (2012). Specific

antibiotics and nematode trophic groups agree in assessing fungal:bacterial activity in agricultural soil.

Soil Biology and Biochemistry, 55, 17–19. doi:10.1016/j.soilbio.2012.05.018

Christensen, S., Dam, M., Ransijn, J., Arndal, M. F., Beier, C., & Vestergård, M. Soil nematodes under grass

increase at elevated CO2 when moisture is low. Manuscript in preparation

Dam, M., Zaytsev, A., Ehlers, B. K., Georgieva, S., & Holmstrup, M.. Community analysis of soil fauna

under D. flexuosa in a changing climate. Manuscript in preparation

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General introduction My research focus is, and has been, soil organisms – nematodes especially. I pursued an understanding of

how the interactions of the soil micro food web respond to impact on the plant-soil system. I have studied

this in temperate grassland and shrubland and have focused on the impacts of certain global change factors.

Soil organisms The majority of Earth’s species may be found in the soil (Wardle 2002). Soil organisms regulate major

ecosystem processes, such as organic matter turnover (Bradford et al. 2007) and nutrient cycling (Standing et

al. 2006), and act as drivers of vegetation change (De Deyn et al. 2003). The soil organisms studied for this

dissertation include fungi, bacteria, nematodes in particular, and oribatid mites.

Microorganisms – fungi and bacteria

The two most abundant groups of microorganisms are bacteria and fungi. Fungi are eukaryotic with a

mycelial morphology comprising a mass of hyphae that enclose multi-nucleated cytoplasm. Networks of

hyphae enable the fungi to grow into new substrates and transport materials through the soil over distances of

centimeters to meters. Fungi have enzymes capable of breaking down virtually all classes of organic

compounds, and have the competitive advantage over bacteria in decomposing substrates of low nutrient

concentration because of their ability to import nitrogen and phosphorus via their hyphal network. Fungi

comprise 60-90 % of the microbial biomass in forest soil where the litter has a high lignin and low nitrogen

concentration (Chapin III et al. 2002). They have the competitive advantage over bacteria at low pH (Rousk

et al. 2010), common in forest and heathland soils. In grasslands with a higher pH, fungi usually make up

half the microbial biomass. Mycorrhiza is a symbiotic association between plant roots and fungi in which the

plant gains nutrients from the fungus in return for carbohydrates. Arbuscular mycorrhiza fungi are

widespread in herbaceous grassland species whereas ericaceous plants such as Calluna vulgaris found in

temperate heathland are often colonized by ericoid mycorrhiza (Michelsen et al. 1998).

Bacteria are single-celled prokaryotes that in some cases search for food by movement with flagella

but primarily rely on substrate diffusion through water. The small size and high surface to volume ratio

enable them to absorb soluble substrates rapidly and grow and divide quickly in substrate-rich environments.

Bacteria therefore often dominate in the rhizosphere where labile substrates are abundant. Certain groups of

bacteria perform specialized functions of great ecological significance, such as the chemoautotrophic

nitrifiers that are of importance for nutrient cycling. An important consequence of the relative immobility of

bacteria is that the colony eventually exhausts the substrate in its immediate environment (Chapin III et al.

2002). When this happens, they become inactive and reduce their respiration to negligible rates. Bacteria

may be inactive for years. Between 50 and 80% of the bacteria in soils are metabolically inactive (Norton &

Firestone 1991). Consequently, DNA and culturing techniques are better indices of potential activity than of

the actual metabolism.

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Microfauna – nematodes and protozoa

Nematodes, together with protozoans make up the soil microfauna. Both are formally aquatic organisms that

move, feed, and reproduce in the water film surrounding soil particles. They ingest microorganisms or feed

on other animals within the same general size category (Hunt et al. 1987; Ferris 2010).

Nematodes are unsegmented worms (roundworms) with the gut and body wall being concentric

tubes. Soil nematodes are typically 1 mm in length, with a 50 um

diameter. The nematodes are a diverse group in which the species

specialize on bacteria, fungi, roots or other soil animals. They can

be classified into trophic groups based on their mouthparts

(Yeates et al. 1993). Nematodes are the most numerous

multicellular animals in soil, with a biomass of 0.2 ton per ha in

some soils (Killham 1994), and are critical to the carbon and

nutrient dynamics of soils. Microbes contain 70 to 80% of the

labile carbon and nitrogen in soils (Chapin III et al. 2002), so

variation in predation/grazing rates of microbes by animals

dramatically alter carbon and nitrogen turnover in soils. Nematode

species have generation times ranging from weeks (sometimes days

under optimal conditions) to months or even years, depending on

their life-history strategy. Fast-growing species are characterized by

a short life-cycle, high colonization ability and their tolerance to

disturbance (Bongers 1990). These species are often bacterivores and live in ephemeral habitats such as the

rhizosphere of a growing root tip (Bongers 1990), and thus most closely mirror the bloom of bacteria or

respond most rapidly to active plant growth. Species with a low reproduction rate in general have long life-

cycles, low colonization ability and are sensitive to disturbance. They never belong to the dominant species

in a sample, they hardly fluctuate in number during the year, and they usually belong to the higher trophic

levels and live as omnivores or predators (Bongers 1990; Yeates et al. 1993).

Acari – oribatids

Soil mites (acari) are generally 0.5-1 mm long and unlike

nematodes, they are true terrestrial animals living in air-filled

soil pores. Oribatidae is the dominant order of acari in most

soils (Petersen & Luxton 1982), and they consume mainly dead

plants parts or fungi, however there are some exceptional

predators and scavengers (Gercócs & Hufnagel 2009). Unlike

other groups of soil microarthropods, the feeding preferences of

oribatids are thought to be relatively species-specific (Gjelstrup

& Petersen 1987). They have low metabolic rates, slow development and low fecundity, and thus are

Helicotylenchus sp., root feeding nematode.

Photo by Ulrich Zunke, nemapix

Acrobeles sp., Bacterial feeding nematode Photo by Paolo Vieira, nemapix

Pergalumna nervosa, oribatid mite. Photo by Marie Dam

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considered the K-strategists of microarthropods. They are therefore sensitive to disturbance, and show

diversity and abundance responses to such (Lindberg & Bengtsson 2005).

Soil food web

Chains and pyramids

Trophic dynamics control the movement of C, nutrients and energy among organisms in an ecosystem. A

group of organisms that are linked together by a transfer of energy and nutrients represents a food chain, and

organisms that obtain their energy with the same number of transfers from plant or detritus belong to the

same level. In the detritus-based system, bacteria and fungi directly break down detritus and absorb the

products for their own growth and maintenance. These primary decomposers are the first trophic level in the

detritus-based food chains and are the foundation for series of trophic transfers among the soil animals. The

transfers are interconnected in complex food webs (Hunt et al. 1987), and energy losses at each trophic

transfer limit the production of higher trophic levels. Not all the biomass that is produced at one trophic level

is consumed at the next level, and only some of the assimilated energy is converted to animal production.

Consequently, a relatively small fraction of the energy available as food at one trophic level is converted into

production at the next link of the chain. This has profound consequences for the trophic structure of

ecosystems because each link in the food chain has less energy available to it than did the preceding link,

creating an ecological pyramid first introduced by Elton (1927). The amount of organic input to the soil thus

constrains the amount of energy that is available at successive trophic levels and could influence the number

of trophic levels that an ecosystem can support. Consequently, decomposer food chains tend to be longer in

more productive ecosystems (Moore & de Ruiter 2000).

Figure 1 Soil food web. Adapted from de Ruiter et al. 1995

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Top-down and bottom-up controls

The quantitative energy flows through decomposer food webs are generally poorly known, and the regulation

of energy and nutrient flow is complicated and varies among ecosystems. There are two opposite

mechanisms between which the controls of flow range: Bottom-up and top-down. The food availability at the

base of the food chain limits the production of higher trophic levels through bottom-up controls. In contrast,

predators that regulate the abundance of their prey exert top-down control. Most food webs are regulated by

some combination of bottom-up and top-down controls (Wardle et al. 1998), and the relative importance of

these controls varies both temporally and spatially (Polis 1999). A trophic cascade occurs when changes in

abundance at one trophic level alter the abundance of other trophic levels across more than one link in the

food web. If predation by organisms at one trophic level reduces the density of their prey, this in turn

releases the prey at the level below from consumer control (Pace et al. 1999). Mites, for example, can control

the decomposition of buried leaf litter by consuming bacterivore nematodes (Killham 1994). Trophic

cascades cause an alteration among trophic levels in biomass of organisms.

Nematodes as indicators

The quantification of the abundance of active organisms in the diverse soil food web requires many different

techniques. Soluble organic substrates are absorbed by bacteria and fungi, while fungi usually also degrade

more recalcitrant sources, so these organisms are potential indicators of the nature of incoming substrate.

However, methods of microbial biomass determination do not necessarily indicate their actual metabolic

activity, and furthermore, organisms at higher trophic levels exert important top-down controls on microbes.

Assessment of microbial biomass alone may therefore miss ecologically important changes in energy flow

within the soil food web (Kampichler et al. 1998). Guilds of nematodes that feed on bacteria and fungi are

responsive to changes in abundance of their food (Bongers & Ferris 2006) and through direct herbivory,

plant-feeding nematodes also contribute to soil food web resources. Thus, analysis of the active nematode

community live extracted from a sample provides indication of carbon flow through an important herbivore

channel and through channels mediated by bacteria and fungi (Ferris & Bongers 2006). In Christensen et al.

(2012) we confirmed that ratios of fungivore nematodes to bacterivore nematodes reflects the ratio of fungal

to bacterial activity in soil.

Decomposition and soil fauna The physical and chemical degradation of dead organic matter by decomposition produces CO2, mineral

nutrients in forms that can be used for plant and microbial production, and a remnant pool of recalcitrant

organic compounds that are resistant to further microbial breakdown. Net mineralization occur when

microbial growth is limited more strongly by carbon than by N, whereas net immobilization occurs when

microbial communities are nitrogen limited (Chapin III et al. 2002). It is the need for the energy locked up in

the C-H bond of organic molecules that drives the decomposition process, and hence the cycling of nutrients

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in soil. Any distinctive change in the carbon cycle caused by e.g. by climate change will inevitably bring

about fundamental changes to other nutrient cycles of the plant-soil system. If climate change causes a

lasting alteration of the balance between NPP and decomposition, the CO2 concentration in the atmosphere

would be further altered and therefore so would the rate of climate change.

Decomposition is largely a consequence of the feeding activity of soil animals and microbes, and the

forces that shape it are those that maximize the growth, survival and reproduction of those soil organisms.

Soil organisms at higher trophic levels, such as protozoans, nematodes, mites, and springtails, influence the

activity and turnover of the microflora through grazing (Ingham et al. 1985; Hedlund & Öhrn 2000).

Furthermore, these organisms excrete nutrients that are in excess when they metabolize much of the

microbial carbon from their food to CO2 to support their energetic costs of maintenance, growth and

reproduction (Chapin III et al. 2002). These nutrients become available for plant and microbial uptake.

Hence, the ecological importance of grazing on the rate of substrate utilization is not only that disappearance

of litter would occur at a faster rate with e.g. nematodes in the system, but also that potentially more

nutrients would be mineralized (Ingham et al. 1985). Loss of soil invertebrates can reduce decomposition

rate and nutrient cycling substantially (Swift et al. 1979; Verhoef & Brussaard 1990). The review by Seastedt

(1984) showed that although the amount of soil metabolism that can be attributed to soil animals themselves

is 10% or less, microarthropods increase litter decomposition rates by an average 23%. Effects such as those

of global change on these higher trophic levels of the soil food web could thus have important consequences

for carbon and nutrient cycling in terrestrial ecosystems, through effects on population size and turnover of

the microflora (Brussaard 1997; Lavelle et al. 1997).

The plant-soil system Primary production provides the organic material that fuels heterotrophic respiration, and heterotrophic

respiration releases minerals that support primary production (Harte & Kinzig 1993). The fundamental biotic

components of ecosystems thus include plants, which bring energy into the system, decomposers, which

break down dead organic matter and release CO2 and nutrients, and animals, which transfer energy and

materials and modulate the activity of plants and decomposers. About half of the gross primary production

(GPP) is respired by plants to support their growth and maintenance (Chapin III et al. 2002), leaving the net

primary production (NPP). Plants loose NPP carbon through several pathways. The largest of these is the

transfer of carbon from plant to soil. This occurs through litterfall, root exudation (the secretion of soluble

compounds by roots to the soil), and carbon transfers to root symbionts (e.g. mycorrhiza and N-fixing

bacteria).

Plant allocation of resources

Plants exhibit a consistent pattern of internal allocation that maximizes growth in response to the balance

between aboveground and belowground supply rates (Enquist & Niklas 2002). Plants allocate new biomass

preferentially to roots when water or nutrients are limiting. They allocate new biomass preferentially to

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shoots when in need of greater leaf area to absorb light and meet photosynthesis capacity (Reynolds &

Thornley 1982). A plant can increase nitrogen uptake by altering root morphology, by increasing root

biomass, or by increasing extent of mycorrhizal colonization (Chapin III et al. 2002). Plant leaves begin to

senesce and reduce their rates of photosynthesis when day length and other environmental cues signal the

end of the season. Before senescence, plants allocate carbohydrates and nutrients to storage organs to prevent

their loss during senescence (Chapin III et al. 1990).

Root input to soil

Initial decomposition and nutrient release from litter occurs at the ground surface, and roots therefore tend to

grow in surface soils to access these nutrients. Thus, root litter is also primarily produced in surface soils,

reinforcing the surface localization of most decomposition. The rhizosphere is considered to be the cylinder

of soil that root hairs exploit and into which they release exudates (Bais et al. 2006). It makes up virtually all

the topsoil in fine-rooted grasslands (avg. root distance = 1 mm) (Newman et al. 1985). Rhizosphere

decomposition may be more sensitive to factors influencing plant carbohydrate status than the soil

environment (Kuzyakov 2002), so the global change effects on decomposition in root-dense ecosystems such

as grasslands may to a large extent be mediated by plant responses. Through root exudation, the plant

releases carbon fixed from the atmosphere, for which the microbial uptake provides a ready sink (Bais et al.

2006). The growth of bacteria in the zone of exudation (Norton & Firestone 1991) is supported by abundant

carbon availability (20 to 40% of NPP), and bacteria must acquire nutrients by breaking down SOM. In other

words, carbon-rich exudates “prime” the decomposition process in the rhizosphere (Kuzyakov 2002). As

protozoa and nematodes graze populations of rhizosphere microorganisms, excess nutrients are excreted

(Clarholm 1985).

Quality of organic input

The quantity and quality of soil organic matter is the major determinant of the quantity of energy that flows

through the decomposer system. Soil microbes vary in growth efficiency, and bacteria typically have lower

growth efficiency than do fungi. All microbes convert substrates into biomass less efficiently with greater

environmental stress or with less labile substrates (Chapin III et al. 2002). The C:N ratio has frequently been

used as a proxy for substrate quality, because a low C:N generally causes a faster decomposition (Berg 2000;

Taylor et al. 1989). Experimental additions of high nutrient substrates have resulted in increased microbial

abundances (Bååth et al. 1978; Jonasson et al. 1996), signifying a nutrient limitation of soil microbes. Under

some circumstances, carbon lability rather than nitrogen may be the primary control over decomposition rate.

Carbon limitation of soil microbes is supported by models of plant–decomposer interactions (Harte & Kinzig

1993) as well as by the responses of microbes to experimental additions of labile carbon (Mikola & Setälä

1998; Nieminen & Setälä 2001). In recalcitrant litter, the concentration of lignin can be a good predictor of

decomposition rate (Taylor et al. 1989). Difference in litter quality between plant species make up an

important mechanism by which vegetation affect ecosystem processes (Wardle et al. 2006). Elevated CO2

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has been seen to increase C:N ratios and lignin content of C3 plant tissues, with possible effects on

decomposition rates (Ball 1997; Cotrufo et al. 1998, but see Norby et al. 2001).

Disturbance of the plant-soil system

Net carbon accumulation by an ecosystem depends strongly on time since disturbance (Chapin III et al.

2002). The greatest causes of variation in net ecosystem production among ecosystems are cycles of

disturbance and succession. Herbivory remove carbon from plants, and often accounts for 5-10% of NPP in

terrestrial ecosystem, and up to 50% in some grasslands (Chapin III et al. 2002). Grasslands with an

evolutionary history of intensive grazing, however, are often more productive when moderately defoliated

than in the absence of aboveground grazers (Frank 1998). In the absence of grazers, plant species

composition shifts to species that are less productive and have lower litter quality (Augustine & Mcnaughton

1998; Bardgett & Wardle 2003). Fire releases carbon directly by combustion and leaves a pyrogenic organic

matter. Fire also removes vegetation that otherwise transpires water and shades the ground surface, and

account for large gaseous losses of N. The amount and forms of nitrogen volatilized during fire depend on

the temperature generated by the fire (Chapin III et al. 2002). Burning impacts decomposition substrates and

rates, alters abiotic soil conditions, and impacts the soil biotic community (O’Lear et al. 1996; Johnson &

Matchett 2001). There is substantial evidence that soil disturbance in many ecosystems impacts

decomposition rates and soil food web trophic structure (O’Lear et al. 1996; Neher et al. 2005).

Temperate grassland and shrubland ecosystems The natural habitats subject to experimental work for this dissertation

are temperate grassland and shrubland. Temperate grasslands occur

naturally in water limited ecosystems, where the annual rainfall is too

light to support heavy forest, too great to result in desert (generally

between 250-800 mm) (Smith & Smith 2003). Temperate grasslands

usually have a high rate of evaporation, and experience periodic severe

droughts. Primary production of North American grasslands decreases

with increasing temperature, but mainly because of indirect temperature

controls by altering water demand and consequently water availability

(Sala et al. 2001). Grasslands, even though adapted to periods of

drought, still do the poorest where precipitation is low and temperature

is high, because reduced soil moisture increases water stress and decreases nutrient uptake. Thus, globally

increased temperatures and altered precipitation patterns with decreasing growing season precipitation (IPCC

2013), could substantially affect the state of temperate grasslands.

Preserving grasslands requires some measure of disturbance

Tallgrass prairie, Konza LTER Photo by Marie Dam

11

The most visible feature of grassland is the tall, green ephemeral

herbaceous growth that develops in spring and dies back in autumn.

Grasslands that are unmowed, unburned or ungrazed accumulate a thick,

organic detritus layer (mulch) (Smith & Smith 2003). The oldest bottom

layer, humic mulch, consists of decomposed remains of fresh mulch, the

top layer consists of fresh litter, deposited throughout the season. The

turnover time is 3-4 years on average (Smith & Smith 2003), but grazing

and burning reduces the mulch. Heavy mulches can suppress growth of

grasses and allow woody encroachment, but when accumulating to a

proper degree, mulch increases soil moisture through effects on

evaporation, decreases erosion, improves seed germination conditions,

and helps grasslands maintain themselves. Many grasslands require periodic fire or other types of

management for maintenance, renewal and removal of woody encroachment. In particular anthropogenically

determined grasslands that are located in areas where potential natural vegetation is forest requires frequent

mowing, grazing or burning to persist.

Growth strategy of grasses

Grasses have a mode of growth that adapts them to grazing and fire. Critical growth tissues are below ground

surface (tillers grow from short underground stems), protected from grazing and fire. In the growing season,

as grazers defoliate plant aboveground, grasses respond by increasing the photosynthetic rate in the

remaining tissue, stimulating new growth, and reallocating nutrients and photosynthates from one part of the

plant to another, especially from roots to stems (Smith & Smith 2003). The root layer is highly developed,

and the large biomass of roots and other underground organs in grasslands and the high concentration of

organic matter provide substrate for a large variety of bacterial, fungal and nematode (Sala et al. 2001). A

soil invertebrate study in tallgrass prairie showed more than 200 nematode species, of which fungivores

constituted 40% (Ransom et al. 1998) and the nematode biomass was exceeded only by that of bacterial and

fungal groups.

Dry heathland is a distinct type of shrubland.

Grazing elk Photo by Marie Dam

12

As for grasslands, dry heathlands are characterized by a low level of plant-available nutrients and there is

often a strong competition for nutrients between

plants, fungi and bacteria (Jonasson et al. 1996).

Further it is, at least periodically, limited by

water. Heathlands are dominated by ericaceous

shrubs (a vegetation of dense to mid-dense

growth of primitive genera adapted to fire). In

Denmark dry heathlands are mainly associated

with the ericaceous Calluna vulgaris. The

diversity of higher plants is low, and the

vegetation is adapted to cope with stressful

conditions. Heathlands, too, often depend on

removal of organic input by mowing, grazing or fire. In the absence of this, and in combination with the

antropogenically derived increased nutrient load, plant species composition changes: Primarily by grass

encroachment, but also by invasion of trees, all causing decomposition rates to increase, and leading to

changes in soil structure and exclusion of indigenous species (Terry et al. 2004). As an ericaceous shrub,

Calluna strongly affects soil formation by decelerating mineralization and maintaining nutrient-poor

conditions. This is to a greater disadvantage to other species than Calluna itself, as Calluna has the

possibility of using organically bound nitrogen via their association with ericoid mycorrhiza (Latham 2003).

However, with increasing nitrogen deposition and climate change, this initial advantage may be of less use.

This susceptibility to global change along with the simplicity of the low species diversity makes the dry

heathlands quite suitable for studying global chance effects on aboveground-belowground interactions.

Global change The work presented in this dissertation revolves around how different factors of global change affects the

abundance and interactions of the soil micro food web in temperate grassland and shrubland, and the possible

implications for the ecosystem functioning. Global change encompasses a vast array of planetary-scale

changes in the Earth system. In this dissertation, it primarily includes increased atmospheric CO2 and derived

climate change: global warming and altered precipitation patterns. It also includes land cover and land use

changes by way of biomass burning.

Fire

Approximately 80–86% of the global area burned occurs in grassland and savannas while the remainder

occurs in forested regions of the world (Mouillot & Field 2005). Fire seasons are lengthening for temperate

and boreal regions and this trend should continue in a warmer world (Flannigan et al. 2009). It has been

estimated that the global area burned increased from 500 to 608 Mha year−1 during the second half of the last

century (Flannigan et al. 2009). Savanna and grassland fires increase primarily in the tropics, but also in

Temperate heathland Photo by Marie Dam

13

temperate regions. For example, area burned has increased in southern Europe as a result of changes in

agricultural policy causing rural exodus and establishment of forest and shrubland on abandoned land

(Mouillot & Field 2005). There is a possible positive feedback mechanism, whereby a warmer and drier

climate will create conditions favoring more fire. This in turn will increase carbon emissions from fires,

which would feed the global warming (Lavorel et al. 2007). Because fire has a very strong effect on

vegetation, in many ecosystems changes in fire regime due to global changes may affect plant distribution

and ecosystem function more than the direct effect of changes in climate (Pausas & Fernández-Muñoz 2011).

CO2, warming and precipitation

Fossil fuel burning and land use changes has increased the atmospheric CO2 concentration by some 30 %

since pre-industrial times (IPCC 2013). This may have a fertilizing effect on the plant-soil system,

particularly in drylands (Donohue et al. 2013), and it induces climatic changes further shaping ecosystems

through contribution to the greenhouse effect. Global warming is predicted to raise the global mean

temperature by 1 to 3.5 °C within the 21st century (IPCC 2013), and the predicted scenarios also include

more extreme weather events and altered precipitation patterns. In NW Europe, winter precipitation (when

the plant-soil system is generally not water limited) is expected to increase by 20-40 %. In summer, a

reduction of 85-90% of current precipitation is expected. Together with a higher frequency of heavy rain

falls, the reduction is expected to result in longer drought periods during the growing season (Danish

Meteorological Institute http:// www.dmi.dk). Possible feedback mechanisms between these changes and

ecosystem functioning are multiple. CO2 emissions from terrestrial ecosystems may e.g. increase as a result

of increased SOM decomposition (Carney et al. 2007), while other mechanisms may have mitigating

influences on climate change effects. Investigating the potential feedbacks in nature is therefore imperative

to improving climate change models.

Elevated CO2, increasing temperature and summer drought can affect soil communities directly and

indirectly. Warming and changes in soil moisture can directly impact belowground organisms. Thus,

warming often has positive effects on nematode abundance (Blankinship et al. 2011) and affects the

microbial composition (Allison & Martiny 2008). Reduced precipitation has been found to decrease the

abundance of fungi, enchytraeids and collembola (Blankinship et al. 2011). CO2 effects on soil communities

are often indirect via plant responses such as increased allocation of resources belowground ((Drigo et al.

2013; Jones et al. 2009). Elevated CO2 generally increases the total microbial biomass (Zak et al. 2000;

Blankinship et al. 2011) and the abundance of mycorrhizal fungi due to enhanced plant mutualism

(Klironomos et al. 1996; Treseder 2004). Eisenhauer et al. (2012) found that elevated CO2 had direct positive

effects on soil water content, shoot biomass, microbial biomass, and soil microarthropod taxa richness. In a

greenhouse CO2 study of a grassland soil, CO2 effects were larger for higher trophic groups of nematodes

(approximately 30% for herbivores and bacterivores, and 110% for predators); perhaps because higher

trophic levels were regulated by resource limitation, while lower levels remained limited by predation

(Yeates et al. 1997).

14

Research objectives and aims The experimental work has revolved around the quantity and quality of aboveground input to the

belowground system. The organic input is affected by defoliating aboveground shoots (Paper II) and by

burning the aboveground biomass (Paper III). I compare soil communities under plant types with different

litter recalcitrance (Paper I) and trace the carbon input to the soil food web from an extremely recalcitrant

source (Paper III).

More specifically, the aims of each paper are:

Paper I: Elevated CO2 increases fungal decomposer channel in rhizospheres of contrasting plant species

To examine the relative strength of plant species effects and global change effects on the micro food web.

We wanted to know, if strong plant species effects could be found in a natural ecosystem, and whether

effects of global change on the plant-soil interaction differed under functionally different plant species.

Paper II: Elevated CO2 stimulates soil biota – defoliation modifies effects depending on season

Here, we added another layer of treatment to our global change experiment, and investigated how global

change affected the plant-soil association response to a disturbance such as significant defoliation.

We also aimed at understanding whether plant growth phase determines the response to defoliation – both

above- and belowground.

Paper III: Fire in the tallgrass prairie: effects on soil communities and trophic transfers of carbon during

litter and pyrogenic organic matter decomposition”

Here we looked at the role the altered character of the organic input plays in the effects of grassland burning.

We wanted to trace pyrogenic organic matter into the soil food web to determine if it is decomposed. We

also investigated the effect of annual burning on the trophic structure of the soil food web and how fire

affects litter decomposition ability.

Paper IV Are there links between responses of soil microbes and ecosystem functioning to elevated CO2,

nitrogen deposition and warming? A global perspective

This meta-analysis aims at establishing if global change effects on microorganisms in general determine the

extent of global change effects on ecosystem function such as carbon cycling. Or if the correlation is e.g.

only with one group of organisms (fungi or bacteria). We also tried to outline major research gaps in

improving understanding of soil organism control of ecosystem responses to global change with intent to

integrate this into large-scale models.

15

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21

Paper I

22

Elevated CO2 increases fungal decomposer channel in rhizospheres of

contrasting plant species

Marie Dam1, Lasse Bergmark2, Mette Vestergård1

1Terrestrial Ecology Section, Department of Biology, University of Copenhagen, Copenhagen, Denmark 2Division for Epidemiology and Microbial Genomics, National Food Institute, Technical University of

Denmark

Abstract As plants of different life-forms generate different soil communities, assessment of climate change impacts

on soil communities and soil services must consider how these impacts are conditioned by the vegetation

above. We investigated how soil communities under different plant species in the same ecosystem respond to

global change. In a heathland FACE-experiment, we modelled projected global changes in the field i.a. by

increasing CO2 and temperature. We assessed trophic composition of nematode communities and abundance

of bacteria and fungi (by qPCR) in rhizospheres of the two dominant plant species Calluna vulgaris (dwarf

shrub) and Deschampsia flexuosa (grass). Fungivores dominated nematode communities under C. vulgaris,

whereas relative abundances of bacterivores, herbivores, and omnivores were all significantly higher under

D. flexuosa. Elevated CO2 stimulated fungivores and increased their abundance relative to bacterivores under

both plant species. The pattern was similar for microorganisms; fungal dominance generally increased

relative to bacterial under elevated CO2.This corresponds with the hypothesis that when CO2 available for

photosynthesis increases without a corresponding increase of soil nitrogen, more carbon will flow through

fungal rather than bacterial channels. Warming modified the effect of elevated CO2 on the fungivore

abundance, but with a different outcome under the two plant species. Together, these results indicate that

decomposition at future CO2 levels may to a greater extent be mediated by fungi. With the corresponding

increase in temperature, however, this effect will depend on plant species

Introduction Soil organisms drive soil processes, and these are conditioned by abiotic and biotic factors, particularly the

production and quality of organic input. Global change such as increased atmospheric CO2 availability for

photosynthesis and climate change including altered precipitation regimes and increased temperatures is

expected to affect both the organic matter input to (Hungate et al. 1997; Rustad et al. 2001; Shaw et al. 2002;

Wardle et al. 2004; Morgan et al. 2004; García-Palacios et al. 2014) and the physical conditions of the

habitat belowground (Merrild et al. 2013a; Niklaus et al. 2003; Porporato et al. 2004). Many soil food-web

23

responses to global change are likely to be indirect and reflect changes in plant productivity and litter quality.

However, these indirect responses are in turn likely to affect carbon and nutrient dynamics at the ecosystem-

level in the longer term (Wardle et al. 1998; García-Palacios et al. 2014). Global change may alter plant

community composition, thus indirectly altering the soil communities that depend on their inputs. Souza et

al. (2010) found that elevated CO2 favored woody species, which have distinct litter chemistries, thereby

impacting microbial communities and carbon dynamics.

Eisenhauer et al. (2013) found that plant diversity effects on the abundance and functioning of soil

organisms far exceeded effects of elevated CO2 and N deposition. However, several authors have found that

plant identity rather than plant species richness per se determines plant community effects on decomposition

rates (Wardle et al. 1997), microhabitat conditions (Waldrop & Firestone 2006), nematode community

composition (De Deyn et al. 2004; Viketoft et al. 2009) and microbial community composition (Pinay et al.

2007). In cases where plant cover responses to global change treatments are very modest, global change

treatments can still affect fundamental plant physiological processes in a way that depends on the growth

strategy of the species considered (Albert et al. 2011).

Hence, with global change we expect both quantitative and qualitative changes in the organic input

with implications for resource fluxes in the decomposer food web. Resource fluxes in the decomposer food

web are thought to be compartmentalized. Moore and Hunt (1988) proposed that the degradation of plant-

derived material depends on the chemical nature of the resources (recalcitrant or labile), and that the

decomposition is divided into two separate resource pathways: a bacterial and a fungal channel (Moore &

Hunt 1988; Moore et al. 1996). Thus, carbon flows along two distinctive routes (Wardle & Yeates 1993), and

the dominance between the channels varies between ecosystems (Wasilewska 1979), along successional

gradients (Ruess & Ferris 2004), and under individual plant species (Witt & Setälä 2010). Energy channel

dominance can have implications e.g. for C sequestration, as fungi are considered more efficient in this

regard (Six et al. 2006), and can be an indicator of ecosystem responses to elevated CO2 and climate change

(Klironomos et al. 1996; Haugwitz et al. 2013). The C:N ratios of plant tissues generally increase at elevated

CO2 (Cotrufo et al. 1998), which has also been found for the vegetation at the research site of the present

study (Larsen et al. 2011). Such shifts in litter quality are likely to mainly influence fungi, with probable

diminishing consequences for decomposition rates (Wardle et al. 1995). In the present heathland study site,

Arndal et al. (2013) found that CO2, warming, and summer drought increased root growth of both heather

and wavy hair-grass, and that arbuscular mycorrhiza fungi associated with wavy hair-grass but not ericoid

mycorrhiza increased at elevated CO2.

The microbial and the nematode community together represent first, second and third level

consumers in the decomposer food web. Nematodes regulate the size and function of fungal and bacterial

populations in the soil (Ingham et al. 1985) and rates of carbon and nitrogen turnover (Griffiths 1994; Neher

2001; Osler & Sommerkorn 2007). Therefore, their community structure can provide important insights

24

regarding many aspects of soil food web function (Ferris 2010), e.g. to which extent fluxes of C and energy

are directed through the fungal or bacterial decomposition channel.

We want to examine how global change affects the soil communities and the compartmentalization

of energy fluxes - and whether these effects are general or specific for the plant type defining the local

environment. We investigated global change effects on the soil communities under two plant species

belonging to different functional groups (a perennial grass and a dwarf shrub), growing together in a

heathland ecosystem. In our analysis, we included several trophic levels of the micro food web.

We hypothesize that plants with very different ecological strategies (and functional traits) have

differently structured food webs. Specifically, we expect that the herbivore load on the more palatable grass

roots is higher than on the dwarf shrub roots. And we expect the fungal decomposition pathway to be more

prevalent under the dwarf shrub than under the grass, due to differences in C:N ratios between the plant

inputs.

We expect elevated CO2 (via increased plant production, Hungate et al. 1997; Morgan et al. 2004)

and warming to increase primary decomposition and thereby influence the abundance of the higher trophic

levels and the length of food chains. Further, with increased C:N ratios of plant and litter at elevated CO2, we

expect fungal dominated decomposition to play a relatively larger role at elevated CO2 (de Vries et al. 2006;

Rousk & Bååth 2007).

We expect drought to decrease nematode abundance through direct effects as well as soil-physical

(Merrild et al. 2013a) and plant-mediated effects (Ransijn 2014), and we expect that warming will intensify

drought effects. On the other hand, we expect elevated CO2 treatments to be less drought-affected due to

better plant water use efficiency (Field et al. 1995).

Methods Site description The experimental site is a dry, temperate heathland in eastern Denmark (55°53’N, 11°5’8E). The site is a

hilly nutrient-poor sandy deposit with a pH of 4–5, dominated by the evergreen dwarf shrub Calluna

vulgaris, Heather, and the perennial grass Deschampsia flexuosa, Wavy Hair-grass. The yearly mean

temperature is 8 °C and yearly mean precipitation is 607 mm (Danish Meteorological Institute, 2013).

Experimental design The experimental setup consists of 12 octagons (7 m in diameter) laid out pairwise in 6 blocks. In each

block, one octagon is exposed to the ambient atmospheric CO2 and the other to an elevated CO2

concentration. Each octagon is subdivided into 4 treatment plots: warming, summer drought,

warming+summer drought, and an untreated control. The experiment is thus full-factorial with individual

treatments of CO2, warming and summer drought, their combinations, and a control. In total 48 plots, with 6

25

replicates of the 8 treatment combinations, in a split-plot design (Mikkelsen et al. 2008). The CO2

concentration is elevated by Free-Air Carbon Enrichment (FACE) technique (Mikkelsen et al. 2008) to 510

ppm. The warming treatment increases temperature 1 °C in 2 cm depth by passive night-time warming with

reflectance curtains. Drought is generated in the early summer by automatic rain-out curtain shelters. The

drought continues for 2-5 weeks or until soil water content drops below 5 vol.% (Mikkelsen et al. 2008).

Soil sampling Soil was sampled by coring directly under randomly selected wavy hair-grass and heather plants in each of

the 48 subplots, yielding 96 samples. Three cores (2 cm diameter, 8 cm deep) were retrieved per sample and

mixed. The soil samples were transported in coolers and kept at 5 °C until processed. The soil was

homogenized by hand and large roots (diameter > 2 mm) removed. 10 g fresh soil was dried at 80 °C for 48

h for soil moisture determination.

Microorganisms DNA was extracted from 0.5 g of fresh soil, using a genomic mini spin kit for universal DNA isolation

(A&A biotechnology, Gdynia, Poland) with a standard protocol (Yu and Mohn, 1999). The DNA extracts

were used for quantitative abundance analysis of the fungal ITS and bacterial 16S region.

Quantitative PCR (qPCR)

Via qPCR the bacterial 16S copy number as well as the fungal ITS2 region were quantified in the extracted

DNA, using the Eub338F/Eub518R primers for bacteria (Fierer et al., 2005) and the ITS primers fITS9

(Ihrmark et al., 2012) and ITS4 (White et al., 1990) for fungi. The reactions were run in technical duplicates

and amplified by adding 2 µL DNA sample (100x diluted to avoid soil inhibitors) to a 23 μL mastermix (1X

Stratagene brilliant III ultra fast master mix (Cedar Creek, Texas. USA), 385 nM forward primer, 385 nM

reverse primer and ddH2O).

All reactions were performed in the Mx-3000 qPCR system (Stratagene, Cedar Creek, Texas, USA). The

same 2-step PCR program was used for both primer sets, combining the annealing and extension: 95 °C for 3

min followed by 40 cycles of 95 °C for 10 s, 60 °C for 20 s, and a final dissociation curve. Standard curves

for each assay were made from a 10-fold dilution series where the fungal standard curve was generated from

a plasmid containing Pilidium concavum (Desm.) Hoehm., and the bacterial standard curve was made from a

pure culture of Pseudomonas putida kt2440.

Nematodes To deal with the large number of samples, soil for nematode extraction was processed on three different days

within two weeks after sampling. Soil from the different treatments was distributed evenly between the days

to avoid bias. This staggered processing is accounted for in our statistical models, and day of processing did

26

not have a significant effect on nematode abundance or composition. Nematodes were extracted from 40 g

(fresh weight) of soil by a combination of the Baermann pan and the Whitehead tray (Whitehead &

Hemming 1965) extraction methods. Samples were extracted for 72 h, and nematodes were counted and

sorted to trophic groups (see Yeates et al. 1993) while live, in an inverted microscope. The abundance of

each trophic group is given as nematodes per g soil dry weight.

Statistics We analyzed effects of global change manipulations and plant species and all possible interactive effects on

the soil community: With the three climate change factors (CO2, temperature and precipitation/ drought) as

well as plant species as fixed factors, we used mixed linear models to test the effect on nematode trophic

group abundances, fungal and bacterial abundances, as well as the derived fungi:bacteria and

fungivore:bacterivore ratios . The statistical model was extended with a random statement to account for

random variation introduced by the experimental design. As random factors we used block (representing

pairs of octagons including all treatments), CO2 nested within block, warming nested within CO2 and block,

and drought nested within CO2 and block. For the analyses of nematode abundance, day of soil processing

was also introduced as a random factor. We applied log-transformation when necessary to obtain normality.

All data was analyzed in R (R_Development_Core_Team, 2013) using the lmer function from the lme4

package (Bates et al. 2014). The anova function from the LmerTest package was used to obtain p-values.

Models were reduced based on evaluation of F values using the step function

(LMERConvenienceFunctions). In the results reported below, only the factors kept in the model after

reduction are shown for each analysis.

Results The plant species strongly affected the composition of the nematode community (Fig. 1) as the abundance of

three trophic groups was significantly influenced by the plant species defining the rhizosphere. Fungivores

were significantly more abundant under heather (Fig. 1a), whereas the abundance of herbivores was higher

under wavy hair-grass (Fig. 1c). The higher trophic level, the omnivores, was also more abundant under

wavy hair-grass than under heather (Fig. 2d). The abundance of microorganisms (gene copies g-1) was lower

under heather (Fig. 2), primarily due to lower abundance of fungi (Fig. 2a). This is reflected in the

fungi:bacteria ratio (Fig. 3b).

27

Figur 1 Abundance of nematode feeding groups (individuals g-1 dry soil). (a) Fungivores, (b) bacterivores, (c) herbivores and (d)

omnivores). Data for all samples are aggregated under the treatment factors kept in the most comprehensive statistical models after

reduction based on F values: Plant species (Calluna vulgaris, Heather; Deschampsia flexuosa, Grass), CO2 (Ambient level, Ambient;

Elevated level, CO2 ) and warming (No warming; Warming). Means with SE bars (n= 12). Significant and near-significant P-values

shown.

The microorganisms did not show strong responses to the global change treatments, but there was a tendency

to greater fungal abundance under elevated CO2 (Fig. 2a). The CO2 effects manifest more clearly in the next

trophic level, the microbivore nematodes. Bacterivore abundance is reduced under heather at elevated CO2

(Fig. 2b), whereas fungivores are stimulated by elevated CO2 under wavy hair-grass (Fig. 2a).

a)

c)

b)

d)

Plant sp. P<1e-07

Plant sp.*CO2 P=0.053

Plant sp.*CO2*Warming P=0.022

Plant sp. P=0.108

CO2 P=0.035

Plant sp.*CO2 P=0.024

Plant sp. P=0.108 Plant sp. P=0.047

28

Figur 2 Abundance of (a) fungi (ITS copies g-1 dry soil) and (b) bacteria (16S copies g-1 dry soil). Plant species (Calluna vulgaris,

Heather; Deschampsia flexuosa, Grass), CO2 (Ambient level, Ambient; Elevated level, CO2) and warming (No warming; Warming).

Means with SE bars (n= 12). Significant and near-significant P-values shown.

Ratios of microorganisms and of the microbivore nematodes reveals effect of global change on energy

channel flow. There is a general increase of both the fungi:bacteria ratio (Fig. 3b) and of the

fungivore:bacterivore ratio (Fig. 3a) which indicates an increased flow through the fungal feeding channel.

The response is most evident under heather, where relative nematode abundances suggest that the fungal

channel is intrinsically prevailing compared to wavy hair-grass.

The fungivore:bacterivore data suggest that warming further enhances the flow through the fungal channel

at elevated CO2 under heather, whereas under wavy hair-grass , warming seems to reduce the flow through

the fungal channel at elevated CO2 (Fig. 3b). This microbivore warming response was the only effect of the

climatic manipulations. The yearly summer drought did not affect any measures of feeding group abundance

or feeding channel flow at this April sampling, and was repeatedly eliminated from the statistical models

when they were reduced based on evaluation of F values.

a) b)

Plant sp. P<10-7

CO2 P=0.10

Plant sp. P=0.027

29

Figur 3 Relationship between abundances of microbivores and between abundances of microorganisms. (a) Nematode

fungivore:bacterivore ratio (individuals g-1 soil: individuals g-1 soil), (b) microbial fungi:bacteria ratio (ITS gene copies g-1 soil:16S

gene copies g-1 soil). Data for all samples are aggregated under the factors kept in the most comprehensive statistical models after

reduction based on F values: Plant species (Calluna vulgaris, Heather; Deschampsia flexuosa, Grass), CO2 (Ambient level, Ambient;

Elevated level, CO2 ) and warming (No Warming; Warming). Means with SE bars (n= 12). Significant P-values are shown.

Discussion Effects of plant species exceed of global change effects In this heathland study, plant identity effects on soil fauna abundance and composition exceeded global

change effects on all measured parameters. This supports previous findings by Pinay et al. (2007), that under

controlled experimental conditions with monocultures of an annual and a perennial species, plant cover type

rather than elevated CO2 strongly affects microbial activity and drives the development of different soil

heterotrophic community structures. Here, sampling directly under functionally different species, we find

stronger effects of plant type than of the global change manipulations of the entire system.

We can confirm our hypothesis that herbivorous nematode seem more abundant under wavy hair-

grass than under heather (Fig. 1c). This is what we expected based on greater palatability of roots and litter

from wavy hair-grass compared to heather, due to lower lignin content (Beier et al. 2004) and lower C:N

ratio (Larsen et al. 2011). Nematode herbivore populations differ depending on resource quality (Wardle et

al. 2003) and Witt & Setälä (2010) also found a higher abundance of plant parasitic nematodes associated

with grass and forb roots than with heather and spruce roots. Furthermore, at our research site wavy hair-

grass roots are less colonized by mycorrhiza than heather roots (Merrild et al. 2013b), reducing the negative

mycorrhiza control on nematode root-feeders (Peña et al. 2006; Elsen et al. 2008; Koricheva et al. 2009).

Bacteria and bacterivores are also more abundant under wavy hair-grass than under heather (Fig. 2), which

we also ascribe to a higher quality of plant inputs.

a) Plant sp. P<10-7

CO2 P=0.0128

Plant sp.*CO2*Warming P=0.0123

Fungivore:Bacterivore

b) Plant sp. P<10-7

CO2 P=0.0155

Fungi:Bacteria

30

Fungal abundances are lower under heather than under wavy hair-grass (Fig. 2a), which is not what

we had expected. However, Christensen et al. (2012) showed a close correlation between abundance of

nematode microbivores and the activity and contribution to decomposition by their respective microbial food

source – fungi or bacteria. The greater dominance of fungal feeding nematodes (Fig. 1a) thus clearly

indicates that the fungal decomposition channel is more important under heather than under wavy hair-grass,

and the fungal abundance appears to be under topdown control. In contrast, the bacterial abundances do not

appear to be top-down regulated by nematodes, and bacterial abundance under heather is numerically higher

than fungal abundance (Fig. 2), which translate to a lower fungi:bacteria ratio under heather compared to

wavy hair-grass (Fig. 3b). However, since the bacterivore abundance is lower than the fungivore, the heather

rhizosphere has a higher fungivore:bacterivore ratio (Fig. 2a), similarly to what Witt & Setälä (2010) find

when comparing a perennial grass and heather. This confirms the hypothesized greater fungal channel

dominance under heather than under to wavy hair-grass.

Elevated CO2 is the major global change impact The global change agents – CO2, warming and altered precipitation – caused less pronounced differences

than plant species, but elevated CO2 did have significant impact on the plant-soil system. We consider the

CO2 effect on soil organisms to be indirect via plant-soil interactions, as direct effects of CO2 enrichment on

soil organisms are probably of little consequence because of the high CO2 levels already present in soils

(Veen et al. 1991). We did not see much difference in absolute abundances that could indicate increased

belowground production in the elevated CO2 plots, as we originally hypothesized and also see in other

studies (Dam et al. 2014, in prep.). We did however see a shift in the relative abundances of both

microorganisms and microbivores. At elevated CO2 the tendency to increased fungal abundance (Fig.2a), the

decreased abundance of bacterivore nematodes under heather (Fig. 1b), the increased abundance of fungivore

nematodes under wavy hair-grass (Fig. 1a), and the increased fungal:bacteria (Fig. 3b) ratio show that

compartmentalization of energy fluxes shifted towards greater fungal dominance at elevated CO2. Thus, in

accordance with our hypothesis, these results indicate that the organic input to the soil is to a greater extent

processed via the fungal decomposition pathway, when more CO2 is available for photosynthesis in a N-

limited (Larsen et al. 2011; Stevnbak et al. 2012) system. Increased C:N ratios of the plant material seen at

elevated CO2 (Larsen et al. 2011) were expected to increase fungal decomposition (Bossuyt et al. 2001;

Nilsson et al. 2012). Furthermore, enhanced nutrient demand at elevated CO2 is also expected to increase the

mycorrhizal association (Treseder 2004), which has indeed been found for the arbuscular mycorrhiza

colonization of the wavy hair-grass at the present research site (Arndal et al. 2013). This would increase the

fungi:bacteria ratio and likely also the fungivore:bacterivore ratio at elevated CO2. Klironomos et al. (1996)

and Rillig et al. (1999) both found fungal abundance to increase relatively more than bacterial abundance

under elevated CO2. Yeates & Newton (2009) found that elevated CO2 stimulated fungivore nematodes more

31

than bacterial-feeding nematodes, and Yeates et al. (1997) found increased fungivore nematode abundance

concurrent with elevated increased root biomass and -necromass at elevated CO2.

It is interesting that elevated CO2 causes similar shifts towards fungal decomposition pathway

dominance in the fundamentally different soil communities under the two plant species. Judged by the

abundances (Fig. 1), the CO2 effect on the fungivore:bacterivore ratio (Fig. 3a) is primarily driven by a

stimulation of fungivores under wavy hair-grass and a reduction of bacterivores under heather. The fungal

abundance and thus the fungi:bacteria ratio generally increase at elevated CO2 (Fig.1a and 3b), but this seems

to primarily increase the fungivore:bacterivore ratio in the already fungivore dominated community under

heather. The reduction of bacterivores at elevated CO2 under heather combined with the unaffected bacterial

abundance may indicate that the bacterial community composition shifted towards a community of overall

poorer food source quality for the bacterivorous nematodes.

The consequence of a relatively greater flow through the fungal based part of the soil food web at

elevated CO2 is unresolved (Niklaus et al. 2003). Several authors have suggested that the higher growth

efficiency of fungi will cause an analogous increase in carbon sequestration (Beare et al. 1992; Klironomos

et al. 1996; Treseder & Allen 2000; Six et al. 2006), while others have reported that increased fungal

production increases depolymerizing enzyme production, which in turn would increase soil organic matter

degradation (Lipson et al. 2005; Carney et al. 2007). We should keep in mind though, that elevated CO2 will

not occur alone as a global change agent. In this multifactorial global change experiment, we did not see

much effect of climatic changes (warming and summer drought), but we did see an interaction between CO2

and warming. Hence, elevated CO2 and warming affected fungivorous nematodes differently under the two

plant species, which manifests in the fungivore:bacterivore ratio. Warming increased the stimulating effect

of CO2 on the fungivores under heather, but reduced the CO2 effect under wavy hair-grass. The interaction

could be related to the arbuscular mycorrhiza colonization, which is increased at elevated CO2, but reduced

in the full-factorial treatment in 5-10 cm depth (Arndal et al. 2013). If warming stimulates microbial nutrient

turnover, there might be less need for the mycorrhizal association, which then reduced. This could reduce

part of the fungivore food source. Generally, however, the relationship between temperature and soil fauna is

weak (Petersen & Luxton 1982) so it is unlikely that the temperature increases will directly induce large

effects on the biomass of broad taxonomic groups (Wardle et al. 1998).

Conclusions Relative abundances of microorganisms and nematode trophic groups indicate a shift towards the fungal

decomposition channel when CO2 is increased. This is concurrently seen in two otherwise substantially

different soil communities under two functionally different plant species, although by way of different

community dynamics. The effect is greater in the heather soil community, which is already dominated by the

fungal pathway, and warming further increases the trend. Under wavy hair-grass warming diminishes the

32

CO2 effect. Hence, we might expect an increased fungal decomposition at future CO2 levels, with possible

consequences for ecosystem functioning. However, changes in plant community composition in future

climate will be of great importance as different plant species are associated with significantly different

rhizosphere communities, and because other climate factors may have more plant specific effects.

Acknowledgements The CLIMAITE project is funded by the Villum Kann Rasmussen foundation, and we thank Sven Danbæk,

Andreas Fernquist, Preben Jørgensen and Nina Wiese Thomsen for keeping the field facility running. We

also thank students David Byriel and Marie Richter Flyger, for assistance both in the field and in the

laboratory. Finally we thank Søren Christensen for valuable comments to the manuscript

33

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Paper II

40

Elevated CO2 stimulates soil biota – defoliation modifies effects depending on

season

Marie Dam1 and Søren Christensen1

1Terrestrial Ecology Section, Department of Biology, University of Copenhagen, Copenhagen, Denmark

Abstract To understand the extent and nature of the responses and possible feedback mechanisms to global change in

terrestrial ecosystems, it is necessary to examine the effects on aboveground-belowground interactions.

We studied a temperate heathland system subjected to experimental climate and atmospheric factors based

on prognoses for year 2075. By impacting the aboveground plants with significant defoliation, we were able

to study how global change modifies the interactions of the plant-soil system. Shoot production, root

biomass, microbial biomass and nematode abundance were assessed in the rhizosphere of manually

defoliated patches of Deschampsia flexuosa in June in a full-factorial FACE-experiment with the treatments:

increased atmospheric CO2, increased nighttime temperatures, summer droughts, and all of their

combinations. We found positive effects of CO2 on root density and microbial biomass and a tending

positive effect on nematode abundance. We found a negative effect of defoliation on microbial biomass that

was not apparently affected by global change. The negative effect of defoliation cascades through to soil

nematodes as dependent on CO2 and drought. At ambient CO2, drought alone reduced nematodes, and a

negative defoliation effect is only seen without drought. In contrast, at elevated CO2 defoliation only

affected nematodes negatively under concurrent influence of drought. We discuss that global change effects

on aboveground-belowground interactions found here and in a previous publication from the same site

(Stevnbak et al. 2012) are related to plant growth phase, and that the proposed mechanism of plants feeding

their belowground microbial loop when in immediate need of nutrients is not present in this natural system

during active grass growth.

Introduction Soil biota plays a significant role in biogeochemical cycling and physical conditions and their responses to

global change are likely important at the ecosystem scale (Lavelle et al. 1997; Brussaard 1998; Bradford et

al. 2002), but is remarkably understudied (West et al. 2006; Bardgett et al. 2013). The interactions between

the aboveground and the belowground spheres are complex relationships affected by biotic as well as abiotic

factors.

Elevated CO2 generally results in an increase of abundance and activity at the bottom of the food

web, i.e. of bacteria, fungi and microfauna (protozoa and nematodes) as found in a meta-analysis of soil biota

41

response to global change (Blankinship et al. 2011). Now, elevated CO2 will not occur alone but in

combination with climatic changes such as elevated temperature and altered precipitation pattern. Since

responses of soil biota to global change are unique for each global change factor (Blankinship et al. 2011),

and interactions between different global change factors may create responses not predicted by single-factor

experiments (Shaw et al. 2002; Norby & Luo 2004), multi-factor experiments are needed. One of the few

recordings of multi global change factors that impact soil biota revealed significant effects involving

elevated CO2, N deposition and summer drought (Eisenhauer et al. 2012). Here, CO2 was the global change

factor affecting most soil biota groups, with increasing abundances at micro-, meso-, and macrofauna level.

Furthermore, CO2 turned out to be the only global change variable playing a role when building a SEM

model of global change effects on the soil food web (Eisenhauer et al. 2012). The likely explanation as

already stated by Ostle et al. (2007) is that environmental changes affecting the quantity and quality of

photosynthate-C inputs to the soil impacts the biology that regulates the soil C cycle.

Defoliation is an aboveground disturbance where effects belowground depend on abiotic factors such

as climate as well as biotic factors such as plant growth phase (Guitian & Bardgett 2000; Wilsey 2001;

Yeates et al. 2003; Ilmarinen et al. 2005; Lau & Tiffin 2009; Yeates & Newton 2009; Stevnbak et al. 2012).

Defoliation effects on the plant-soil interactions probably relates to whether plants stimulate decomposition

through the exudation of low-molecular-mass carbon compounds when in apparent need of nutrients

(Griffiths & Robinson 1992), and thereby feed the microbial loop and increase the microbial grazers and

higher trophic levels of the soil food web (Bonkowski et al. 2000). There have been several studies with both

negative (Holland & Detling 1990; Northup et al. 1999; Nguyen & Henry 2002) and positive (Holland et al.

1996; Hamilton & Frank 2001) effects on carbon release from roots upon defoliation, so the question

remains unresolved. Defoliation has been found to decrease microbial biomass (Williamson & Wardle 2007;

Guitian & Bardgett 2000). In contrast, Mawdsley & Bardgett (1997) found that number of bacteria increased

following defoliation, but this did not result in an increased soil microbial activity.

Defoliation effects on soil biota will depend greatly on plant responses such as changes in litter

quality (Ball 1997), water use efficiency (Field et al. 1995), root biomass, and rhizodeposition (Jones et al.

2009). Defoliation effects have been observed on resource allocation within the plant and in the rhizosphere.

In grassland sampled in the middle of the growing season, defoliation by grazing resulted in increased

aboveground production (Frank 1998), and an increased transport of N and P from roots to shoots (Mikola et

al. 2009). Mikola et al. (2009) found no stimulation of either mineralization or soil fauna by defoliation of

plants in active growth. Similar results were obtained in microcosms with newly established grass in active

growth: at field nutrient levels, defoliation altered allocation of C and N, but did not stimulate either

microbial activity or abundance of microbial grazers (Ilmarinen et al. 2008). In another microcosm

experiment with defoliation of grasses in different phases of growth, Ilmarinen et al. (2005) suggest that the

reduced root C concentration they find in defoliated plants could be due to increased C allocation to growing

42

shoots at the expense of roots following defoliation, as found in Caldwell et al. (1981), Briske et al. (1996),

and Strauss & Agrawal (1999). However, they also find that plants defoliated in the later stages of the

growing season increased the root mass relative to plant mass. Stimulating effects of defoliation on soil biota have been reported, but often in studies done under

less favorable conditions for plant growth, after the most productive part of the growing season. Defoliation

in a cool Scottish upland in September (Ostle et al. 2007) or in water-limited grassland of Yellowstone in the

driest month of July (Hamilton et al. 2008) both resulted in transfer of more photosynthate to soil biota. The

present study follows up on a study from a Danish heathland done in September at the end of the growing

season, where defoliation also resulted in increased carbon flow through the soil biota, and more so at

elevated CO2 (Stevnbak et al. 2012). Based on the above-mentioned studies it seems as if defoliation of

actively growing grass does not induce carbon release from plants to soil biota whereas carbon exudation

may increase when the active growth phase is over. In line with this, an experiment where defoliation of

grass was performed in both early and late growth phase resulted in a reduced microbial biomass early but

increased microbial biomass in the late growth phase (Guitian & Bardgett 2000).

In this study, we defoliated grass in a field site in a multifactor FACE experiment where CO2,

temperature and precipitation are manipulated to simulate predicted global change (IPCC 2013). This

allowed us to test how this disturbance affected aboveground-belowground interactions under influence of

elevated CO2 as well as predicted climatic changes. If defoliation causes plants to actively increase

rhizodeposition in order to gain nutrients, we would expect at least the same response belowground as

reported for the same grass species in the same experimental set-up by Stevnbak et al. (2012) or more likely

an even larger response because the nutrient need for the plant is higher before seed-set than in the late

season study of Stevnbak et al. (2012).

Methods Site description The experiment took place at the CLIMAITE experimental site (55°53’ N, 11°58’ E) – a FACE facility

approximately 50 km northwest of Copenhagen, Denmark. The site is a dry, temperate heathland, dominated

by the dwarf shrub Calluna vulgaris (L.) and the perennial grass Deschampsia flexuosa (L.). The soil is a

well-drained, nutrient-poor sandy deposit with a pH of 4–5 and an organic top layer ranging from 2 to 5 cm

in depth. Long-term annual mean air temperature is 8.0°C, annual mean precipitation is 607 mm (Danish

Meteorological Institute, http://www.dmi.dk).

Experimental design The set-up consists of twelve 7 m diameter octagons. Each octagon is divided into four plots receiving either

(1) summer drought (D) by automatic rain-out shelters , (2) passive nighttime warming (T) of air and soil by

43

reflectance curtains 50cm above ground, (3) a combination of drought and warming (TD) or (4) neither

drought nor temperature treatment. Furthermore six of the twelve octagons are under ambient (A)

atmospheric CO2 concentrations and the other six subjected to an elevated (CO2) CO2 concentration (510ppm

in a free air CO2 enrichment setup, FACE). The experiment thus has a full-factorial design arranged in

blocks of pairwise octagons representing all combinations of D, T and CO2, including an untreated control

for reference (A). Hence eight treatments with six replicates, in total 48 plots, arranged in a split plot design

(Mikkelsen et al. 2008). The warming treatment elevates the air and soil temperature by 1–2 °C. The drought

continues for 2–5 weeks or until soil water content falls below 5 vol.% water content in the top 20 cm of the

soil (Mikkelsen et al. 2008). The drought effect lasts into the fall, but by October, the soil moisture is only

1% lower in the drought treated plots (Christensen et al. in prep.). The experimental area is protected from

large herbivores by fencing.

Defoliation treatment The entire climate manipulation design was overlain with a +/- defoliation treatment on areas with

Deschampsia flexuosa. In each plot, two circular units of 0.07 m2 were marked off in segments where D.

flexuosa was dominant. Two of the plots did not have a sufficient area of grass leaving us with 46 plots (six

treatments with six replicates, two treatments with five replicates, n=92). The vegetation in the grass units

was either left non-defoliated as a control or defoliated by cutting. Cutting was done manually four times,

every six-eight days starting June 1st when the annual drought treatment had ceased. The cuttings were

removed from the plots. Before the first cutting, the grass height of the units was assessed. The average of all

units was 14.4 cm ±3.9 with no treatment differences. At each defoliation event the vegetation was cut down

by 1/6 of the pre-treatment median height in each individual defoliation unit. Thus, by the end of the

treatment, the defoliation had removed 2/3 of the original vegetation, and the median height was

approximately 8-10 cm above the soil, depending on the original median height.

Soil sampling At June 27th, soil samples were randomly collected in all 92 units by coring. One larger core (4 cm diameter,

15 cm deep) was sampled for root biomass determination. Three cores (2 cm diameter, 8 cm deep) were

retrieved and mixed to cover spatial variability. The soil from the 2 cm cores was analyzed for soil moisture

content, SOM, nematode numbers and microbial biomass by chloroform fumigation. Root C:N, substrate

induced respiration (SIR) and protozoan numbers were estimated, too, but these data showed no significant

response and are not presented here. The soil samples were transported in coolers and kept at 5 °C until

processed. To deal with the large number of samples, they were processed over 8 days, with the different

treatments distributed evenly between the days to avoid bias. This staggered processing is furthermore

44

accounted for in the statistical model. When processed, aboveground plant biomass was removed and the

rhizosphere carefully cut up for homogenization.

Shoot and root analyses The grass height of all units (+/- defoliation) was assessed at each defoliation treatment as well as at the end

of the defoliation treatment. These values were used for estimation of shoot production. The roots from each

4 cm core were carefully washed over a 2 mm sieve, all root material was collected and dried at 80 °C and

weighed. From the mixed 2 cm cores, subsamples of 5 g soil were dried at 80 °C for 48 h for soil moisture

determination and combusted at 550 °C for 6 h for SOM determination by loss-on-ignition.

Soil microbial biomass A subsample of 10 g of soil mixture was fumigated in ethanol-free chloroform (CHC13) for 24 h to release

the nutrients in the soil microbial biomass (Jenkinson & Powlson 1976; Tate et al. 1988). After fumigation,

the soil was extracted in 50 ml 0.5 M K2SO4 for 1 h and filtered. Simultaneously, another subsample was

extracted in the same manner but without fumigation to recover the soil inorganic nutrients. Total organic

carbon (TOC) (fumigated samples) and dissolved organic carbon (DOC) (non-fumigated samples) was

measured on Shimadzu TOC-5000A total organic C analyzer using the infrared gas detector (IRGA) method.

Microbial carbon was calculated using the extractability factor KEC = 0.45, to account for the microbial

biomass C that is not released by fumigation and extracted by K2SO4 (Jonasson et al. 1996): Microbial C =

(TOC – DOC) / KEC.

Soil fauna Nematodes were extracted from 5 g (fresh weight) of soil by modified combination of the Baermann pan and

the Whitehead tray (Whitehead & Hemming 1965) extraction methods. Samples were extracted for 48 h,

and nematodes were then counted at x40 magnification using a dissecting microscope. After counting the

samples were fixed in a 4 % formaldehyde solution. They were later analyzed for nematode community

composition of trophic groups. Based on mouth part morphology, the nematodes were identified to one of

five feeding groups (Yeates et al. 1993) under a dissecting microscope at x40 magnification.

Statistics We analyzed effects of global change manipulations and the defoliation and all possible interactive effects on

the plant-soil system: With the three climate change factors (CO2, temperature and precipitation/ drought) as

well as defoliation as fixed factors, we used mixed linear models to test the effect on every measured plant

and soil variables. The statistical model was extended with a random statement to account for random

variation introduced by the experimental design. As random factors we used block (representing pairs of

45

octagons including all treatments), CO2 nested within block, warming nested within CO2 and block, and

drought nested within CO2 and block. We applied log-transformation when necessary to obtain normality.

All data was analyzed in R (R_Development_Core_Team, 2013) using the lmer function from the lme4

package (Bates et al. 2014). The anova function from the LmerTest package was used to obtain p-values.

Models were reduced based on evaluation of F values using the step function

(LMERConvenienceFunctions). In the results reported below, only the factors kept in the model after

reduction are shown for each analysis.

Results The elevated CO2 led to increases of three important links in the belowground food chain: The model

showed statistically significant main effects of CO2 on root density (PCO2 = 0.0026) and microbial biomass

(PCO2 = 0.041), and a tending main effect on nematode abundance (PCO2 = 0.096). Fig. 1, 2 and 3,

respectively, show how belowground pools all increase at elevated CO2. Drought and temperature had less

pronounced effects on the system, and primarily affected root density. Root density was reduced by drought

(Pdrought = 0.0095) and temperature enforced drought effects, further decreasing density (Pdrought*temperature =

0.0024) (Fig. 1).

Figure 1 Root density. All data is shown aggregated into a figure showing only the factors kept in the statistical model after reduction based on F values: CO2, drought and warming. Means with SE bars (n= 12). Significant effects at P < 0.01 are displayed.

root

(g d

w)/

soil

(g d

w)

CO2: P=0.003

Drought: P=0.009

Drought*Warming: P=0.002

46

For nematode abundance, the interaction between CO2 treatment, defoliation and drought (PCO2*Defoliation*Drought

= 0.013) was significant. Drought and defoliation each reduces nematode abundance at ambient CO2,

whereas only the combination of the two reduces nematode abundance at elevated CO2 (Fig. 3). The relative

abundance of nematode feeding groups was not affected by the treatments. The average distribution was

45% bacterivores, 30% herbivores, 15% fungivores and 5% omnivores and predators (5% were

unidentified). Microbial biomass showed a numerically small, but statistically significant increase at

elevated CO2 (Fig. 2).

Figure 2 Microbial biomass. All data is shown aggregated into a figure showing only the factors kept in the statistical model after reduction based on F values: CO2 and defoliation. Means with SE bars (n= 24). Significant effects at P < 0.05 are displayed.

Mic

robi

al b

iom

ass

(mg

mic

C g

-1 d

sw)

CO2: P=0.041

Defoliation: P=0.008

47

Figure 3 Nematode abundance. All data is shown aggregated into a figure showing only the factors kept in the statistical model after reduction based on F values: CO2, drought and defoliation and. Means with SE bars (n= 12). Effects at P < 0.1 are displayed.

We found the global change effects on the aboveground-belowground interactions to be countered by

defoliation of the plant shoots. Aboveground defoliation reduced belowground biota as seen in Fig. 2 and 3.

The model showed statistically significant main effects of defoliation on both microbial biomass (P = 0.011)

and nematode abundance (P = 0.048). At the same time, there is a considerable regrowth of the defoliated D.

flexuosa (Fig. 4) - comparably larger than the growth of the non-defoliated plants in the same time span (P <

10-7). The results also show that the plants were indeed in active growth when defoliated, as there is a

considerable growth of the non-defoliated plots, too (Fig. 4). The defoliation was not just numerically but

also statistically (P < 10-7) the most significant effect on shoot growth. The model did show some

interactions between global change treatments on shoot growth as well, but due to the relatively imprecise

method of measurement, we chose to include only the defoliation effect, as the other effects were

numerically smaller and statistically less strong. Root density was not affected by defoliation (Fig. 1).

Nem

atod

e ab

unda

nce

(# g

-1 d

sw)

CO2: P=0.096 Defoliation: P=0.048 CO2* Defoliation*Drought: P=0.013

48

Figure 4 Shoot productivity: Growth of D. flexuosa in treatment units since first defoliation date. For defoliation treated units, the cuttings are included in cumulative values for growth. Values are means with SE bars (n= 6). Significant effects at P < 0.01 are displayed.

Discussion CO2 increases soil biota and belowground plant biomass All measured belowground pools tended to increase at elevated CO2: Root density, microbial biomass and

nematode abundance increase in agreement with previous results from the experimental sites, observing

increases in plant net photosynthesis at light saturation (Albert et al. 2011), biomass of roots (Arndal et al.

2014), and in soil respiration (Selsted et al. 2012). These components were all stimulated either by elevated

CO2 alone or in interaction with drought or temperature. This is most likely due to the increased input of C

into the belowground food chain from the increased CO2 available to aboveground photosynthesis. This

result is in line with Eisenhauer et al. (2012) who found elevated CO2 to increase root and shoot biomass,

and found the root biomass to be a determining factor for the soil food web. Hence, as hypothesized we

might see more organisms in the decomposer food web at future CO2 levels. Elevated CO2 even seems to

create a new robust level of carrying capacity for nematodes in the system, which it takes a combination of

two stressors (drought and defoliation) to reduce. It takes only one stressor (drought or defoliation) to reduce

nematode numbers under present day ambient CO2. This difference is confirmed by a significant interaction

between CO2 treatment, defoliation and drought. The drought effect seen in the nematodes could be a result

of the reducing effect of drought on root density (which is further strengthened by warming). Although we

see no effect on microbial biomass, this could still reflect a treatment effect on the system originating from

the summer drought. It seems reasonable that effects of episodic stress such as summer drought are no

longer seen in the quickly turning over microbial biomass with generation times measured in hours or days a

month after the treatment has ceased. In the longer lived organisms such as the nematodes (with generation

Shoo

t gro

wth

Defoliation: P<1*10-7

49

times of months or years) and especially the plants with generation times measured in years, the effects is

still evident after a month. The CO2 treatment, however, is continuous and is therefore seen at all organism

levels, as is the defoliation treatment that is concluded days before the final harvest. It is also interesting to

note, in reference to the hierarchy proposed by Bardgett et al. (2013), that after 7 years of treatment, the

global change effects are not seen aboveground by way of community reordering of the vegetation. Rather,

the effects are seen in the belowground community. At the given timescale, organisms with different

generation times are therefore at different levels in the response hierarchy of Bardgett et al. (2013). Thus, we

have demonstrated individual adjustment aboveground and community reordering sensu Bardgett (2013)

belowground.

Belowground response to defoliation depends on plant growth phase In Stevnbak et al. (2012) a comparable amount of aboveground biomass was removed by grasshopper

defoliation and was done in September (after flowering and seed-set, at the end of the growing season). The

present study was done in June, before seed-set, where we expected that the plants were still investing a

considerable amount of resources aboveground. Indeed, there was a considerable regrowth of the grasses

contrary to the September results from Stevnbak et al. (2012), where there was no compensatory growth in

the defoliated grasses. However, contrary to Stevnbak et al. (2012), we found that aboveground defoliation

reduced belowground biota – both microbial biomass and nematode abundance. The system is presumably C

limited as there are stimulating effects of elevated CO2, which through increased photosynthesis is likely to

also increase C allocation belowground. Since the defoliation does not have a stimulating effect, it seems

reasonable to assume, that the difference in the results of the two studies are at least in part caused by

differences in allocation of resources belowground and the derived changes in root exudation, determined by

growth phase and the need for resources aboveground for production of biomass, photosynthesis and

flowering/seed-set. This difference in plant growth phase is confirmed by the lack of growth of non-

defoliated vegetation during the September study, while the June study shows a considerable growth even in

the non-defoliated units. In accordance with this, Frank (1998) finds a positive relationship between forage

consumption and plant production in the growing season, and Wilsey (1996) finds an increased shoot

production of grass defoliated soon after having been brought out of mimicked winter dormancy. Also in

support of our results on investment of resources aboveground instead of in root-exudation in plants

defoliated before seed-set, Ilmarinen et al. (2008) finds a reduced allocation of C to roots and an increased

allocation of N to shoots – without a corresponding increase in N uptake – upon defoliation. The study was

done on relatively young plants still in active growth, and indicates an altered internal allocation of C and N

in the plant rather than increased uptake and shows no stimulation of soil biota at defoliation (Ilmarinen et al.

2008). A study on defoliation of 8 week old plants in microcosms (Stanton 1983) and a grassland field study

50

of defoliation effects in spring (Todd 1996) showed reduced nematode abundances comparable to our

findings.

Effects through the growing season In Stevnbak et al. (2012) the defoliation-induced stimulation of belowground biota and nutrient availability is

greater under elevated CO2 where photosynthetic capacity of grass plants is increases (Albert et al. 2011),

where they grow more roots (Arndal et al. 2014), and thus contain more resources. In the present study, CO2

stimulated soil biota. Even when significantly reduced by defoliation and drought, the nematode abundance

was higher at elevated CO2 than under ambient CO2. Hence, in the two otherwise contrasting parts of the

growing season, increased CO2 stimulate soil biota and thereby likely the decomposer capacity (Blankinship

et al. 2011; Eisenhauer et al. 2012) partly due to increased rhizodeposition (Eisenhauer et al. 2012). Further,

it seems that the often proposed mechanism of plants feeding their belowground microbial loop when in

immediate need of nutrients (Bardgett et al. 1998; Bonkowski 2004) is not present in this natural system,

even though it is indeed relatively nitrogen limited (Larsen et al. 2011). Rather, the joined results of the

present investigation and the Stevnbak et al. (2012) study show the opposite: exudation and belowground

allocation of resources to the advantage of the soil biota mainly occur when the perennial plant is not in need

of resources for shoot growth and nutrients are stored in and released from roots. In the face of predicted

increased CO2 levels in the atmosphere and the derived increased C input belowground and more abundant

decomposer community demonstrated in this study it therefore stands to reason to consider that management

of grazing intensity of natural areas during the season could help modify the effects, as defoliation and CO2

worked antagonistically in the productive part of the season (June) whereas the effects were synergistic later

in the season (September).

Conclusion With this study, we wanted to explore if the global change effects on aboveground-belowground interactions

found in Stevnbak et al. (2012) are related to plant growth phase. Among elevated CO2, warming and

summer drought, we find CO2 to be the most distinctly influential global change factor, stimulating all

measured belowground pools, reflecting a more carbon rich environment with higher decomposition. We

find this stimulation to be counteracted by a general defoliation effect, but find only little interaction between

global change factors and defoliation. Instead, we find defoliation effects to depend on plant growth phase:

If the results from Stevnbak et al. (2012) and other studies showing a stimulation of belowground biota by

defoliation (Ostle et al. 2007; Hamilton et al. 2008; Mikola et al. 2001) were indeed due to plants releasing

carbon to feed the microbial loop when in need of nutrients, we would expect a more pronounced response

when plants are in active growth than when the growth conditions are less favorable and the growing season

is terminating. However, when we compare defoliation impact on soil biota before seed-set (this study) with

51

impact after seed-set (Stevnbak et al. 2012), we find that the aboveground-belowground interactions upon

defoliation seem to depend on prioritizing of resources related to aboveground growth rather than on the

plants induction of the rhizosphere associated biota. These results emphasize the need for further

investigation into whether plants are strategically regulating the life around its roots or if their inputs into the

soil simply reflect differences in flow of resources depending on varying needs within the plant.

Acknowledgements We would like to thank Gosha Sylvester, Annette Spangenberg, Michelle Schollert, David Byriel and Karna

Heinsen for assistance in the laboratory. We would also like to especially thank Mette Vestergård for

valuable discussions and comments to the manuscript. The CLIMAITE project has been funded by the

Villum Kann Rasmussen foundation and further supported by Air Liquide Denmark A/S.

52

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Paper III

58

Fire in the tallgrass prairie: effects on soil communities and trophic transfers of

carbon during litter and pyrogenic organic matter decomposition

Marie Dam1, 2, Jennifer L. Soong1, Diana H. Wall1, M. Francesca Cotrufo1

1 Natural Resource Ecology Laboratory, Colorado State University, Fort Collins, CO, USA 2Terrestrial Ecology Section, Department of Biology, University of Copenhagen, Copenhagen, Denmark

Abstract Grassland ecosystems are frequently disturbed by anthropogenic fire management and wildfires. During a

fire biomass is partially combusted and litter inputs to the soil are substituted with inputs of pyrogenic

organic matter. This is a considerably more recalcitrant organic input than fresh litter. In order to determine

the effect of burning and the altered organic input on the trophic structure of the soil biota and the

decomposition process, we traced both litter and pyrogenic organic matter into the soil food web. We

incubating 13C-labeled litter and pyrogenic OM for 11 months at tallgrass prairie sites and analyzed uptake

into microbial PLFAs and nematodes. To separate the effect of pyrogenic organic matter from other effects

of burning, we did the same experiment at both an annually burned and site burned only once in a decade.

Litter-derived C was incorporated into both microorganisms and nematodes while pyrogenic organic matter

was left largely undecomposed by the microbes, and the soil communities in the plots amended with

pyrogenic organic matter were comparable to those in bare soil plots. Annual burning generates a soil micro

food web with significantly reduced abundance of predators and omnivores and with increased abundance of

microbial feeders and greater uptake of carbon from the offered bioavailable litter source. Together with

unavailability of the pyrogenic organic matter, this altered structure could be contributing to the observed

decoupling of C and N cycles in frequently burned ecosystems.

Introduction The role of soil food web dynamics in belowground processes of decomposition and nutrient cycling is a

complex component of the soil system (Carrillo et al. 2011; Nielsen et al. 2011; Petersen & Luxton 1982;

Mikola & Setälä 1998; Bardgett et al. 1999; Ekschmitt et al. 1999; Bonkowski et al. 2000; Setälä 2002; de

Vries et al. 2013). Soils contain approximately 80% of global terrestrial carbon stocks and exhibit strong

feedbacks to global climate through decomposition, soil organic matter formation and support of primary

plant productivity (IPCC 2013). Soil biota are the main actors responsible for the transformation and

recycling of carbon and nutrients in the soil, and their abundance and activity can be greatly impacted by

land management practices and climate controls (Bardgett et al. 1997; Blankinship et al. 2011; Frey et al.

59

2008; Garcia-Palacios et al. 2013). In grasslands, which cover nearly one fifth of the Earth’s land surface

and store an estimated 30% of the total world soil carbon (Anderson 1991; Grieser et al. 2006), burning is

often used as a land management practice to promote grassland productivity (Hall & Scurlock 1991;

Mouillot & Field 2005). Pyrogenic organic matter (pyrogenic OM) produced from fires is estimated to

comprise up to 80% of soil organic matter (SOM) in some soils (Schmidt & Noack 2000). The impact of

grassland fires on soil biological processes through the removal of aboveground litter inputs and the addition

of pyrogenic OM to the soil could thus alter the flow of C and nutrients through the soil and the ecosystem.

Soil microbes are responsible for the majority of plant material decomposition in the soil (Paul

2007), however soil nematodes have been found to stimulate decomposition by aiding redistribution of and

applying top down controls on microorganisms (Bardgett 2005) and to play a substantial role in nutrient

mineralization, particularly of nitrogen (Osler & Sommerkorn 2007; Neher et al. 2012). Nematodes can

occur at densities of approximately 1 million to 10 million m-2 in grasslands (Yeates et al. 1997; Bardgett et

al. 1997), and thus play a fundamental role in decomposition dynamics (Griffiths 1994; Neher 2001; Nielsen

et al. 2011; Osler & Sommerkorn 2007). Nematodes as a soil fauna group are of particular utility because

they occupy a range of consumer trophic levels within the soil food web. Therefore, their community

structure can provide important insights regarding many aspects of ecosystem function (Ferris 2010). Annual

burning of the tallgrass prairie has been found to alter soil nematode community composition compared to

unburned prairie (Todd 1996). Soil microbial biomass has been found to be lower in annually burned

tallgrass prairie in comparison to unburned prairie (Ajwa et al. 1999). However, annual burning has been

found to promote higher rates of C mineralization (Knapp et al. 1998b, Johnson & Matchett 2001) but N

availability than the tallgrass prairie (Johnson & Matchett 2001).

Pyrogenic OM remaining from fires has recently been highlighted as a previously unidentified yet

significant component of persistent SOM in burned soils (Singh et al. 2012; Schmidt et al. 2011; Knicker et

al. 2012). Due to its highly aromatic chemical structure (Knicker 2011; Masiello 2004), pyrogenic OM has

been found to be largely resistant to microbial decomposition (Kuzyakov et al. 2014). However, studies on

pyrogenic OM and biochar have found that soil microbes, particularly gram-negative bacteria, can utilize an

initial labile fraction of charred material (Gomez et al. 2014; Foereid et al. 2011; Kuzyakov et al. 2014;

Soong and Cotrufo Submitted), as revealed by the use of isotope tracing. However, the effect of pyrogenic

OM inputs on soil fauna and trophic level nutrient cycling is one of the least studied components of

pyrogenic OM and biochar research (Lehmann et al. 2011). The removal of aboveground litter inputs to the

soil by burning, and its replacement with partially combusted pyrogenic OM, could alter the energy and

nutrient inputs from plants to the soil food web. A history of annual litter removal and replacement with

pyrogenic OM could thus play a role in altering C and N cycling in annually burned grasslands, such as areas

of the tallgrass prairie in the central great planes region of the USA.

60

Our research aims to understand how fire affects soil microbial and nematode communities through

the removal of aboveground litter and their replacement with pyrogenic OM inputs to the soil. We are

specifically interested in how litter or pyrogenic OM inputs, as well as history of burning, affects soil

biological activity and community composition. If pyrogenic OM itself or a history of annual burning does

alter the activity or community composition of soil biota it could reveal some important and previously

unrecognized effects of fire on soil C and nutrient cycling. We hypothesize (H1a) that aboveground inputs

to the soil in the form of litter is decomposed while the soil food web is largely incapable of decomposing

aboveground inputs to the soil in the form of pyrogenic OM. But (H1b) that to the extent that any

decomposition does happen, it is primarily seen at the annually burned site, due better adaptation of

decomposer food web at this site to extract energy from pyrogenic OM. Additionally we hypothesize (H2a)

that a history of annual burning alters the community composition of soil microbes and nematodes, and

(H2b) that any burning derived differences in soil community composition is due to replacement of

biologically available litter with recalcitrant pyrogenic OM.

To test these hypotheses, we conducted a field experiment incubating 13C labeled Andropogon

gerardii above-ground litter and pyrogenic OM produced from the partial combustion of the same litter in an

annually burned (AB) and unburned (UB) tallgrass prairie site for 11 months. Using the 13C label, we traced

litter and pyrogenic OM decomposition into microbial phospholipid fatty acids (PLFAs) and nematodes.

Materials and methods 13C labeled litter and pyrogenic organic matter Dual and uniformly 4 atom % 13C and 7 atom % 15N labeled Andropogon gerardii was grown in a 13C

continuous labeling chamber as described in Soong et al. (2014). A. gerardii was started from seeds and

grown in the labeling chamber to maturity for 15 weeks. Then the plants were removed from the chamber

and the aboveground senesced biomass (litter) was harvested by cutting at the crown, and air-dried. Half of

the labeled A. gerardii litter was pyrolyzed for four hours at 400ºC in a muffle furnace with ultra-high purity

nitrogen flow, as described by Rutherford et al. (2012) to produce our labeled pyrogenic OM amendment.

Three replicates of the initial litter and pyrogenic OM were analyzed for %C and δ13C on an elemental

analyzer connected to an isotope ratio mass spectrometer (EA-IRMS, Carlo Erba NA 1500 coupled to a VG

Isochrom continuous flow IRMS, Isoprime inc.).

Experimental site and design The study was conducted at the Konza Prairie long-term ecological research station in Kansas, USA. This is

a tallgrass prairie, dominated by A. gerardii. Climate at the site is temperate-continental, with average

annual precipitation of 835 mm and a mean annual temperature of 12.8ºC. Two sites were chosen for this

study. One site was burned annually from 1972 to 2000, when burning treatments ceased, and was then left

61

unburned from 2000 onwards, except for one wild fire in 2008. This site is hereafter referred to as unburned.

This site was chosen since it did not have a recent history of burning, but due to its previous burning regime

A. gerardii still dominates the system. The other site has been burned in the springtime annually from 1972

until present, and is referred to as the annually burned site. Both sites are on topographically located on

footslopes, with silty-clay textured mollisol soils (Knapp et al. 1998a).

The experiment consisted of two treatments: a site treatment with two levels (annually burned and

unburned), and a soil surface amendment treatment with three levels: litter, pyrogenic OM, and a bare soil

control. At each of the two sites, the surface amendment treatments were replicated 4 times in a randomized

block design. The experimental unit was a PVC collar (20 cm diameter and 10 cm tall) inserted in the ground

to 5 cm. All collars were inserted 24 hours prior to the start of the experiment, when the native litter was

removed. The 13C labeled A. gerardii litter was added to the collars at the rate of 400 g/m2, which is the

estimated above ground net primary annual production (ANPP) at the site (Knapp et al. 1998a). The

pyrogenic OM was added at the rate of 132 g/m2, corresponding to a 30% burning recovery of the ANPP.

The litter was placed on the surface of the soil in the collars and the pyrogenic OM was sprinkled on the

surface of the soil and tilled in to 2 cm to limit wind erosion. The bare soil collars (with no above ground

inputs) served as the natural soil control end member for the isotope-mixing model (See Data Analysis). 11

months after starting the experiment, soil was sampled at 0-5 cm depths in each PVC collar. The experiment

is thus treated as a split plot on a randomized complete block design.

At the unburned site, a time treatment was added with two sampling times: 4 and 11 months.

Consequently, a separate set of 4 replicate blocks was sampled at this site 4 months after the start of the

experiment. This was to ensure that we did not miss any potential short-term pyrogenic OM decomposition

and C uptake in the soil community. Analyses methods and results from this sampling can be found as

supplementary material.

Soil sampling Soil was sampled at the unburned site after 4 months of treatments (Sep. 8, 2012) and at both sites after 11

months (April 4, 2013). Soil and litter samples were collected from each of the four replicate collars of each

treatment. First, where present, the litter was collected by hand and stored in plastic bags. Then, the soil

within the collar was gently excavated with the use of hand shovel by incremental depths of 0-2 and 2-5 cm

and stored in plastic bags. All soil and litter samples were stored with ice in coolers before being brought to

the laboratory the following day. There they were stored at 4ºC until they were processed within two weeks

of collection.

62

Litter and pyrogenic OM recovery As reported by Soong and Cotrufo (Submitted), after 11 months in the field the litter had lost 55% of its

initial mass. Only 60-75% of the initially applied pyrogenic OM mass was recovered after 11 months,

possibly due to wind erosion of the topsoil where the lightweight pyrogenic OM was applied (Soong and

Cotrufo Submitted), despite measures taken to prevent this. The total amount of litter and pyrogenic OM

derived C in the top 5 cm of the soil presented in Table S1 accounts for the depths of the soil where we

extracted PLFAs and nematodes. Litter and pyrogenic OM C and N were recovered down to 40 cm, however

(Soong and Cotrufo Submitted).

PLFAs extractions and 13C-PLFA measurements A sub sample of the sieved soils from the 0-2 and 2-5 cm depth layer was picked clean of all visible roots,

frozen (-20ºC) and lyophilized for 48 h prior to PLFA extraction. PLFAs were extracted on these samples

using conventional methods (Bligh & Dyer 1959; Denef et al. 2007). In brief, for the extraction, 6 g of

freeze-dried soil were mixed with a 0.1 M potassium phosphate buffer:chloroform:methanol solution (0.8:1:2

ratio volume, ml g-1 of soil). Neutral, glyco- and phospholipids were separated over SPE silica columns

eluting respectively with chloroform, acetone and methanol. Phospholipids were saponified to obtain free

fatty acids, which were subsequently methylated using 0.2 M methanolic KOH to form fatty acid methyl

esters (FAMEs). FAMEs were quantified and analyzed for 13C by capillary gas chromatography combustion

isotope ratio mass spectrometry (GC-c-IRMS, Trace GC Ultra, GC Isolink and DeltaV IRMS, Thermo

Scientific). A capillary GC column type DB-5 was used for FAME separation (length 30 m, i.d. 0.25 mm,

film thickness 0.25 µm; Agilent). The GC temperature program proceeded at 60ºC with a 0.10 min hold,

followed by a heating rate of 10ºC min-1 to 150ºC (2 min hold), 3ºC min-1 to 220ºC, 2ºC min-1 to 255ºC, and

10ºC min-1 to 280ºC with a final hold of 1 min. Individual fatty acids were identified based on relative

retention times to an internal standard (12:0), which was added to the FAME extract prior to gas

chromatography, and cross referenced with several standards: a mixture of 37 FAMEs (37 component FAME

Mix, 47885-U, Sigma-Aldrich, USA). FAME identification was verified by analyzing a few samples on a

capillary GC-mass spectrometer (Shimadzu QP-2010SE) with a SHRIX-5ms column (30 m length x 0.25

mm i.d., 0.25 µm film thickness) using the NIST 2011 mass spectral library.

Quantification was performed using relative response factors (RRF) relative to an internal standard

(19:0), added to the FAME extract prior to GC analysis. RRFs were determined in advance by using a

dilution series of the 37-component FAME mix, to which the 19:0 standard was added. The abundance of

individual PLFAs was calculated in absolute C amounts (ng PLFA-C g-1 soil) based on the PLFA-C

concentrations in the liquid extracts.

The biomarker PLFAs analyzed within this dataset included: 18:1ω9c and 18:2ω6,9c (indicative of

saprophytic fungi), 16:1ω5 (indicative of arbuscular mycorrhizal fungi-AMF), i15:0, a15:0, i16:0, i17:0,

63

a17:0 (indicative of Gram-positive bacteria), cy17:0, cy19:0, 16:1ω7c and 18:1ω7c (indicative of Gram-

negative bacteria), 14:0, 15:0, and 18:0 (indicative of non-specific (ns) bacteria) and 10Me PLFAs

(indicative of actinobacteria) (Kroppenstedt 1985; Olsson et al. 1995; Zelles 1997). Total fungi, bacteria,

gram-positive bacteria and gram-negative bacteria PLFA quantities were calculated as the sum of all PLFA-

C from each biomarker group. 13C values were corrected using the working standards (12:0 and 19:0)

calibrated on an EA-IRMS. To obtain δ13C values of the PLFAs, measured δ13C FAME values were

corrected individually for the addition of the methyl group during transesterification by simple mass balance

(Denef et al. 2007).

Nematode extractions and 13C isotope analysis For nematode community analysis, soil from the 0-2 and 2-5 cm depth layer was mixed in a 2:3 ratio. 75 g of

soil mixture was extracted by the Baermann funnel method for 72 hours. A couple of grams of the re-

collected litter were extracted by the same method to include the litter-colonizing nematodes in the analysis

(see S4). However, they represented an insignificant part of the overall abundance, and were left out of the

analysis, to be able to report nematode abundances per g soil. Within days of extraction, nematodes were

counted and sorted to trophic groups (see Yeates et al. 1993) while live, for community composition

analysis. Both the total abundance and abundance of each trophic group is given in nematodes per g soil (or

per g litter).

For isotopic analysis of the nematode community as a whole, 100 individuals from each sample were

randomly? picked into tin capsules (8x5mm, Elemental Microanalysis BN/170056). The capsules were sent

to Kansas State University's Stable Isotope Mass Spectrometry Laboratory (SIMSL) and analyzed for C and

N composition as well as isotopic values of 13C and 15N using a CE-1110 CHN element analyzer (EA) for

sample combustion and separation. The EA was coupled to a Finnigan Delta Plus mass spec via a Conflo II

interface. A 3m gas chromatography column was packed with poraplot Q to separate N2 and CO2. Average

nematode biomass %C values were obtained from an average of 360 nematodes extracted from soils at the

site (E.A. Shaw et al., unpublished data), and the unit reported is ng input-derived nematode-C/g dry soil.

Data analysis

The litter and pyrogenic OM carbon contribution to PLFA and nematode biomasses was assessed for

the litter and pyrogenic OM treated plots as compared to the bare soil plots. The isotopic mixing model was

applied as follows:

𝑓𝑓𝑏𝑏𝑏𝑏𝑏𝑏𝑏𝑏 = 𝛿𝛿𝑆𝑆 − 𝛿𝛿𝐵𝐵𝛿𝛿𝑏𝑏𝑏𝑏𝑏𝑏𝑏𝑏 − 𝛿𝛿𝐵𝐵

where fblue is the litter or pyrogenic OM derived C fraction of a PLFA or nematode sample, δS and δB is the

δ13C of the specific PLFA or nematode sample from a litter or pyrogenic OM plot (δS) and the corresponding

64

bare (δB) plot; and δblue is the δ13C of the initial litter or pyrogenic OM. First, the size of the PLFA or

nematode C pool was determined for the litter or pyrogenic OM amended plots, then it was multiplied by the

fblue to quantify the amount of litter or pyrogenic OM derived C in each pool.

We tested the effect of burning history on PLFA biomarker abundance and the amount of pyrogenic

OM and litter incorporation into the PLFAs between the unburned and annually burned sites using a general

linear mixed model with site (unburned, annually burned), input (no input (bare soil), litter, pyrogenic OM),

depth (0-2 cm, 0-5 cm) and their interactions as categorical fixed effects, with block and the interaction

between block and site as categorical random effects. We used the SAS software, version 9.3 and used type

III tests of fixed effects. We checked for normality of the data and homogeneity of variances of the residuals

and used a log transformation when necessary.

There was no depth factor for the nematodes, as they were all extracted from soil from 0-5 cm depth.

Therefore, we tested amount of pyrogenic OM and litter derived C incorporation into the nematodes using

mixed linear models with input, site, and their interaction as fixed factors in R

(R_Development_Core_Team, 2013) using the lmer function from the lme4 package (Bates et al. 2014). We

applied log-transformation to obtain normality. For the litter-extracted nematodes, site was the only fixed

factor.

To test the effect of organic matter input and site on the overall nematode community based on the

abundance of the individual trophic groups, we analyzed the multivariate abundance data in R

(R_Development_Core_Team 2013), using the manyglm function from the mvabund package (Wang et al.

2013), which fits a separate generalized linear model to each trophic group, using a common set of

explanatory variables. Here, with input (no input/bare soil, pyrogenic OM or litter), site (unburned or

annually burned), and their interaction as fixed factors. Furthermore, we tested the effect of burning history

on total nematode abundance as well as on the abundance of each trophic group in the nematode community

between the unburned and annually burned sites at the 11 month sampling using mixed linear models with

input (no input/bare soil, pyrogenic OM or litter), site (annually burned or unburned), and their interaction as

fixed factors. All nematode abundance data was analyzed in R (R_Development_Core_Team 2013) using the

lmer function from the lme4 package (Bates et al. 2014). We applied log-transformation when necessary to

obtain normality.

To test the effect of litter or pyrogenic OM input and site on the overall microbial community based

on the relative contribution of all of the individual PLFAs to the entire extractable PLFA pool (mol %), we

utilized a distance-based redundancy analysis (dbRDA) using the R: Vegan package (Oksanen et al. 2013),

following the approach described in Bell et al. (2014). Briefly, we chose the dbRDA analysis over other

multivariate statistical approaches due to its non-linear distance-metric options, which have robust multi-

dimensional resolution to assess categorical variables. Distance based RDA is a three step ordination

technique that tests the effects of response parameters (i.e. mol %) on defined groups (i.e. input, site or time).

65

First, a dissimilarity or distance matrix is calculated for the different treatments. We chose the Bray-Curtis

dissimilarity (non-linear) measure to model the species matrix as suggested by Legendre & Anderson (1999).

For steps two and three of the dbRDA, a principal coordinate analysis (PCoA) is calculated based on the

distance matrix, from which the eigenvalues (obtained in the PCoA) were applied to a redundancy analysis

(RDA).

Results Litter and pyrogenic OM uptake by microbes and nematodes The pyrogenic OM and litter derived 13C was recovered in all of the examined PLFAs, but distributed

differently across the groups of micro-organisms (Fig. 1). The magnitude of PLFA uptake of litter C far

exceeded that of pyrogenic OM C by over 50 fold (Fig. 1). The gram-negative bacteria were the group

responsible for the greatest uptake of both litter and pyrogenic OM C (Fig. 1 and Fig. S1) in accordance with

the greater abundance of gram-negatives (Table 1). The amount of input taken up by fungi relative to

bacteria did not differ between the different input treatments. Time did not have a significant effect on the

overall amount of litter or pyrogenic OM derived C built into the PLFAs: the total amounts of input-derived

PLFA-C were the same after 11 months as after 4 months for both litter and pyrogenic OM (P>0.05,

unburned site) (Fig. 1 and Fig. S1). Litter derived 13C was recovered from higher trophic levels of the

decomposer food web, the nematodes (Fig. 2). We did not recover any pyrogenic OM derived 13C in the

nematode community at either site (Fig. 2) or at either time (Fig. 2 and Fig. S2), indicating that the pyrogenic

OM derived C is not distributed in the part of the soil food web represented by the nematodes. We saw a

numerical albeit not statistically significant increased incorporation of litter derived C in the nematode

biomass with time at the unburned site (Fig. 2 and S2).

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Figure 1 Incorporation of a) litter derived C and b) pyrogenic OM (py-OM) derived C in the microbial phospholipid fatty acid (PLFA) after 11 months of incubation in the 0-2 cm soil at the unburned (UB) and annually burned (AB) site, respectively. PLFAs biomarkers were summed by microbial group, error bars are standard error, n=4.

Burning history effects on microbe and nematode uptake of litter and pyrogenic OM At the 11-month harvest, there was significantly more litter C incorporated into the PLFAs of the annually

burned site than the unburned site (P<0.05, Fig. 1a). The amount of litter derived 13C recovered in the

nematode biomass did not differ significantly between the two sites (Fig. 2). Burning history did not have a

significant effect on the overall PLFA incorporation of the pyrogenic OM C (P>0.05, Fig. 1b). However, of

the very limited pyrogenic OM C uptake that occurred, the gram-positive:gram-negative ratio of C

incorporation had a tendency (P=0.072) to be higher at the annually burned site (=0.73) than at the unburned

site (=0.11), meaning that gram-positives are incorporating pyrogenic OM C to a relatively greater extent at

the annually burned site than at the unburned site (see Fig. 1b). A corresponding difference between sites

was not seen for litter C incorporation.

* *

*

*

*

*

*

a)

b)

*PBurning<0.05,

litter C uptake

Total py-OM-C incorporation UB: 245±15 AB: 188±25

Total litter-C incorporation UB: 10559±90 AB: 17859±874

67

Figure 2 Incorporation of pyrogenic OM derived and litter derived carbon in the nematode biomass after 11 months of incubation in the 0-5 cm soil. Results from Two-way ANOVA on input and site are shown. Error bars are standard error, n=4.

Litter and pyrogenic OM input effects on microbial and nematode abundance and community Litter and pyrogenic OM inputs affected the abundance of different microbial PLFAs alike for at the two

burning histories after 11 months (Table 1).Compared to bare soil, addition of pyrogenic OM did not affect

fungi, total bacteria or the gram negative fraction, whereas gram-positive bacteria were stimulated. In

contrast, all four groups were reduces in the litter amended plots compared to bare soil (Table 1). There was

an interaction between time and input affecting the development of the microbial abundances at the unburned

site (Table S2). In the pyrogenic OM amended and bare soil plots, the microbial abundances increased with

time, whereas they decreased in the litter amended plots (Table S2). At the unburned site, the nematode

abundance generally decreased in all plots with time (P<0.05) (Fig. 4 and Fig. S3).

Input: P<0.0001 Burning: NS Burning*Input: NS

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Table 1 Abundances of PLFAs from different fuctional groups of soil microbes. Values are averages for topsoil (0-5 cm depth) with standard error in parentheses (n=8). In the Site|Input model, only the main effect of input was significant. Letters indicate statistical similarity or difference within columns.

Input Site

Fungi

(ng C/g soil)

Total Bacteria

(ng C/g soil)

Gram+

Bacteria

(ng C/g soil)

Gram-

Bacteria

(ng C/g soil)

Bare soil

Unburned 2487±164 a

12663±756 a

4680±257 b

7135±464 a Annually

burned 2533±239 12799±631 4728±423 7121±206

Pyrogenic OM added

Unburned 2126±71 a

13130±433 a

5354±193 a

6809±234 a Annually

burned 2221±289 15086±1495 7285±1820 6513±713

Litter added

Unburned 1625±26 b

8274±128 b

3186±103 c

4624±41 b Annually

burned 1739±88 9297±326 3406±158 5338±264

Input (litter, pyrogenic OM or bare soil) had a significant effect (P<0.005) on the microbial community

composition, as assessed by the dbRDA analysis of the relative contribution of the individual PLFAs (Fig.

3). There was a time effect on both the gram-positive:gram-negative bacteria ratio and the fungi:bacteria

ratio The bacteria increased relatively more than the fungi and the gram-positives bacteria increased

relatively more than the gram-negatives with time in the bare soil/pyrogenic OM plots (Table S2). In the

litter plots where the microbial abundances decreased with time, the ratios remained unchanged.

The multivariate analysis of the nematode community showed that input (litter, pyrogenic OM or

bare soil) had significant impact on the community composition (Fig. 4). Looking more into each trophic

group with univariate tests, the litter amended plots had higher abundances at both the bottom and top of the

trophic group range (Fig. 4), which generated a significantly higher total abundance of nematodes compared

to both the bare soil and the pyrogenic OM plots (P<0.001 and P=0.003, respectively). The bacterivores and

omnivores/predators were significantly more abundant in the litter amended samples than in both the bare

soil and the pyrogenic OM amended samples (Fig. 4, Table 2). Neither total abundance nor abundance of any

feeding group differed significantly between bare soil plots and pyrogenic OM amended plots (Table 2).

69

Figure 3 DbRDA results of the microbial community composition using relative abundance of the individual PLFAs after 11 months analyzed by input, where B= bare soil, L= litter and P= pyrogenic OM.

Burning history effects on microbe and nematode abundance and community History of burning did not have a significant effect on the abundance of fungi or bacteria (nor gram-positive

bacteria or gram-negative bacteria) (Table 1). Furthermore, annual burning did not have a statistically

significant effect on the overall microbial community composition, as assessed by the dbRDA analysis of the

relative contribution of the individual PLFAs grouped by functional group (Data not shown). The total

abundance of nematodes was significantly higher at the annually burned site (P=0.003) (Fig. 4), and the

annual burning also changed the structure of the community. The multivariate analysis of the assembly of

nematode trophic groups showed a significant effect of burning history on the community composition

(P=0.004). The univariate analyses confirm this, with different effects of burning on the abundances of the

lower and upper part of the trophic range in the community. The bacterivores (P<0.001) and fungivores

(P=0.007) were significantly more abundant at the annually burned site than at the unburned site (Fig. 4).

The omnivores and predators, however, were reduced by the annual burning, with significantly lower

abundances at the annually burned site (P<0.001) (Fig. 4).

a

70

Figure 4 Number of nematodes per g dry soil after 11 months of field incubation. The four trophic groups are stacked for each input treatment (bare soil, pyrogenic OM and Litter) at each site (Unburned, Annually burned), and the total abundance represented by the dashed line. Significant effects at P<0.05 from multivariat glm of community composition are shown. Error bars are standard error, n=4. Table 2 Differences in trophic group abundances between treatment inputs. The Two-Way ANOVA on data from 11 months sampling analysis showed Site and Input, but not their interaction to be significant. Letters therefore indicate statistical differences within a column, within each site.

Site Input Bacterivores Fungivores Herbivores Omnivores/ Predators

Unburned Bare soil a a a a Pyrogenic OM a a a a Litter b a a b

Annually burned

Bare soil p p p p Pyrogenic OM p p p p Litter q p p q

Discussion Litter and pyrogenic OM input effects on soil biota

We hypothesized (H1a) that soil microbes and nematodes biologically decompose aboveground

plant inputs to the soil in the form of litter but not pyrogenic OM. Evidence from 13C incorporation into

microbial PLFAs and nematodes was used to track the biological decomposition of litter-derived C through

the soil food web, and likewise reveal the lack biological decomposition and food web incorporation of

pyrogenic OM during the first 11 months of incubation in the field. Approximately 50 times more litter C

Site: P=0.001 Input: P=0.001 Site*Input: n.s.

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was incorporated into PLFAs than pyrogenic OM C (Fig. 1). Although use of isotope labeling reveals that a

small fraction of lipids and polysaccharides from fresh pyrogenic OM can be taken up by microbes into

PLFAs (Gomez et al. 2014; Foereid et al. 2011), the extremely low rates of utilization compared to litter and

SOM reaffirm the overall low microbial use of pyrogenic OM as a C source (Kuzyakov et al. 2014; Singh et

al. 2014). This is further supported by comparing the results from the 4 and 11 months sampling at the

unburned site, showing no significant time effect, and the same lack of uptake as after 11 months (P>0.05,

Fig. S1). The tallgrass prairie soils had a general dominance of bacteria relative to fungi, with a dominance

of gram-negatives (Table 1 and S2). The limited uptake of pyrogenic OM that did occur was predominantly

by the gram-negative bacteria (Fig. 1 and S1), supporting previous findings (Gomez et al. 2014), but a

stimulation of abundance by pyrogenic OM amendment was found only for gram-positives, along the lines

of Santos et al. (2012). In accordance with their generally greater abundance, gram-negative bacteria were

also the dominant users of litter.

Nematodes and soil fauna in general are known to be important regulators of C cycling (Nielsen et

al. 2011) and ecosystem productivity (Neher 2001) through their role as top down grazers of soil microbes

and as mineralizers of N. The transfer of 13C derived from litter through both soil microbes and nematodes

clearly demonstrates this function at our tallgrass prairie site. The already very limited incorporation of

pyrogenic OM in the microbial PFLAs is not found to be further distributed in the soil food web for which

the nematode community serves as a proxy (Bongers & Ferris 1999, Neher 2001). Litter derived C was

found to be incorporated in the nematode biomass, but no pyrogenic OM derived C was found in the

nematode biomass (Fig. 2). This suggests that the nematodes did not ingest pyrogenic OM or microbes that

had incorporated pyrogenic OM - or, alternatively, that the dilution of the isotopic signature from the small

amount of microbially incorporated pyrogenic OM could have been too great to detect any pyrogenic OM

derived C in the nematodes. The inability of the pyrogenic OM to serve as a food source for the decomposer

food web is further confirmed by the results showing that only the litter, not the pyrogenic OM increases the

abundance of nematodes relative to the un-amended bare soil (Fig. 4).

Neither the microbial nor the nematode community differed markedly in composition between the

bare soil plots and the pyrogenic OM amended plots at any of the sites. Thus, we cannot conclude, that the

pyrogenic OM addition alone changes the soil communities (H2b). To the extent the pyrogenic OM

represents a removal of litter, however, it has substantial implications, as the communities and carbon uptake

differs significantly between the litter amended plots and the bare soil/pyrogenic OM plots. As described, the

litter amended plots had a significantly higher uptake of C in both microbes and nematodes, and the

increased energy input manifested in a significantly higher abundance of all trophic groups of nematodes.

The microbial abundance was lower in the litter plots at the 11 month harvest, but at the sampling done after

4 months in the unburned site, the microbial abundance was significantly higher in the litter plots compared

to the bare soil/pyrogenic OM plots. This is likely due to trophic interactions of the decomposition process,

72

where the initial boost of microbial abundance upon litter addition, is since grazed down by the nematodes.

As the litter resource is gradually depleted, the microbial growth stagnates and abundance is thus reduced by

grazing, which eventually also reduces nematode abundance after 11 months. The C cycle in the bare

soil/pyrogenic OM plots is bound to be more slow as this process is forced to switch to the much more

recalcitrant C sources present in the soil organic matter pool, when the plants are removed from the plots at

the beginning of the experiment. We see an increase in microbial abundance from 4 to 11 months at the

unburned site. This could be a combination of acclimatizing to a different C source and a reduced grazing

pressure, as the low abundance of nematodes in these plots are further reduced with time. We see an

increased gram-positive:gram-negative ratio in the bare soil/pyrogenic OM plots from 4 to 11 months in

accordance with Fierer et al. (2003), Bossio & Scow (1998) and Griffiths et al. (1999), who found gram-

positive bacteria rather than gram negative bacteria to be associated with lower carbon availability. In a

biochar addition study by Zhang et al. (2013), biochar additions were found to have no effect on total

nematode abundance; however, the community composition shifted toward more fungivorous nematodes and

less plant parasites. In our study the pyrogenic OM treated nematode community did not significantly differ

from the bare soil, but our addition rates of pyrogenic OM were much lower than the biochar additions by

Zhang et al. (2013) to simulate natural fire derived inputs, and more importantly, we had no living plants in

our plots.

Although other studies have traced isotopes from decomposing root litter (Shaw et al. Submitted)

and protozoans (Crotty et al. 2012), this is the first study to our knowledge that has traced decomposing

aboveground litter and pyrogenic OM into the decomposer food web in situ. The lack of pyrogenic OM

uptake by soil nematodes reveals a potential impact of aboveground litter removal by burning on soil

biological functioning and trophic interactions.

Burning history effects on soil biota

We hypothesized (H2a) that annual burning alters the community composition of soil microbes and

nematodes, and in fact found significant differences in the nematode community between the sites of

different burning history: At the site that had been burned annually for decades, the abundance of

microbivores was significantly increased (Fig. 4), whereas the omnivores and predators were markedly

reduced by approx. a factor 3. This confirms the previous findings on the effects of annual burning of the

tall-grass prairie on nematode community assemblages found by Todd (1996) and Shaw et al. (Submitted).

It is likely that these higher trophic levels were limited by the annual burning, as they are generally longer-

lived animals with a persister life strategy rather than a fast colonizer strategy (Bongers 1990). Burning-

derived physical conditions such as extreme soil temperatures and xeric post-fire conditions is likely to be

lethal to topsoil organisms (Ahlgren 1974), and with their longer generation times (up to a year or more), the

omnivores and predators may struggle to reach the population densities seen at the unburned site. This

release from predators could be what allows the abundance of microbivores (particularly bacterivores) to

73

increase, which in turn means that we did not see any effects of burning history on the abundance of the

microbial groups (Table 2). Microorganisms are often found to be under substantial top-down regulation in

terrestrial ecosystems (Wardle & Yeates 1993; Bonkowski et al. 2009), and the greater litter C uptake seen at

the annually burned site coupled with higher CO2 fluxes found from decomposing litter (Soong and Cotrufo

Submitted) with no concurrent increase in abundances could be a result of this. The reductions of omnivores

and predators could have implications for the pool of plant available N, due to reduced mineralization (de

Ruiter et al. 1993). We had hypothesized (H1b), that annual burning would cause the microbial community

to develop towards greater ability to utilize pyrogenic OM. However, during our 11-month field study we

saw no difference in the overall pyrogenic OM use by the microbes or the nematodes between the unburned

site and the annually burned site (Fig. 1b and Fig. 2). We did see a tending difference in the gram-

positive:gram-negative pyrogenic OM uptake ratio between the sites. The two groups were both contributing

to the limited pyrogenic OM decomposition at the annually burned site, whereas primarily the gram-

negatives were able to take up pyrogenic OM C at the unburned site. This might represent an adaptation of

the gram-positives by annual burning, but the overall pyrogenic OM utilization remains as low as at the

unburned site. The discrepancy between bacteria stimulated by (gram-positives) and bacteria utilizing

pyrogenic OM (gram-negatives) may relate to population interactions and needs further study.The increased

decomposition and C mineralization of litter we see at the burned site could be due to a greater

decomposition capacity in this soil. Burning is known to increase grass productivity in tallgrass prairie

(Towne & Owensby, 1984; Abrams et al. 1986), and prevent woody encroachment. Burned sites are

therefore likely to have had a greater input of palatable organic matter, and the soils may have a more active

microbial community, which supports higher microbivore abundance consistently seen in all burned site

plots but particularly in the litter amended plots. The differences we see between nematode communities at

sites of different burning history may be an effect of the reoccurring direct lethal effect of the fire,

particularly affecting the populations of the longer-lived predators and omnivores that are relatively confined

to their microhabitats and slow colonizers. The resulting cascading affect down the food web through

decades could be the reason for the changes in community structure, which may furthermore have

implications for N mineralization: In a landmark short-grass steppe study, Hunt et al. (1987) showed that

predacious and omnivorous nematodes were among the groups for which ten-fold increase in abundance

nearly doubled the already significant fraction of total N mineralization attributable to fauna. A three-fold

reduction as we see in this study is therefore likely to affect N mineralization negatively. The increased C

mineralization we observe and the potentially reduced N mineralization could play a role in the decoupling

of C and N cycles that has been raised as a concern regarding high intensity burning (Asner et al. 1997).

74

Conclusions With this novel tracing of pyrogenic OM and litter into the soil micro food web in the field, we wanted to

analyze how frequent burning affects the soil biota and the trophic carbon transfer and specifically

understand the effect of the role of burnt residues. We conclude that to the extent that pyrogenic OM

represents a removal of litter, the fire-derived input significantly changes the functioning of the soil food

web. Microbes do not actively decompose pyrogenic OM inputs to the soil, and what remains in the soil is

unprocessed and mostly unavailable to the soil decomposers. Annual inputs of such bio-unavailable

pyrogenic OM could be inferring limited carbon availability to soil microbes at the annually burned site. To

this end, we see that when given an input of bioavailable litter, the microbes decomposes it at a greater rate

than at the unburned site. This increased decomposition capacity at the annually burned site could be due to

the more bottom-heavy soil food web, created by the often favorable effects of fire on NPP and the harmful

effects of fire on the higher trophic levels of the soil food web. We find several indications of top-down

control of the trophic interactions of the decomposition process, and the initial boost of microbial abundance

upon litter addition, is since grazed down by the nematodes. The overall effect of long term pyrogenic OM

inputs to the soil from fire on soil carbon stocks does not include adaptation of the decomposer community

to this source of organic matter; we believe that the pyrogenic OM that remains in the soil will have a long

residence time due to its resistance to biological decomposition. Our study demonstrates how frequent

burning fundamentally alters biological processes and trophic interactions.

Acknowledgements Special thanks to the staff at Konza Prairie LTER, D. Rutherford at USGS in Denver for producing the py-

OM, and D. Reuss and C. Pinney at the EcoCore analytical facilities (http://ecocore.nrel.colostate.edu/). We

would also like to thank E. A. Shaw for assistance with the stable isotope analysis of nematodes, and our

many field and laboratory assistants including M.L. Haddix, A.J. Horton, D. Cox, E. Bernier, J. Betzen, M.

Jurich, J. Botte, T. Gravina, A. Valente, S. Marciano, C. Larned, J. Lavallee, S. Fulton-Smith, I. Leoni, E. J.

Foster, B. Osborne, X. Jiang and K. Guilbert. The Villum Kann Rasmussen foundation, NSF DEB grant

#0918482 and NSF Graduate Research Fellowship Program funded this work

75

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Supplementary 4 month sampling at unburned site Soil and litter samples were collected from each of the four replicate collars of each treatment at the 4 month

block set at the unburned site. PLFA and nematode extractions were done as decribed in the methods section.

We tested the effect of litter or pyrogenic OM input at the unburned site on the amount of input incorporation

into the bulk soil and PLFAs (H1) using a general linear mixed model with input (pyrogenic OM or litter),

depth, time and their interactions as categorical fixed effects, with block and the interaction between block

and time as categorical random effects. We used the SAS software, version 9.3 and used type III tests of

fixed effects. We checked for normality of the data and homogeneity of variances of the residuals and used a

log transformation when necessary. We tested PLFA biomarker abundance and amount of pyrogenic OM

and litter derived C incorporation into the nematodes at the unburned site using mixed linear models with

input, time, and their interaction as fixed factors in R (R_Development_Core_Team, 2013) using the lmer

function from the lme4 package (Bates et al. 2014). We applied log-transformation to obtain normality. To

test the effect of organic matter input and time on the overall nematode community based on the abundance

of the individual trophic groups, we analyzed the multivariate abundance data in R

(R_Development_Core_Team, 2013), using the manyglm function from the mvabund package (Wang et al.

2013), which fits a separate generalized linear model to each trophic group, using a common set of

explanatory variables. Here, with (1) input (no input, pyrogenic OM or litter), time (4 months, 11 months),

and their interaction as fixed factors. We tested the effect of organic matter input and time on total nematode

abundance as well as on the abundance of each trophic group in the nematode community using mixed linear

models with input, time, and their interaction as fixed factors. All nematode abundance data was analyzed in

R (R_Development_Core_Team, 2013) using the lmer function from the lme4 package (Bates et al. 2014).

We applied log-transformation when necessary to obtain normality.

81

Table S1 Total amount of litter and pyrogenic OM derived C in the top 5 cm of the soil at the annualy burned (AB) and unburned (UB) site. Averages with standard errors in parentheses (n=4)

Litter AB

Pyrogenic OM AB

Litter UB

Pyrogenic OM UB

Initial input C (mg) 7127.07 2483.34 7127.07 2483.34

4 months litter mass remaining

(mg) NA NA 11005.13

(263.94) NA

4 months input derived C in top

5 cm (mg) NA NA 213.97

(11.58) 1265.52 (60.60)

11 months Litter mass remaining

(mg)

8649.31 (116.7) NA 8972.60

(344.31) NA

11 months Input derived C in top

5 cm (mg)

400.88 (64.00)

1824.15 (319.93)

260.30 (9.77)

1363.39 (85.20)

Table S2 Effect of input on microbial PLFA abundances after 4 and 11 months of incubation at the unburned watershed. Values are averages for topsoil (0-5 cm depth) with standard error in parentheses (N=8). Letters indicate statistical similarity or difference within columns.

Time Input

Fungi

(ng C/g

soil)

Total

Bacteria

(ng C/g soil)

Gram+

Bacteria

(ng C/g

soil)

Gram-

Bacteria

(ng C/g soil) Fungi:Bacteria Gram+:Gram-

4

mhs

Bare soil 1983±24b 17609±236a 5590±107a 10765±119ab 0,519c 0,226a

Pyrogenic

OM 1894±97ab 19277±493ab 7255±200ab 10437±251ab 0,695b 0,197b

Litter 2247±122bc 22219±608bc 8307±218b 12049±352bc 0,691b 0,203b

11

mths

Bare soil 2487±164c 25326±756c 9360±257bc 14270±464c 0,657b 0,196b

Pyrogenic

OM 2126±71bc 26259±433c 10709±193c 13618±234c 0,787a 0,162c

Litter 1625±26a 16547±129a 6371±103a 9247±41a 0,689b 0,196b

82

Figure S1 Incorporation of pyrogenic OM (py-OM) derived and litter derived carbon in the PLFA biomass after 4 months of incubation in the 0-5 cm soil at the unburned site. Error bars are standard error, n=4.

Figure S2 Incorporation of pyrogenic OM derived and litter derived carbon in the nematode biomass after 4 months of incubation in the 0-5 cm soil at the unburned site. Results from Two-way ANOVA on input and time are shown (see Fig. 2 for comparison). Error bars are standard error, n=4.

Total C incorporation Py-OM: 175±2 Litter: 10408±259

Input: P<0.0001 Time: n.s. Input*Time: n.s.

83

Figure S3 Number of nematodes per g soil after 4 months of field incubation. The four trophic groups are stacked for each input treatment (bare soil, pyrogenic OM and Litter), and the total abundance represented by the dashed line. Significant effects at P<0.05 from multivariat glm of community composition at 4 months and 11 months at the unburned site are shown (see Fig. 4 for comparison). Error bars are standard error, n=4.

Time: P=0.017 Input: P=0.001 Time*Input: n.s.

84

Nematodes extracted from litter residues To investigate the differences in litter-colonization and look into the community composition in the litter, we

analyzed the litter-extracted nematodes separately. Here, there was no effect of time at the unburned site on

either the abundance or composition of the community. Litter could sustain the same number of nematodes

pr. g dw, and the community remained dominated by fungivores (Fig. S4). There was an effect of site (P=

0.006), however, and after 11 months, the nematodes extracted from aboveground litter residues at the

annually burned site was much lower pr. g litter dw than at the unburned site (Fig. 4).

Figure S4 Number of nematodes per g dry litter. The four feeding groups are stacked for each treatment, and the total abundance represented by the dashed line. Error bars are standard error, n=4.

85

Paper IV

86

Are there links between responses of soil microbes andecosystem functioning to elevated CO2, N deposition andwarming? A global perspectivePABLO GARC�IA - PALAC IOS 1 , 2 , MART I JN L . VANDEGEHUCHTE 2 , 3 , E . A SHLEY SHAW2 ,

MAR IE DAM2 , 4 , KE I TH H . POST 2 , KELLY S . RAMIREZ 5 , 6 , ZACHARY A . SYLVA IN 2 , 7 ,

CEC I L IA MILANO DE TOMASEL 2 and DIANA H. WALL2 , 5

1Centre d’Ecologie Fonctionnelle & Evolutive, CEFE-CNRS, 1919 route de Mende, Montpellier 34293, France, 2Department of

Biology and Natural Resource Ecology Laboratory, Colorado State University, Fort Collins, CO 80523, USA, 3Research Unit

Community Ecology, Swiss Federal Institute for Forest, Snow and Landscape Research WSL, Z€urcherstrasse 111, Birmensdorf

CH-8903, Switzerland, 4Terrestrial Ecology Section, Department of Biology, University of Copenhagen, Universitetsparken 15,

Copenhagen DK-2100, Denmark, 5School of Global Environmental Sustainability, Colorado State University, Fort Collins, CO

80523, USA, 6Netherlands Institute of Ecology, Wageningen 6708 PB, The Netherlands, 7Natural Resources Canada, Canadian

Forest Service, Fredericton, NB E3B 5P7, Canada

Abstract

In recent years, there has been an increase in research to understand how global changes’ impacts on soil biota trans-

late into altered ecosystem functioning. However, results vary between global change effects, soil taxa, and ecosystem

processes studied, and a synthesis of relationships is lacking. Therefore, here we initiate such a synthesis to assess

whether the effect size of global change drivers (elevated CO2, N deposition, and warming) on soil microbial abun-

dance is related with the effect size of these drivers on ecosystem functioning (plant biomass, soil C cycle, and soil N

cycle) using meta-analysis and structural equation modeling. For N deposition and warming, the global change effect

size on soil microbes was positively associated with the global change effect size on ecosystem functioning, and these

relationships were consistent across taxa and ecosystem processes. However, for elevated CO2, such links were more

taxon and ecosystem process specific. For example, fungal abundance responses to elevated CO2 were positively cor-

related with those of plant biomass but negatively with those of the N cycle. Our results go beyond previous assess-

ments of the sensitivity of soil microbes and ecosystem processes to global change, and demonstrate the existence of

general links between the responses of soil microbial abundance and ecosystem functioning. Further we identify criti-

cal areas for future research, specifically altered precipitation, soil fauna, soil community composition, and litter

decomposition, that are need to better quantify the ecosystem consequences of global change impacts on soil biodi-

versity.

Keywords: bacteria, carbon cycling, fungi, global change, meta-analysis, microorganisms, nitrogen cycling, plant biomass

Received 14 May 2014; revised version received 8 October 2014 and accepted 20 October 2014

Introduction

Elevated atmospheric carbon dioxide (CO2), nitrogen

(N) deposition, and climate change (e.g., elevated tem-

peratures and altered precipitation regimes) are

among the major drivers of ongoing global change for

terrestrial ecosystems worldwide (IPCC 2007). Increas-

ing concern about the impacts of these drivers has

boosted research on ecosystem processes such as

plant productivity and global biogeochemical cycles

(MEA, 2005). Although much of the ecosystem

research has been devoted to understanding the role

of plants as agents of ecosystem functioning (EF)

responses to global change (Zavaleta et al., 2003; Reich

et al., 2004; Cardinale et al., 2012; Hooper et al., 2012),

in recent years there has been an increase in the num-

ber of studies focusing on soil communities (Allison

& Martiny, 2008; Bardgett & Wardle, 2010). Like

plants, soil biota are sensitive to global change, and

the alterations in their abundance or diversity from

climate change, N deposition, and elevated CO2 can

feedback to affect the ecosystem processes they gov-

ern (van der Heijden et al., 2008; Bardgett & Wardle,

2010). Thus, a remaining challenge is to assess

whether the effects of particular global change drivers

on belowground communities are associated with the

responses of multiple ecosystem processes.

Soil biota are structured in complex, highly diverse

communities and their responses to global changeCorrespondence: Pablo Garc�ıa-Palacios, tel. +33 467 613 236, fax

+33 467 613 336, e-mail: pablogpom@yahoo.es

1© 2014 John Wiley & Sons Ltd

Global Change Biology (2014), doi: 10.1111/gcb.12788

Global Change Biology

87

include abundance, compositional, and physiological

shifts (Eisenhauer et al., 2012; Wall et al., 2012; Frey

et al., 2013). However, abundance measurements are

most consistently used across taxa and studies (Tresed-

er, 2004, 2008; Blankinship et al., 2011) and are simpler

to standardize. Recent literature syntheses have shown

unique and predictable effects of different global

change drivers on soil biota abundance. For example, N

deposition has been found to decrease microbial bio-

mass (Treseder, 2008), whereas elevated CO2 has been

found to increase it (Blankinship et al., 2011). Further-

more, warming often has positive effects on nematode

abundance (Blankinship et al., 2011) but negative effects

on total numbers of enchytraeids (Briones et al., 2007).

Reduced precipitation has been found to decrease the

abundance of fungi, enchytraeids and collembola (Blan-

kinship et al., 2011). Whether these global change

effects on belowground organisms’ abundance help to

explain those on ecosystem functioning is still unclear,

although many studies have found concurrent effects

of global change on the abundance of particular soil

taxa and the rates of magnitude of specific ecosystem

processes (Allen et al., 2000; Allison & Treseder, 2008;

Lamb et al., 2011).

Previous literature screenings have synthesized how

the effect size of N deposition on the abundance of par-

ticular soil taxa (total microbial community) relates

with the effect size of N deposition on carbon (C) cycle

variables such as soil CO2 efflux (Treseder, 2008) or

mineral soil C (Liu & Greaver, 2010). However, a syn-

thesis of relationships between global change effect

sizes across several soil taxa and ecosystem processes

as we present here has been lacking. The evaluation of

such linkages, and the identification of knowledge

gaps, is fundamental to integrate soil organisms’

control of ecosystem responses to global change into

large-scale models (Wall et al., 2012), guide future

research efforts and mitigation plans (Jeffery et al.,

2010), and support emerging soil policy initiatives

(Koch et al., 2013).

Here we compiled data from individual global

change studies, and used meta-analytical tools and

structural equation modeling (SEM) to evaluate

whether the abundance responses of soil microbes to a

particular global change driver explain variation in the

EF responses to this driver. We argue that our SEM,

which is a regression-based modeling technique and

thus not causal in its essence, allows causal interpreta-

tion of these linkages based on prior ecological knowl-

edge. We focused on the most commonly studied

drivers (elevated CO2, N deposition, and warming), soil

taxa (total microbial community, fungi, and bacteria)

and ecosystem processes (plant biomass, soil C, and N

cycle). Different metrics were used to measure the

abundance of microbial taxa (e.g., PLFA biomass and

chloroform-fumigation extractions for total microbial

community; Table 1) and the ecosystem processes (e.g.,

NO3�-N, total N or N mineralization rate for N cycle;

Table 1). Specifically, this study aimed to: (i) quantify

the effect size of elevated CO2, N deposition, and

warming on total microbial community, bacterial and

fungal abundance, and on plant biomass, soil C cycle

and soil N cycle, (ii) assess whether the effect size of

elevated CO2, N deposition, and warming on soil

microbes (across taxa) is related with the effect size of

such drivers on EF (across processes), as well as evalu-

ate the relative contribution of study length, magnitude

of global change and climatic conditions, and (iii)

analyze the correlations between the effect size of each

global change driver on each microbial taxon and

ecosystem process separately. Finally, the scope of our

literature synthesis was drastically constrained by the

infrequency of studies that examined both soil biota

and EF responses to a global change driver, so we are

using this study as an opportunity to identify

knowledge gaps and discuss future research directions

that will advance our understanding of how soil biota

control ecosystem responses to global change.

Materials and methods

Data collection

We quantitatively synthesized studies that evaluated the

effects of global change on the abundance or biomass of at

least one soil taxon and one ecosystem process. Briefly, ‘global

change’ was regarded as elevated CO2, warming, N deposi-

tion, and/or altered precipitation; ‘soil biota’ was designated

to include fungi, bacteria, total microbial community, and

numerous other taxa of soil fauna (Fig. 1); and ‘EF’ included

several processes (N, C and P cycles, plant productivity and

litter decomposition). Any global change study addressing

only soil biota or only EF was omitted. Searches were con-

ducted using the ISI Web of Knowledge (http://apps.isi-

knowledge.com) on 26 April 2013, with no restriction on

publication year, and were supplemented with references

from previous reviews on the topic. See Appendix S1 for

details on the term combinations used in the literature search,

which yielded 8662 references.

To be included in our database, studies had to utilize global

change rates no more extreme than those predicted under

future scenarios, because they have the potential to be more

informative to land managers and for parameterization of eco-

system models (Marshall et al., 2008). Maximum warming and

CO2 levels were established based on estimates from the A1B

Scenario of the Fourth IPCC Assessment (IPCC 2007). Elevated

CO2 and warming studies were included if they provided less

than 850 ppm of CO2 and increased air temperature up to

2.8 °C. N deposition studies were included if they provided

less than 150 kg N ha�1 yr�1 (Dentener et al., 2006) from

© 2014 John Wiley & Sons Ltd, Global Change Biology, doi: 10.1111/gcb.12788

2 P. GARC�IA-PALACIOS et al.

88

inorganic N sources (e.g., NH4NO3) or urea, which are the

two most studied N forms in global change research and affect

microbial activity in the same way (Ramirez et al., 2010). As

predictions of changes in precipitation levels are not very

accurate due to considerable regional variation (IPCC 2007),

studies were only included if treatments represented a

(a) (b) (c)

Fig. 1 Percentage (%) of case studies addressing different global change drivers (a), soil taxa (b), and ecosystem processes (c). Data

from 75 articles representing 330 cases studies. Gray bars represent the case studies used in further analyses.

Table 1 Target variables selected and metrics used to evaluate the effects of global change on microbial abundance and EF

Target variable Metric n (Elevated CO2) n (N deposition) n (Warming)

Bacteria Bacterial biomass 1 2 1

Bacteria PLFA 6 3 1

Bacteria MPN 4 2 1

Bacteria DNA fingerprint 2 NA NA

Bacteria qPCR 1 NA 5

Bacteria Biolog NA NA 2

Fungi Fungal biomass NA 2 1

Fungi PLFA 5 3 1

Fungi MPN 3 1 1

Fungi Ergosterol 3 NA 3

Fungi Root colonization 10 6 1

Fungi qPCR NA NA 4

MC PLFA 1 2 NA

MC MB (soil weight) 17 27 21

MC MB (soil surface) 2 5 NA

Plant biomass Shoot biomass 4 8 NA

Plant biomass Root biomass 12 4 8

Plant biomass Total biomass 12 NA 2

C cycle Soil respiration 5 6 9

C cycle TOC 12 19 13

C cycle DOC 4 4 3

N cycle DIN 8 7 12

N cycle DON NA 1 NA

N cycle N flux 12 5 7

N cycle NH4+ -N NA NA 6

N cycle NO3‒-N 1 NA NA

N cycle Total N 2 7 NA

n represents the sample size in terms of case studies.

MC, total microbial community; MPA, most probable number counts; MB, microbial biomass; TOC, total organic carbon; DOC, dis-

solved organic carbon; DIN, dissolved inorganic nitrogen (NH4‒N + NO3‒N); DON, dissolved organic nitrogen; N flux (N mineral-

ization, nitrification or ammonification rates), NA (not available).

© 2014 John Wiley & Sons Ltd, Global Change Biology, doi: 10.1111/gcb.12788

SOIL MICROBES, FUNCTIONING, AND GLOBAL CHANGE 3

89

maximum of 50% increase or decrease in precipitation with

respect to the control, as has been done in previous climate

change experiments (Zavaleta et al., 2003). Other study selec-

tion criteria are described in Appendix S1. Only field studies

were selected for warming and N deposition experiments, but

controlled environments such as growth chambers and/or

greenhouses were also included for CO2 studies, because they

represent an important proportion of the conducted body of

research (Blankinship et al., 2011).

Data extraction

A total of 75 articles, representing 330 cases studies, met the

established criteria. However, not all topics were evenly repre-

sented, and therefore underrepresented global change drivers,

soil taxa, and ecosystem processes were subsequently omitted

from all downstream analyses. Thus, we focused our analysis

on the most studied global change drivers (elevated CO2,

warming, and N deposition) and target variables (total micro-

bial community, bacterial and fungal abundance, plant bio-

mass, soil C cycle and soil N cycle; Fig. 1), which rendered a

final dataset of 238 case studies from 70 articles (Appendix

S2).

For each study, the microbial and EF data were recorded

for both the control and global change plots. Mean, standard

deviation and sample size values were extracted directly from

tables or from graphs using Dexter, an online tool provided

by the German Astrophysical Virtual Observatory (http://dc.

zah.uni-heidelberg.de/sdexter/). Microbial abundance mea-

surements were always preferred, but biomass measurements

were also included as a surrogate for abundance (Coleman

et al., 2004), and hence both are referred to as ‘abundance’.

Authors used several metrics to assess microbial abundance

(Table 1). To represent the abundance of the total microbial

community, microbial biomass was measured using chloro-

form-fumigation extractions (mg C kg�1) or measured as

PLFA biomass (Treseder, 2008; Blankinship et al., 2011). Bacte-

rial abundance was determined with incubations using selec-

tive inhibitors, bacteria-specific PLFA biomass, most probable

number counts, DNA fingerprint (% of total clones from 16S

rRNA gene clone libraries), qPCR and Biolog. Fungal abun-

dance was determined with the same metrics, in addition to

ergosterol concentration and root colonization. All of the EF

metrics used (Table 1) were associated with one of the follow-

ing processes: plant biomass (as a measure of productivity;

Scurlock et al., 2002), soil C and soil N cycles. The metrics used

described different aspects of the same process (e.g., root,

shoot and total biomass as metrics for plant biomass, or

NO3�-N, mineralization rate and total N as metrics for N

cycle). All these processes are directly linked to the mainte-

nance of primary production, biomass accumulation and

nutrient cycling. According to this rationale, and taking into

account the spatial extent of our study, we assumed that the

higher the values for the different variables measured at a

given study, the higher the overall rate of functioning at that

site (Maestre et al., 2012). Methodological features of the

experimental design (magnitude of global change, field vs.

controlled environment for CO2 studies, study length, latitude

and longitude) and the metric used (analytical method and

units) were also recorded. We obtained the mean annual tem-

perature and precipitation of each field study site from the

WorldClim database (Hijmans et al., 2005), which provides

average climatic values for the period 1950–2000.

Meta-analytical procedure for data grouping

Data collected were inherently heterogeneous, as the studies

used different metrics to measure the soil taxa and ecosystem

processes (Table 1). Thus, a data grouping procedure was per-

formed to deal with such heterogeneity and to test its influ-

ence on the global change effect sizes. Two method features

were evaluated: (i) environmental conditions (controlled vs.

field environments, only for CO2 studies) and (ii) metric used

(see Table 1 for the different categories analyzed in each target

variable). We ran two separate weighted random-effects meta-

analyses (Gurevitch & Hedges, 1999) for each global change

driver, one for microbial abundance and one for EF. We calcu-

lated Hedge’s d as an estimate of global change effect size,

and assessed its heterogeneity between the categories of the

method feature studied. Positive values indicate that the

response variable (the microbial taxa abundances and the eco-

system processes) in the global change plot has a larger value

than in the control. To test the effects of global change drivers

on each target variable, we assessed whether the bias-cor-

rected 95% bootstrap confidence intervals (CI) from 999 itera-

tions overlapped zero (Rosenberg et al., 2000). We used

estimated nonparametric variances because most of our exper-

imental data did not follow a normal distribution (Adams

et al., 1997). See Appendix S1 for a full description of the

meta-analytical procedure. The results of the random-effects

models confirmed that the responses of each microbial taxon

and ecosystem process to elevated CO2, N deposition, and

warming did not depend on the experimental conditions or

metric used (Table S1; Prandom > 0.05 in all cases).

Relationship between the responses of microbes andecosystem functioning to global change

To assess the relationship between the effect size of global

change on microbial abundance and on rates of EF, we used

structural equation modeling (Grace, 2006). However, the low

number of case studies jointly examining both variables (Table

S2) drastically constrained our ability to build robust models

for each specific combination of microbial taxon and ecosys-

tem process. Thus, we analyzed such relationship among

effect sizes across microbial taxa (total microbial community,

fungal and bacterial abundance) and ecosystem processes

(plant biomass, soil C and N cycles), an approach that has

been previously followed when synthesizing the effects of

multiple global change drivers (Blankinship et al., 2011). We

tested whether the effect sizes (Hedge’s d) of global change on

overall microbial abundance modulated the effect sizes on

overall EF. The contribution of site-specific factors, such as cli-

matic conditions, study length and magnitude of global

change to the effect sizes was also evaluated. We also tested

separate models for each of elevated CO2, N deposition, and

© 2014 John Wiley & Sons Ltd, Global Change Biology, doi: 10.1111/gcb.12788

4 P. GARC�IA-PALACIOS et al.

90

warming. In addition, we calculated Pearson correlations

among the effect sizes of global change on each microbial

taxon and ecosystem process separately. The evaluation of

such relationships, although not directional, allowed us to

interpret whether the two effect sizes were linked at the micro-

bial taxon and ecosystem process level. Correlations were not

performed when microbial abundance and EF were measured

in fewer than eight case studies. Although we conducted a

large number of statistical tests, P values were not adjusted for

multiple testing as this approach is considered overly conser-

vative (Gotelli & Ellison, 2004).

On the basis of current ecological knowledge, we hypothe-

sized a hierarchy of relationships in a path diagram (Grace,

2006). Our conceptual a priori global SEM (Fig. 2), which was

tested separately for elevated CO2, N deposition, and warm-

ing, predicted the effect size of global change on microbial

abundance to explain variation in the effect size of global

change on EF. This path can be interpreted as the ability of

overall microbes (across taxa) to control overall EF (across pro-

cesses) responses to global change, and is the key aspect of

our study. Previous theoretical, modeling and experimental

studies have underlined the important role played by

microbes controlling ecosystem responses to global change

(Allison & Treseder, 2008; Allison et al., 2010a; Bardgett &

Wardle, 2010; Wagg et al., 2014). Thus, the model structure

proposed was supported by current ecological knowledge,

justifying a causal interpretation of the model outputs (Ship-

ley, 2002). Nevertheless, as all SEM, they are contingent on the

structure imposed by the modelers. Therefore, our model only

allows causal interpretation of the directional relations intro-

duced (e.g., effect size on microbes explaining variation in

effect size on EF), but does not deny the possibility of other

type of relations existing within the study system (e.g., effect

size on EF explaining effect size on microbes), which were

beyond the scope of our synthesis and therefore not tested.

We accounted for the fact that data for global change effect

sizes on overall microbes related to different microbial taxa

(bacteria, fungi, and microbes), and those effects on overall EF

related to different processes (plant biomass, soil C cycle, and

soil N cycle). Thus, 2 two-indicator latent variables, ‘Microbial

taxon’ and ‘Ecosystem process’, were introduced in the model.

Latent variables have multiple uses, but here they function to

sum together the effects of the levels of a categorical variable,

which are represented by dummy variables (Grace, 2006).

When modeling categorical dummy variables, it is necessary

to omit one indicator (Grace, 2006), and we omitted the soil

taxon or ecosystem process that allowed us to interpret the

latent variable in a more straightforward way. A significant

individual path coefficient from a category composing ‘Micro-

bial taxon’ (e.g., bacteria) or ‘Ecosystem process’ (e.g., soil C

cycle) means a larger effect size of global change for that

specific category with respect to the reference.

The a priori model also predicted a direct effect of climate,

study length, and magnitude of global change on the effect

size of global change (Hedge’s d) on both microbes and EF.

Study length and magnitude of global change were intro-

duced in the model as exogenous variables. Climate was mod-

eled as a composite variable, which allows an additive

combination of the effects of multiple conceptually related

variables (mean annual temperature and mean annual precipi-

tation) upon a response variable (the effect sizes of global

change). Composite variables are primarily a graphical and

Fig. 2 Generalized a priori conceptual structural equation model depicting the influence of study length, global change (GC) magni-

tude, climate, microbial taxon, and ecosystem process upon the effect sizes (Hedge’s d) of GC on overall (across taxa) microbial abun-

dance and overall (across processes) EF. The same model structure was hypothesized for elevated CO2 (without climate), N deposition,

and warming. Colors highlight site-specific factors (blue), microbial variables (green), and EF variables (red). Single-headed arrows rep-

resent a hypothesized directional influence of one variable upon another. Double-headed arrows represent a correlation in which no

direction is specified. Squares indicate measured variables entered in the model. Hexagons indicate theoretical non-measured variables.

‘Microbial taxon’ and ‘Ecosystem process’ account for the three different categories introduced in each of them. ‘Climate’ is an additive

combination of the effects of mean annual temperature (MAT) and mean annual precipitation (MAP). EF: ecosystem functioning. Total

effects of all variables on GC effect size on microbes and GC effect size on EF are summarized in Fig. 4.

© 2014 John Wiley & Sons Ltd, Global Change Biology, doi: 10.1111/gcb.12788

SOIL MICROBES, FUNCTIONING, AND GLOBAL CHANGE 5

91

numerical interpretation tool, and do not change the underly-

ing model (Grace, 2006). Climatic influence was not included

in the elevated CO2 model, because some studies were per-

formed in controlled environments.

We examined the distributions of the endogenous variables

and tested their normality. To increase the degrees of freedom,

any path with a coefficient <0.10 was removed from the model

when not significant. Overall goodness-of-fit of the models

was tested against the dataset and checked following Scherm-

elleh-Engel et al. (2003). We used the traditional v2 goodness-

of-fit test, but, because of its sensitivity to sample size, the

RMSEA index was also considered (Grace, 2006). SEM analy-

ses were performed with AMOS Software Version 22.0 (Amos

Development Co.).

Results

Summary of selected studies and responses to globalchange drivers

The most studied global change drivers were elevated

CO2, N deposition, and warming (95% of the case stud-

ies collected, Fig. 1). The most studied soil taxa were

fungi, bacteria, and the total microbial community (90%

of the case studies collected, Fig. 1). The most studied

ecosystem processes were plant biomass, soil C cycle

and soil N cycle (86% of the case studies collected,

Fig. 1). A positive effect size of elevated CO2 was

observed for total microbial abundance and plant bio-

mass (Fig. 3a). N deposition had a positive effect on

bacterial abundance and the soil N cycle (Fig. 3b),

meaning that N deposition plots showed higher values

of the variables describing the N cycle (e.g., NO3�-N, N

mineralization rate or total N) than the control plots.

Warming increased fungal abundance, and had posi-

tive, but nonsignificant, effects on plant biomass, soil C

cycle, and soil N cycle (Fig. 3c).

Relationships between the effect sizes of global change onmicrobes and ecosystem functioning

Goodness-of-fit tests for all SEM evaluated indicated

acceptable fits (Figures S1–S3), as the v2 tests were not

significant (P > 0.05 in all models) and the RMSEA fit

measure was <0.08 (P > 0.1 in all the models), indicat-

ing that the data fitted the a priori model hypothesized

for the three global change drivers (Fig. 2).

The effect size of elevated CO2 on overall (across

taxa) microbes was not related with the overall (across

processes) CO2 effect size on EF (Fig. 4a and Figure S1).

The two site-specific factors had contrasting relation-

ships with CO2 effect size on microbes, with a positive

influence of the magnitude of the CO2 treatment and a

negative one of study length. The two latent variables

introduced in the model, ‘Microbial taxon’ and ‘Ecosys-

tem process’, affected both effect sizes, which indicates

that the responses to elevated CO2 were different

between taxa and processes, as also demonstrated in

Fig. 3a. At the microbial taxon and ecosystem process

level, fungal abundance responses to elevated CO2

were positively correlated with those of plant biomass

but negatively correlated with those of the N cycle

(Table 2). The effect size of elevated CO2 on total micro-

bial community abundance and plant biomass varied

in the same direction.

The N deposition effect size on overall microbes was

significantly related (r = 0.31, P = 0.002) with the N

deposition effect size on overall EF (Fig. 4b and Figure

S2). Study length significantly affected the effect size on

EF but not on microbes. Climatic conditions were asso-

ciated with the N deposition effect size on microbes,

with a higher effect size in the warmer sites. As shown

in Fig. 3b, the two effect sizes were different between

taxa and processes. Separate correlations showed that

(a) (b) (c)

Fig. 3 Mean effect size (Hedges’d) of elevated CO2 (a), N deposition (b), and warming (c) on soil bacterial, fungal, and total microbial

community (MC) abundance, plant biomass, soil C cycle, and N cycle. The bars around the means are bias-corrected 95% bootstrap

confidence intervals. Positive mean effect size indicates that the global change plot has a larger value for the target variable than the

control plot. Sample sizes for each target variable are indicated in parentheses. The dotted line separates the soil taxa from the

ecosystem processes.

© 2014 John Wiley & Sons Ltd, Global Change Biology, doi: 10.1111/gcb.12788

6 P. GARC�IA-PALACIOS et al.

92

total microbial community abundance responses to N

deposition were highly related to those of C cycle

(Table 2).

The warming effect size on overall microbes

accounted for an important part of the variance

(r = 0.19, P = 0.041) in the warming effect size on over-

all EF (Fig. 4c and Figure S3). Although not signifi-

cantly, the magnitude of warming and the study length

were associated with both effect sizes. The higher the

contrast between the warmed and control plots, the

higher the warming effect size on microbes, but the

longer the warming study duration, the lower the

warming effect size on EF. The influence of climate on

warming effect size on EF was due to both a positive

effect of total precipitation and a negative effect of

mean temperature. As found in Fig. 3c, the warming

effect size on microbes was different between taxa.

However, the effect sizes of warming on total microbial

community, bacterial, and fungal abundances were not

significantly correlated with the effect sizes of this glo-

bal change driver on soil C and N cycles (Table 2).

Discussion

Soil microbial responses to global change are linked withresponses of ecosystem functioning

Here, we present an assessment of the relationships

between multiple global change drivers, and the

responses of soil microbes and ecosystem processes.

Table 2 Pearson correlation coefficients (r) between the effect size of global change (Hedge’s d value) on microbial abundance and

EF

Global change driver Microbial taxon Ecosystem process n r P

CO2 Fungi Plant biomass 15 0.540 0.038

CO2 Fungi N cycle 9 �0.855 0.003

CO2 MC Plant biomass 14 0.562 0.037

CO2 MC C cycle 14 �0.157 0.592

CO2 MC N cycle 21 �0.165 0.474

N deposition MC C cycle 17 0.824 <0.001

N deposition MC N cycle 12 0.078 0.809

Warming Bacteria C cycle 8 0.635 0.091

Warming Bacteria N cycle 8 0.204 0.629

Warming Fungi C cycle 11 0.476 0.139

Warming Fungi N cycle 8 0.229 0.586

Warming MC C cycle 17 0.419 0.094

Warming MC N cycle 19 0.185 0.449

P values below 0.05 are in bold. n represents the sample size in terms of case studies. MC: total microbial community.

(a) (b) (c)

Fig. 4 Standardized total effects (direct plus indirect effects) derived from the structural equation models for elevated CO2 (a), N depo-

sition (b), and warming (c). These effects describe the influence of the variables depicted in the x axis upon the effect sizes (Hedge’s d)

of each global change driver on overall (across taxa) microbial abundance (white bars) and overall (across processes) EF (black bars).

Note that the climatic variables were not evaluated in the elevated CO2 model. EF: ecosystem functioning, GC: global change, MAP:

mean annual precipitation, MAT: mean annual temperature. ***P < 0.001, **P < 0.01, *P < 0.05. See Fig. 2 for a description of the a priori

model, and Figures S1–S3 for the full graphical representation of the three structural equation models.

© 2014 John Wiley & Sons Ltd, Global Change Biology, doi: 10.1111/gcb.12788

SOIL MICROBES, FUNCTIONING, AND GLOBAL CHANGE 7

93

While such correlative approaches are not the ultimate

test to quantify the ecosystem consequences of global

change impacts on soil biodiversity, they can indicate

general trends and direct future research efforts. When

evaluating the effect size of elevated CO2 on overall

responses across microbial taxa and ecosystem pro-

cesses, the responses of microbes did not explain those

of EF. Nevertheless, at the microbial taxon and ecosys-

tem process level, the increases in total microbial com-

munity and fungal abundance found with elevated

CO2 were positively correlated with the increase in

plant biomass. Elevated CO2 effects on belowground

organisms are more likely to occur through altered

plant root production and exudation than via direct

effects of aboveground CO2 (Paterson et al., 1997; Zak

et al., 2000), due to the high CO2 concentrations in the

soil pore space (Drigo et al., 2008). However, such

increase in root biomass can affect soil C and N cycling

by altering soil microbial biomass and activity (de Gra-

aff et al., 2006). For example, arbuscular mycorrhizae,

which are linked with low rates of N cycling (van der

Heijden et al., 2008), are typically stimulated by such

root growth increases under elevated CO2 (Treseder,

2004). Since all metrics used to measure the effects of

elevated CO2 on the N cycle were related with rapid

increases in soil N availability (DIN or N flux rates;

Table 1), the negative correlation found between the

effect sizes of elevated CO2 on fungal abundance and

on soil N cycle supports such plant-mediated mecha-

nism. Thus, increased mycorrhizal biomass as a conse-

quence of root growth with elevated CO2 may decrease

soil N transformation rates compared with control

plots, which are likely more bacterial-dominated

(Fig. 3a). The fact that different microbial taxa may

have different effects on EF could explain the absence

of an important functional role for soil microbes, when

responses to elevated CO2 were evaluated across micro-

bial taxa and ecosystem processes.

Overall responses of microbes to N deposition were

explained by the responses of EF to this global change

driver. The functional role of microbial abundance was

larger than the one played by site-specific factors such

as the duration of the study and the magnitude of N

enrichment. Total microbial community and fungal, but

not bacterial, abundance have been found to decrease

as N load and duration of the N deposition treatment

increase (Treseder, 2008). To simulate realistic future N

deposition rates, we limited our studies to 150 kg

N ha�1 (Dentener et al., 2006). This restrictive rate

excluded unrealistically high N loads and long-term

agricultural studies that may have promoted the pat-

tern found by Treseder (2008). The effects of N deposi-

tion on total microbial community abundance were

highly correlated with the responses of the soil C cycle

to such N enrichment. This is an interesting result

because it indicates that N deposition effects on soil C

cycling, but not on N cycling, are associated with those

of soil microbial biomass. Soil C cycle was positively

affected by N deposition, although the confidence inter-

vals barely included zero (Fig. 3b), which indicates an

increase in the variables describing such ecosystem pro-

cess (total organic C in the 65% of the case studies;

Table 1). The enhancement of soil organic C with N

deposition has also been found in previous meta-analy-

ses (Nave et al., 2009; Liu & Greaver, 2010), and reduc-

tions of litter decomposition rates via changes in either

plant litter quality (Knorr et al., 2005), microbial com-

munities (Sinsabaugh et al., 2002) or the extent of litter

decay (Whittinghill et al., 2012) have been hypothesized

as potential mechanisms. Our correlative approach sup-

ports the microbial-driven hypothesis. Although plant

biomass increases have been identified as one of the

main N deposition contributions to the C cycle (LeBa-

uer & Treseder, 2008; Liu & Greaver, 2010), we did not

find enough case studies to facilitate addressing the

links between any microbial taxon and plant biomass.

Thus, we cannot elucidate whether the relationship

found between the abundance of the total microbial

community and C cycle responses to N deposition will

explain ecosystem C sequestration. However, our study

does highlight that more research investigating the lit-

ter decomposition-microbial mechanism is needed to

better understand the N deposition effects on C cycling.

The effect sizes of warming on microbes and EF eval-

uated across taxa and ecosystem processes, respec-

tively, were also positively and significantly associated,

suggesting that microbial abundance and EF respond

in parallel to elevated temperatures. We found an inter-

esting matching in the temporal and treatment rate

responses to warming, where short-term studies and

higher temperature treatments promoted larger posi-

tive warming effects on both microbes and EF, which

may have facilitated the previous link found. This pat-

tern may be a product of long-term microbial thermal

adaptation to elevated temperature (Bradford et al.,

2008), but should be interpreted with caution as the

relationships were not significant. Warming generally

has weak effects at a global scale on net ecosystem C

exchange, due to the offset of plant production with C

losses (Lu et al., 2013). However, the absence of signifi-

cant correlations between specific microbial taxa and

soil C cycle, and the low number of case studies mea-

suring plant biomass, prevented us from demonstrating

whether changes in microbial abundance with warming

contribute to the balance between ecosystem C efflux

and influx. In general, our elevated CO2, N deposition,

and warming results, which are based on a review of

field studies, provide empirical support to theoretical

© 2014 John Wiley & Sons Ltd, Global Change Biology, doi: 10.1111/gcb.12788

8 P. GARC�IA-PALACIOS et al.

94

and modeling efforts advocating for an explicit inclu-

sion of the microbial component of soils into ecosystem

models (Allison & Martiny, 2008; Allison et al., 2010a;

Treseder et al., 2012; Wieder et al., 2013).

Research gaps and guidelines for future ecosystem studies

The ambitious scope of this literature screening

allowed us to identify major standardization and

research gaps on the linkages between global change,

soil communities and EF. The study selection criteria

were based on predicted rates for global change driv-

ers available from the IPCC (2007; Appendix S1),

which resulted in the exclusion of many altered pre-

cipitation (e.g., Van Gestel et al., 1992; Williams &

Rice, 2007) and N deposition studies (e.g., Allison

et al., 2008; Zheng et al., 2008; Allison et al., 2010b)

due to the magnitude of treatments falling beyond

our specified cutoffs. While the importance of stan-

dardized research methods to facilitate cross-site

comparisons is increasingly recognized (e.g., Wall

et al., 2008; Powers et al., 2009; Sylvain et al., 2014),

there is little consensus on the magnitude of treat-

ments to simulate responses of ecosystems to global

change. If we hope to prediction future ecosystem

scenarios, experiments should be designed to account

for changes in global change rates predicted over the

next 50–100 years. Altered precipitation studies were

underrepresented in our assembled database. This

constitutes a major gap to understand how soil biota

modulates ecosystem responses to global change,

because altered precipitation has a larger influence

on soil biota abundance across taxa than elevated

CO2 or warming, as found by the most recent review

on the topic (Blankinship et al., 2011). Soil fauna were

also infrequently measured in the available literature,

despite their functional role (Bardgett & Chan, 1999;

Eisenhauer et al., 2011; Garcia-Palacios et al., 2013)

and sensitivity to warming and altered precipitation

at global scales (Blankinship et al., 2011), hindering a

full assessment of the functional implications of soil

biodiversity under global change. Regarding the eco-

system processes measured, the current underrepre-

sentation of studies assessing litter decomposition

complicates the understanding of how soil biota

mediates global change effects on nutrient dynamics

and C cycling.

Here, we evaluated the responses of each ecosys-

tem process (e.g., C cycle) to global change across

different metrics (e.g., soil respiration, dissolved

organic C or total organic C), as the low number of

studies found prevented us from conducting a spe-

cific analysis for each metric, and acknowledge that

our analysis cannot discriminate among particular

outcomes of each process (e.g., soil C losses vs. soil C

accumulation). An appropriate procedure to over-

come this issue, and scale up from particular ecosys-

tem processes to whole EF, would be the use of

multifunctionality indexes, which address the ability

to maintain multiple functions simultaneously (Zaval-

eta et al., 2010). However, current global change

research lacks the homogeneity needed to calculate

such indexes across studies. Finally, our literature

review only focused on soil microbial abundance

measurements. Similar meta-analytical approaches

will greatly benefit from the inclusion of community

compositional metrics because this will allow exami-

nation of changes in diversity/function relationships

in response to changing environmental pressures. A

good example of how to quantitatively synthesize

bacterial community diversity and compositional data

derived from the sequencing of the 16S rRNA gene

is the meta-analysis by Shade et al. (2013), techniques

from which could be implemented when more soil

biodiversity data become available in global change

studies. High-throughput sequencing is also opening

promising avenues for process-based ecosystem mod-

els by linking soil organisms’ phylogeny, physiologi-

cal traits, and responses to global change

disturbances (Fierer et al., 2013; Luo et al., 2013;

Evans & Wallenstein, 2014).

Strengths and limitations of the approach followed torelate global change effect sizes on microbes andecosystem functioning

Our synthesis effort goes beyond the assessment of

global change effects on soil microbes or EF sepa-

rately, which has already been done (Treseder, 2008;

Liu & Greaver, 2010; Blankinship et al., 2011; Lu

et al., 2013). Specifically, we studied whether the

responses of soil microbial abundance (fungi, bacte-

ria, and total microbial community) explained varia-

tion in those of plant biomass, soil C cycling and soil

N cycling. We acknowledge that our analysis is

based on statistical associations from the structural

equation modeling, and that it does not enable us to

estimate ultimate causality such as in controlled

experimental designs. However, structural equation

modeling allows the assessment of multivariate

hypotheses predicting multiple drivers of a treatment

effect size (e.g., climate, methodological features, the

effect size upon other variables), and thus its use in

ecological meta-analysis is growing (Grace et al.,

2007; Eldridge et al., 2011; Garcia-Palacios et al., 2013).

Our approach allowed us to synthesize current litera-

ture, find general patterns and identify key areas for

future global change research.

© 2014 John Wiley & Sons Ltd, Global Change Biology, doi: 10.1111/gcb.12788

SOIL MICROBES, FUNCTIONING, AND GLOBAL CHANGE 9

95

Conclusions

Our literature synthesis demonstrated the existence of

strong general links between the responses of soil

microbes and EF (plant biomass, soil C cycling and soil

N cycling) to global change (elevated CO2, warming,

and N deposition). This suggests that soil microbes are

important mediators of global change effects on EF. In

the case of N deposition and warming, these links were

strong and consistent across taxa and ecosystem pro-

cesses, whereas for elevated CO2 such links were more

taxon and ecosystem process specific. The links found

between global change, EF and soil microbes support

the explicit consideration of soil organisms in ecosys-

tem models. To do so, we need to understand the

mechanisms underlying, for example, the effects of

plant-soil interactions on N cycle responses to elevated

CO2, or how changes in soil C sequestration with N

deposition are modulated by microbially driven shifts

in litter decomposition. Important gaps (altered precipi-

tation, soil invertebrates, soil community composition,

and litter decomposition) prevented us to conduct a

broader assessment of the soil biodiversity-ecosystem

functioning relationship under global change, and

deserve attention in future studies.

Acknowledgements

We thank Barbara Fricks for her help extracting data from thepapers. We also thank four anonymous reviewers for improvingthe manuscript. PGP was supported by a Fulbright postdoctoralcontract from the Spanish Ministerio de Educaci�on and by aEuropean Commission’s FP7 Marie Curie IEF grant (DECOM-FORECO-2011-299214).

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Supporting Information

Additional Supporting Information may be found in theonline version of this article:

Appendix S1. Materials and Methods.Figure S1. Structural equation models depicting the influ-ence of study length, CO2 magnitude, microbial taxon andecosystem process upon the effect sizes (Hedge´s d) of ele-vated CO2 on overall (across taxa) microbial abundance andoverall (across processes) EF (ecosystem functioning). Colorshighlight site-specific factors (blue), microbial variables(green) and EF variables (red).Figure S2. Structural equation models depicting the influ-ence of study length, N deposition magnitude, climate,microbial taxon, and ecosystem process upon the effect sizes(Hedge’s d) of N deposition on overall (across taxa) micro-bial abundance and overall (across processes) EF (ecosystemfunctioning).Figure S3. Structural equation models depicting the influ-ence of study length, warming magnitude, climate, micro-bial taxon, and ecosystem process upon the effect sizes(Hedge´s d) of warming on overall (across taxa) microbialabundance and overall (across processes) EF (ecosystemfunctioning).Table S1. Results from the random-effect models tested toevaluate the effects of global change (elevated CO2, N depo-sition, and warming) on soil microbial abundance (bacteria,fungi, and total microbial community) and ecosystem func-tioning (plant biomass, N cycle and C cycle).Table S2. List of references, case studies and descriptorsused in the meta-analysis. Global change and control ratesin ppm (elevated CO2), Kg N ha�1 yr�1 (N deposition) and°C (warming). Study length in days.

© 2014 John Wiley & Sons Ltd, Global Change Biology, doi: 10.1111/gcb.12788

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General discussion The effects of global change on the soil micro food web found in this dissertation do seem to be strongest

when mediated by plants (i.e. quantity and quality of belowground input), as also discussed by Wardle et al.

(1998) - specifically the effects caused by elevated CO2 and burning. The most consistent global change

impact on the soil community we see in the CLIMAITE experiment is of CO2 (Paper I and II). In Paper I,

plant types representing different substrate quality furthermore had stronger effects on the soil community

composition than any global change factors. In Paper III changing the character of the organic input to

pyrogenic matter was prima facie equivalent to removing the litter all together. However, the recurring

renewal of grass growth by annual burning did seem to have created a more active and abundant decomposer

food web, although we also saw effects on the trophic structure of the soil micro community possibly

originating from direct deadly effects of burning, creating a more “bottom-heavy” food-web.

Our meta-analysis suggests that CO2 effects on microorganism impact on ecosystem function is – to a greater

degree than nitrogen deposition and warming - context dependent. This may indeed be because the CO2

effect is primarily mediated by plants, via a fertilizer-effect increasing belowground translocation of carbon

(Jones et al. 2009; Drigo et al. 2010) and effect on soil moisture due to reduced plant stomatal conductance

(Field et al. 1995). Nitrogen deposition and warming is also likely to affect plants (Schimel 1995; Vitousek

et al. 1997; Rustad et al. 2001; Wardle et al. 2004; Cleland et al. 2007), but probably also has a more direct

effect on soil microorganisms (Carriero et al. 2000; Frey et al. 2004; Rousk et al. 2012). When organisms in

the soil food web is under top-down control , effects manifest at higher trophic levels (as we see indications

of in Paper I and III), but otherwise, a detectable effect on microorganisms of continuous global change

factors such as warming and elevated CO2 is in principle to expect at any given sampling time. Periodic

exposure, however, is more likely to be detectable in longer-lived organisms at a given sampling time after

the exposure has ceased, while microorganisms although susceptible e.g. to drought have a fast turnover rate

and may level out a reduction of biomass or growth rate within days of returning to normal conditions (de

Vries et al. 2012). A possible example of this is seen in Paper II, where samples are taken one month after

conclusion of the annual summer drought treatment but under continuous CO2 enrichment. A drought effect

is detectable in the nematode abundance, but is not seen in the microbial biomass. The CO2 effect on the

other hand is seen in both groups of organisms, and so is the effect of defoliation treatment, which was

concluded a few days prior to sampling. Hence, the hierarchical resolution (Bardgett et al. 2013) with which

we study the plant-soil interactions matters.

Plant species effects in Paper I are stronger than effects of elevated CO2 and climate change. Thus,

we might deduce that global change is going to have the greatest impact on natural systems to the extent it

alters plant community compositions, which is also what Eisenhauer et al. (2013) concludes. However, in

this long-term global change experiment, where CO2 concentration, temperature and drought have been

98

manipulated through 8 years, no effects on plant cover have been detected (Kongstad 2012; Ransijn 2014).

Global change effects are seen on photosynthetic capacity (Albert et al. 2011), root growth (Arndal et al.

2014), microbial and microfaunal biomass (Paper II), microbial processes (Bergmark 2013),

compartmentalization of energy flow (Paper I) and mycorrhizal colonization (Merrild et al. 2013). Thus, the

plant-soil system does respond, but not in a way that changes the competitive balance between the

dominating plant species aboveground. For such a shift to develop a major disturbance such as fire would

probably be required - or observation of the system for an entire regeneration cycle of the plants (Ransijn

2014). Hence, the hierarchy of responses pertains to both aboveground and belowground organisms, due to

the different timescales over which changes occur (Bardgett 2013).

The CLIMAITE experiment (Paper I and II) is designed as a multifactorial experiment, to better

understand the consequences of impending global change by studying varying agents in concert, and abate

the paucity of multi-factor long-term studies, particularly on belowground processes (Eisenhauer et al. 2012).

The possibility of observing interactive effects has thus been at hand. We do find a few interactions with

drought (Paper II) and warming (Paper I), but CO2 is the consistent effect. This is further established by an

analysis of all nematode abundances sampled during the first 6 years of the CLIMAITE experiment showing

a general increase of nematode abundances at elevated CO2, although primarily at the lower range of soil

moisture (Table 1, Fig.1 and Christensen et al. in prep). The reason may in the experimental design:

prolonged summer drought and mean temperature fluctuations of 1-2 ° C, although not common, are still

within the range of possible inter-annual variation, to which the temperate heathland is adapted. An

immediate 30% increase of CO2 availability, however, is more severe and novel to the system, and therefore

likely to have greater impact.

Figure 1 Nematode abundance at ambient (grey legend) and elevated (black legend) CO2, as affected by soil moisture (weight %) during last month before sampling (n=236). From Christensen et al. in prep.

Table 1Nematode abundances at ambient and elevated CO2 during the first six years of the CLIMAITE experiment. Nematode density g-1 soil, means with S.E. in parenthesis. From Christensen et al. in prep.

99

In contrast to nematodes, oribatid mites show a weak but consistent tendency to be more diverse under

warming and/or drought in this temperate heathland (fig. 2 and Dam et al. in prep). The difference in impact

of CO2 vs. warming and summer drought seen between microfauna and mesofauna may be related to another

subdivision of studying the plant-soil system: rhizosphere or litter. Elevated CO2 has been shown to increase

root exudation (Hungate et al. 1997; Eisenhauer et al. 2012), microbial biomass and composition (Zak et al.

2000; Carney et al. 2007), and have positive effects on microbes and nematodes but not on mesofauna such

as mites (Blankinship et al. 2011).We have studied the rhizosphere of the plants in the system, and the

associated microorganisms and nematodes respond to CO2 effects on root deposition and root biomass (Paper

II , Arndal et al. 2013). Thus, the apparent fertilizer-effect of CO2 is translated to the rhizosphere organisms.

The CLIMAITE experiment has not seen equivalent CO2 effects on aboveground production (Kongstad

2012; Ransijn 2014). Microarthropods such as oribatid mites are associated with litter and soil carbon more

than with the rhizosphere and effects on litter decomposition in the CLIMAITE experiment are primarily

influenced by drought, which interestingly increases the turnover (Haugwitz et al. 2013).

Figure 2 Simpsons index of Diversity for oribatid mites in the dry heathland of the CLIMAITE experiment. Means with SE bars (n=

12).From Dam et al. in prep.

Conclusions and perspectives Some general conclusions drawn from the present dissertation are:

- Observed global change effects on terrestrial ecosystems act via the plant carbon input to the

soil. Carbon input from living plants are seen to be modified by CO2, which affects both the

abundance and composition of the soil biota (Paper I and II). The effects on the soil system may vary

because the effects are indirect via another living organism, whereas effects of factors such as N-

Sim

pson

s ind

ex o

f Div

ersi

ty (1

-D)

100

deposition that more directly impacts soil organisms are more universal (Paper IV). We also see

indications of frequent burning altering the subsequent living plant input to the soil, also affecting

soil food web structure (Paper III). The soil organisms and processes related to the dead plant input

are affected by fire-derived transformation of litter to pyrogenic organic matter and more directly by

global change factors drought and temperature, as described for oribatid mites in the above review.

- For the plant-soil interactions in the systems I have studied, abundance and community effects are

most clearly seen on organisms with medium-long generation times. After 5-10 years of altered

management and climate change treatment, there are few changes in the aboveground plant

community of the prairie and heathland sites. The organism with shorter generation times responds,

however, affected by changes in plant physiological responses (Paper I, II, II). When the

experimental time frame is years or months, the longer-lived organisms of the soil micro food web

are more suitable, as they reflect more general patterns than the fast microbes, responding to daily

changes in conditions. Furthermore, to the extent that the food web is regulated by top-down control,

the response is transferred to the higher trophic levels, which we also see (Paper I and III)

- Nature of vegetation is of overriding importance for soil food web composition (Paper II), so

when forecasting effects of global change, effects on the established ecosystems should be related to

global changes on plant community compositions on longer time scales.

Organisms involved directly in plant-soil interaction in the rhizosphere (microorganisms and microfauna)

and organisms more associated with litter and soil organic matter (meso- and macrofauna), are surveyed in

newly released comprehensive study on the community changes of soil fauna related to land use

intensification and derived soil organic matter loss (Tsiafouli et al. 2014). When studying the same general

type of land cover (grass) under increasing intensity of farming, they find that community composition and

diversity of microarthropods and earthworms responds significantly to the different land uses and thereby

different soil organic matter contents. In the Tsiafouli et al. (2014) study, nematodes – more dependent on

the living plant inputs than soil organic matter per se - are not responsive. In contrast, I find that nematode

community analysis is a great tool for studying impacts on terrestrial ecosystems, in particularly effects on

the dynamics of plant-soil interactions, whether determined by different plant types or by various

disturbances of the primary producers. Thus, it is imperative that researchers are aware of linking the

appropriate organisms to the questions they ask, when studying soil systems.

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