Introduction
Reductions in biodiversity have been linked with anthropogenic global change
drivers such as climatic change, land cover change, reduction and
fragmentation of natural areas, and human dependence on synthetic
fertilizers (Vitousek and Mooney, 1997; Sanderson et al., 2002; Stevens,
2004; Phoenix et al., 2006; Fischer and Lindenmayer, 2007; Clark and Tilman,
2008). Over the past few hundred years, human activities have driven the
species extinction rate to ca. 1000 times the typical background extinction
rate (MEA, 2006). This global decline in biodiversity impacts important
ecosystem functions, such as net primary production (NPP) and biogeochemical
cycles of carbon (C) and nutrients, threatening the services that ecosystems
provide to the human population (Wardle et al., 2011; Cardinale et al.,
2012).
The C cycle is a particularly important ecosystem service because the
dynamic balance between C stored in ecosystems and in the atmosphere plays a
key regulatory role in the global climate. Although vegetation stores a
significant amount of C, soils are the largest terrestrial C reservoir,
containing more C globally than the biotic and atmospheric pools combined
(Lal, 2004). As such, soil C dynamics, and the processes that influence
them, have the potential to impact atmospheric carbon dioxide (CO2)
concentrations and subsequent global change. Perturbations in terrestrial
ecosystems that influence soil C dynamics could help mitigate the current
rise in atmospheric CO2 and associated climate change by promoting soil
C storage (e.g., Cramer et al., 2001; Johnson and Curtis, 2001).
Alternatively, they could exacerbate climate change by causing increased
soil CO2 efflux rates through increased decomposition rates of soil
organic C (SOC) (e.g., Mack et al., 2004; Bardgett et al., 2008).
Loss of biodiversity has the potential to influence climate change if it
alters SOC pools by reducing ecosystem C uptake or by increasing CO2
outputs from terrestrial ecosystems to the atmosphere (Jastrow et al.,
2007). Greater plant species diversity can increase C uptake by promoting
biomass production (Tilman et al., 2006; Cardinale et al., 2012), which can
enhance SOC storage (Fornara and Tilman, 2008; Steinbeiss et al., 2008;
Cong et al., 2014). Conversely, declines in plant species diversity can
reduce SOC storage (Hooper et al., 2012). Despite the growing body of
evidence suggesting strong links between plant species diversity and soil C
cycling, there is a dearth of information on whether similar relationships
exist between biodiversity of soil organisms (hereafter “soil biodiversity”)
and C cycling (Nielsen et al., 2011). With ongoing losses in diversity
belowground (Hooper et al., 2000), understanding relationships between soil
biodiversity and C cycling is critical for projecting how loss of diversity
under continued human alteration of the environment will impact global C
cycling processes.
Soil communities typically have high levels of both taxonomic and functional
diversity (e.g., De Deyn and Van der Putten, 2005). High taxonomic
diversity, small sizes of organisms, and large population sizes make
characterization of soil communities much less straightforward than that of
plant communities. As such, characterization of soil organisms is often
based on body size (e.g., Bradford et al., 2002), grouping organisms into
macrofauna (> 2 mm) such as earthworms, mesofauna (100 µm-2 mm) such as mites and springtails, microfauna (< 100 µm) such
as nematodes and protozoa, and soil microorganisms including bacteria and
saprophytic and mycorrhizal fungi. These body size classes typically are
useful functional groupings as they correlate with metrics such as metabolic
rate, generation time, and food size (Peters, 1983). Estimates suggest that
1 g of soil can harbor tens of thousands of bacterial taxa; up to 200 m of
fungal hyphae; and a wide range of micro-, meso-, and macrofauna (Roesch et
al., 2007; Bardgett, 2005). This complex soil community plays an important
role in determining the magnitude and direction of C fluxes between the
atmosphere and terrestrial ecosystems, controlling soil C mineralization and
promoting plant growth by regulating soil nutrient availability (e.g., De
Deyn and Van der Putten, 2005; Fitter et al., 2005; Wall et al., 2010; de
Vries et al., 2013). Despite a general consensus that the soil community is
integral to the global C cycle, the impact of soil community diversity on
ecosystem function is still little understood (Nielsen et al., 2011;
Briones, 2014).
The positive impact of plant species diversity on soil C cycling processes
can be mirrored in the soil community, with reported positive relationships
between soil biodiversity and C cycling processes (e.g., Setälä,
2002; Heemsbergen et al., 2004). However, evidence suggests that this
positive relationship is not universal, as other studies have found no
significant impacts of soil biodiversity on C cycling (e.g., Griffiths et
al., 2000). Understanding the relationship between soil biodiversity and C
cycling is thus not so much a question of “does diversity matter?” but
rather “under which circumstances does soil diversity affect C cycling?”. One
possibility for addressing this question is to consider the role of
functional similarity among taxa, because relatively small or no responses
of ecosystem processes to loss or gain of soil biodiversity would be
expected in the case of functional redundancy among soil organisms (Bengtsson,
1998; Andrén and Balandreau, 1999; Setälä et al., 2005).
To date, studies have assessed soil community diversity impacts on soil C
cycling by manipulating diversity within or across multiple organismal
groups (specifics of grouping criteria differ among studies, but are often
taxonomic, functional, or body size groups). For example, studies have
manipulated the diversity within groups of microorganisms (e.g., bacteria
(Bonkowski and Roy, 2005; Griffiths et al., 2000) and mycorrhizal fungi (van
der Heijden et al., 1998; Maherali and Klironomos, 2007)), soil mesofauna
(e.g., microarthropods; Liiri et al., 2002), and macrofauna (Heemsbergen et
al., 2004; Zimmer et al., 2005). Other studies have manipulated the
diversity across groups of soil organisms that differ in body size (i.e.,
microbes, and micro-, meso-, or macrofauna) or, alternatively, have
manipulated diversity of trophic or functional groups (Hedlund and Ohrn,
2000; Ladygina et al., 2010). Although different taxa within soil microbial
(Cox et al., 2001; Hanson et al., 2008; Orwin et al., 2006) or faunal
(Bignell and Eggleton, 2000; Milcu et al., 2008; Heemsbergen et al., 2004)
groups can have unique impacts on the C cycle, functional redundancy among
taxa would be expected to be reduced when a more complex food web of
organisms is manipulated (e.g., across different size classes or feeding
guilds) (Setälä, 2002). Thus, studies assessing biodiversity impacts
on ecosystem processes across multiple groups of soil organisms may yield
very different answers than studies that probe for biodiversity impacts
within single groups.
Although our knowledge of relationships between soil biodiversity and soil C
cycling processes has increased with expanding research emphasis in this
area, a comprehensive understanding to date is hampered by a lack of
quantitative synthesis of existing studies. Nielsen et al. (2011) performed
the most extensive synthesis on this topic to date, with a qualitative
analysis. They found that diversity is often positively correlated with
ecosystem function (e.g., soil respiration), although they cautioned that
negative relationships between soil biodiversity and C cycling may be
related to experimental limitations. In particular, Nielsen et al. (2011)
found that strong relationships between soil biodiversity and C cycling were
most common under unrealistically low levels of diversity. Further, their
synthesis showed that the soil community composition, rather than species
richness per se, had significant impacts on C cycling processes. This indicates
high levels of functional redundancy among soil organisms and suggests that
a loss of soil biodiversity may not necessarily impact the C cycle.
We aimed to quantitatively assess how soil biodiversity affects soil C
cycling pools and processes using meta-analysis. We tested the general
hypothesis that soil biodiversity positively impacts the soil C cycle, where
reductions in diversity decrease soil C pools and process rates. Further, we
tested the hypothesis that biodiversity manipulations across groups of
organisms with different body sizes more strongly affect C cycling processes
than manipulations within groups, due to a higher degree of functional
redundancy within than across groups (Andrén and Balandreau, 1999;
Setälä, 2002). In addition, we tested whether diversity of soil
microbes versus soil fauna (including micro-, meso-, and macrofauna) impacts C
cycling differently. Finally, since “biodiversity” is a metric that differs
greatly in absolute numbers for soil organisms that differ in body size, we
evaluated how the relative loss of diversity (in percent) within body size
groups (i.e., microbes, soil fauna) affects soil C cycling.
Methods
Data compilation
We compiled published studies that explicitly manipulated soil biodiversity
and measured responses of soil C cycling pools and/or processes. We
systematically searched ISI Web of Science using all possible combinations
of one soil C search term (plant biomass, soil C, decomposition,
respiration, or NPP), one soil organism search term (microbes, bacteria,
fungi, microorganism*, soil fauna, soil biota, soil organism*, microfauna,
mesofauna, macrofauna, nematode*, collembola, acari, termite*, earthworm*),
and the term “*diversity”. We used “*” as a wildcard character such that
papers using either singular or plural terms were returned. Additional
relevant studies referenced in those returned by the search engines were
also included in the literature compilation. While biodiversity sensu stricto includes
both species richness and abundance (Magurran, 2005), we follow the recent
nomenclature used in plant and soil studies and assume that the number of
species present in a community represents the diversity of the community.
Each study included in our analysis presented data on one or more commonly
measured biogeochemical C pool and/or process. Biogeochemical pool
measurements were plant biomass and soil C pools (either total soil C,
dissolved organic C (DOC), or microbial biomass). Measured biogeochemical
processes were soil C respiration and plant tissue decomposition. The
duration of manipulative experiments included in this analysis ranged from
14 days to 3 years. More studies were conducted under controlled laboratory
and greenhouse conditions rather than under field conditions (37 and 8 studies,
respectively). When extracting data from these studies, we took values
directly from published tables or the text whenever possible. If necessary,
we estimated values from graphical data with image analysis software
(ImageJ, National Institutes of Health, Bethesda, MD, USA).
Overview of studies used in the analyses.
Response
Body size
Size
Taxonomic
Number of
Reference
variable
group
class
group
studies
Total plant biomass
Microbes
< 5 µm
Ectomycorrhizal fungi
1
Baxter and Dighton (2001)
Mesofauna
100 µm–2 mm
Microarthropods
2
Liiri et al. (2002), Eisenhauer and Schädler (2011)
Microbes + microfauna
< 5–100 µm
Microflora, nematodes
1
Bezemer et al. (2005)
Microbes + macrofauna
< 5 µm, > 2 mm
Fungi, earthworms
2
Eisenhauer et al. (2010), Eisenhauer and Schädler (2011)
Mesofauna + macrofauna
100 µm–> 2 mm
Collembola, enchytraeids, earthworms
1
Eisenhauer and Schädler (2011)
Microbes + microfauna + mesofauna
< 5 µm–2 mm
Bacteria, fungi, protozoa, nematodes, microarthropods, enchytraeids
1
Sulkava et al. (2001)
Shoot biomass
Microbes
< 5 µm
Ectomycorrhizal fungi
1
Baxter and Dighton (2001)
Mesofauna
100 µm–2 mm
Microarthropods
3
Liiri et al. (2002), Cole et al. (2004), Eisenhauer and Schädler (2011)
Microbes + macrofauna
< 5 µm, > 2 mm
Fungi, earthworms
2
Eisenhauer et al. (2010), Eisenhauer and Schädler (2011)
Mesofauna + macrofauna
100 µm–> 2 mm
Collembola, enchytraeids, earthworms
1
Eisenhauer and Schädler (2011)
Microbes + microfauna + mesofauna
< 5 µm–2 mm
Bacteria, fungi, protozoa, nematodes, microarthropods, enchytraeids
1
Sulkava et al. (2001)
Microbes + microfauna + mesofauna
< 5 µm–2 mm
Bacteria, fungi, protozoa, microarthropods, enchytraeids
1
Laakso and Setälä (1999)
Microbes + microfauna + mesofauna + macrofauna
5 µm–> 2 mm
Fungi, nematodes, enchytraeids, microarthropods, wireworms
1
Ladygina et al. (2010)
Root biomass
Microbes
< 5 µm
Ectomycorrhizal fungi
1
Baxter and Dighton (2001)
Mesofauna
100 µm–2 mm
Microarthropods
3
Liiri et al. (2002), Eisenhauer et al. (2011), Eisenhauer and Schädler (2011)
Microbes + macrofauna
< 5 µm, > 2 mm
Fungi, earthworms
2
Eisenhauer et al. (2010), Eisenhauer and Schädler (2011)
Mesofauna + macrofauna
100 µm–> 2 mm
Collembola, enchytraeids, earthworms
1
Eisenhauer and Schädler (2011)
Microbes + microfauna + mesofauna
< 5 µm–2 mm
Bacteria, fungi, protozoa, nematodes, microarthropods, enchytraeids
1
Sulkava et al. (2001)
Microbes + microfauna + mesofauna + macrofauna
5 µm–> 2 mm
Fungi, nematodes, enchytraeids, microarthropods, wireworms
1
Ladygina et al. (2010)
Respiration
Microbes
< 5 µm
Bacteria
1
Wertz et al. (2006)
Microbes
< 5 µm
Bacteria, fungi
3
Griffiths et al. (2000, 2001, 2004)
Microbes
< 5 µm
Fungi
5
Wilkinson et al. (2010, 2011, 2012), Tiunov and Scheu (2005), Setälä and McLean (2004)
Macrofauna
> 2 mm
Earthworms
1
Scheu et al. (2002)
Microbes + mesofauna
< 5 µm, 100 µm–2 mm
Microflora, enchytraeids, microarthropods
1
Edsberg (2000)
Mesofauna + macrofauna
100 µm–> 2 mm
Not specified
1
Risch et al. (2013)
Microfauna + mesofauna + macrofauna
5 µm–> 2 mm
Nematodes, enchytraeids, earthworms
1
Bradford et al. (2007)
Decomposition
Microbes
< 5 µm
Bacteria
1
Bell et al. (2005)
Microbes
< 5 µm
Fungi
3
Progar et al. (2000), Toljander et al. (2006), LeBauer et al. (2010)
Microbes
< 5 µm
Bacteria, fungi
3
Griffiths et al. (2000, 2001), Liebich et al. (2007)
Mesofauna
100 µm–2 mm
Collembola
2
Cragg and Bardgett (2001), Eisenhauer and Schädler (2011)
Macrofauna
> 2 mm
Woodlice, millipedes
1
Collison et al. (2013)
Macrofauna
> 2 mm
Woodlice, earthworms
1
Zimmer et al. (2005)
Microbes + microfauna
< 5 µm–100 µm
Bacteria, fungi, nematodes
2
Mikola and Setälä (1998a, b)
Microbes + mesofauna
< 5 µm, 100 µm–2 mm
Microflora, enchytraeids, microarthropods
1
Edsberg (2000)
Microbes + mesofauna
< 5 µm, 100 µm–2 mm
Fungi, collembola, mites
1
Hedlund and Ohrn (2000)
Microbes + macrofauna
< 5 µm, > 2 mm
Fungi, ants, termites
1
Warren and Bradford (2012)
Microbes + microfauna + mesofauna
< 5 µm–2 mm
Bacteria, fungi, protozoa, nematodes, microarthropods, enchytraeids
1
Sulkava et al. (2001)
Microbes + microfauna + mesofauna
< 5 µm–2 mm
Bacteria, fungi, nematodes, protozoa, collembola, enchytraeids, mites
1
Cortet et al. (2003)
Microbes + microfauna + mesofauna
< 5 µm–2 mm
Not specified
2
Heneghan et al. (1999), Wall et al. (2008)
Microfauna + mesofauna + macrofauna
5 µm–> 2 mm
Protozoa, nematodes, enchytraeids, arthropods, earthworms
1
Bradford et al. (2002)
Microbes + microfauna + mesofauna + macrofauna
< 5 µm–> 2 mm
Fungi, arthropods
1
Araujo et al. (2012)
Microbes + microfauna + mesofauna + macrofauna
5 µm–> 2 mm
Fungi, bacteria, protozoa, nematodes, microarthropods
1
Carrillo et al. (2011)
Microbes + microfauna + mesofauna + macrofauna
< 5 µm–> 2 mm
Not specified
1
Slade and Riutta (2012)
Microbes + microfauna + mesofauna + macrofauna
< 5 µm–> 2 mm
Not specified
1
Yang and Chen (2009)
Soil C
Microbes
< 5 µm
Bacteria, fungi
1
Liebich et al. (2007)
Mesofauna
100 µm–2 mm
Collembola
1
Cragg and Bardgett (2001)
Macrofauna
> 2 mm
Woodlice, earthworms
1
Zimmer et al. (2005)
In total we analyzed 45 published studies, of which 8 examined the effects
of soil biodiversity on total plant biomass, 10 examined effects on
aboveground plant biomass, 9 examined effects on root biomass, 13 examined
effects on C respiration, 25 examined effects on decomposition, and 3 used
laboratory microcosms to examine effects on soil C pools (Table 1). For soil
C respiration, we included data that were obtained from either laboratory or
in situ incubation studies in which the substrate was soil only or soil with organic
amendments other than plant tissue (e.g., glucose). Laboratory studies
typically estimated potential C mineralization rates, using temperature and
moisture conditions assumed to be optimal for microbial activity. These
measurements were made in closed microcosms with flux rates estimated from
two or more repeated measurements of headspace gas concentrations. In situ studies
used static or flow-through chambers to measure CO2 flux rates from the
soil surface, and thus would include both microbial heterotrophic and root
(autotrophic) respiration (Holland et al., 1999). Plant tissue decomposition
data were obtained from studies that measured either litter mass loss
through time or C respiration from plant tissues decomposed under controlled
laboratory conditions. Litter mass loss analyses used mesh litterbags and
measured mass at two or more points in time (Harmon et al., 1999). For
studies in which the source of decomposed material (i.e., soil or plant C)
could be partitioned, data were separated and included in soil C respiration
or plant tissue decomposition data groupings. For all biogeochemical pool
and process studies in which data were available from multiple measurement
times, we calculated the mean value for all measurement times and used only
that value in the meta-analysis.
Soil biodiversity impacts on C respiration and decomposition were assessed
by manipulating biodiversity either within a single body size group (i.e.,
microbes (including bacteria and fungi), micro-, or meso-, or macrofauna) or
across multiple body size groups (e.g., micro-, meso-, macrofauna; e.g.,
Bradford et al., 2002). We treated the within-body size and across-body size
groupings as two separate categories for the analysis. For plant biomass,
however, there were not enough studies to run meta-analyses for individual
categories. We also categorized the studies by soil microorganisms or soil
fauna (micro-, meso-, and macro fauna grouped together due to inadequate
numbers of studies to split these up). Categorizing studies in this manner
allowed us to assess whether species diversity within or across body size
groups affected C cycling differently, while also enabling us to compare the
relative impacts of diversity within the soil microbial community versus soil
biodiversity within the soil faunal community.
Many of the studies we compiled assessed soil biodiversity impacts on C
cycling by quantifying responses to a diversity gradient (e.g., > 2
diversity levels). Inherent to this design is the possibility for multiple
comparisons among diversity treatments. For example, an analysis of how
diversity of a three species community (species a, b, and c) affects
ecosystem processes could yield a comparison of each single-species
community (a, b, or c) with the three-species community (a+b+c). This
comparison yields three observations: (1) “a” versus “a+b+c”, (2) “b” versus
“a+b+c”, and (3) “c” versus “a+b+c”. In our meta-analysis we did not
consider these three comparisons to be independent, but we calculated the
mean of the three single-species treatments and then calculated one response
variable based on the comparison between that one single-species mean and
the three-species community. Intermediate levels of diversity were excluded
from the meta-analysis to avoid any individual study from unduly weighting
the analysis. This method prevented studies with a large species diversity
gradient (i.e., a multitude of species included in the analysis) from
dominating our meta-analysis.
Statistical analyses
To test how soil microbial and/or soil faunal diversity affects ecosystem C
pools (plant biomass) and processes (C respiration and decomposition), and
to test whether biodiversity manipulations across multiple body size groups
affected C cycling differently from manipulations within groups, we analyzed
the data set with meta-analysis (Curtis and Wang, 1998; Hungate et al.,
2009), using the statistical software MetaWin 2.0 (Rosenberg et al., 2000).
We were unable to use meta-analysis for soil C pools because the number of
studies available (3) was inadequate for a meaningful analysis. The response
ratio (R) was calculated as the value of a particular response variable at
low diversity divided by the value at high diversity. The natural log of the
response ratio R (lnR) was used as a metric for all of the response
variables (de Graaff et al., 2006; van Groenigen et al., 2006). To ease
interpretation of figures, the results for the analyses on lnR were
back-transformed to response ratios and reported as percentage change under
a reduction in diversity (that is, 100×[R-1]). Thus, for response
variables where there was no change between higher and lower diversity
communities, the change would equal 0. For cases with greater values for
response variables in low-diversity communities than high-diversity
communities, the percent change would be would be positive, and lower values
for response variables in low-diversity communities than high-diversity
communities would yield negative values for the percent change.
Conventional meta-analyses weight each individual observation by the
reciprocal of the mixed model variance (Curtis and Wang, 1998). However,
such an analysis requires that the standard deviations of individual studies
be known. These data were not available for a large proportion of the
studies used in our analysis. Thus, we weighted individual values included
in the analysis by experimental replication (Hedges and Olkin, 1985; Adams
et al., 1997), assuming that better-replicated experiments resulted in data
with lower variance. We choose this metric because well-replicated studies
provide more reliable estimates of the response of individual variables
(e.g., Hungate et al., 1996, 2009). We used bootstrapping to
calculate confidence intervals on mean effect size estimates for the whole
data set and for individual categories (Adams et al., 1997). We considered
diversity effects significant if the 95 % confidence intervals did not
overlap with zero. In addition, we considered diversity effects for
individual categories different from each other if they varied significantly
at the p≤0.05 level.
Further, we tested how a loss of belowground species diversity is linked to
changes in C pools and processes by performing linear regressions with
percent change in species diversity and the effect size (lnR) of each of the
response variables. Percent change in diversity was calculated as (low
diversity – high diversity/high diversity) ×100. Since the absolute number
of species typically manipulated for diversity gradient studies varies among
species that differ in body size in absolute terms (i.e., many more species
are usually present in studies of microbial diversity than in studies of
faunal diversity), we calculated relative differences in species diversity
for each treatment. Thus manipulation of microbial diversity might include a
low-diversity treatment of 100 versus a high-diversity treatment of 1000
species, while manipulation of soil fauna might span from low diversity of 1
species to high diversity of 10 species. Calculated as relative differences
in diversity, both examples would be the same (i.e., low diversity is 10 %
of the number of species present in high diversity). We performed two sets
of regressions. The first included all soil biodiversity levels, and the
second included the highest and lowest biodiversity levels only. We used
linear regression (SPSS v. 20) to regress lnR against relative change in
species diversity. We performed regressions in which we considered lnR (the
effect size) between every diversity level, and also regressions in which we
only considered lnR between the highest and lowest diversity levels,
omitting intermediate diversity levels.
The percent response of total plant biomass, shoot biomass,
and root biomass to a change in soil community diversity (i.e., “high” vs.
“low” diversity). Studies included in the analysis manipulated diversity of
the soil microbial community, the soil micro-, meso, or macrofaunal
community or a combination of these trophic groups. Data represent means
with 95 % confidence intervals; numbers in brackets represent the total
number of data points included in the analysis.
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Discussion
Changes in biodiversity have been linked with changes in ecosystem
functioning, but so far studies have largely focused on plant diversity
(e.g., Isbell et al., 2011). Whether or not similar patterns exist for soil
biodiversity remains largely unknown. Here, we provide the first
quantitative synthesis of studies testing effects of soil biodiversity on C
cycling. Using meta-analysis and regression analysis, we showed that loss of
soil biodiversity can have negative consequences for the soil C cycle but
that relationships between C cycling processes with soil biodiversity vary
across groups of soil organisms and are process-dependent. Below we explore
how our findings contribute to our knowledge of how soil biodiversity drives
ecosystem functions. We also discuss experimental shortcomings and
methodological challenges and suggest directions for future research.
Biodiversity impacts on C pools
Few studies have assessed the relationship between soil biodiversity and
soil C pools. We found just three studies in our literature search, and
these studies all used different indices of soil C pools: ergosterol, which
is a measure of fungal biomass (Liebich et al., 2007); dissolved organic C
(Cragg and Bardgett, 2001); and soil organic C concentration (Zimmer et al.,
2005). All three studies were of short duration (range: 42 to 70 days) and were
conducted in microcosms in which diversity of the microbial community
(Liebich et al., 2007), microfauna (collembola; Cragg and Bardgett, 2001),
or macrofauna (earthworms and woodlice; Zimmer et al., 2005) was
manipulated. Due to the small number of studies, we were unable to conduct a
quantitative analysis. However, none of the individual studies showed an
impact of soil biodiversity on soil C pools. It is probably not surprising
that very few studies attempted to relate soil community diversity to soil C
pools, due to the difficulty of maintaining soil biodiversity manipulations
in microcosms for long time periods. Bulk soil organic C pools are typically
stable on the order of years to decades due the large pool sizes and the
relatively slow rates of biological processes (Conen et al., 2003; Smith,
2004), so short-term effect of soil biodiversity on this pool would be
expected to be low and undetectable. We expect that some metrics of the
relatively labile fraction in the soil C pool (e.g., particulate organic
matter, DOC) will be more temporally dynamic than the bulk pool (Six et al.,
2002) and would be better target response variables for assessing
biodiversity impacts. Even more likely to provide information on soil
biodiversity impacts on soil C cycling are studies assessing diversity
effects on short-term C fluxes.
Plant biomass, assessed as a whole or partitioned into root and shoot
biomass, was not significantly affected by soil biodiversity. Our analysis
included studies that manipulated diversity of mycorrhizal fungi (Baxter and
Dighton, 2001), microarthropods (Liiri et al., 2002), meso- and macrofaunal
decomposers (Eisenhauer and Schädler, 2011), or more complex foodwebs
consisting of multiple groups of different body sizes (Sulvaka et al., 2001;
Laakso and Setälä, 1999; Ladygina et al., 2010; Eisenhauer et al.,
2010; Eisenhauer and Schädler, 2011). The lack of a response of plant
biomass production to soil biodiversity results from the contradicting
results generated across a number of studies, and indicates that soil
biodiversity does not unequivocally promote plant production (reviewed in
van der Heijden et al., 2008). With the exception of rhizosphere organisms
such as mycorrhizal fungi, rhizobia, and root pathogens or herbivores,
linkages between soil organisms and plant biomass are indirect, i.e.,
decomposer organisms break down organic compounds and make nutrients
available for plant uptake (Wardle et al., 2004). This indirect link between
plant growth and soil organisms may result in a rather weak relationship
between soil biodiversity and plant production (Balvanera et al., 2006). In
line with this, we did find a trend of decreased plant production with loss
of soil biodiversity. However, the limited number of studies reduced our
statistical power, restricting our ability to quantify soil biodiversity
impacts on plant biomass production. Another complication in assessing
biodiversity impacts on plant production is that to date studies have
exclusively been conducted in laboratory and greenhouse settings. While
laboratory manipulations can provide useful information about potential
controls over ecological processes, these manipulations are by necessity
short-term (range: 35 days–52 weeks for the studies we compiled) and may
include only a subset of the complex food webs and biogeochemical processes
that occur in natural field settings (Hol et al., 2010). Furthermore,
diversity effects may become more apparent over time as functional
redundancy declines (Reich et al., 2012).
Although soil biodiversity did not conclusively impact soil C pools or
affect plant biomass production, biodiversity as a whole appears to play an
important role in maintaining and enhancing plant biomass production and
soil C pools. For example, plant diversity can promote plant biomass
production and soil C storage (Tilman et al., 2006; Cardinale et al., 2012;
Fornara and Tilman, 2008; Steinbeiss et al., 2008), and these benefits of
aboveground biodiversity on ecosystem functions are often attributed to
increases in plant nutrient uptake resulting from belowground spatial and
temporal differentiation in resource use (e.g., McKane et al., 1990; Tilman
et al., 1996; Casper and Jackson, 1997; Schenk et al., 1999; van Ruijven and
Berendse, 2005; van der Heijden et al., 2003). As such, a single limiting
resource (e.g., soil N) may be spatially partitioned among co-occurring
plant species, which reduces inter-specific competition and thereby
facilitates coexistence (McKane et al., 1990). It is reasonable to expect
that similar mechanisms occur for soil organisms (Prosser and Nicol, 2012;
Sechi et al., 2015), and effects of plant and soil organism diversity on
ecosystem functions may not be independent of each other, as increased plant
diversity may be accompanied by increased soil biodiversity (Scherber et al.,
2010; Eisenhauer et al., 2011a, b). If this is the case, soil biodiversity could
explain, at least in part, the observed positive relationship between plant
diversity and ecosystem C pools and processes. It is worthy of
mention here that soil fungal pathogens have been found an important driver of
observed positive relationships between plant diversity and productivity
(Maron et al., 2011).
Soil biodiversity impacts on soil C processes
Results from our meta-analysis show that loss of soil biodiversity
significantly reduces soil C respiration (-27.5 %). This is a strong
reduction in soil C mineralization that could have important ecosystem-level
consequences for the soil C cycle. However, some caution is warranted in
interpreting these results as the experimental design of many of the studies
included in the analyses may have inadvertently over- or underestimated soil
biodiversity impacts on processes important to the soil C cycle (Nielsen et
al., 2011). The response of C respiration to a loss in soil biodiversity was
overwhelmingly driven by studies manipulating soil microbial diversity, and
when we categorized the analysis by studies that manipulated the soil
microbial community only, the average response to a reduction in
biodiversity was even greater (-41 %). In addition, the regression
analysis revealed that a loss in soil biodiversity was significantly related
to a loss in soil C respiration only when soil microbial diversity was
included in the analysis. This suggests that these studies contributed in
large part to the strong response of soil C respiration to a reduction in
soil biodiversity. Many of these studies used a relatively low number of
microbial species when compared to soil microbial diversity in natural
ecosystems. For example, Setälä and McLean (2004) used 43 taxa of
saprophytic fungi, a large number relative to most manipulative experiments,
but a small number relative to the estimated number of fungi in natural
soils. In addition, the majority of the studies were conducted under highly
controlled and short-term laboratory conditions. Some studies used
fumigation or dilution methods to alter soil microbial diversity (Griffiths
et al., 2000, 2001, 2004; Wertz et al., 2006), and although it appears that
microbial diversity decreases with increased dilution or fumigation, the
main impacts of these treatments may be on the community structure by
favoring taxa that physiologically withstand the pressures of dilution or
fumigation. Studies using this technique showed that the response of C respiration to an increase in biodiversity leveled
off (Bell et al., 2005) with increasing species
number, and that responses to these treatments are often
idiosyncratic, which suggests that they are driven by the soil microbial
community structure rather than by diversity. Thus, although our synthesis
indicates that the response of soil C respiration to a reduction in soil
biodiversity can be significant, we contend that the response may be an
overestimation of what would happen in soils with natural communities.
Although many studies have assessed the impact of soil microbial diversity
on soil C respiration, only one study evaluated effects of soil faunal
diversity (earthworms) on C respiration (Scheu et al., 2002). This study
indicated that soil faunal richness impacts soil microbial community
functioning, which may in turn alter soil C respiration. Soil fauna can
strongly affect microbial processes and community composition by enhancing
the availability of resources to the microbial community (Edwards, 2000;
Heal et al., 1997; Petersen and Luxton, 1982). For example, invasive
earthworms in North America have been associated with changes in soil
respiration, although these effects may be mediated through changes in the
abundance of other organisms, such as microbes (Szlavecz et al., 2011) or
availability of soil and litter resources (Huang et al., 2010). Thus, a
change in the diversity of soil fauna is expected to alter soil C
respiration, but more studies that manipulate soil fauna are required to
confirm this hypothesis.
Results from the meta-analysis and the regression analysis show that loss of
soil biodiversity significantly reduces plant tissue decomposition
(-18 %). Unlike the other response variables, soil biodiversity impacts on
plant tissue decomposition were not dominated by studies that manipulated
the soil microbial community. Rather, ca. 84 % of decomposition studies in
our compilation manipulated soil fauna or multiple groups of soil organisms;
soil faunal biodiversity effects on plant tissue decomposition were
significant (-37 %). The significant impact of soil faunal manipulations
on litter decomposition may be due to the strong direct effect of soil fauna
on litter decomposition, particularly in the early stages of decomposition
(Heemsbergen et al., 2004; Berg and Laskowski, 2006; Milcu and Manning,
2011). By contrast, soil microbial diversity reductions alone did not
significantly suppress decomposition rates. This finding is despite
individual observations that the diversity of litter-associated microbes
increases as decomposition proceeds (e.g., Dilly et al., 2004). Nonetheless,
based on our observation that soil faunal diversity has a strong impact on
plant tissue decomposition, we propose that diversity of the soil community,
and particularly soil faunal diversity, is an important factor driving rates
of litter decomposition. This notion is supported by a recent article
showing that a reduction in the diversity of detritivores (both microorganisms
and invertebrates) slows the rate at which litter is decomposed, regardless
of the location of the experiment (Handa et al., 2014).
We hypothesized that diversity across multiple organismal groups composed of
different body sizes would impact soil C cycling processes to a greater
extent than diversity within groups of organisms with a single body size,
with the assumption that within-group diversity would be accompanied by a
greater degree of functional redundancy (Laakso and Setälä, 1999;
Wardle, 1999; Cragg and Bardgett, 2001). For example, a higher diversity of
feeding guilds has been linked to more effective substrate use
(Setälä and Huhta, 1991). Our analysis, however, suggests that, for
plant tissue decomposition, diversity across multiple body size groups has
similar impacts on soil C cycling to diversity within body size groups. For
soil C respiration, the effect of reduced diversity within groups was even
stronger than that of smaller diversity across groups. This may result from
the approaches taken to assess community impacts on soil C cycling. First,
there is a lack of consistency in approaches taken to study effects of soil
biodiversity on C cycling, both for the within- and the across-body size
group approach. As such, the level of functional diversity between “high-” and “low-” diversity treatments may have varied across studies, and
it is unclear whether shifts in functional diversity were greater for
across-group manipulations than for within-group manipulations. Except for
Heemsbergen et al. (2004), no studies have explicitly tested for the functional
dissimilarity among the species manipulated. Second, populations of soil
organisms at lower trophic levels may show compensatory growth responses to
loss of biomass predation by organisms of higher trophic levels (e.g.,
Ingham et al., 1985), resulting in no net effect of manipulation of trophic
diversity on the processes regulated by lower-trophic-level soil organisms.
Third, effects of functional or trophic groups of organisms may have
opposing effects on the C cycling pools and process rates. For example,
Ladygina et al. (2010) showed that arbuscular mycorrhizal fungi and
decomposer (enchytraeids and collembola) canceled each other out in
affecting plant community biomass.
Methodological concerns
While manipulating diversity of any group of organisms is fraught with
challenges, manipulation of soil organism diversity is particularly
challenging. A more comprehensive assessment of soil diversity impacts on C
cycling will require that some of these challenges are addressed. Nielsen et al. (2011) found that the response to a reduction in diversity was greater
if diversity levels were low (i.e., < 10 species included in the
analysis) and conducted under more controlled experimental conditions,
rather than under high diversity (i.e., > 10 species included in
the analysis) and more natural experimental conditions. Our analysis set out
to quantify how the design of the study affected soil C cycling processes,
by comparing studies that manipulated soil biodiversity within a single body
size group with studies that manipulated biodiversity across multiple
groups. Across-body size group manipulations approach the natural complexity
of soil food webs to a much greater degree than within-group manipulations.
However, even the most complex manipulations accounted only for a fraction
of the diversity likely under natural field conditions. As such, to further
enhance our understanding of soil community diversity impacts on soil C
cycling, studies should incorporate more natural conditions in their design
and manipulate more complete soil communities. A recent study by Wagg et al. (2014) used a method for manipulating a broad size range of soil organisms
by inoculating sterilized soils with soil communities derived through a
series of different-sized filters. This method allowed the researchers to
successfully obtain a broad soil biodiversity gradient within and across
groups of soil organism that span a gradient in body sizes in their
grassland microcosms, and showed that soil biodiversity loss and
simplification of soil community composition impairs multiple ecosystem
functions, including litter decomposition and soil C sequestration. However,
filtering groups of organisms based on body size does not allow for separating
between effects of functional dissimilarity from effects of biodiversity
(species richness) per se. This underscores the pervasive challenge to truly
measuring biodiversity effects on ecosystem processes. A parallel concern is
that soil biodiversity typically cuts across multiple trophic groups (e.g.,
manipulation of nematodes would potentially alter both herbivores and
predators in the soil). Aboveground diversity–ecosystem function studies
have typically been limited to the primary producer trophic level, but
results may yield very different relationships if consumer trophic levels
are included in diversity manipulations (Borer et al., 2014). Finally, we
caution that measuring soil biodiversity is exceedingly difficult, and in
many cases treatments were assumed to affect biodiversity for the duration
of the experiment, but this was often not measured. It is also possible that
a change in the relative abundance of organisms is an important component of
biodiversity and studies in our compilation typically equated species
richness with diversity, whereas biodiversity sensu stricto includes both species richness
and abundance (Magurran, 2005). To fully comprehend how biodiversity impacts
ecosystem function, an attempt should be made at manipulating and
maintaining different levels of soil community diversity, in the strict
sense of the definition.