Environmental Conservation: page 1 of 11 C
Foundation for Environmental Conservation 2016 doi:10.1017/S0376892916000011
How are soil carbon and tropical biodiversity related?
DOUGLAS SHEIL1∗, BRENTON LADD2,3 , LUCAS C. R. SILVA4, SHAWN W. LAFFAN5AND
MIRIAM VAN HEIST1
1Department of Ecology and Natural Resource Management, Norwegian University of Life Sciences, P.O. Box 5003 NO-1432 As, Norway,
2Evolution and Ecology Research Centre, School of Biological, Earth and Environmental Sciences, University of New South Wales, Sydney, NSW
2052, Australia, 3Facultad de Ciencias Ambientales, Universidad Cientíﬁca del Sur, Lima 33, Perú, 4Department of Land, Air and Water Resources,
University of California Davis, USA and 5School of Biological, Earth and Environmental Sciences, University of New South Wales, Sydney,
NSW 2052, Australia
Date submitted: 13 July 2015; Date accepted: 3 December 2015
This article discusses how biological conservation
can beneﬁt from an understanding of soil carbon.
Protecting natural areas not only safeguards the
biota but also curtails atmospheric carbon emissions.
Opportunities for funding biological conservation
could potentially be greater if soil carbon content
is considered. In this article current knowledge
concerning the magnitude and vulnerability of soil
carbon stocks is reviewed and the relationship of these
stocks to biological conservation values is explored.
Looking at two relatively well-studied tropical regions
we ﬁnd that 15 of 21 animal species of conservation
concern in the Virunga Landscape (Central Africa),
and nine of ten such species in the Federal District
of Brazil (Central Brazil), rely on carbon-rich habitats
(alluvial and/or wetlands). At national scales, densities
of species, endemics and threatened taxa (plants,
mammals, birds, reptiles, amphibians and ﬁsh) show
positive and signiﬁcant relations with mean soil carbon
content in all but two cases (threatened amphibians and
threatened ﬁsh). Of more than 1000 threatened species
in 37 selected tropical nations, 85% rely on carbon-
rich habitats. This tendency is observed in plants,
mammals, reptiles, amphibians and crustaceans, while
birds appear more evenly distributed. Research to
clarify and explore these relationships is needed. Soil
carbon offers major opportunities for conservation.
Keywords: carbon stocks, conservation, detritus, organic
matter, peat, REDD, tropical forests, wetlands
Stemming the increase in atmospheric carbon dioxide and
loss of tropical biodiversity are both global challenges. Major
international efforts are made, but much more needs to
∗Correspondence: Professor Douglas Sheil Tel: +47 67231783
∗Supplementary material can be found online at http://dx.doi.org/
be done (McCarthy et al. 2012; Fearnside 2013). The
projected annual cost of reaching 2050 carbon emission targets
ranges from US$350 billion to several trillion (Loftus et al.
2015). The total yearly expenditure on global biodiversity
conservation between 2001–2008 was approximately US$21.5
billion (Waldron et al. 2013), which is less than a third of the
US$76.1 billion (McCarthy et al. 2012) or perhaps more (Sheil
et al. 2013), needed to slow extinctions and achieve United
Nations biodiversity targets. Given the shortfall, available
resources should be used effectively.
Protecting natural areas can maintain carbon stocks and
biodiversity values simultaneously. Such synergies mean that
the limited funds available for carbon sequestration and nature
conservation could, if used wisely, achieve more of both
(Venter et al. 2009 a; Gardner et al. 2012;Phelpset al. 2012).
However, few studies have explored the complete carbon
beneﬁts arising from protecting natural habitats (forests and
non-forests) and how more complete carbon accounting
may increase resources for biological conservation. Reasons
likely include limited awareness and appreciation of the
Here, we ﬁrst identify the critical role of habitat protection
in addressing both atmospheric carbon emissions and
biodiversity conservation. Next we highlight the signiﬁcance
of soil carbon, its spatial variation and its vulnerability. We
then consider the distribution of threatened biodiversity. Next
we show that habitat categories with high soil carbon stocks
are associated with species-rich regions and also with many
threatened species. We then identify various uncertainties
and knowledge gaps. Despite many unknowns, our results
highlight important soil carbon-biodiversity relationships and
indicate that incentives to retain vulnerable soil carbon could
beneﬁt biodiversity conservation.
Conversion and degradation
Loss and degradation of forests and other natural habitats
threaten biodiversity (Newbold et al. 2015) and also release
about 0.9 Mg (C) to the atmosphere each year (GCP
2014). Efforts to reduce atmospheric carbon have generated
various initiatives to lower emissions from land cover
change. These initiatives include activities under nationally
2D. Sheil et al.
appropriate mitigation action (NAMA) and governmental
as well as market-led projects for reducing emissions from
deforestation and forest degradation (REDD). Although
aspects are debated, many such projects aim to incentivize the
protection of tree cover and associated carbon stocks. REDD
in particular has been the subject of many pilot projects,
evaluations and forecasts (Agrawal et al. 2011; Angelsen et al.
2012). Substantial opportunities appear possible; one forecast
suggests that REDD might generate as much as US$32.1
billion annually by 2020 (Streck & Parker 2012).
Much has been written about how efforts to reduce
carbon emissions could inﬂuence biodiversity conservation
(Venter et al. 2009 a; Gardner et al. 2012;Phelpset al.
2012; Armenteras et al. 2015; Beaudrot et al. in press).
Potential synergies offer an attractive basis for action.
Governments, NGOs and businesses like to ‘double count’
their contributions to the global environment. In some cases
funds spent on biodiversity conservation may even result
in greater carbon storage than equal spending on carbon
payments (Busch 2013). Several global initiatives seek to
reduce atmospheric carbon emissions in ways that also
contribute to biological and environmental conservation, for
example the Climate, Community & Biodiversity Alliance
initiative and REDD+Social & Environmental Standards
initiative (CCBA 2015;REDD+SES 2015).
Before these schemes can operate on a large scale, various
issues must be resolved. These include deﬁning baselines and
reference-level estimates of carbon stocks, selecting the best
methods for monitoring, reporting and veriﬁcation of carbon
stocks, and determining who will get paid how much, by whom
and for what (Agrawal et al. 2011;Batjes2011; Angelsen et al.
2012; Gardner et al. 2012;Smithet al. 2012; van Noordwijk
et al. 2014). Thus, while the ultimate format is uncertain, the
search for cost-effective ways to reduce atmospheric carbon
emissions looks likely to continue.
Carbon assessments and budgets have, by convention,
recognized ﬁve pools of carbon: above-ground biomass;
below-ground biomass; deadwood; litter; and soil organic
matter (Eggleston et al. 2006). Efforts are focused on
assessments of above-ground biomass (Clark & Kellner 2012;
Thomas & Martin 2012;Mitchardet al. 2013;Asneret al.
2014) but this is a streetlight effect: a case of looking at the
most visible carbon rather than at where the largest stocks
occur, which is in the soil.
Total soil carbon is greater than that found in global
vegetation and the atmosphere combined (Houghton 2007).
Published estimates are 504–3000 Gt with a median of 1460.5
Gt; the uncertainty itself is greater than the total carbon (816
Gt) in the atmosphere (Scharlemann et al. 2014). Considerable
variation in local carbon stocks adds to the complexity in
generating global estimates for total soil carbon. For example,
wetlands cover about 3% of the global land surface but
likely possess between one-quarter and one-third of the total
organic soil carbon (Gumbricht 2012). The different soil
depths considered by various assessments further complicate
While soil carbon content typically declines with depth, the
total stocks in deep soil are considerable. Soil proﬁles typically
indicate more than half as much carbon again if extended
from 1 to 3 m deep: namely 56% is one global summary’s
average, with 86% for deserts, 84% for tropical deciduous
forests and 74% for tropical grassland/savannah (Jobbágy &
Jackson 2000). The cumulative global totals are 615 Gt C in
the top 0.2 m, 1502 Gt to 1 m and 2344 Gt C to 3 m (Jobbágy &
Jackson 2000). Deeper proﬁles would locate additional carbon.
Many global estimates only consider soil to 1 m, while current
measurement standards only consider the top 0.3 m.
Where above- and below-ground biomasses have been
compared in natural terrestrial systems, soils tend to possess
the greater stocks when measured to sufﬁcient depth in
savannahs, forests and wetlands (Chen et al. 2003; Murdiyarso
et al. 2009;Silva et al. 2013 a). For example, in central Brazil’s
tropical wooded savannahs, and in tropical forests in riparian
zones within this biome, mean soil carbon stocks (to 1 m depth)
are markedly greater (c. 150 to more than 1500 t C ha−1)
than the associated above-ground biomass (c. 5to
100 t C ha−1;Silva et al. 2013 a). This is observed in natural
forest-savanna gradients, where plant diversity modulates the
effect of climate and disturbance regime on ecosystem C input
to soils (Franco et al. 2014; Silva 2014;Paivaet al.2015),
as well as in degraded soils undergoing restoration, where
a direct link between diversity of colonizing species and C
accumulation has been recently described (Silva et al. 2015).
Similarly in riverine forests, peat forests and mangroves, the
soils are especially carbon-rich. For example, in mangroves
in Tanjung Puting National Park (Central Kalimantan) mean
above-ground and soil carbon values were respectively 142.9
and 1220.2 t C ha–1 (Murdiyarso et al. 2009).
Soil carbon stocks vary across landscapes, inﬂuenced by many
factors (Gleixner 2013) and the underlying relationships are
complex to assess and extrapolate (Scharlemann et al. 2014).
For example, no relationship was detected between ﬁeld
observations and one interpolated international soil carbon
map (Ladd et al. 2013). Nevertheless, predictable relations
can be used to improve surveys and maps of soil carbon
stocks; at continental scales, relationships between climate,
plant productivity and soil carbon are statistically robust but
noisy (Ladd et al. 2013). Within a given biome, geology,
topography and drainage are important (Webster et al. 2011)
and within most landscapes soil carbon increases when moving
from dry upland sites to wet alluvial sites (van Noordwijk et al.
1997; Silva et al. 2008; Webster et al. 2011).
Soil carbon and tropical biodiversity 3
Conversion of old-growth forest to agriculture is the primary
threat to tropical biodiversity and above-ground biomass
(Ghazoul & Sheil 2010). Conversion is also a threat to the
carbon stored in soils.
Soil carbon stocks represent a dynamic balance between
inputs and outputs (Gleixner 2013). Carbon ﬂows from soils
to the atmosphere are several times larger than the emissions
from fossil-fuel combustion and small imbalances could have
major consequences (Stockmann et al. 2013). Soil carbon
can decline substantially when forest is cleared (Tiessen
et al. 1994). Measured values vary (Smith et al. 2012)but
typically 25–30% of soil carbon is lost in the ﬁrst two to three
decades after tropical forests are converted to cropland (Don
et al. 2011). Much of the carbon in wetland soils ultimately
decomposes when such soils are drained (Hooijer et al. 2012),
and carbon-rich sediments are often dispersed by currents
when riverine forests or mangroves are damaged or removed
(Murdiyarso et al. 2009).
When these losses are recognized in carbon accounting, the
values used are often not measured but rather based on generic
values, for example, in IPCC carbon accounting methods
(Eggleston et al. 2006;Batjes2011). The actual changes in soil
carbon stocks when natural land is converted to cultivation or
otherwise modiﬁed depend on the site and the management
practices. Many of the world’s cultivated soils have lost most
of their original carbon (often 30–40 t C ha−1), with greater
percentage losses through erosion from the most carbon-rich
tropical soils (Lal 2004). Erosion associated with agriculture
may remove an additional 0.404 (±0.202 SE) Gt of global
soil carbon annually (Doetterl et al. 2012). The fate of this
carbon remains uncertain although more than 30% likely
returns to the atmosphere (Regnier et al. 2013). Selective
timber harvesting typically has little impact on soil carbon
(Berenguer et al. 2014). In some cases, conversion to pasture
leads to little apparent change (Twongyirwe et al. 2013), or
even perhaps an ultimate increase (after an initial decline)
as observed in Amazonia (Fujisaki et al. 2015). Agroforestry
proponents highlight the beneﬁts of mixed cropping for soil
carbon (Young 1997). Nonetheless, soil carbon loss is the
typical outcome when land is converted to agriculture.
We know little about soil processes at depths greater than
a few tens of centimetres but we know that deep soil organic
carbon is dynamic (Balesdent et al. 2014) and can be inﬂuenced
by land use change (Fontaine et al. 2007;Xianget al. 2008). We
have almost no relevant data from the tropics. Nonetheless,
some soil researchers are advocating increased attention to
changes in deep soil carbon due to the considerable changes
sometimes observed with land use change (Boddey et al. 2010;
Shi et al. 2013).
LOCATING THREATENED BIODIVERSITY
Species diversity is positively correlated with moisture
availability at multiple scales. Moving along a gradient from
desert to forest as well as within habitats, wetter sites
usually have more species (Hawkins et al. 2003). Various
explanations exist, for example, wetter areas are more likely
to maintain species with ﬁner niches and strong niche
conservatism, maintain the productivity needed to support
higher species densities, and serve as (or are near to) refugia
during drought and past episodes of climate change (Hawkins
et al. 2003; Ghazoul & Sheil 2010; Romdal et al. 2013).
Dead biomass can play a signiﬁcant role in nutrient ﬂows,
community stability, trophic specializations and resulting
species diversity, though most observational studies have
focused on aquatic food webs (Hairston Jr and Hairston Sr
1993; Moore et al. 2004). Various theories also link diversity to
productivity (Cardinale et al. 2009; Willig 2011; Tilman et al.
Lowland alluvial soils, comprising parent materials
transported to their location by water, often show high
levels of nutrient accumulation and possess intrinsically
higher moisture-holding capacity due to their texture and
high organic matter content (van Noordwijk et al. 1997).
These areas are thus attractive for agriculture and can
support dense human populations. Intact ecosystems in wet
habitats are thus comparatively rare; species that persist are
often severely affected by habitat fragmentation and face
multiple additional challenges from human proximity (e.g.
exploitation, pollution). Many protected area networks are
dominated by land with low agricultural value. Based on
these considerations, we hypothesize that more species of
conservation signiﬁcance are associated with wet lowland
habitats than with drier ones, and are also at greater risk of
endangerment and extinction.
The criteria and measures allowing for an objective
exploration of this hypothesis are debatable. For example,
biologists and soil scientists would distinguish wet (but
well-drained) carbon- and species-rich lowlands from water-
logged wetlands, where a few specialized plant and animal
species dominate some anoxic environments (Junk et al.
2006; Ghazoul & Sheil 2010). But fractal-like drainage
patterns make such distinctions difﬁcult to apply consistently
for broad-scale assessments. Freshwater ecosystems harbour
many specialized and restricted species (Naiman et al. 1993),
so while only 0.8% of the world’s surface is covered by
freshwater, these water bodies host approximately one-third
of all vertebrate species (Dudgeon et al. 2006). Therefore,
protecting wet areas in general brings beneﬁts, even when
speciﬁc locations may show a reduced number of species. This
simpliﬁes the area selection process for combined soil carbon
and rare species conservation. In this context, we propose
that including wetlands within large landscapes generally
raises conservation values, especially when vulnerable aquatic
species are included. Wetlands may only represent a small
portion of total land area, but they are highly productive and
thus a strong continuous sink for atmospheric carbon (Mitsch
et al. 2013). Wetlands are also subject to the same threats
from human impacts and proximity as many other lowland
4D. Sheil et al.
Introduction and approach
Our ﬁrst evaluation considered the Virunga Landscape in
Central Africa. This is the highlands area of the Bwindi
ImpenetrableNational Park ofUganda and VirungaVolcanoes
of Rwanda, Congo and Uganda that provides the sole habitat
of the less than 1000 remaining Mountain gorillas (Gorilla
beringei beringei;Plumptreet al. 2007). Wetlands and wet
valley bottom habitats cover around one-quarter of the overall
landscape. We repeated this evaluation of habitat preference
and conservation status for a contrasting tropical landscape
in South America’s Cerrado biome of the Federal District
of Brazil. Alluvial gallery forests occupy only 5% of the
area (Felﬁli et al. 2001). We list as ‘threatened’ those species
recorded as ‘vulnerable, endangered, or critically endangered’
in the IUCN Red List (IUCN 2015).
In the Virunga Landscape of Central Africa, 15 of the 21 of
the region’s threatened animals (all vertebrates) are associated
with wetlands and/or valley bottoms for some or all of their
lives. This list of 15 is conservative as we know too little to
assess the habitat requirements of four of the other six species
(two rodents and two shrews; Appendix S1).
In the Federal District of Brazil, three of four of the
region’s threatened animals (all invertebrates) are associated
with wetlands. If we include a further six species, denoted
‘lower risk/conservation dependent’ and ‘needs updating’ in
the IUCN Red List 2015 (each known only from a single
collection at a single location), the summary rises to nine out
of ten species (all invertebrates; Appendix S1).
Introduction and approach
We sought to examine the relationship between soil carbon
and species of conservation signiﬁcance at larger scales using
national data. We anticipated that both biodiversity and soil
carbon would be positively associated with rainfall, so we also
explored annual rainfall data to capture and explain variation
in these indicators. We estimated mean topsoil carbon density
for the portion of each country that lies within the tropics
(% by weight in the top 0.3 m, source data described in
Nachtergaele et al. 2008). We used a similar procedure for
generating mean annual rainfall (Hijmans et al. 2005). Species
information is derived from the summary compilations of
country-level counts and densities in Roberts (1998). We
minimize the effects of country area on our comparisons of
country-level species counts by analysing species densities
(number of species/10 000 km2)aswellasresidualsfromthe
linear regression between the log of species count and the log
of country land area. We then use the resulting residuals as
area-corrected counts (Table 1 and Appendix S2).
Species of plants per 10000 km
Soil carbon (%)
0 2000 4000
Mean soil carbon (% by weight)
Mean annual rain (mm)
Figure 1 Example plots of (a) estimated plant species density
(species per 10 000 km2) versus mean topsoil carbon density (% by
weight for the top 0.3 m) for tropical countries by region and (b)
country mean topsoil carbon density (% by weight for the top 0.3
m) versus country mean-annual-rainfall for tropical areas of
countries with some portion in the tropics. The relation is positive
and signiﬁcant (p<0.0005, n =124, y =0.0008x +0.4409, R2=
0.23). Residuals were negatively correlated with country area
indicating some effects of scale and heterogeneity (Kendall’s tau =
-0.160, p=0.008, n =124).
We also assessed IUCN data (2013) for threatened species
(here threatened, critically endangered and extinct) country
by country until we had at least 1000 threatened species
(and stopped when every country included at that point
was completed). As we wanted to focus on the tropics we
included only countries with at least 70% of their area between
23°2616Nand23°2616 S. This process yielded summaries
for 37 tropical countries (n =16 for Africa, n =8forthe
Americas, n =6 for Asia and n =7 for Island states). For
each threatened species we determined which were speciﬁcally
associated with wetlands and/or valley bottoms for some or all
of their lives (i.e. swamps, mangroves and riparian forest), and
which species were listed as being either habitat generalists
or restricted to dry upland habitat (i.e. Terra Firme forest,
savanna and grassland). See Appendix S3 for more detail on
Variation in total species densities (per 10 000 km2)forplants,
mammals, birds, reptiles and amphibians is positively related
to variation in mean topsoil carbon density (Table 1). These
rank correlations are also observed within each continent (e.g.
plants; Fig. 1 a). Similar rank correlations with carbon density
were found for area-corrected counts of species including ﬁsh
and for area-corrected counts of endemic species. The results
for area-corrected counts of threatened species versus carbon
density were also positive for plants, mammals, birds and
reptiles though not for amphibians or ﬁsh (Table 1). All but
two of the country level species measures were also positively
correlated with country mean annual rainfall (averaged over
the same tropical area used for our topsoil carbon assessment).
Soil carbon and tropical biodiversity 5
Table 1 Kendall’s tau_b rank correlation (tau), associated probabilities (p) and number of countries included (n) for country mean topsoil
carbon (% by weight) and country mean-annual-rainfall versus estimated species density per 10 000 km2, and regression residuals from
log-log regressions of (log) total species counts, endemic and threatened taxa versus (log) country area. ∗Asterisks denote low probabilities
(∗ࣘ0.05, ∗∗ࣘ0.01, ∗∗∗ࣘ0.001). See Appendix S1 for more details on source data and data handling. na =Data were insufﬁcient.
Plants Mammals Birds Reptiles Amphibians Fish
Species per 10 000 km2
Carbon tau 0.454∗∗∗ 0.293∗∗∗ 0.388∗∗∗ 0.390∗∗∗ 0.369∗∗∗ na
Rain tau 0.451∗∗ 0.320∗∗ 0.409∗∗ 0.326∗∗ 0.249∗∗ na
Residuals on area: all species
Carbon tau 0.417∗∗∗ 0.297∗∗∗ 0.384∗∗∗ 0.400∗∗∗ 0.454∗∗∗ 0.279∗
n 8185845453 37
Rain tau 0.401∗∗∗ 0.333∗∗∗ 0.356∗∗∗ 0.322∗∗∗ 0.338∗∗∗ 0.345∗∗
n 8185845453 37
Residuals on area: endemic species
Carbon tau 0.344∗∗∗ 0.264∗∗ 0.280∗∗ 0.289∗∗ 0.330∗∗∗ na
Rain tau 0.247∗∗ 0.266∗∗ 0.251∗0.251∗∗ 0.282∗∗∗ na
Residuals on area: threatened species
Carbon tau 0.233∗∗ 0.263∗∗ 0.280∗∗ 0.374∗∗ −0.067 0.126
n 7885847923 35
Rain tau 0.160∗0.362∗∗ 0.329∗∗ 0.315∗∗ −0.099 0.082
p0.0379 9.40×10−79.55×10−64.09×10−50.509 0.486
n 7885847923 35
The strength of these rank correlations follows a pattern
similar to the rank correlations with mean topsoil carbon
density (Table 1), and country carbon density is positively
related to country mean-annual-rainfall (Fig. 1 b).
Our more detailed assessment of the habitat requirements
of listed threatened species included 1048 species, of which
85% are speciﬁcally associated with wetter habitats. While
the pattern appears general it shows some regional variation;
threatened species are strongly associated with alluvial habitat
in Asia and the Americas and less so in Africa (Fig. 2 a), and
are positively associated with rainfall (Fig. 2 b). Threatened
plants, mammals, reptiles, amphibians and crustaceans all
showed some association with wet lowland alluvial habitat,
but birds appeared more evenly distributed among habitats
(Fig 2 c).
Our broad-brush review and analyses raises many questions.
Better data on vulnerable soil carbon and its distribution
will be required to answer most of them. Nonetheless our
provisional analyses strongly suggest that natural habitats
with more soil carbon often possess more species and more
threatened species. These patterns hold for several groups of
aquatic and terrestrial organisms. Our results are consistent
with wetter, more carbon-rich, areas typically possessing
more species and more species of conservation concern; and
typically being under greater threat, than are drier less carbon-
There are two levels of synergy associated with considering
soil carbon in habitat protection: ﬁrst that consideration
of soil carbon greatly increases the total carbon storage
associated with most natural habitats and second that soil
carbon stocks are frequently greater where conservation needs
are also greater. While the ﬁrst is self-evident, the second
has not previously been noted. The closest is the study by
Venter et al. (2009 b), which showed how carbon payments
could substantially reduce the opportunity costs of protecting
natural forests in Borneo from conversion – especially in the
carbon-rich peat forests – if soil carbon were included in the
calculations (these relationships are also explored by Murray
et al. 2015).
If protection of soil carbon stocks can be translated into
habitat preservation the beneﬁts for species conservation could
6D. Sheil et al.
Africa Americas Asia Islands
Mean annual rain (mm)
AmphibiansBirds Crustacea InsectsPlants Mammals Reptiles
Figure 2 The percentage of species in
the IUCN categories vulnerable,
threatened, critically endangered and
extinct associated with low land alluvial
lowland habitat (a) by continent (b)by
rainfall and (c) by taxonomic group. Each
plotted value represents one of 37 tropical
countries (Africa , the Americas ,
Asia  and Island states ) that met
these criteria (Appendix S1). 1048 species
were examined. Horizontal bars are mean
values with standard errors. (b)y=
0.011x +57, R²=0.27. The outlier
arrowed near the bottom, the Solomon
Islands, is not included in the regression.
Not all countries provide data for all taxa,
n=14 for plants, n =28 for mammals, n
=36 for birds, n =24 for reptiles, n =28
for amphibians, n =4 for crustaceans and
n=15 for insects. Note only nine species
were in the extinct category and make
little difference to the overall pattern of
be considerable. Although our results appear promising we
view them as provisional: better data are required. If only
3% of the global land surface possess about 30% of the total
organic soil carbon (Gumbricht 2012) focusing on such areas
(and on forests for their above- and below-ground carbon)
offers a useful basis for setting priorities for wider habitat
Here we brieﬂy consider our results, their reliability and
In Central Africa and Central Brazil, most threatened animal
species are associated with habitats with comparatively
carbon-rich soils. The pattern is also apparent at larger scales:
in most tropical territories those with a greater percentage
of carbon in their topsoil also tended to have more species,
more endemic species and more threatened species. Of 1048
threatened species from 37 selected tropical countries, most
(85%) are associated with wetlands or alluvial habitats; this
preference is not only observed for taxa such as amphibians
and crustaceans, where it might be anticipated, but also
for mammals, plants and reptiles. Threatened birds are an
exception; this is striking given that birds are so often used
as indicators in conservation priority setting exercises (e.g.
McCarthy et al. 2012).
Available information is inadequate to conﬁdently assess all
the key relationships. Indeed our larger-scale assessments rely
on data that we ourselves have found to have limited accuracy
(Ladd et al. 2013). We know too little about soil carbon stocks
across landscapes, let alone how deep and how vulnerable
they are. We do not even know the extent of deep peat soils
in the tropics (Ghazoul & Sheil 2010). Our identiﬁcation of
annual rainfall data as a potential proxy for both soil carbon
and conservation values requires further exploration. We also
note that improved characterization and measurement of wet
versus dry habitat, and possibly of intermediate classes, may
help prioritize habitats suitable for species of conservation
Our argument assumes large-scale funding to reduce carbon
emissions. Here is not the place to examine such schemes,
but we can brieﬂy respond to those who claim too much
carbon is already available to support a viable market, namely
that oversupply means prices are too low to encourage trade
(Fearnside 2013). For sceptics, adding soil carbon would
exacerbate the oversupply problem and further decrease the
viability of any payment schemes. Such claims ignore the
scale of activities demanded. To make a sufﬁcient difference a
range of approaches is required. This necessitates inclusion of
more costly options thus ensuring a reasonable carbon price
Initiatives to protect soil carbon in natural habitats would offer
multiple opportunities. These include improved protection
Soil carbon and tropical biodiversity 7
of wetlands with migratory species (values captured in the
RAMSAR Convention) and of mangrove forests that stabilize
and protect coastlines and support ﬁsheries (Murdiyarso et al.
2009; Murdiyarso et al. 2015). Developments within the
IPCC, such as their Wetlands Supplement, suggest that such
opportunities are gaining traction (IPCC 2013).
Incentives for maintaining carbon in soils would also
encourage good land management practice more generally,
such as maintaining natural vegetation along watercourses
(Castelle et al. 1994). Protecting such vegetation beneﬁts
water quality, offers habitat for biodiversity and maintains
landscape connectivity. Such sites often play a special role in
supporting wildlife by producing food (young leaves, ﬂowers
and fruits), when water is scarce elsewhere due to seasonal
effects or drought (MacNally et al. 2009). Considering soil
carbon beneﬁts could also help offset the cost of habitat
restoration; evidence shows that restoration often increases
both soil carbon and species diversity (Silva et al. 2013 b).
Consideration of soil carbon could also guide and improve
local scale choices in conservation planning. For example, if a
habitat corridor is proposed connecting the two Mountain
gorilla populations in the Virunga Landscape of Central
Africa, then accounting for soil carbon might encourage
incorporation of low-lying terrain that, despite its higher
cost, can provide better habitat and result in improved
connectivity for a greater number of threatened species.
Further insight may be gained from examining how drainage
patterns inﬂuence soil properties and how these patterns relate
to carbon stocks and their vulnerability as well as the needs of
species of conservation signiﬁcance (Lowe et al. 2006).
The technical capacity and ﬁnancing required to protect
carbon stocks in the tropics brings challenges (Angelsen et al.
2012). A more complete accounting for soil carbon would
raise multiple technical issues including assessment methods.
Consider depth: many studies have only considered the top 20
or 30 cm, whereas deeper carbon stocks are often substantial
(Harper & Tibbett 2013). In tropical savannahs and forests
the carbon stocks do not always decline with depth and the
ﬁrst metre of the soil proﬁle may have several times as much
carbon as the top 0.2 m (Silva et al. 2013 a). Tropical peats
(soils dominated by organic matter) can reach more than 15 m
deep (Rieley et al. 1997). Though less rich in organic matter,
mineral soils and subsoils can be much deeper, for example, in
Surinam, bed rock may be more than 100 m below the surface
(FAO 2001). More needs to be learned about deep soil carbon.
The IUCN assessments themselves contain various
uncertainties including data-deﬁcient species and possible
biases in coverage, while the habitat descriptions are broad and
unsuitable for ﬁne-scaled analysis. The correlation between
soil carbon and conservation values is a general pattern, a
scatter of points rather than a tight linear relationship. There
will be sites with high carbon soils and low biodiversity
values, and sites with low carbon soils and high conservation
values (Murray et al. 2015). Also, different considerations
may yield other priorities. Consider the equatorial Asian
rainforests: those on mineral soil typically possess more
threatened terrestrial plant and animal species per hectare
than nearby carbon-rich peat swamp forest (Slik et al. 2009).
We suspect that examples such as these have persuaded many
conservation biologists that the most carbon-rich sites are
poor in biodiversity values. However, the biodiversity of
these forests is only low when considered relative to those
with the highest values on the planet and far surpasses
most. Furthermore the rapid conversion of peat forest for
oil palm cultivation and the poorly known aquatic biota in
these systems might also inﬂuence our conservation value
weightings (Venter et al. 2009 b; Meijaard & Sheil 2013).
An evaluation of stenotopic ﬁsh associated with peat forests
predicts 16 extinctions by 2050 due to habitat loss (Giam et al.
2012). Similar patterns may occur elsewhere.
Carbon payments can help protect habitat but they are
no panacea (Angelsen et al. 2012; van Noordwijk et al.
2014). Achieving net conservation beneﬁts from synergies
with carbon stocks requires that carbon ﬁnance does not
substitute conventional conservation funding. Threatened
species outside carbon-rich sites (e.g. many birds) also
require conservation and resources will still be needed to
address hunting, over-harvesting, invasive species and other
Addressing information needs
Many uncertainties exist in quantifying the distribution of
soil carbon vulnerable to habitat change and in gaining
accurate measures of conservation value. We highlight the
value of biologists and soil scientists working together to
better characterize these patterns and their inter-relationships.
Developing the capacity to improve knowledge of soil carbon
in natural and modiﬁed habitats appears a surmountable
problem: most countries provide training in soil sciences with
a focus on agriculture, and if more funds are directed to
quantifying and monitoring soil carbon more generally, this
will provide an incentive to develop the necessary skills.
Despite shortcomings in available data, it seems clear that
natural habitats with greater soil carbon stocks are often
associated with more species and also more threatened species
than those with less soil carbon. We have looked at this in
multiple ways and each indicates this relationship.
Fifteen of twenty one and nine of ten animal species of
conservation concern in Central Africa and in Central Brazil,
respectively, rely on carbon-rich wetland habitats. At the
country level the densities of species of plants, mammals,
birds, reptiles, amphibians and ﬁsh, as well as the densities
of endemics and threatened taxa in these same groups, tend
to be positively correlated with mean soil carbon content
(threatened amphibians and threatened ﬁsh are exceptions).
8D. Sheil et al.
Looking at what we know about 1048 threatened species we
ﬁnd evidence that 85% rely on wetlands or carbon-rich alluvial
habitats to a signiﬁcant degree. This tendency is observed for
plants, mammals, reptiles, amphibians and crustaceans but
not for birds.
In total our results indicate that wetter, more carbon-rich,
habitats harbour more species of conservation signiﬁcance,
than do drier less carbon-rich habitats. These carbon-rich
habitats, and their biota, are also under greater threat from
human activities, which further accentuates the conservation
signiﬁcance of these areas and their species. Current data
appear inadequate to fully explore these relationships with
We note that annual rainfall measurements have potential
as indicators of soil carbon densities and also of conservation
values at larger scales. High soil fertility and adequate water
supply have led to the conversion of many lowland-alluvial
habitats to agriculture or plantations. Including soil carbon
in funding schemes to reduce global carbon emissions could
increase the funds available to protect natural habitats. The
spatial correlation between patterns of soil carbon and patterns
of threatened biodiversity suggest signiﬁcant opportunities
for biodiversity conservation if soil carbon protection was a
marketable beneﬁt. Protecting sites with high conservation
value will safeguard large stocks of carbon from being emitted
into the atmosphere and often protecting sites with high soil
carbon will contribute signiﬁcantly to biological conservations.
The scale of the possible synergies and beneﬁts highlights that
the relationship between tropical soil carbon and biodiversity
values deserves recognition and evaluation.
DS and BL conceived the study, developed the analyses and
led on writing the paper. LS, SL and MvH contributed data
and suggestions. All authors contributed to and approved
the ﬁnal text. We declare no conﬂicts of interest. We thank
Les Christadis, Xingli Giam, Meine van Noordwijk, Nicholas
Polunin and reviewers for their comments on earlier texts. No
speciﬁc grants were involved in writing this article though this
paper contributes to the outputs of the Australian Research
Council project LP130100498 and no conﬂicts of interest were
To view supplementary material for this article, please visit
Agrawal, A., Nepstad, D. & Chhatre, A. (2011) Reducing emissions
from deforestation and forest degradation. Annual Review of
Environment and Resources 36: 373–396.
Angelsen, A., Brockhaus, M., Sunderlin, W. D. & Verchot, L. V.
(2012) Analysing REDD+: challenges and choices. Bogor,
Armenteras, D., Rodríguez, N. & Retana, J. (2015) National and
regional relationships of carbon storage and tropical biodiversity.
Biological Conservation 192: 378–386.
Asner, G. P., Knapp, D. E., Martin, R. E., Tupayachi, R., Anderson,
C. B., Mascaro, J., Sinca, F., Chadwick, K. D., Higgins, M.,
Farfan, W. & Llactayo, W. (2014) Targeted carbon conservation
at national scales with high-resolution monitoring. Proceedings of
the National Academy of Sciences of the United States of America
Balesdent, J., Basile-Doelsch, I., Chadoeuf, J., Cornu, S.,
Derrien, D., Fekiacova, Z. & Hatté, C. (2014) Dynamics
of carbon in deep soils inferred from carbon stable isotopes
signatures: a worldwide meta-analysis. In: EGU General
Assembly Conference Abstracts, p. 9052. [www document]. URL
Batjes, N. H. (2011) Soil organic carbon stocks under native
vegetation – revised estimates for use with the simple assessment
option of the Carbon Beneﬁts Project system. Agriculture,
Ecosystems & Environment 142(3): 365–373.
Beaudrot, L., Kroetz, K., Alvarez-Loayza, P., Amaral, I.,
Breuer, T., Fletcher, C. D., Jansen, P. A., Kenfack, D.,
Lima, M. G. M., Marshall, A. R., Martin, E. H., Ndoundou-
Hockemba, M., O’Brien, T. G., Razaﬁmahaimodison, J. C.,
Romero-Saltos, H., Rovero, F., Roy, C. H., Sheil, D., Silva,
C. E., Spironello, W. R., Valencia, R., Zvoleff, A., Ahumada,
J. & Andelman, S. Limited carbon and biodiversity co-beneﬁts
for tropical forest mammals and birds. Ecological Applications.
Berenguer, E., Ferreira, J., Gardner, T. A., Aragão, L. E. O. C., De
Camargo, P. B., Cerri, C. E., Durigan, M., Oliveira, R. C. D.,
Vieira, I. C. G. & Barlow, J. (2014) A large-scale ﬁeld assessment
of carbon stocks in human-modiﬁed tropical forests. Global Change
Biology 20(12): 3713–3726.
Boddey, R. M., Jantalia, C. P., Conceicao, P. C., Zanatta, J. A.,
Bayer, C., Mielniczuk, J., Dieckow, J., Dos Santos, H. P.,
Denardin, J. E., Aita, C. & Giacomini, S. J. (2010) Carbon
accumulation at depth in Ferralsols under zero-till subtropical
agriculture. Global Change Biology 16(2): 784–795.
Busch, J. (2013) Supplementing REDD+with biodiversity
payments: the paradox of paying for multiple ecosystem services.
Land Economics 89(4): 655–675.
Cardinale, B. J., Hillebrand, H., Harpole, W., Gross, K. &
Ptacnik, R. (2009) Separating the inﬂuence of resource
‘availability’from resource ‘imbalance’on productivity–diversity
relationships. Ecology Letters 12(6): 475–487.
Castelle, A. J., Johnson, A. W. & Conolly, C. (1994) Wetland
and stream buffer size requirements – a review. Journal of
Environmental Quality 23(5): 878–882.
CCBA (2015) The Climate, Community & Biodiversity Alliance.
[www document]. URL http://www.climate-standards.org/
Chen, X., Hutley, L. & Eamus, D. (2003) Carbon balance of a
tropical savanna of northern Australia. Oecologia 137(3): 405–
Clark, D. B. & Kellner, J. R. (2012) Tropical forest biomass
estimation and the fallacy of misplaced concreteness. Journal of
Vegetation Science 23(6): 1191–1196.
Doetterl, S., Van Oost, K. & Six, J. (2012) Towards constraining
the magnitude of global agricultural sediment and soil organic
Soil carbon and tropical biodiversity 9
carbon ﬂuxes. Earth Surface Processes and Landforms 37(6): 642–
Don, A., Schumacher, J. & Freibauer, A. (2011) Impact of tropical
land-use change on soil organic carbon stocks – a meta-analysis.
Global Change Biology 17(4): 1658–1670.
Dudgeon, D., Arthington, A. H., Gessner, M. O., Kawabata,
Z.-I., Knowler, D. J., Lévêque, C., Naiman, R. J., Prieur-
Richard, A.-H., Soto, D., Stiassny, M. L. J. & Sullivan, C. A.
(2006) Freshwater biodiversity: importance, threats, status and
conservation challenges. Biological Reviews 81(2): 163–182.
Eggleston, H., Buendia, L., Miwa, K., Ngara, T. & Tanabe, K.
(2006) IPCC guidelines for national greenhouse gas inventories.
Hayama, Japan: Institute for Global Environmental Strategies.
FAO (2001) Lecture notes on the major soils of the World. Rome,
Fearnside, P. M. (2013) What is at stake for Brazilian Amazonia in
the climate negotiations. Climatic Change 118(3): 509–519.
Felﬁli, J. M., Mendon¸
ca, R. C. & Walter, B. S. J. M. C.(2001) Flora
fanerogâmica das Matas de Galeria e Ciliares do Brasil. In: Cerrado:
cão e Recupera¸
cão de Matas de Galeria. Planaltina, eds.
J. F. Ribeiro, C. E. L. Fonseca & J. C. Sousa-Silva, pp. 195–263.
Fontaine, S., Barot, S., Barré, P., Bdioui, N., Mary, B. & Rumpel,
C. (2007) Stability of organic carbon in deep soil layers controlled
by fresh carbon supply. Nature 450(7167): 277–280.
Franco, A., Rossatto, D., de Carvalho Ramos Silva, L. & da Silva
Ferreira, C. (2014) Cerrado vegetation and global change: the
role of functional types, resource availability and disturbance in
regulating plant community responses to rising CO2levels and
climate warming. Theoretical and Experimental Plant Physiology
Fujisaki, K., Perrin, A. S., Desjardins, T., Bernoux, M., Balbino,
L. C. & Brossard, M. (2015) From forest to cropland and pasture
systems: a critical review of soil organic carbon stocks changes in
Amazonia. Global Change Biology.21(7): 2773–2786.
Gardner, T. A., Burgess, N. D., Aguilar-Amuchastegui, N., Barlow,
J., Berenguer, E., Clements, T., Danielsen, F., Ferreira, J.,
Foden, W. & Kapos, V. (2012) A framework for integrating
biodiversity concerns into national REDD+programmes.
Biological Conservation 154: 61–71.
GCP (2014) The Global Carbon Project [www document]. URL
Ghazoul, J. & Sheil, D. (2010) Tropical Rain Forests Ecology, Diversity
and Conservation. Oxford, UK: Oxford University Press.
Giam, X., Koh, L. P., Tan, H. H., Miettinen, J., Tan, H. T. W. &
Ng, P. K. L. (2012) Global extinctions of freshwater ﬁshes follow
peatland conversion in Sundaland. Frontiers in Ecology and the
Environment 10(9): 465–470.
Gleixner, G. (2013) Soil organic matter dynamics: a biological
perspective derived from the use of compound-speciﬁc isotopes
studies. Ecological Research 28(5): 683–695.
Gumbricht, T. (2012) Mapping global tropical wetlands from earth
observing satellite imagery. Bogor, Indonesia: CIFOR.
Hairston, N. G. Jr & Hairston, N. G. Sr (1993) Cause-effect
relationships in energy ﬂow, trophic structure, and interspeciﬁc
interactions. American Naturalist 142(3): 379–411.
Harper, R. & Tibbett, M. (2013) The hidden organic carbon in deep
mineral soils. Plant and Soil 368(1): 641–648.
Hawkins, B. A., Field, R., Cornell, H. V., Currie, D. J., Guégan,
J.-F., Kaufman, D. M., Kerr, J. T., Mittelbach, G. G., Oberdorff,
T., O’Brien, E. M. & Porter, E. E. (2003) Energy, water, and
broad-scale geographic patterns of species richness. Ecology
Hijmans, R. J., Cameron, S. E., Parra, J. L., Jones, P. G. & Jarvis, A.
(2005) Very high resolution interpolated climate surfaces for global
land areas. International Journal of Climatology 25(15): 1965–
Hooijer, A., Page, S., Jauhiainen, J., Lee, W., Lu, X., Idris, A. &
Anshari, G. (2012) Subsidence and carbon loss in drained tropical
peatlands. Biogeosciences 9(3): 1053–1071.
Houghton, R. A. (2007) Balancing the global carbon budget. Annual
Review of Earth and Planetary Sciences 35: 313–347.
IPCC (2013) Supplement to the 2006 IPCC Guidelines for
National Greenhouse Gas Inventories: Wetlands (Wetlands
Supplement), eds. T. Hiraishi, T. Krug, K. Tanabe, N. Srivastava,
J. Baasansuren, M. Fukuda & T. G. Troxler. Switzerland: IPCC
IUCN (2013; updated 2015) The IUCN Red List of threatened
species. [www document]. URL http://www.iucnredlist.org
Jobbágy, E. G. & Jackson, R. B. (2000) The vertical distribution
of soil organic carbon and its relation to climate and vegetation.
Ecological Applications 10(2): 423–436.
Junk, W. J., Brown, M., Campbell, I. C., Finlayson, M., Gopal, B.,
Ramberg, L. & Warner, B. G. (2006) The comparative biodiversity
of seven globally important wetlands: a synthesis. Aquatic Sciences
Ladd, B., Laffan, S. W., Amelung, W., Peri, P. L., Silva, L. C. R.,
Gervassi, P., Bonser, S. P., Navall, M. & Sheil, D. (2013)
Estimates of soil carbon concentration in tropical and temperate
forest and woodland from available GIS data on three continents.
Global Ecology and Biogeography 22(4): 461–469.
Lal, R. (2004) Soil carbon sequestration to mitigate climate change.
Geoderma 123(1): 1–22.
Loftus, P. J., Cohen, A. M., Long, J. & Jenkins, J. D. (2015) A critical
review of global decarbonization scenarios: what do they tell us
about feasibility? Wiley Interdisciplinary Reviews: Climate Change
Lowe, W. H., Likens, G. E. & Power, M. E. (2006) Linking scales
in stream ecology. BioScience 56(7): 591–597.
MacNally, R., Bennett, A. F., Thomson, J. R., Radford, J. Q.,
Unmack, G., Horrocks, G. & Vesk, P. A. (2009) Collapse of an
avifauna: climate change appears to exacerbate habitat loss and
degradation. Diversity and Distributions 15(4): 720–730.
McCarthy, D. P., Donald, P. F., Scharlemann, J. P., Buchanan,
G. M., Balmford, A., Green, J. M., Bennun, L. A., Burgess,
N. D., Fishpool, L. D. & Garnett, S. T. (2012) Financial costs of
meeting global biodiversity conservation targets: current spending
and unmet needs. Science 338(6109): 946–949.
Meijaard, E. & Sheil, D. (2013) Oil palm plantations in the context
of biodiversity conservation. In: Encyclopedia of Biodiversity (2nd
Edition), ed. S. A. Levin, pp. 600–612. The Nethelands; Elsevier.
Mitchard, E. T., Saatchi, S. S., Baccini, A., Asner, G. P., Goetz,
S. J., Harris, N. L. & Brown, S. (2013) Uncertainty in the spatial
distribution of tropical forest biomass: a comparison of pan-
tropical maps. Carbon Balance and Management 8(10): 1–13.
Mitsch, W. J., Bernal, B., Nahlik, A. M., Mander, Ü., Zhang, L.,
Anderson, C. J., Jørgensen, S. E. & Brix, H. (2013) Wetlands,
carbon, and climate change. Landscape Ecology 28(4): 583–
Moore, J. C., Berlow, E. L., Coleman, D. C., Ruiter, P. C., Dong,
Q., Hastings, A., Johnson, N. C., McCann, K. S., Melville, K.,
Morin, P. J. & Nadelhoffer, K. (2004) Detritus, trophic dynamics
and biodiversity. Ecology Letters 7(7): 584–600.
10 D. Sheil et al.
Murdiyarso, D., Donato, D., Kauffman, J. B., Kurnianto, S.,
Stidham, M. & Kanninen, M. (2009) Carbon storage in mangrove
and peatland ecosystems: a preliminary account from plots in
Indonesia. Working Paper 48. Bogor, Indonesia: CIFOR.
Murdiyarso, D., Purbopuspito, J., Kauffman, J. B., Warren, M. W.,
Sasmito, S. D., Donato, D. C., Manuri, S., Krisnawati, H.,
Taberima, S. & Kurnianto, S. (2015) The potential of Indonesian
mangrove forests for global climate change mitigation. Nature
Climate Change 5(12): 1089–1092.
Murray, J. P., Grenyer, R., Wunder, S., Raes, N. & Jones, J. P. (2015)
Spatial patterns of carbon, biodiversity, deforestation threat, and
REDD+projects in Indonesia. Conservation Biology.29(5): 1434–
Nachtergaele, F., Van Velthuizen, H., Verelst, L., Batjes, N.,
Dijkshoorn, K., Van Engelen, V., Fischer, G., Jones, A.,
Montanarella, L., Petri, M. & Prieler, S. (2008) Harmonized world
soil database. Rome, Italy: FAO.
Naiman, R. J., Decamps, H. & Pollock, M. (1993) The role of
riparian corridors in maintaining regional biodiversity. Ecological
Applications 3(2): 209–212.
Newbold, T., Hudson, L. N., Hill, S. L., Contu, S., Lysenko, I.,
Senior, R. A., Börger, L., Bennett, D. J., Choimes, A., Collen,
B., Day, J., De Palma, A., Diaz, S., Echeverria-Londono, S.,
Edgar, M. J., Feldman, A., Garon, M., Harrison, M. L. K.,
Alhusseini, T., Ingram, D. J., Itescu, Y., Kattge, J., Kemp, V.,
Kirkpatrick, L., Kleyer, M., Correia, D. L. P., Martin, C. D.,
Meiri, S., Novosolov, M., Pan, Y., Phillips, H. R. P., Purves,
D. W., Robinson, A., Simpson, J., Tuck, S. L., Weiher, E.,
White, H. J., Ewers, R. M., Mace, G. M., Scharlemann, J. P. W.
& Purvis, A. (2015) Global effects of land use on local terrestrial
biodiversity. Nature 520(7545): 45–50.
Paiva, A. O., Silva, L. C. R. & Haridasan, M. (2015) Productivity-
efﬁciency tradeoffs in tropical forest-savanna transitions: linking
plant and soil processes through litter input and composition.
Plant Ecology 216, 775–787.
Phelps, J., Webb, E. L. & Adams, W. M. (2012) Biodiversity co-
beneﬁts of policies to reduce forest-carbon emissions. Nature
Climate Change 2(7): 497–503.
Plumptre, A. J., Kujirakwinja, D., Treves, A., Owiunji, I. &
Rainer, H. (2007) Transboundary conservation in the greater
Virunga landscape: its importance for landscape species. Biological
Conservation 134(2): 279–287.
REDD+SES (2015) REDD+Social & Environmental Standards
[www document]. URL http://www.redd-standards.org/
Regnier, P., Friedlingstein, P., Ciais, P., Mackenzie, F. T.,
Gruber, N., Janssens, I. A., Laruelle, G. G., Lauerwald,
R., Luyssaert, S., Andersson, A. J., Arndt, S., Arnosti, C.,
Borges, A. V., Dale, A. W., Gallego-Sala, A., Godderis, Y.,
Goossens, N., Hartmann, J., Heinze, C., Ilyina, T., Joos, F.,
LaRowe, D. E., Leifeld, J., Meysman, F. J. R., Munhoven,
G., Raymond, P. A., Spahni, R., Suntharalingam, P. &
Thullner, M. (2013) Anthropogenic perturbation of the carbon
ﬂuxes from land to ocean. Nature Geoscience 6(8): 597–
Rieley, J., Page, S. & Shepherd, P. (1997) Tropical bog forests
of South East Asia. In: Conserving Peatlands,eds.L.Parkyn,
R. Stoneman & H. Ingram, pp. 35–41. Wallingford, UK:
Roberts, L. (1998) World Resources 1998–99. Washington D.C.,
USA: World Resources Institute; United Nations Environment
Programme; United Nations Development Programme; World
Romdal, T. S., Araújo, M. B. & Rahbek, C. (2013) Life on a tropical
planet:niche conservatism and theglobal diversity gradient. Global
Ecology and Biogeography 22(3): 344–350.
Scharlemann, J. P., Tanner, E. V., Hiederer, R. & Kapos, V. (2014)
Global soil carbon: understanding and managing the largest
terrestrial carbon pool. Carbon Management 5(1): 81–91.
Sheil, D., Meijaard, E., Angelsen, A., Sayer, J. & Vanclay, J. K.
(2013) Sharing future conservation costs. Science 339(6117): 270–
Shi, S., Zhang, W., Zhang, P., Yu, Y. & Ding, F. (2013) A synthesis
of change in deep soil organic carbon stores with afforestation of
agricultural soils. Forest Ecology and Management 296: 53–63.
Silva, L.C.R. (2014) The importance of climate-driven forest-
savanna biome shifts in anthropological and ecological research.
Proceedings of the National Academy of Sciences of the United States
of America 111(37) E3831–E3832.
Silva, L. C. R., Doane, T. A., Corrêa, R. S., Valverde, V., Pereira,
E. I. P. & Horwath, W. R. (2015) Iron-mediated stabilization
of soil carbon ampliﬁes the beneﬁts of ecological restoration in
degraded lands. Ecological Applications 25(5): 1226–1234.
Silva, L. C. R., Sternberg, L. S. L., Haridasan, M., Hoffmann,
W. A., Miralles-Wilhelm, F. & Franco, A. C. (2008) Expansion
of gallery forests into central Brazilian savannas. Global Change
Biology 14(9): 2108–2118.
Silva, L., Hoffmann, W. A., Rossatto, D. R., Haridasan, M., Franco,
A. C. & Horwath, W. R. (2013 a) Can savannas become forests? A
coupled analysis of nutrient stocks and ﬁre thresholds in central
Brazil. Plant and Soil 373(1): 829–842.
Silva, L., Corrêa, R., Doane, T. A., Pereira, E. & Horwath, W. R.
(2013 b) Unprecedented carbon accumulation in mined soils: the
synergistic effect of resource input and plant species invasion.
Ecological Applications 23:1345–1356.
Slik, J., Raes, N., Aiba, S. I., Brearley, F. Q., Cannon, C. H.,
Meijaard, E., Nagamasu, H., Nilus, R., Paoli, G., Poulsen,
A. D., Sheil, D., Suzuki, E., Van Valkenburg, J. L. C. H., Webb,
C. O., Wilkie, P. & Wulffraat, S. (2009) Environmental correlates
for tropical tree diversity and distribution patterns in Borneo.
Diversity and Distributions 15(3): 523–532.
Smith, P., Davies, C. A., Ogle, S., Zanchi, G., Bellarby, J., Bird, N.,
Boddey, R. M., McNamara, N. P., Powlson, D., Cowie, A., van
Noordwijk, M., Davis, S. C., Richter, D. D. E. B., Kryzanowski,
L., van Wijk, M. T., Stuart, J., Kirton, A., Eggar, D., Newton-
Cross, G., Adhya, T. K. & Braimoh, A. K. (2012) Towards an
integrated global framework to assess the impacts of land use and
management change on soil carbon: current capability and future
vision. Global Change Biology 18(7): 2089–2101.
Stockmann, U., Adams, M. A., Crawford, J. W., Field, D. J.,
Henakaarchchi, N., Jenkins, M., Minasny, B., McBratney, A. B.,
de Courcelles, V. d. R., Singh, K. & Wheeler, I. (2013) The
knowns, known unknowns and unknowns of sequestration of soil
organic carbon. Agriculture, Ecosystems & Environment 164: 80–
Streck, C. & Parker, C. (2012) Financing REDD+.In:Analysing
REDD+: Challenges and Choices, eds. A. Angelsen, M. Brockhaus,
W. Sunderlin & L. Verchot, pp. 111–127. Bogor, Indonesia:
Thomas, S. C. & Martin, A. R. (2012) Carbon content of tree tissues:
a synthesis. Forests 3(2): 332–352.
Soil carbon and tropical biodiversity 11
Tiessen, H., Cuevas, E. & Chacon, P. (1994) The role of soil organic-
matter in sustaining soil fertility. Nature 371(6500): 783–785.
Tilman, D., Reich, P. B. & Isbell, F. (2012) Biodiversity impacts
ecosystem productivity as much as resources, disturbance, or
herbivory. Proceedings of the National Academy of Sciences of the
United States of America 109(26): 10394–10397.
Twongyirwe, R., Sheil, D., Majaliwa, J. G. M., Ebanyat, P.,
Tenywa, M. M., van Heist, M. & Kumar, L. (2013) Variability of
soil organic carbon stocks under different land uses: a study in an
afro-montane landscape in southwestern Uganda. Geoderma 193:
van Noordwijk, M., Agus, F., Dewi, S. & Purnomo, H. (2014)
Reducing emissions from land use in Indonesia: motivation,
policy instruments and expected funding streams. Mitigation and
Adaptation Strategies for Global Change 19(6): 677–692.
van Noordwijk, M., Woomer, P., Cerri, C., Bernoux, M. & Nugroho,
K. (1997) Soil carbon in the humid tropical forest zone. Geoderma
Venter, O., Laurance, W. F., Iwamura, T., Wilson, K. A., Fuller,
R. A. & Possingham, H. P. (2009 a) Harnessing carbon payments
to protect biodiversity. Science 326(5958): 1368.
Venter, O., Meijaard, E., Possingham, H., Dennis, R., Sheil, D.,
Wich, S., Hovani, L. & Wilson, K. (2009 b) Carbon payments as
a safeguard for threatened tropical mammals. Conservation Letters
Waldron, A., Mooers, A. O., Miller, D. C., Nibbelink, N.,
Redding, D., Kuhn, T. S., Roberts, J. T. & Gittleman, J. L.
(2013) Targeting global conservation funding to limit immediate
biodiversity declines. Proceedings of the National Academy of
Sciences of the United States of America 110(29): 12144–
Webster, K., Creed, I., Beall, F. & Bourbonnière, R. (2011) A
topographic template for estimating soil carbon pools in forested
catchments. Geoderma 160(3): 457–467.
Willig, M. R. (2011) Biodiversity and productivity. Science
Xiang, S.-R., Doyle, A., Holden, P. A. & Schimel, J. P. (2008)
Drying and rewetting effects on C and N mineralization and
microbial activity in surface and subsurface California grassland
soils. Soil Biology and Biochemistry 40(9): 2281–2289.
Young, A. (1997) Agroforestry for Soil Management. Wallingford,