全 文 ：An experimental approach to addressing ecological questions related
to the conservation of plant biodiversity in China
Roy Turkington a, b, *, William L. Harrower a, b
a Botany Department, University of British Columbia, Vancouver, BC, V6T 1Z4, Canada
b Biodiversity Research Centre, University of British Columbia, Vancouver, BC, V6T 1Z4, Canada
a r t i c l e i n f o
Article history:
Received 4 November 2015
Received in revised form
5 December 2015
Accepted 6 December 2015
Available online 27 May 2016
Keywords:
Biodiversity
Climate change
Community structure
Conservation
Ecosystem function
Ecosystem services
Experiments
Gradients
a b s t r a c t
We briefly introduce and describe seven questions related to community structure and biodiversity
conservation that can be addressed using field experiments, and provide the context for using the vast
geographic diversity, biodiversity, and network of Nature Reserves in China to perform these experi-
ments. China is the worlds third largest country, has a diverse topography, covers five climatic zones
from cold-temperate to tropical, has 18 vegetation biomes ranging from Arctic/alpine tundra and desert
to Tropical rain forest, and supports the richest biodiversity in the temperate northern hemisphere (>10%
of the world total). But this tremendous natural resource is under relentless assault that threatens to
destroy biodiversity and negatively impact the services ecosystems provide. In an attempt to prevent the
loss of biodiversity, China has established 2729 nature reserves which cover 14.84% of the nations area.
Unfortunately underfunding, mismanagement, illegal activities, invasive species and global climate
change threaten the effectiveness of these protected areas. Attention has focused on protecting species
and their habitats before degradation and loss of either species or habitats occur. Here we argue that we
must move beyond the simple protection of ecosystems, beyond their description, and by using exper-
iments, try to understand how ecosystems work. This new understanding will allow us to design con-
servation programs, perform restoration of damaged or degraded areas, and address resource
management concerns (e.g., agriculture, logging, mining, hunting) more effectively than with the current
approach of ad hoc reactions to ecological and environmental problems. We argue that improving our
understanding of nature can best be done using well designed, replicated, and typically manipulative
field experiments.
Copyright © 2016 Kunming Institute of Botany, Chinese Academy of Sciences. Publishing services by
Elsevier B.V. on behalf of KeAi Communications Co., Ltd. This is an open access article under the CC BY-
NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/).
1. Introduction
China is the worlds third largest country. It stretches 5,200 km
from east to west, and the mainland spans 35 degrees of latitude,
and covers five climatic zones: cold-temperate, temperate, warm-
temperate, subtropical and tropical. China has a great physical di-
versity. The east and south of the country consists of fertile low-
lands and foothills, the west and north of the country is dominated
by sunken basins (e.g. the Gobi and the Taklamakan), plateaus (e.g.
QinghaieTibetan Plateau), and high peaks. The country is dissected
by deep river valleys including the Yellow, Yangtze and Mekong
Rivers. This wide range of environmental conditions supports a vast
biodiversity that is under relentless assault and threat, from both
within and outside China. Without proper understanding of how
ecological processes act to maintain this diversity, the degradation
and loss of Chinas tremendous natural resource will have signifi-
cant and far reaching effects.
Biodiversity provides significant ecosystem goods and services
to humans (Costanza et al., 1997), and that biodiversity generates
considerable economic benefits for humans (Balmford et al., 2002;
He et al., 2005). It is these services and the resultant economic
benefits that play an important role in ensuring the well-being of
people (Xu et al., 2008). Biodiversity is one of Chinas tremendous
natural resources and it must not only be preserved, but it should
be promoted. We argue that a productive way to achieve this is to
move beyond simple descriptions of nature, and move towards a
more comprehensive understanding provided by scientific inquiry.
* Corresponding author. Botany Department, University of British Columbia,
Vancouver, BC, V6T 1Z4, Canada.
E-mail address: roy.turkington@botany.ubc.ca (R. Turkington).
Peer review under responsibility of Editorial Office of Plant Diversity.
Contents lists available at ScienceDirect
Plant Diversity
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http://dx.doi.org/10.1016/j.pld.2015.12.001
2468-2659/Copyright © 2016 Kunming Institute of Botany, Chinese Academy of Sciences. Publishing services by Elsevier B.V. on behalf of KeAi Communications Co., Ltd. This
is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/).
Plant Diversity 38 (2016) 2e9
We can identify patterns in nature (generalization) and even
attribute causal explanations for these patterns in the form pro-
cesses, mechanism, interactions and conditions. However, it is only
by testing these descriptions through controlled, repeatable ex-
periments that we are able to confirm our understanding of nature.
This knowledge is essential to help solve conservation, restoration,
and resourcemanagement problems currently facing China and the
world.
1.1. Chinas plant biodiversity
Chinas great physical diversity produces a range of environ-
mental (abiotic) conditions that support a vast plant biodiversity. Ni
(2001) describes 18 biomes and The Vegetation of China (Editorial
Committee for Vegetation of China, 1980) describes 28 vegetation
types (see Ni, 2001). These vegetation types range from various
types of forest (e.g. cold-temperate deciduous broadleaved, warm-
temperate evergreen coniferous, subtropical mixed decid-
uouseevergreen broadleaved, tropical rain forest) tropical scrub or
coppicewood, Tamarix scrub, montane scrub, various types of
grassland (e.g. temperate formegramineous meadow steppe,
temperate desert steppe), to temperate sandy desert, and alpine
desert. Within these vegetation types, China holds the richest
biodiversity in the temperate northern hemisphere, and China is
one of the most biodiverse countries in the world (Harkness, 1998).
The country has at least 4357 species of terrestrial vertebrates
(Ministry of Environmental Protection, 2015) (world total: ~40,000
named species) and ~32,067 (Wang et al., 2015) species of vascular
plants (world total: >368,000; Chapman, 2009). This means that
China has slightly <10% of theworlds total in terrestrial vertebrates
and vascular plants (Liu et al., 2003). Many of Chinas species are
unique. Perhaps half of Chinas species are endemic and include
many archaic and distinctive evolutionary lines such as Ginkgo
biloba (Ginkgoes), Wisteria sinensis (Chinese Wisteria), Pinus bun-
geana (Lacebark Pine), Magnolia denudata (Yulan Magnolia), Juni-
perus chinensis (Chinese Juniper), Cunninghamia lanceolata (China-
fir), Taxus cuspidata (Japanese Yew) and Metasequoia glyptos-
troboides (Dawn Redwood), the latter often considered the oldest
and rarest plant in the world.
1.2. Threats to Chinas biodiversity
Many protected areas are being degraded and destroyed (Liu
et al., 2001; Hockings, 2003; Quan et al., 2011) because of poor or
overly bureaucratic management. The loss of species, in both
species-poor (e.g., deserts) and species-rich ecosystems (e.g.,
tropical forests), can reduce ecosystem functions such as primary
productivity, nutrient cycling, decomposition rates, water filtration,
and species interactions such as mutualisms, facilitation, compe-
tition, and predation. Ecosystems provide services that support
people and societies, and services arising from basic ecosystem
functions include forage production for livestock, soil stabilization,
carbon storage, climate regulation, pollination of crops, and natural
pest control. Thus, the loss of biodiversity imposes a high economic
cost.
The current loss of biodiversity has been the most dramatic
change humans have induced on ecosystems in the past century,
and the current global extinction rate is up to 100
times faster
than pre-human levels (Liu et al., 2003; Barnosky et al., 2011;
Ceballos et al., 2015). There is a growing concern that this loss of
species will have important effects on ecosystem functioning
(Cardinale et al., 2012; Hooper et al., 2012), whereby species-poor
ecosystems may perform differently or less efficiently than the
species-rich systems fromwhich they are derived. As in many parts
of the world, Chinas biodiversity is facing a critical situation, as it is
suffering from an increase in the extent and intensity of human
activities. Illegal harvesting of plants, overgrazing, vegetation har-
vesting, soil erosion, agriculture, pollution and firewood collection
are some of the primary threats to biodiversity in this region.
Overgrazing and desertification have been identified as severe
problems (Convention on Biological Diversity, 2010). The great
physical and biological diversity of China also makes the country
especially vulnerable to the establishment of invasive species.
Potentially invasive species from most areas of the world may find
suitable habitat somewhere in China. The International Union for
Conservation of Nature (IUCN, 2002) has identified the worlds 100
worst invasive species. Fifty of them are found in China and 23 have
invaded Chinas natural reserves and threaten the local biological
diversity.
1.3. Climate change
Any efforts to preserve biodiversity or understand ecosystem
functionmust be considered in the context of human-caused global
climate change. The pollution of earths atmosphere with abnormal
amounts of carbon dioxide will have profound effects on China
(IPCC, 2014). Most descriptions of climate change focus on how it
will alter the abundance and distribution of abiotic resources in
ecosystems. Because of its great diversity these changes will not be
uniform across China. Since 1961, NW China has experienced a
0.7 C rise in temperature and 22%e33% increase in rainfall (IPCC,
2014). There has been an increase in short duration heat waves.
While the north has seen a decrease in extreme rains, the south and
west have experienced an increase. There has also been an increase
in the number of floods and drought (Zhai et al., 2005). Thus,
because the climate in China is changing, it prudence demands that
we understand how ecosystems work rather than simply
describing what we see.
1.4. Protecting biodiversity
Nature reserves receive protection because of their recognized
natural, ecological, and cultural values. Nature reserves are essen-
tial for biodiversity conservation by providing habitat and protec-
tion from hunting or fishing for both common and rare species.
Protection also helps maintain ecological processes that cannot
persist in intensely managed human dominated landscapes.
Nature Reserves can be used as a baseline against which to
compare other ecosystems where degrading and damaging human
activities have occured. China has established 2729 nature reserves
which cover 14.84% of the nations area (Zheng and Cao, 2014;
Ministry of Environmental Protection, 2015). China has set ambi-
tious goals of increasing the number of reserves. Nonetheless, the
nature reserve system faces serious challenges that often involve
under-funding, inefficient management, lack of education, size and
location of reserves, and illegal activities (Anonymous, 2000; Liu
et al., 2003; Quan et al., 2011; Wu et al., 2011). Many research sci-
entists have expressed concern about the effectiveness of nature
reserves.
2. After description
The high level of physical diversity in China coupled with the
enormous level of biological diversity provides an almost unpar-
allelled opportunity to conduct experiments on the processes that
maintain biodiversity and ecosystem function. Ironically, the ability
to do experiments is enhanced by the altered or degraded state of
many ecosystems. This in no way implies that current research is of
low quality nor does it questionmuch of the excellent experimental
research on biodiversity already being done in China; research such
R. Turkington, W.L. Harrower / Plant Diversity 38 (2016) 2e9 3
as by Jiang et al. (2003),Wang and Ba (2008), Gao et al. (2009), Yang
et al. (2013), Bruelheide et al. (2014), Zhao et al. (2013) in addition
to those cited throughout the paper and many others. However, the
protection of Chinas biodiversity is crucial to both China and the
world. Understanding the processes the give rise to and maintain
biodiversity is essential if we are to develop a comprehensive and
sustainable strategy for the protection of biodiversity. China has
made enormous efforts to protect its biodiversity resources and its
successes and failures have provided a platform for experimental
field research that can advance both knowledge and theory. One of
the most fundamental questions in ecology to which we still have
no definitive answer is a biodiversity question: How do plant
species, which can reach densities of 50m2 in grasslands and up to
500 tree species ha1 in forests, stably coexist?
Nature Reserves and other protected areas are chiefly designed
to protect habitat, and of course, sometimes individual species.
Sadly, we need much more than a line on a map, a fence on the
ground, or laws banning the transport or sale of species. Even with
adequate funding and expert management the existence of a pro-
tected area or persistence of an endangered species does not
guarantee the long-term survival of biodiversity. A problem with
Nature Reserves and other protected areas is the false hope in
believing that the enclosed species are protected. However, com-
munities and ecosystems are not static and continually undergo a
process of natural change which eventually may lead to the
exclusion of the species and habitat that the area was initially
designed to protect. In addition, it is rarely practical to protect all
species or important ecological processes (e.g., water and air
filtration) in protected areas because these processes dont respect
the boundaries of the reserve.
Ofmore critical importance is that we try to understand how the
ecosystems work, and if they are broken (i.e., degraded) take steps
to restore them. We argue that this can best be done using well
designed, replicated, and often manipulative field experiments
(Bruelheide et al., 2014), although gradients are useful tools as well.
Experiments are conducted for the purpose of answering specific
questions about nature. These questions are usually stated as hy-
potheses about howwe think nature works and thereby contain an
implied prediction. Confirmation of our predictions is a powerful
means to demonstrate the accuracy of our understanding of the
biological world. In principle, manipulative experiments are an
acceptable procedure, because their planning implies that we are
making some level of prediction of the outcome. Biodiversity
problems are recognized through observations made in the field,
and therefore such observations are vital to the science. But after
that observation or description has been made we need to seek
explanations, usually by doing experiments. A robust prediction
requires a clear statement of outcomes that must occur for the
prediction to be confirmed. Therefore, to carry out a satisfactory
experiment, one most have full knowledge of the conditions
existing before the experiment is begun- and this is precisely what
is provided by many of Chinas protected areas.
Recently Bruelheide et al. (2014) established a major biodiver-
sity experiment in subtropical Xingangshan, Jiangxi Province,
China. The mountainous region of Southwest China provides ideal
conditions for many of the experiments we will describe. It is a
biodiversity hotspot that stretches over 250,000 km2 of temperate
to alpine mountains between the eastern edge of the Tibetan
Plateau and the Central Chinese Plain. It has a complex topography,
ranging from <2000 m in some valley bottoms to 7556 m at the
summit of Mount Gongga and this results in a wide range of cli-
matic conditions. The mountain ridges are oriented in a generally
northesouth direction. Temperatures range from frost-free
throughout the year in southern parts of Yunnan to permanent
snow and ice on some of the high mountain peaks of Sichuan and
Yunnan. At higher altitudes in Yunnan, annual average rainfall can
be >1000 mm on SW slopes, while NW areas of the region rarely
receive 400 mm annually. This variability in climatic and topo-
graphic conditions results in a wide variety of vegetation types
across the hotspot, including broad-leaved and coniferous forests,
bamboo groves, scrub communities, savanna, meadow, prairie and
alpine scrub. There are of course many other locations scattered
throughout China that are perfect for doing biodiversity experi-
ments. These locations include areas such as the extensive scrub-
land of the high Qinghai Tibetan plateau, and the seemingly endless
expanse of the grasslands in Inner Mongolia.
3. Experiments
The number and range of ecological questions that could be
asked about biodiversity in China is vast and it is not our intent to
be comprehensive in our coverage. We seek to provoke some
thought and stimulate more research activates focused on under-
standing Chinas biodiversity (Duffy, 2009). Biodiversity experi-
ments currently provide some of the most exciting and timely
research options in ecology and addressing these questions in
China will provide valuable knowledge to both China and the
world. To promote and stimulate more experimental ecological
research in China we will highlight seven major questions and
hypotheses (also see Bruelheide et al., 2014 ‘Major questions
addressed by BEF experiments’). While many of the experimental
designs are quite simple in principle, in practice, manymajor design
decisions have to be made based on the type of vegetation,
topography, and availability of appropriate sites (Bruelheide et al.,
2014). Perhaps one of the most critical design decisions is the
size, or the scale, of the experiment. Multi-scale experimental
studies are rare, yet ecologists have long recognized the importance
of spatial scale in ecological processes and patterns (Sandel and
Smith, 2009). Many patterns have been shown to be, or have the
potential to be, scale-dependent (e.g. Crawley et al., 2005; Levine,
2000). The grasslands of Inner Mongolia provide vast areas of
relatively uniform conditions that would permit testing for scale-
dependency of responses to treatments, specifically animal (graz-
ing) density, productivity (fertilizer addition), and predator addi-
tions or removals (Zhang et al., 2013a,b).
3.1. How are natural communities structured?
More than 50 years ago, Hairston et al. (1960) asked a simple
question about the fundamental limiting processes in biological
communities, and whether they were more likely to come from
below (nutrient availability) or above (herbivores or carnivores).
Their question was: “why dont herbivores increase in numbers to
such levels as to deplete or devastate vegetation?” This question
spawned a number of additional questions and hypotheses with
fundamentally different implications on vegetation regulation,
such as: the “Exploitation Ecosystem Hypothesis (EEH)” (Fretwell,
1977, 1987; Oksanen et al., 1981, Oksanen and Oksanen, 2000),
“The World is Prickly and Tastes Bad” (Pimm, 1991), the “Green
Desert” (White, 1978; Moen et al., 1993), “The World is White,
Yellow, and Green” (Oksanen et al., 1981; Pimm,1991), “The Cruddy
Ingredient Hypothesis” (Hartley and Jones, 1997), and “Large Parts
Of The World Are Brown Or Black” (Bond, 2005). These hypotheses
address whether ecosystems are structured from the top-down, or
from the bottom-up, with much debate and review (Polis, 1999;
Crawley, 1997; Sinclair et al., 2000; Turkington, 2009; Shurin,
2015). There are two main schools of thought and these have
been reviewed by Turkington (2009), Shurin (2015) and OConnor
et al. (2015).
R. Turkington, W.L. Harrower / Plant Diversity 38 (2016) 2e94
In the “bottom-up” hypothesis it is assumed that communities
are regulated by nutrient availability or environmental conditions
from below; higher trophic levels (in our case herbivores) have no
regulating or limiting effect on productivity or biomass on the
levels below them. In the “top-down” hypothesis, top predators are
self-regulating and each level then controls the level below. In the
top-down case, plants are not limited by nutrient availability, but
rather, by herbivores. Many other models involve variations of the
top-down and/or bottom-up hypotheses and there is a growing
recognition that both process operate simultaneously in most
communities.
Understanding whether communities are structured from the
top-down or the bottom-up is not only a theoretical debate, but has
management and conservation implications, especially in National
Parks, Nature Reserves, and other protected areas. The relative ef-
fects of these forces is becoming increasingly important as humans
alter ecosystems by overharvesting of consumers, increasing nu-
trients especially nitrogen over large areas, and an altering the
distribution of primary producers by clear cutting of forests, and
promoting desertification. These anthropogenic influences are
changing the strengths of top-down and bottom-up forces in eco-
systems. Therefore, identifying the relative strengths of these forces
and how they differ across different types of ecosystems is
increasingly important for understanding how communities are
structured, and especially how this structure is maintained or
changed across different spatial (Sandel and Smith, 2009) and
temporal scales. The EEH (Fretwell, 1977, 1987; Oksanen et al., 1981,
Oksanen and Oksanen, 2000) makes different predictions for eco-
systems having different productivity levels. Turkington (2009)
develops three hypotheses leading to 13 testable predictions. This
same approach can be extended to ask questions about the role of
consumers in ecosystems (see question #7).
The experimental testing of these hypotheses is quite simple in
principle, although forested systems may not be amenable to these
tests. The experiments would involve the addition (or reduction) of
fertilizer, water, or some other potentially limiting resource com-
bined with a reduction of herbivore or predator density, or their
exclusion. The interacting effects of these treatments with other
treatments such as warming (using open-top chambers or heaters),
water availability (added, or reduced using rain-out shelters) and
clipping (simulated herbivory) can be done quite easily, and can be
repeated at various points along natural productivity gradients.
Potential difficulties with these experiments are in selecting spe-
cies pools, identifying controls, deciding on spatial and temporal
scales, etc.
3.2. How does biodiversity determine the function of ecosystems?
Biodiversity is now the primary focus of both scientists and
natural resource agencies because of its potential importance for
the adequate functioning of the earths ecosystems (Isbell, 2014) so
that they can continue to provide services to humans. Ecosystem
services are benefits received by humans such as food production,
pollination, clean water, flood control, carbon sequestration, soil
formation and preservation, among other things. Consequently, the
preservation of ecosystem function is essential to humans.
Theories about the relationships between biodiversity and
ecosystem processes have a long history in theoretical ecology (e.g.
Elton, 1958; MacArthur, 1972; Pimm, 1984). Early experimental and
theoretical studies focused on building communities and observing
changes in function (references in Tilman, 1996; Naeem and Li,
1997; Lawton et al., 1998). These early experiments identified
some fundamental issues, such as, which “ecosystem function” is
the relevant indicator of ecosystem change (e.g. plant growth, soil
fertility and decomposition, or nutrient cycling etc.) or whether we
need to look at ecosystem properties such as “stability” (i.e.,
resistance to change, or resilience following disturbance) to un-
derstand the maintenance of biodiversity. A novel approach to this
question is to start with whole ecosystems having numerous plant
species and observe what happens to various ecosystem functions
when biodiversity changes as a result of manipulative disturbances
that resemble human impacts. Simplification of ecosystems
through experimentation, such as removing individual native
species or groups of species, adding of nitrogenous fertilizer, or
subjecting areas to a range of livestock grazing intensities is
perhaps the only way to know how declines in species number, loss
of functional groups, or absence of whole trophic levels will impact
ecosystem functions. Likewise, approaches that simulate nitrifica-
tion or grazing in controlled replicated experiments are the best
way to understand how humans impact ecosystem function. A
better understanding of how human activities impact essential
ecosystem functions is critical to design either conservation areas
or industrial and agricultural practices that preserve the services
provided by ecosystems.
Innovation, in either conservation or industrial sectors, comes
from examining real, complex systems that have been systemati-
cally and deliberately simplified and comparing them with parallel
unaltered (natural) systems. This approach allows us to identify
which ecosystems functions differ between species-rich and
species-poor systems. The results will also indicate which functions
are most sensitive to loss of species, the stability consequences that
result from loss of species, and the approximate time scale that may
be involved. In practice, variables such as primary productivity,
decomposition, nutrient cycling, soil development, and soil fertility
have, by default, come to represent function and the impact of
species loss, nutrient deposition, and land use change on these and
other functions should be examined.
We recognize three hypotheses of the possible relationship
between biodiversity and ecosystem function (Vitousek and
Hooper, 1993). First, is the rivet hypothesis in which each addi-
tional species has a constant (i.e. all species considered to have
equal effects) additional effect on function (Ehrlich and Wilson,
1991). This hypothesis implies that all species play some role in
maintaining ecosystem function. Progressive loss of species,
therefore, results in a loss of some function and at some point there
is a catastrophic change in function. The second hypothesis is called
the keystone hypothesis (Lawton and Brown, 1993) and this is
another version of the rivet hypothesis. This hypothesis is asymp-
totic (i.e. species are considered unequal in their effects) so that
each species that is added has a decreasing impact on function. It
suggests that some species are keystones, dominants or foundation
species and are essential to the structure of the community. The
loss of these disproportionately important species results in rapid
and catastrophic change in ecosystem function, whereas the loss of
other species can have little or no effect on ecosystem function. The
third hypothesis, the redundancy hypothesis (Walker, 1992) sug-
gests there should be no impact of biodiversity loss on ecosystem
function, until almost all species are lost. This hypothesis proposes
that species have overlapping roles in their communities so that the
loss of a species is compensated by an increase in activity or density
of the remaining species. Only after considerable species loss occurs
do compensatory processes break down and function declines.
Since each hypothesis makes a different set of predictions in terms
of function an experimental perturbation could test which hy-
pothesis is the most relevant.
There is a growing understanding that the effects of biodiversity
loss require recognition of the non-random nature of species loss
(Wardle et al., 2011; Estes et al., 2011). But, it is important to note
that “species” may not be the best unit by which to discuss
ecosystem function, and the classification of organisms into
R. Turkington, W.L. Harrower / Plant Diversity 38 (2016) 2e9 5
functional groups (e.g. nitrogen fixers, perennial grasses, forbs,
trees), guilds, or even whole trophic levels (e.g., top predators,
herbivores) may be more relevant ecologically. It may be that the
identity of a species, or group of species, may play an even larger
role than that of the diversity of the species themselves. Despite
this realization, virtually no studies specifically examine the inde-
pendent effects of species composition and species diversity on the
functioning of ecosystem processes such as productivity, decom-
position rates, and nutrient cycling. In addition, different measures
of ecosystem functionmay fit different hypotheses (see question #6
below).
3.3. How does the loss of biodiversity alter the stability of
ecosystems?
A potential problem with National Parks, Nature Reserves and
other protected areas is the false hope in believing that the
enclosed species are protected. However, communities and eco-
systems are not static; they undergo a process referred to as the
exploitation, conservation, release and reorganization cycle
(Holling, 1986). This process is not continuous and in the mature
successional stages most of the nutrients and energy are locked-up
in biomass, and the system gradually becomes more vulnerable to
disturbances. Until recently, most disturbances were natural and
included fire, wind storms, pest outbreaks, etc., but now human
activities such as excesses of harvesting, exploitation, grazing, and
exploitation are more common, intense and operate over larger
spatial scales. The intensification of disturbance can lead to pro-
gressively faster processes of release and reorganization (Holling,
1986).
An important question here is whether the stability of a com-
munity (defined by variance in some factor such as biomass) is
related to biodiversity. This is not a new debate (e.g. Elton, 1958),
but recent work suggests statistical averaging of variances across
species inevitably leads to a joint community variance that declines
with increasing biodiversity. This is the “portfolio effect” (Doak
et al., 1998). Therefore the question: How does the loss of biodi-
versity alter the stability of ecosystems? May be reworded as: Does
decreasing biodiversity lead to more variability in community
biomass (i.e. less stability)? In effect, we can predict that a greater
number of species, functional groups, feeding guilds, or trophic
levels will lead to greater stability if: (i) variance of individual
species biomasses increases more than linearly with mean indi-
vidual biomass (portfolio hypothesis); or (ii) there is compensation
between species through interspecific competition (competition
hypothesis). Monitoring community biomass in perturbed and
intact systems over time can test these predictions (see question #6
below).
3.4. How does the loss of biodiversity alter the integrity of
ecosystems?
Ecosystems are complex systems and thus have a number of
important properties (May, 1977; Kay, 1991). These properties
include: (1) non-linearity in that the whole system behaves as a
unit more than the sum of its constituents; (2) multiple scales since
local and regional processes differ but affect each other; (3) mul-
tiple states, the combination of species, and population sizes, can
switch as a consequence of disturbance; (4) dynamic stability in the
sense that systems fluctuate around potential stable points that are
rarely observed; (5) resilience in that within a range of environ-
mental perturbations they maintain their characteristic species
assemblages and rate of return from a given intensity of distur-
bance; (6) fragility as measured by the limits that a system can
tolerate perturbations (resistance); (7) catastrophic behaviour
meaning that they exhibit sudden change if disturbances push the
system beyond the limits that determine the systems fragility
(Regier, 1992). These properties emerge because ecosystems are
thermodynamically open systems that are never in equilibrium
(Kay and Schneider, 1992). Therefore our question: How does the
loss of biodiversity alter the integrity of ecosystems? May be
reworded as: What features of ecosystem integrity change as
biodiversity changes? Some of the measurable factors are those
listed above. Again, a comparison of unaltered natural systemswith
systems showing varying degrees of degradation or simplification
can provide an adequate field situation to test these hypotheses
(see question #6 below).
3.5. Diversity and species composition
Many studies have concluded that in addition to the role that
diversity plays in controlling ecosystem function, stability and
integrity, the identity or type of speciesmay play an even larger role
than that of diversity (number of species) itself. There are almost no
studies that specifically examine the independent effects of species
composition and species diversity on the functioning of ecosystem
processes such as productivity, decomposition rates and nutrient
cycling. However, there is good evidence that functional diversity,
especially in plants, is an important predictor of ecosystems func-
tion (Díaz and Cabido, 2001; Díaz et al., 2003).
Climate change, one of the drivers of biodiversity loss, is likely to
have different effects on different functional groups, resulting in
the loss of some functional groups before others. For example, the
loss of predators can have pervasive effects on lower trophic levels
(Schmitz et al., 2010) and have even been shown to indirectly in-
crease carbon sequestration in plants (Strickland et al., 2013).
Likewise, herbivores can play a major functional role in ecosystems
altering nutrient cycles in ways that change with abiotic conditions
(Metcalfe et al., 2014). Different functional groups can have
compensatory effects, as trophic processes such as herbivory
compensating for the negative effects of nitrification on diversity
and subsequently function (Borer et al., 2014). By understanding
the role that each group (functional groups, feeding guilds, or tro-
phic level) plays in determining how the entire ecosystem responds
to this change, we can predict with greater accuracy the overall
effects of climate-changed induced species loss.
3.6. How does the loss of species determine the ability of
ecosystems to respond to disturbances?
The frequency and intensity of fires, storms, drought etc. will
undoubtedly change with current global warming trends and with
increased human pressure encroaching ecosystems. A likely
outcome of these changing disturbances for both flora and fauna is
a loss of species or changes in species composition. The objective
here is to test whether the loss of some of the plant species will
result in significant changes in ecosystem processes and whether
the remaining species will be able to compensate. This question is a
natural follow up to some of the previous questions and is a com-
parison of degraded ecosystems. Two or more ecosystems could be
degraded to an “equal” degree in that both may have lost the same
percentage of their biomass, species richness, or functional di-
versity of their undegraded natural state. However, those losses
may have come about in the degraded ecosystems by losses of
different species or species functional groups. In other words, one
system may have lost many graminoids, another may have lost
mostly forbs, and another may have lost representatives of both
groups. This question now investigates the role of plant functional
groups by asking how these different degraded ecosystems respond
to imposed stresses or disturbance.
R. Turkington, W.L. Harrower / Plant Diversity 38 (2016) 2e96
The essence of this experiment is to remove different plant
species or functional groups, then compare the stability, produc-
tivity and other functional properties of this degraded system un-
der experimental stresses (e.g. fire, nutrient addition, drought),
with systems that have their full biota. Recent reviews (Hooper
et al., 2005, Hooper and Vitousek, 2011; Spehn et al., 2005) and
meta-analyses (Balvanera et al., 2006; Cardinale et al., 2006) pro-
vide clear evidence that biodiversity has significant effects on
ecosystem function, and research is shifting to exploring the un-
derlying mechanisms (Díaz et al., 2003). If the traits that determine
extinction risk are correlated with the traits that control ecosystem
function, the order in which species are lost may have critical ef-
fects on ecosystem function and services. Therefore, we need to
predict the effect that the loss of particular species or groups of
species will have on ecosystem function.
This question is of particular interest in China. The Hohhot
Declaration (Peart, 2008) stated that “temperate indigenous
grasslands are critically endangered and urgent action is required
to protect and maintain the service they provide to sustain human
life.” This is especially critical because drought and over-grazing are
threatening much of the worlds grasslands, particularly those in
China. In addition, anthropogenic disturbances, including climate
change (most models predict that temperate grasslands will
experience increased temperatures, drier summers, and wetter
winters), are expected to cause an increase in nitrification and ni-
trogen deposition, increased levels of CO2, produce longer growing
seasons, and potentially cause many species extinctions, all of
which have the potential to drastically alter grassland commu-
nities. However, a theoretical basis for decision making on con-
servation andmanagement policy in grasslands is often lacking and
many decisions are based on expediency, politics, and on a case-by-
case basis.
3.7. How does food web complexity and productivity influence the
relative strength of tropic interactions, and how do changes in
trophic structure influence ecosystem function?
This question extends question #1 by asking if predators can
initiate a trophic cascade whereby predator effects propagate
downward through the food chain, crossing a number of trophic
levels and ultimately influencing the biodiversity of the plant or soil
levels. The effect of predators may cascade through ecosystems to
impact disturbance regimes, the prevalence of diseases, soil
nutrient cycling, water quality, species invasions, and biodiversity
(Estes et al., 2011). The recognition that predators can have
important effects on species, communities, and ecosystems even in
systemswith strong resource-control is increasing. Though a classic
trophic cascade is not the only way predators affect ecosystems,
food webs, or biodiversity, it is a clear outcome demonstrating the
role predators can play in ecosystems. We have good examples of
ecosystem-level (energy, nutrients), community-level (biomass),
and species-level (diversity) cascades in many ecosystems (Dyer
and Letourneau, 2002, 2013) and our understanding of
ecosystem-level cascades is continuing to grow (Schmitz et al.,
2000; Strickland et al., 2013).
Consumers can exert indirect control on ecosystems through a
variety of mechanisms, resulting in very different impacts. The
classic example of top-down control is how the removal or addition
of predators can change the distribution of biomass between tro-
phic levels (community-level cascade). This occurs through either
changing herbivore abundance or behaviour. Second, the addition,
removal, or alteration of the predator community composition can
change the composition, diversity or behaviour of herbivores
resulting in changes in the biomass, biodiversity, or composition of
the producer community (species-level cascade). A third way that
consumers can exert control is through changing the function of
ecosystems. There is growing recognition that predators can exert
control on ecosystems processes such as decomposition rates,
nutrient flux and storage (including carbon), primary productivity,
and water cycling without obvious visual changes to the herbivore
community (Strickland et al., 2013).
We propose that by adding or removing predators from an
ecosystem we will see changes in the productivity, diversity, and
function of both herbivores and primary producers. In addition, we
propose that human-caused changes to the distribution and
abundance of resources can change the strength of consumer-
control in any particular ecosystem and possibly elicit unwanted
trophic cascades. We suggest that it may be possible to bring about
a trophic cascade not simply through the addition or removal of
predators in a system but by changing the amount, distribution or
quality of resources in an ecosystem. Likewise, changing the
number of species that occupy lower trophic levels (through in-
vasions, extinctions etc.) could alter the transfer of energy through
the food web making ecosystems either more or less susceptible to
inherit instability of trophic cascades. Understanding that trophic
structure can be altered in indirect ways, ways other than direct
perturbations, harvesting, or exotic invasions, is important when
we think about conserving and restoring ecosystems.
4. Conclusions
We have briefly introduced and described seven questions that
could be addressed using the vast geographic diversity, biodiver-
sity, and network of Nature Reserves and other protected sites in
China. The aim of the account is introductory and thus it is not a
comprehensive list or literature review. Our goal is to stimulate
experimental research into how ecosystems function and how to
preserve and promote biodiversity. For example, each of the seven
questions we outlined can be asked in at least five different
contexts:
(1) How do the observed responses change across the 28 vege-
tation types in China identified by the Editorial Committee
for Vegetation of China (1980)?
(2) How do the observed responses change from the low pro-
ductivity grasslands of the Qinghai Plateau to higher pro-
ductivity grasslands in other parts of China, or temperate
steppe versus Alpine steppe, or along a transect across any of
the major biomes?
(3) How do the observed responses change along a gradient in
the intensity of human-use or degradation.
(4) How long should an experiment be conducted given that the
immediate results are seldom indicative of longer-term
outcomes?
(5) How does the scale of the experiment influence treatment
responses?
China is one of the worlds biodiversity hotspots, and not unlike
many parts of the world much attention has been focused on
describing, listing, legislating and protecting habitat and species in
various types of protected areas including Botanical Gardens, Na-
tional Parks, and Nature Reserves. While this is both necessary and
commendable we have tried to make the case that understanding
how these habitats and ecosystems work, how they function and
how they respond to both natural and anthropomorphic pertur-
bations, is the next important level of investigation. Simply con-
taining ecosystems in preserves will not help protect them. We
must understand how ecosystems function, first to know what
components are essential to preserve, second to ensure that eco-
systems continue to function and thus provide humans services.
R. Turkington, W.L. Harrower / Plant Diversity 38 (2016) 2e9 7
Chinas diverse physical geography, its 18 biomes and 34 vegetation
types, vast biodiversity and network of protected areas provides an
ideal opportunity to test many of these vital questions because
these areas can provide the contrasts necessary to examine how
human impacts influence ecosystems.
Acknowledgements
The writing of this paper was funded by the Natural Sciences
and Engineering Research Council of Canada. The Xishuangbanna
Tropical Botanical Garden provided funding to RT as an invited
scholar. The stimulus to write this paper came through a casual
conversation with Professor Zhou Zhe-kun. We also thank two
anonymous referees for valuable comments on the manuscript.
References
Anonymous, 2000. Chinas National Committee on Man and Biosphere, Study of
Sustainable Management Policies for Chinas Nature Reserves (in Chinese).
Scientific and Technical Documents Publishing House, Beijing, China.
Balmford, A., Bruner, A., Cooper, P., et al., 2002. Economic reasons for conserving
wild nature. Science 297 (5583), 950e953. http://doi.org/10.1126/science.
1073947.
Balvanera, P., Pfisterer, A.B., Buchmann, N., et al., 2006. Quantifying the evidence for
biodiversity effects on ecosystem functioning and services. Ecol. Lett. 9,
1146e1156. http://dx.doi.org/10.1111/j.1461-0248.2006.00963.x.
Barnosky, A.D., Matzke, N., Tomiya, S., et al., 2011. Has the Earths sixth mass
extinction already arrived? Nature 471, 51e57.
Borer, E.T., Seabloom, E.W., Gruner, D.S., et al., 2014. Herbivores and nutrients
control grassland plant diversity via light limitation. Nature 508, 517e520.
http://doi.org/10.1038/nature13144.
Bond, W.J., 2005. Large parts of the world are brown or black: a different view on
the ‘Green World’ hypothesis. J. Veg. Sci. 16, 261e266.
Bruelheide, H., Nadrowski, K., Assmann, T., et al., 2014. Designing forest biodiversity
experiments: general considerations illustrated by a new large experiment in
subtropical China. Meth. Ecol. Evol. 5, 74e89. http://dx.doi.org/10.1111/2041-
210X.12126.
Cardinale, B.J., Duffy, J.E., Gonzalez, A., et al., 2012. Biodiversity loss and its impact
on humanity. Nature 486, 59e67.
Cardinale, B.J., Srivastava, D.S., Duffy, J.E., et al., 2006. Effects of biodiversity on the
functioning of trophic groups and ecosystems. Nature 443, 989e992.
Ceballos, G., Ehrlich, P.R., Barnosky, A.D., et al., 2015. Accelerated modern human-
induced species losses: Entering the sixth mass extinction. Sci. Adv. 1 (5)
http://dx.doi.org/10.1126/sciadv.1400253 e1400253.
Chapman, A.D., 2009. Numbers of Living Species in Australia and the World, second
ed. Australian Biological Resources Study, Canberral, pp. 1e80.
Convention on Biological Diversity, 2010. .https://www.cbd.int/countries/profile/
default.shtml?country¼cn (accessed 1.12.15.).
Costanza, R., Arge, R., De Groot, R., et al., 1997. The value of the worlds ecosystem
services and natural capital. Nature 387, 253e260. http://doi.org/10.1038/
387253a0.
Crawley, M.J., 1997. Plant-herbivore dynamics. In: Crawley, M.J. (Ed.), Plant Ecology,
second ed. Blackwell, Oxford, pp. 401e474. http://dx.doi.org/10.1080/
10643389.2010.502645.
Crawley, M.J., Johnston, a E., Silvertown, J., et al., 2005. Determinants of species
richness in the park grass experiment. Am. Nat. 165 (2), 179e192. http://doi.org/
10.1086/427270.
Díaz, S., Cabido, M., 2001. Vive la diff
erence: plant functional diversity matters to
ecosystem processes. Trends Ecol. Evol. 16 (11), 646e655. http://doi.org/10.
1016/S0169-5347(01)02283-2.
Díaz, S., Symstad, A.J., Chapin III, F.S., et al., 2003. Functional diversity revealed by
removal experiments. Trends Ecol. Evol. 18, 140e146.
Doak, D.F., Bigger, D., Harding, E.K., et al., 1998. The statistical inevitability of
stability-diversity relationships in community ecology. Amer. Natur. 51 (3),
264e276.
Duffy, J.E., 2009. Why biodiversity is important to the functioning of real-world
ecosystems. Front. Ecol. Environ. 7, 437e444. http://dx.doi.org/10.1890/070195.
Dyer, L.A., Letourneau, D., 2002. Top-down and bottom-up diversity cascades in
detrital vs. living food webs. Ecol. Lett. 6, 60e68.
Dyer, L., Letourneau, D., 2013. Can climate change trigger massive diversity cascades
in terrestrial ecosystems? Diversity 5, 479e504.
Editorial Committee for Vegetation of China, 1980. Vegetation of China. Science
Press, Beijing.
Elton, C.S., 1958. The Ecology of Invasions by Animals and Plants. University of
Chicago Press, Chicago IL.
Ehrlich, P.R., Wilson, E.O., 1991. Biodiversity studies: science and policy. Science 253,
758e762.
Estes, J.A., Terborgh, J., Brashares, J.S., et al., 2011. Trophic downgrading of planet
Earth. Science 333, 301e306.
Fretwell, S.D., 1977. The regulation of plant communities by food chains exploiting
them. Perspect. Biol. Med. 20, 169e185.
Fretwell, S.D., 1987. Food chain dynamics: the central theory of ecology? Oikos 50,
291e301.
Gao, Q.Z., Li, Y., Wan, Y.F., et al., 2009. Significant achievements in protection and
restoration of alpine grassland ecosystem in Northern Tibet, China. Restor. Ecol.
17, 320e323. http://dx.doi.org/10.1111/j.1526-100X.2009.00527.x.
Hairston, N.G., Smith, F.E., Slobodkin, L.B., 1960. Community structure, population
control, and competition. Am. Nat. 421e425, 94.879.
Harkness, J., 1998. Recent trends in forestry and conservation of biodiversity in
China. China Quart. 156, 911e934. http://dx.doi.org/10.1017/
S0305741000051390.
Hartley, S.E., Jones, C.G., 1997. Plant chemistry and herbivory, or why the world is
green. In: Crawley, M.J. (Ed.), Plant Ecology, second ed. Blackwell, Oxford,
pp. 284e324.
Hockings, M., 2003. Systems for assessing the effectiveness of management in
protected areas. Bioscience 53, 823e832.
Holling, C.S., 1986. The resilience of terrestrial ecosystems: local surprise and global
change. In: Clark, W.C., Munn, R.E. (Eds.), Sustainable Development of the
Biosphere. Cambridge Univ. Press, Cambridge, pp. 297e317.
Hooper, D.U., Chapin, F.S., Ewel, J.J., et al., 2005. Effects of biodiversity on ecosystem
functioning: a consensus of current knowledge. Ecol. Monogr. 75, 3e35.
Hooper, D.U., Vitousek, P.M., 2011. The effects of plant composition and diversity on
ecosystem processes. 277(5330), 1302e1305.
Hooper, D.U., Adair, C.E., Cardinale, B.J., et al., 2012. A global synthesis reveals
biodiversity loss as a major driver of ecosystem change. Nature 486, 105e108.
He, H., Pan, Y., Zhu, W., et al., 2005. Measurement of terrestrial ecosystem service
value in China. Chin. J. Appl. Ecol. 16, 1122e1127.
IPCC, 2014. In: Core Writing Team, Pachauri, R.K., Meyer, L.A. (Eds.), Climate Change
2014: Synthesis Report. Contribution of Working Groups I, II and III to the Fifth
Assessment Report of the Intergovernmental Panel on Climate Change. IPCC,
Geneva, Switzerland, 151.
Isbell, F., 2014. Ecosystem services. In: Gibson, D. (Ed.), Oxford Bibliographies in
Ecology. Oxford University Press, NY. http://dx.doi.org/10.1093/OBO/
9780199830060-0052.
IUCN, 2002. www.iucn.org/themes/ssc/redlist2002/rl_news.htm.
Jiang, Y., Kang, M., Gao, Q., et al., 2003. Impact of land use on plant biodiversity and
measures for biodiversity conservation in the loess plateau in Chinaea case
study in a hilly-gully region of the Northern Loess Plateau. Biodivers. Conserv.
12, 1212e2133.
Kay, J.J., 1991. A nonequilibrium thermodynamic framework for discussing
ecosystem theory. Environ. Manag. 15, 483e495.
Kay, J.J., Schneider, E.D., 1992. Thermodynamics and measures of ecosystem integ-
rity. In: McKenzie, D.H., Hyatt, D.E., Mc Donald, V.J. (Eds.), Ecological Indicators,
Proceedings of the International Symposium on Ecological Indicators, vol. 1.
Elsevier, Fort Lauderdale, Florida, pp. 159e182.
Lawton, J.H., Naeem, S., Thompson, L.J., et al., 1998. Biodiversity and ecosystem
function: getting the Ecotron experiment in its correct context. Funct. Ecol. 12,
843e856.
Lawton, J.H., Brown, V.K., 1993. Redundancy in ecosystems. In: Schulze, E.D.,
Mooney, H.A. (Eds.), Biodiversity and Ecosystem Function. Springer, Berlin,
pp. 255e270.
Levine, J.M., 2000. Species diversity and biological invasions. Science 288, 852e854.
Liu, J., Linderman, M., Ouyang, Z., et al., 2001. Ecological degradation in protected
areas: the case of Wolong nature reserve for giant pandas. Science 292, 98e101.
Liu, J., Ouyang, Z., Pimm, S.L., et al., 2003. Protecting Chinas biodiversity. Science
300 (5623), 1240e1241.
MacArthur, R.H., 1972. Geographical Ecology. Harper and Rowe, New York.
May, R.M., 1977. Thresholds and breakpoints in ecosystem with multiplicity of
stable states. Nature 269, 471e477.
Metcalfe, D.B., Asner, G.P., Martin, R.E., et al., 2014. Herbivory makes major contri-
butions to ecosystem carbon and nutrient cycling in tropical forests. Ecol. Lett.
17 (3), 324e332. http://doi.org/10.1111/ele.12233.
Ministry of Environmental Protection, 2015. The 2014 Report on the State of the
Environment in China. http://www.gzhjbh.gov.cn/images/dtyw/tt/gndttt/2014/
6/5/7e4df35f-d266-429e-8269-6eba756c24f8.pdf (in Chinese, last accessed
30.11.15.).
Moen, J., Gardefjell, H., Oksanen, L., et al., 1993. Grazing by food-limited microtine
rodents on a productive experimental plant community: does the green desert
exist? Oikos 68, 401e413.
Naeem, S., Li, S., 1997. Biodiversity enhances ecosystem reliability. Nature 390,
507e507.
Ni, J., 2001. A biome classification of China based on plant functional types and the
BIOME3 model. Folia Geobot. 36, 113e129.
OConnor, M.I., Bernhardt, J.R., Caulk, N.C., 2015. Food webs. In: Gibson, D. (Ed.),
Oxford Bibliographies in Ecology. Oxford University Press, NY. http://dx.doi.org/
10.1093/OBO/9780199830060-0138.
Oksanen, L., Oksanen, T., 2000. The logic and realism of the hypothesis of exploi-
tation ecosystems. Am. Nat. 155, 703e723.
Oksanen, L., Fretwell, S.D., Arruda, J., et al., 1981. Exploitation systems in gradients of
primary productivity. Am. Nat. 118, 240e261.
Peart, B., 2008. Life in a working landscape: towards a conservation strategy for the
Worlds temperate grasslands. In: Report of the XXI International Grasslands
Congress/VIII International Rangeland Congress. Hohhot, Inner Mongolia, China.
Pimm, S.L., 1984. The complexity and stability of ecosystems. Nature 307, 321e326.
R. Turkington, W.L. Harrower / Plant Diversity 38 (2016) 2e98
Pimm, S.L., 1991. The Balance of Nature: Ecological Issues in the Conservation of
Species and Communities. The University of Chicago Press, Chicago.
Polis, G.A., 1999. Why are parts of the world green? Multiple factors control pro-
ductivity and the distribution of biomass. Oikos 86, 3e15.
Quan, J., Ouyang, Z., Xu, W., et al., 2011. Assessment of the effectiveness of nature
reserve management in China. Biodivers. Conserv. 20, 779e792.
Regier, H.A., 1992. Ecosystem integrity in the Great Lakes Basin: an historical sketch
of ideas and actions. J. Aquat. Ecosyst. Health 1, 25e37.
Sandel, B., Smith, A.B., 2009. Scale as a lurking factor: incorporating scale-
dependence in experimental ecology. Oikos 118, 1284e1291.
Schmitz, O.J., Hamb€ack, P.A., Beckerman, A.P., et al., 2000. Trophic cascades in
terrestrial systems: a review of the effects of carnivore removal on plants. Am.
Nat. 155, 141e153.
Schmitz, O.J., Hawlena, D., Trussell, G.C., 2010. Predator control of ecosystem
nutrient dynamics. Ecol. Lett. 13, 1199e1209.
Shurin, J.R., 2015. Top-down and bottom-up regulation of communities. In:
Gibson, D. (Ed.), Oxford Bibliographies in Ecology. Oxford University Press, NY.
http://dx.doi.org/10.1093/OBO/9780199830060-0029.
Sinclair, A.R.E., Krebs, C.J., Fryxell, J.M., et al., 2000. Testing hypotheses of trophic
level interactions: a boreal forest ecosystem. Oikos 89, 313e328.
Spehn, A.E.M., Hector, a, Joshi, J., Schmid, B., et al., 2005. Ecosystem effects of
biodiversity manipulations in European grasslands. Ecol. Monogr. 75 (1), 37e63.
Strickland, M.S., Hawlena, D., Reese, A., et al., 2013. Trophic cascade alters
ecosystem carbon exchange. PNAS U. S. A. 110, 11035e11038. http://doi.org/10.
1073/pnas.1305191110.
Tilman, D., 1996. Biodiversity: population versus ecosystem stability. Ecology 77,
350e363.
Turkington, R., 2009. Top-down and bottom-up forces in mammalian herbivore-
vegetation systems: an essay review. Botany 87 (8), 723e739. http://doi.org/
10.1139/B09-035.
Vitousek, P.M., Hooper, D.U., 1993. Biological diversity and terrestrial ecosystem
biogeochemistry. In: Schulze, E.D., Mooney, H.A. (Eds.), Biodiversity and
Ecosystem Function. Springer-Verlag, Berlin, pp. 3e14.
Walker, B., 1992. Biodiversity and ecosystem redundancy. Conserv. Biol. 6, 18e23.
Wang, D., Ba, L., 2008. Ecology of meadow steppe in northeast China. Rangel. J. 30
(2), 247e254. http://dx.doi.org/10.1071/RJ08005.
Wang, L., Jia, Y., Zhang, X., et al., 2015. Overview of higher plant diversity in China.
Biodivers. Sci. 23, 217e224.
Wardle, D.A., Bardgett, R.D., Callaway, R.M., et al., 2011. Terrestrial ecosystem re-
sponses to species gains and losses. Science 332 (6035), 1273e1277. http://
dx.doi.org/10.1126/science.1197479.
White, T.C.R., 1978. The importance of relative shortage of food in animal ecology.
Oecologia 33, 71e86.
Wu, R., Zhang, S., Yu, D.W., et al., 2011. Effectiveness of Chinas nature reserves in
representing ecological diversity. Front. Ecol. Environ. 9 (7), 383e389. http://
doi.org/10.1890/100093.
Xu, H.G., Wu, J., Liu, Y., et al., 2008. Biodiversity congruence and conservation
strategies: a national test. BioScience 58, 632e639.
Yang, X.F., Bauhus, J., Both, S., et al., 2013. Establishment success in a forest biodi-
versity and ecosystem functioning experiment in subtropical China (BEF-
China). Eur. J. For. Res. 132, 593e606.
Zhai, P., Zhang, X., Wan, H., et al., 2005. Trends in total precipitation and frequency
of daily precipitation extremes over China. J. Clim. 18, 1096e1108.
Zhang, H., Gilbert, B., Wang, W., et al., 2013a. Grazer exclusion alters plant spatial
organization at multiple scales, increasing diversity. Ecol. Evol. 3, 3604e3612.
Zhang, H., Gilbert, B., Zhang, X., et al., 2013b. Community assembly along a suc-
cessional gradient in sub-alpine meadows of the Qinghai-Tibetan Plateau,
China. Oikos 122, 952e960.
Zhao, M., Gao, X., Wang, J., et al., 2013. A review of the most economically important
poisonous plants to the livestock industry on temperate grasslands of China.
J. Appl. Toxicol. 33, 9e17. http://dx.doi.org/10.1002/jat.2789.
Zheng, H., Cao, S., 2014. Threats to Chinas biodiversity by contradictions policy.
Ambio 23e33. http://doi.org/10.1007/s13280-014-0526-7.
R. Turkington, W.L. Harrower / Plant Diversity 38 (2016) 2e9 9