Andrew Balmford1, *, Aaron Bruner2,
Philip Cooper3, Robert Costanza4, Stephen Farber5,
Rhys E. Green1, 6, Martin Jenkins7, Paul Jefferiss6,
Valma Jessamy3, Joah Madden1, Kat Munro1,
Norman Myers8, Shahid Naeem9, Jouni Paavola3,
Matthew Rayment6, Sergio Rosendo3, Joan Roughgarden10,
Kate Trumper1, and R. Kerry Turner3
1.
Conservation
Biology Group, Department of Zoology, University of Cambridge, Cambridge CB2
3EJ, UK
2.
Center
for Applied Biodiversity Science at Conservation International, 1919 M Street,
N.W., Suite 600, Washington DC
20036, USA
3.
CSERGE,
School of Environmental Sciences, University of East Anglia, Norwich NR4 7TJ,
UK
4.
Center
for Environmental Science, Biology Department and Institute for Ecological
Economics, University of Maryland, Box 38, Solomons MD 20688, USA;
after 9/02: Gund Institute
of Ecological Economics, The University of Vermont, Burlington, VT 05405 USA
5.
Graduate
School of Public and International Affairs, University of Pittsburgh, Pittsburgh
PA 15260, USA
6.
The
Royal Society for the Protection of Birds, The Lodge, Sandy, Beds SG19 2DL, UK
7.
UNEP-WCMC,
219 Huntingdon Road, Cambridge CB3 ODL, UK
8.
Green
College, Woodstock Road, Oxford OX2 6HG, UK; and Upper Meadow, Old Road,
Headington, Oxford OX3 8SZ, UK
9.
Department
of Zoology, University of Washington, 24 Kincaid Hall, Box 351800, Seattle WA
98195-1800, USA
10. Department of Biological
Sciences, Stanford University, Stanford CA 94305, USA
* To
whom correspondence should be addressed; email: a.balmford@zoo.cam.ac.uk
On the eve of the World Summit on Sustainable Development, it is timely to assess progress over the ten years since its predecessor in Rio de Janeiro. Loss and degradation of remaining natural habitats has continued largely unabated. However, evidence has been accumulating that such systems generate marked economic benefits, which the available data suggest exceed those obtained from continued habitat conversion. We estimate that the overall benefit: cost ratio of an effective global programme for the conservation of remaining wild nature is at least 100 : 1.
Humans benefit from wild nature (1) in very many ways Ð
aesthetically and culturally; via the provision of ecological services such as
climate regulation, soil formation and nutrient cycling; and from the direct
harvest of wild species for food, fuel, fibres and pharmaceuticals (2). In the face of increasing
human pressures on the environment these benefits should act as powerful
incentives to conserve nature, yet evaluating them has proved difficult because
they are mostly not captured by conventional, market-based economic activity
and analysis.
In 1997, Costanza et al. published a synthesis (3) of over 100 attempts to
value ecosystem goods and services using a range of techniques including
hedonic pricing, contingent valuation and replacement cost methods (4). Using case studies to
derive average values per hectare for each of 17 services across 16 biomes and
then extrapolating to the globe by multiplying by each biomeÕs area, the
Costanza team estimated the aggregated annual value of natureÕs services
(updated to 2000 US $) to lie in the range $18 Ð 61 trillion (1012),
around a rough average of ~$38 trillion. These figures are of similar size to
global Gross National Product (GNP), but have been criticised by some in the
economic community (5-9).
One problem is that such macroeconomic extrapolations are
inconsistent with microeconomic theory: extrapolation from the margin to a
global total should incorporate knowledge about the shape of the demand curve (3, 5-8). In practice, it is very
likely that per unit demand for non-substitutable services escalates rapidly as
supply diminishes, so that simple grossing-up of marginal values (as is also
done in calculating GNP from prices) will probably underestimate true total
values. On the other hand, high local values of services such as tourism may
not be maintained if extrapolated worldwide. In addition, while some policy
decisions are made using macroeconomic indicators, many others are made at the
margin, and so are more appropriately informed by marginal rather than total
valuations (9).
Another problem with the original estimate is that
landscapes can yield substantial (albeit rather different) flows of goods and
services after as well as before conversion by humans (which is of course why
people convert them). A clearer picture of the value of retaining habitat in
relatively undisturbed condition might therefore be obtained by estimating not
the gross values of the benefits provided by natural biomes, but rather the
difference in benefit flows between relatively intact and converted versions of
those biomes.
To address these concerns, we reviewed over 300 case
studies, searching for matched estimates of the marginal values of goods and
services delivered by a biome when relatively intact, and when converted to
typical forms of human use. To ensure we did not neglect private benefits of
conversion, studies were only included if they covered the most important
marketed goods, as well as one or more non-marketed services delivering local
social or global benefits. We cross-validated figures for individual goods and
services with other estimates from similar places. Last, we checked that the
comparisons across different states of a biome used the same valuation
techniques for particular goods and services. Our survey uncovered only five
examples which met all these criteria. Here we summarise their findings, with
all figures expressed as Net Present Values (NPVs, in 2000 US $ ha-1),
and using the discount rates considered by the authors (see Fig. 1 and online
material [10]
for further details).
Two studies quantified net marginal benefits of different
human uses of tropical forest areas. Kumari compared the values obtained from
timber plus a suite of Non-Timber Forest Products (NTFPs), as well as the
values of water supply and regulation, recreation, and the maintenance of
carbon stocks and endangered species, for forests under a range of management
regimes in Selangor, Malaysia (11). Compared with two methods of
reduced-impact logging, high intensity, unsustainable logging was associated
with greater private benefits through timber harvesting (at least at high
discount rates and over one harvesting cycle), but reduced social and global
benefits (through loss of NTFPs, flood protection, carbon stocks and endangered
species). Summed together, the Total Economic Value (TEV) of forest was some
14% greater when under more sustainable management (at ~$13,000 cf $11,200 ha-1).
A study from Mount Cameroon comparing low impact logging
with more extreme land-use change again found that private benefits favour
conversion, this time to small-scale agriculture (12). However, a second
alternative to retaining the forest - conversion to oil palm and rubber
plantations Ð in fact yielded negative private benefits, once the effect of
market distortions was removed. Social benefits, from NTFPs, sedimentation
control and flood prevention, were highest under sustainable forestry, as were
global benefits from carbon storage and a range of option, bequest and
existence values. Overall, the TEV of sustainable forestry was 18% greater than
that of small-scale farming (~$2570 cf $2110 ha-1), while
plantations had a negative TEV.
Three other biomes yielded single studies meeting our
criteria. Analysis of a mangrove system in Thailand revealed that conversion
for aquaculture made sense in terms of short-term private benefits, but not once external costs were factored
in (13).
The global benefits of carbon sequestration were considered to be similar in
intact and degraded systems. However, the substantial social benefits
associated with the original mangrove cover - from timber, charcoal, NTFPs,
offshore fisheries and storm protection Ð fell to almost zero following
conversion. Summing all measured goods and services, the TEV of intact
mangroves exceeded that of shrimp farming by around 72% (~$60,400 cf $16,700 ha-1).
van Vuuren and Roy (14) reported that draining freshwater
marshes in one of CanadaÕs most productive agricultural areas yielded net
private benefits (in large part because of substantial drainage subsidies).
However, social benefits of retaining wetlands, arising from sustainable
hunting, angling and trapping, greatly exceeded agricultural gains.
Consequently, for all three marsh types considered, TEVs were higher when the
wetlands remained intact, exceeding figures for conversion by a mean of 58%
(~$8800 cf $3700 ha-1).
Finally, a synthesis of economic studies examining Philippine
reef exploitation demonstrated that despite high initial benefits, destructive
techniques such as blast fishing had a far lower NPV of private benefits than
sustainable fishing (15).
The social benefits of sustainable exploitation Ð from coastal protection and
tourism Ð were also lost upon dynamiting reefs. As a consequence, the TEV of
retaining an essentially intact reef was some 73% higher than that of
destructive fishing (at ~$3300 cf $870 ha-1).
One clear message from our survey is the paucity of empirical
data on the central question of the changes in delivery of goods and services
arising from the conversion of natural habitats for human use. For ten of
Costanza et al.Õs
(3)
largely natural biomes (including rangelands, temperate forests, rives and
lakes, and most marine systems) we found no studies that met all of our
criteria. For the four biomes which were analysed, only a handful of
well-established ecosystem services were considered, and some particularly
valuable services, such as nutrient cycling, waste treatment, and the provision
of cultural values, were not examined at all.
Despite the limited data, our review also suggests a second
broad finding: in every case examined, the loss of non-marketed services
outweighs the marketed marginal benefits of conversion, often by a considerable
amount. Across the four biomes studied, mean losses in Total Economic Value due
to conversion run at roughly one half of the TEV of relatively intact system
(mean = 54.9%; SE = 13.4%; n = 4). This is certainly not to say that conversion has
never been economically beneficial Ð in most instances, past clearance of
forests and wetlands for prime agricultural land and other forms of development
probably benefited society as a whole. But unless the present case studies or
the range of services and biomes examined in the literature is extremely
unrepresentative (and we know of no reason why this should be the case), our
synthesis indicates that nowadays, conversion of remaining habitat for
agriculture, aquaculture or forestry often does not make sense from the
perspective of global sustainability.
These results therefore provide a clear and compelling
economic case, alongside socio-cultural and moral arguments (16-18), for us to strengthen
attempts to conserve what remains of natural ecosystems. Yet when we summarised
available estimates of recent trends in the global status of natural habitats
and free-ranging vertebrate populations, we found that although key data are
again disturbingly scarce, they show that rates of conversion are high across
most biomes (10).
We included in our survey any estimate of global trend in
habitat cover based on a series which began in 1970 or later and included a
period of at least five years after the United Nations Conference on
Environment and Development in 1992.
We supplemented this with biome-specific indices based on time series
data on populations of wild vertebrates, derived from the WWF 2000 Living
Planet Index (LPI) and FAO fisheries data (19, 20). For three biomes we found two
estimates using different methods and either largely or wholly independent
data. In each case the two estimates were remarkably similar (10), and so were averaged to
yield single estimates of rates of change. Data such as these, quantifying
trends in areal coverage and in populations, in some ways provide a more
tractable measure of the scale of the ongoing crisis facing nature than do
estimates of extinction rates, which are harder to document and more difficult
to link to monetary values.
Overall we found that five out of the six biomes measured
have experienced net losses since the 1992 Rio summit, with the mean rate of
change across all measured biomes running at -1.2% per year, or Ð11.4% over the
decade (Fig. 2; see [10]
for details). Hence the capacity of natural systems to deliver goods and
services upon which we depend is decreasing dramatically. Costing the overall
value of these losses is fraught with the problems of extrapolation and data
availability already discussed. Nevertheless, it is sobering to calculate that
if the Costanza et al.
aggregate figures (3)
and our estimate of the proportion of TEV lost through habitat change are
roughly representative, a single yearÕs habitat conversion costs the human
enterprise, in net terms, of the order of $250 billion that year, and every
year into the future (10).
Why, then, is widespread habitat loss still happening, and what can we do about
it?
In economic terms, our case studies illustrate three broad,
inter-related reasons why we are continuing to lose natural ecosystems despite
their overall benefits to society (21). First, there are often failures of
information. For many services, we lack valuations of their provision by
natural systems, and particularly of changes in this provision as human impacts
increase. While this is an understandable reflection of substantial technical
difficulties, we believe that future work needs to compare delivery of multiple
services across a range of competing land uses if it is to better inform policy
decisions. Our examples show that even when only a few ecosystem services are
considered, their loss upon conversion typically outweighs any gains in
marketed benefits.
Second, these findings highlight the fundamental role of
market failures in driving habitat loss. In most of the cases we studied, the
major benefits associated with retaining systems more or less intact are
non-marketed externalities, accruing to society at local and global scales.
Conversion generally makes narrow economic sense because such external benefits
(or related external costs, as in the case of the damage caused by shrimp
farming [13])
have very little impact on those standing to gain immediate private benefits
from land-use change. Hence conserving
relatively intact habitats will often require compensatory mechanisms to
mitigate the impact of private, local benefits foregone, especially in
developing countries. We see the development of market instruments that capture
at a private level the social and global values of relatively undisturbed
ecosystems - for instance through carbon or biodiversity credits or through
premium pricing for sustainably harvested wild-caught fish or timber (22, 23) Ð as a crucial step towards
sustainability.
Third, the private benefits of conversion are often
exaggerated by intervention failures. In the Cameroon study, for example,
forests were cleared for plantations because of private benefits arising from
government tax incentives and subsidies (12).
The same is true for the Canadian wetland example (14), as well as for many other
wetlands across USA and Europe (24). While over the short term these
programs may be rational with respect to public or private policy objectives,
over the longer term many result in both economic inefficiency and the erosion
of natural services. Globally, the subset of subsidies which are both
economically and ecologically perverse totals between $950 and $1950 billion
each year, (depending on whether the hidden subsidies of external costs are
also factored-in [25,
26]).
Identifying and then working to remove these distortions would simultaneously
reduce rates of habitat loss, free up public funds for investing in sustainable
resource use, and save money (25-27).
Tackling these underlying economic problems requires action
on many levels, but should in due course result in public and private decision
makers acting to reduce conversion of remaining habitats worldwide. More
immediately, given concerns about the practicalities of exploiting natural
resources sustainably, one of the most important strategies to safeguard
relatively intact ecosystems is the maintenance of remaining habitats in
protected areas. This costs money, and predictably, our current undervaluation
of nature is reflected in marked underinvestment in reserves. To the best of
our knowledge the world spends (in 2000 US $) ~ $6.5 billion each year on the
existing reserve network (28). Yet half of this is spent in USA alone. Globally, despite
increased expenditure since the Rio summit by both international institutions
and private foundations, available resources for existing reserves fall far
short of those needed to meet basic management objectives (29). Moreover, terrestrial and
marine reserves currently cover only around 7.9% and 0.5% of the earthÕs land and sea area, respectively (30, 31) Ð well below the minimum
safe standard considered necessary for the task of maintaining wild nature into
the future (32-34).
To estimate the resources needed to meet this shortfall on
land, we reworked recent calculations (28, 35) of the costs of properly managing
existing terrestrial protected areas and expanding the network to cover around
15% of land area in each region. We found that a globally effective network
would require around an annual outlay of between ~ $20 and $28 billion
(including payments to meet private opportunity costs imposed by existing and
new reserves, spread out over 10y and 30y respectively [10]). New work derived from the
costs of existing marine reserves suggests that an equivalent initiative for
the worldÕs seas, this time covering 30% of total area (34, 36), would cost at most ~ $23
billion per year in recurrent costs, plus ~ $6 billion per year (over 30 years)
in start-up costs (10).
The estimated mean the total cost of an effective, global reserve programme on
land and at sea is some $45 billion per year. This sum dwarfs the current $6.5
billion annual reserve budget yet could be readily met by redirecting less than
5% of existing perverse subsidies (25, 26). The crucial question is whether this
is a price worth paying.
Although limited data make the answer imprecise, they
indicate that conservation in reserves represents a strikingly good bargain. We
assumed that the mean proportional loss of value upon conversion recorded in
our case studies is representative of all biomes and services, and that
previous gross per hectare values of those services are roughly correct (3). If these assumptions are
valid, then our hypothetical global reserve network would ensure the delivery
of goods and services with an annual value (net of benefits from conversion) of
between ~ $4400 and $5200 billion,
depending on the level of resource use permitted within protected areas, and
with the lower number coming from a network entirely composed of strictly
protected reserves (for working, see [10]). The benefit : cost ratio of a reserve system meeting minimum safe
standards is therefore around 100 : 1.
Put another way, the case studies, Costanza et al.Õs (3) service values or our
reserve costs would have to be off by a factor of 100 for the reserve programme
envisaged to not make economic sense. We consider errors of this size to be
highly unlikely, as most of our assumptions are conservative (for other sensitivity
analyses, see [10]).
For example, in terms of the values of services, we assume that unit values
will not increase as supply diminishes, that nature reserves do not increase
the flow of services beyond their boundaries (whereas some clearly can [34,
37]), and
that all of a biomeÕs services not included in the Costanza et al. survey (3) are worthless. On the
reserve costs side, we assume that management costs do not decrease once local
communitiesÕ private opportunity costs are met, and that expanding reserve
systems yield no cost savings through economies of scale or dissemination of
best practice. Because these assumptions are biased against conservation, we
consider our 100 : 1 ratio as a low estimate of the likely benefits of
effective conservation.
In advocating greatly increased funding for the maintenance of natural ecosystems, we are not arguing against development. Given forecast increases in the human population of over three billion by 2050 (38) and the fact that some 1.2 billion people still live on less than a dollar a day (39), development is clearly essential. However, current development trajectories are self-evidently not delivering human benefits in the way that they should: income disparity worldwide is increasing and most countries are not on track to meet the United NationÕs goals for human development and poverty eradication by 2015 (39). Our findings show one compelling reason why this is the case Ð our relentless conversion and degradation of remaining natural habitats is eroding overall human welfare for short-term private gain. In these circumstances, retaining as much as possible of what remains of wild nature through a judicious combination of sustainable use, conservation, and, where necessary, compensation for resulting opportunity costs (as called for at Rio [40]) makes overwhelming economic as well as moral sense.
References and Notes
1.
By
Òwild natureÓ we mean habitat in which biodiversity, non-biotic components and
ecosystem functioning are sufficiently intact that the majority of ecosystem
services typically derived from such a habitat are still being sustainably and
reliably supplied. Our usage
differs from other usages, such as those adopted in cultural or anthropological
studies. Because our focus is on wild nature, we excluded the Cropland and
Urban biomes when using data from Table 2 of (3).
2.
G.
C. Daily, Ed., NatureÕs Services (Island Press, Washington D. C., 1997).
3.
R.
Costanza et al., Nature
387,
253-260 (1997).
4.
The
hedonic price method values environmental services by comparing market prices
(e.g. for residential housing) across situations which differ in the provision
of those services. Contingent valuation involves asking respondents how much
they would be prepared to pay for a particular environmental benefit (such as
ensuring the survival of a species or habitat) or how much compensation they
would demand for its loss. The replacement cost technique quantifies the cost
of restoring or synthetically replacing an ecosystem service.
5.
M.
Toman, Ecological Economics 25,
57-60 (1998).
6.
R.
K. Turner, W. N. Adger, R. Brouwer, Ecological Economics 25, 61-65 (1998).
7.
P.
Dasgupta, Human Well-Being and the Natural Environment (Oxford University Press,
Oxford, 2001).
8.
P.
A. L. D. Nunes, J. C. J. van den Bergh, Ecological Economics 39, 203-222 (2001).
9.
G.
C. Daily et al., Science
289,
395-396 (2000).
10.
For
further details, see online notes. Many of the numbers reported here are
unavoidably imprecise. To enable readers to follow our working, we generally
present numbers used in calculations to three significant figures, but then
round-off the final results in accord with their precision.
11.
K.
Kumari, thesis, University of East Anglia, Norwich (1994).
12.
G.
Yaron, Journal of Environmental Planning and Management 44, 85-108 (2001).
13.
S.
Sathirathai, Economic Valuation of Mangroves and the Roles of Local
Communities in the Conservation of Natural Resources: Case Study of Surat
Thani, South of Thailand
(unpublished report, Economy and Environment Program for Southeast Asia,
Singapore, 1998).
14.
W.
van Vuuren, P. Roy, Ecological Economics 8, 289-305 (1993).
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A.
T. White, H. P. Vogt, T. Arin, Marine Pollution Bulletin 40, 598-605 (2000).
16.
P.
W. Taylor, Environmental Ethics 3,
197-218 (1981).
17.
D.
Ehrenfeld, in Biodiversity, E. O. Wilson, Ed. (National Academic Press, Washington D.
C., 1988), pp. 212-216.
18.
B.
Norton, Environmental Values 1,
97-111 (1992).
19.
J.
Loh et al., WWF Living Planet Report 2000 (WWF, Gland, 2000).
20.
FAO,
The State of the World Fisheries and Aquaculture (FAO, Rome, 2000).
21.
R.
K. Turner et al.,
Ecological Economics
35, 7-23
(2000).
22.
J.
Hardner, R. Rice, Scientific American 286 (5), 71-95 (2002).
23.
S.
Scherr, A. White, D. Kaimowitz, Policy Brief: Making Markets For Forest
Communities
(Forest Trends, Washington, D.C., and Center for International Forestry
Research, Bogor, 2002).
24.
R.
Turner, T. Jones, Eds. Wetlands: Market and Intervention Failures (Four Case
Studies)
(Earthscan, London, 1991).
25.
C.
P. van Beers, A. P. G. de Moor, Addicted to Subsidies: How Governments Use Your
Money to Destroy the Earth and Pamper the Rich (Institute for Research on Public
Expenditure, The Hague, 1999).
26.
N.
Myers, J. Kent, Perverse Subsidies (Island Press, Washington D. C., 2001).
27.
S.
L. Pimm et al.,
Science 293, 2207-2208 (2001).
28.
A.
James, K. J. Gaston, A. Balmford, BioScience 51, 43-52 (2001).
29.
A.
N. James, M. J. B. Green, J. R. Paine, Global Review of Protected Area
Budgets and Staff (WCMC,
Cambridge, 1999)
30.
IUCN,
1997 United Nations List of Protected Areas (WCMC and IUCN, Cambridge and Gland,
1998).
31.
G.
Kelleher, C. Bleakley, S. Wells, A Global Representative System of Marine
Protected Areas
(The World Bank, Washington D. C., 1995).
32.
IUCN,
Parks for Life: Report of the IVth World Congress on National Parks and
Protected Areas
(IUCN, Gland, 1993).
33.
M.
E. SoulŽ, M. A. Sanjayan, Science 279, 2060-2061 (1998).
34.
J.
Roughgarden, P. Armsworth, in Ecology: Achievement and Challenge, M. Press, N. Huntly, S.
Levin, Eds (Blackwell Science, Oxford, 2001), pp. 337-356.
35.
A.
N. James, K. J. Gaston, A. Balmford, Nature 401, 323-324 (1999).
36.
California
Department of Fish and Game, NOAAÕs Channel Islands National Marine Sanctuary, A
Recommendation for Marine Protected Areas in the Channel Islands National
Marine Sanctuary
(California Dept. of Fish and Game, Santa Barbara, 2001).
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C.
M. Roberts, J. A. Bohnsack, F. Gell, J. P. Hawkins, R. Goodridge, Science 204, 1920-1923 (2001).
38.
UN
Population Division, World Population Prospects: The 2000 Revision (UN Dept. of Economic and Social
Affairs. New York, 2001).
39.
UNDP,
Human Development Report, 2001 (UNDP, New York, 2001).
40.
See
the text of the Convention on Biological Diversity at www.biodiv.org
41.
This
review is the result of a workshop convened by the Royal Society for the
Protection of Birds and sponsored by the RSPB and the UK GovernmentÕs
Department for Environment, Food and Rural Affairs. We thank Neal Hockley,
Pippa Gravestock, Jšrn Scharlemann and Catherine Tiley for help with research,
and Mark Avery, Richard Cowling, Gretchen Daily, Alistair Gammell, David
Gibbons, Jeffrey McNeely, and Callum Roberts for stimulating discussions.
Fig. 1. The marginal benefits of retaining and converting natural
habitats, expressed as Net Present Values (in 2000 US $ ha-1)
calculated using the discount rates and time horizons presented. Values of
measured goods and services delivered when habitats are relatively intact and
when converted are plotted as black and white columns, respectively. (From [11-15]; see [10] for further details.)
Fig. 2. Recent estimates of the annual rate of change in the area
or abundance of vertebrate populations for six biomes. Note that the biomes
that have declined deliver very valuable ecosystem services (3). *Values plotted are the
mean of habitat and population-based estimates; little confidence can be
attached to this value (10).
Online material
All values were converted to 2000 US $ using a GDP deflator
index (S1).
Where case studies gave ranges of values, we took midpoints. All figures were
taken directly from the sources cited, except for the coral reef example, where
the time schedule for yields from destructive fishing was estimated from Figure
2 of ref. [S2]
as 36 tonnes km-2 in year 1, and then 3 tonnes km-2,
rising to 5 tonnes km-2 by year 10.
Rates of loss
We searched the published literature and available databases
for global estimates of recent trends in the area of largely unmodified
habitats in all the relevant biome categories of Costanza et al. (S3) except rock and ice and open
ocean. We supplemented this with biome-specific indices based on time series
data on populations of wild vertebrates, derived from the WWF 2000 Living
Planet Index (LPI) and FAO fisheries data (S4, S5). Unless otherwise stated, annual
percentage rates of change in area or index value were calculated by taking the
values a1,
a2 in
the first year t1 and the last year t2 of the series under
consideration and calculating 100*(1 - (a2 /a1)(1/(t2 - t1))).
Tropical forests: We used the estimate in the FAO Forest
Resources Assessment 2000 (S6) of a global net change of -7% in the area of tropical
forest for the period 1990-2000, yielding an annual decline of 0.8%, although
we are aware that some authorities consider this an underestimate. The LPI Index (S4) for tropical forest
vertebrates showed a decrease of 26% between 1970 and 1999, yielding an average
annual decline of 1.1%.
Temperate and boreal forests: The FAO Forest Resources Assessment
estimates that temperate and tropical forests have increased in extent by 1%
during the period 1990-2000, yielding an annual increase of 0.1% (S6). The LPI Index (S4) for temperate forest
vertebrates showed a change of +4% between 1970 and 1999, yielding a small
annual increase of 0.1%.
Mangroves: Valiela et al.(S7) estimated on the basis of a
comprehensive assessment of mangrove resources that at least 35% of the global
area of mangrove forests has been lost in the past two decades. Their data
yield an annual decline of at least 2.5%.
Grasslands, rangelands, deserts and tundra: There are no global
estimates for rates of change in the extent of these habitats or for overall
changes in their condition. The available data for vertebrate populations are
currently inadequate to allow development of a reliable LPI for any of these
biomes.
Coral reefs: Although Bryant et al. (S8) report that around one quarter of the
worldÕs reefs are believed to be at high risk of degradation, there are no
reliable global estimates for the rate at which coral reefs are actually being
lost or degraded.
Seagrass and algal beds: There are no global estimates for the
extent of algal beds, nor for rates of change in extent. No comprehensive
survey of seagrass beds has been carried out, although it has been estimated
that there may be between 500,000 and 1,000,000 km2 in total (M.
Spalding pers. comm.). Short and Wyllie-Echeverria (S9) stated that perhaps 900 km2
of seagrass beds had been lost globally between 1985 and 1995, although the
basis for this is not clear. Extrapolation would give an annual decline of
0.01-0.02% although little confidence can be attached to this figure.
Estuaries: We found no global assessment of rates of loss or
degradation of estuarine habitats.
Coastal shelf: The only measure of coastal shelf
habitat modification for which we found global estimates was disturbance of the
sea-floor by bottom trawling. However, it is not clear what proportion of
bottom trawling caused long-term habitat degradation, so we have not used this
estimate.
Marine: The marine component of the WWF LPI (S4) does not distinguish between
different marine biomes. Overall it indicates a 36% decline in abundance of
marine fish, mammals, birds and reptiles over the period 1970-1999, yielding an
average annual decline of 1.5%. Further evidence for decline is provided by
fitting a curve to FAO data (S5) on changes since 1974 in the proportion of all the worldÕs
marine fish stocks that are exploitable (i.e. categorised as fully, moderately
or under-exploited). Most fish stocks reduced to unexploitable levels show
little evidence of recovery within 15 years of their decline (S10) and so can be regarded as
effectively lost to exploitation for the foreseeable future. The fitted curve
suggests that exploitable fish stocks have effectively been ÒlostÓ at the rate
of 1.5% per year.
Thus for each of three biomes we have two estimates derived
by different methods and either independent data (tropical forest and
temperate/boreal forest) or largely independent data (marine LPI and fish
stocks). In all cases the two
estimates were remarkably similar.
The rates of change were therefore averaged for these biomes to yield a
single estimate. Five of the six
global biome-specific estimates of change in habitat area or population show
declines, which are distributed about a mean of 1.2% per year (SE = 0.5%; n = 6).
If the Costanza et al. aggregate figures (S3) for largely natural biomes
and our estimate of the proportion of TEV lost upon conversion are roughly
correct, then a single yearÕs average losses in the 1990s cost society
approximately $37.6 trillion x 54.9% x 1.2% Å $250 billion every year into the
future.
The hypothetical terrestrial reserve network would cover
~15% of each region (S11,
S12). The
costs include resources needed for the effective management of existing and
new reserves; the costs of
adequately compensating local residents in developing countries for the unmet
private opportunity costs of existing, strictly protected reserves (spread over
10 years); the costs of surveying and then leasing or acquiring new reserves
(spread over 30 years); and the private opportunity costs of greening forestry
or farming in buffer zones around the margin of reserves, covering an additional
1.5% of each regionÕs total area.
The cost of the hypothetical marine network was derived from
a survey of current and unmet expenditure for 71 Marine Protected Areas (MPAs;
A. Balmford, P. Gravestock and C. Roberts, unpubl. data). Total management
costs of MPAs can be predicted from their size (regression gives log10[annual
cost, in 2000 US $] = 5.00 + 0.20 [log10(area, in km2)],
with r2
= 0.79). The management costs of the hypothetical global network were then
estimated by combining this relationship with the log-normal size distribution
for 991 existing reserves (S13), which together cover 0.50% of the seas, and assuming 30%
coverage is achieved by simple replication of this current network (note that
this will overestimate total costs because plausible spatial patterns of
network expansion inevitably lead to reserve merging and hence economies of
scale). One-off set-up costs of MPAs were estimated at 7.3 times annual
management costs (from n=4
reserves, including [S2]),
and were spread evenly over a 30-year implementation period.
Total costs for both the terrestrial and marine reserve
system varied through the implementation period, from $32 to $54 billion per
year, with a mean of $45 billion.
If Costanza et al.Õs (S3) per hectare values of ecosystem
services and our 54.9% estimate for the relative loss of TEV upon conversion
are approximately correct, the proposed reserve network would safeguard annual
flows worth $23.8 trillion x 54.9% x 30% Å $3900 billion at sea and (because a network covering 15% of
land area would cover ~16.9% of the largely natural biomes [S3]) $13.8 trillion x 54.9% x
16.9% Å $1300 billion on land, or ~$5200 billion in total. However, under
strict protection, those flows accruing from resource extraction would not be available.
Remaining services constitute ~91.4% of all services by value, according to
Table 2 of ref. (S3)
(conservatively assuming all recreation is incompatible with strict
protection). Hence a strict reserve network would safeguard annual flows with a
net worth of ($23.8 trillion x 91.4% x 30%) Ð ($23.8 trillion x 45.1% x 30%) Å
$3300 billion at sea, and ($13.8 trillion x 91.4% x 16.9%) Ð ($13.8 trillion x
45.1% x 16.9%) Å $1100 billion on land, or ~$4400 billion in total.
Our qualitative conclusions remained robust to varying
discount rates in the case studies between 3% and 10%, and to excluding tourism
and live fishing benefits in the reef example. The ~ 100 : 1 ratio was also
robust when (because it was not addressed in the case studies) we excluded all
benefits and costs from open oceans, and decreased only as low as ~ 40 : 1 even
when we made the unlikely assumption that nutrient cycling (the largest service
not examined in the case studies) differed from all measured services in being
delivered equally by intact and converted biomes.
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International
Monetary Fund. International Financial Statistics (International Monetary Fund,
Washington, DC, 2001).
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A.
T. White, H. P. Vogt, T. Arin, Marine Pollution Bulletin 40, 598-605 (2000).
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R.
Costanza et al., Nature
387,
253-260 (1997).
S4.
J.
Loh et al., WWF Living Planet Report 2000 (WWF, Gland, 2000).
S5.
FAO,
The State of the World Fisheries and Aquaculture (FAO, Rome, 2000).
S6.
FAO,
Global Forest Resources Assessment 2000: Main Report (FAO, Rome, 2001).
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J.
L. Valiela, J. K. Bowen, J. K. York, BioScience 51, 807-815 (2001).
S8.
D.
Bryant, L. Burke, J. McManus, M. Spalding, Reefs at Risk (WRI, New York, 1998).
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T. Short, S. Wyllie-Echeverria, Environmental Conservation 23, 17-27 (1996).
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A. Hutchings Nature
406,
882-885 (2000).
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A.
James, K. J. Gaston, A. Balmford, BioScience 51, 43-52 (2001).
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A.
N. James, K. J. Gaston, A. Balmford, Nature 401, 323-324 (1999).
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G.
Kelleher, C. Bleakley, S. Wells, A Global Representative System of Marine
Protected Areas
(The World Bank, Washington D. C., 1995).