Global forest loss disproportionately erodes biodiversity in intact landscapes

Abstract

Global biodiversity loss is a critical environmental crisis, yet the lack of spatial data on biodiversity threats has hindered conservation strategies. Theory predicts that abrupt biodiversity declines are most likely to occur when habitat availability is reduced to very low levels in the landscape (10-30%). Alternatively, recent evidence indicates that biodiversity is best conserved by minimizing human intrusion into intact and relatively unfragmented landscapes. Here we use recently available forest loss data to test deforestation effects on International Union for Conservation of Nature Red List categories of extinction risk for 19,432 vertebrate species worldwide. As expected, deforestation substantially increased the odds of a species being listed as threatened, undergoing recent upgrading to a higher threat category and exhibiting declining populations. More importantly, we show that these risks were disproportionately high in relatively intact landscapes; even minimal deforestation has had severe consequences for vertebrate biodiversity. We found little support for the alternative hypothesis that forest loss is most detrimental in already fragmented landscapes. Spatial analysis revealed high-risk hot spots in Borneo, the central Amazon and the Congo Basin. In these regions, our model predicts that 121-219 species will become threatened under current rates of forest loss over the next 30 years. Given that only 17.9% of these high-risk areas are formally protected and only 8.9% have strict protection, new large-scale conservation efforts to protect intact forests are necessary to slow deforestation rates and to avert a new wave of global extinctions.

Full text

27 JULY 2017 | VOL 547 | NATURE | 441
LETTER doi:10.1038/nature23285
Global forest loss disproportionately erodes
biodiversity in intact landscapes
Matthew G. Betts1,2*, Christopher Wolf1,2*, William J. Ripple1,2, Ben Phalan1,3, Kimberley A. Millers4, Adam Duarte5,
Stuart H. M. Butchart3,6 & Taal Levi1,4
Global biodiversity loss is a critical environmental crisis, yet the lack
of spatial data on biodiversity threats has hindered conservation
strategies1. Theory predicts that abrupt biodiversity declines are
most likely to occur when habitat availability is reduced to very low
levels in the landscape (10–30%)
2–4
. Alternatively, recent evidence
indicates that biodiversity is best conserved by minimizing human
intrusion into intact and relatively unfragmented landscapes5.
Here we use recently available forest loss data6 to test deforestation
effects on International Union for Conservation of Nature Red List
categories of extinction risk for 19,432 vertebrate species worldwide.
As expected, deforestation substantially increased the odds of a
species being listed as threatened, undergoing recent upgrading
to a higher threat category and exhibiting declining populations.
More importantly, we show that these risks were disproportionately
high in relatively intact landscapes; even minimal deforestation
has had severe consequences for vertebrate biodiversity. We found
little support for the alternative hypothesis that forest loss is most
detrimental in already fragmented landscapes. Spatial analysis
revealed high-risk hot spots in Borneo, the central Amazon and the
Congo Basin. In these regions, our model predicts that 121–219
species will become threatened under current rates of forest loss
over the next 30 years. Given that only 17.9% of these high-risk
areas are formally protected and only 8.9% have strict protection,
new large-scale conservation efforts to protect intact forests7,8 are
necessary to slow deforestation rates and to avert a new wave of
global extinctions.
A critical question in global efforts to reduce biodiversity loss is
how best to allocate scarce conservation resources. To what extent
should conservation be focused on modified and fragmented land-
scapes where threats are potentially greatest, versus landscapes that
are largely intact9? Although it is expected that both approaches have
value, in some human-influenced habitats, many species seem sur-
prisingly resilient to habitat loss and fragmentation, and can coexist
with humans in highly modified landscapes
10,11
, provided that habitat
loss does not exceed critical thresholds2. Theory predicts that abrupt
biodiversity declines are most likely to occur when habitat availability is
reduced to very low levels in the landscape (10–30%)
3,4,12
. Alternatively,
recent evidence indicates biodiversity is best conserved by minimizing
human intrusion into intact and relatively unfragmented landscapes,
which implies concentrating the impacts of anthropogenic disturbance
elsewhere5,13. This is because initial intrusion may result in rapid deg-
radation of intact landscapes, not only via the direct effects of habitat
loss, but also the coinciding effects of overhunting, wildfires, selective
logging, biological invasions and other stressors5. Such evidence has
led to recent calls to increase the protection of substantial intact areas
of the Earth’s terrestrial ecosystems14,15. Testing the extent to which
these alternative hypotheses explain patterns of extinction risk globally
can improve the effectiveness of conservation efforts and inform the
formulation of policies, affecting the future of life on Earth.
Recent advances in remote sensing have enabled the development
of a spatially explicit, high-resolution global dataset on rates of forest
change6, which provide the capacity to quantify the effects of contem-
porary global forest loss on biodiversity16. We quantified the association
between global-scale forest loss and gain within the ranges of 19,432
species and their International Union for Conservation of Nature
(IUCN) Red List category of extinction risk, recent genuine changes
in extinction risk, and overall population trend direction. The species
spanned three vertebrate classes, and included 4,396 (22.6%) listed as
threatened (Vulnerable, Endangered, or Critically Endangered) and
15,214 (78.3%) associated with forest habitats. Under the ‘habitat
threshold’ hypothesis, we expected the effects of recent forest loss to
be most detrimental for species that have already lost a substantial
proportion of forest within their ranges. Under the ‘initial intrusion’
hypothesis, we expected species with relatively intact forest within their
ranges to show the most severe effects of deforestation.
We obtained range maps for amphibians and mammals from the
IUCN Red List17 and those for birds from BirdLife International and
NatureServe18. We classified species as ‘non-forest’, ‘forest-optional’,
and ‘forest-exclusive’ based on the IUCN Red List habitat classifica-
tion data17. Within each species’ range, we used fine-resolution forest-
change data (2000–2014)6 to calculate the amount of recent forest
cover, loss, and gain (Fig. 1). Given that many species were assessed
for the Red List in the early period of our recent forest-loss data
(or even before this; Methods), it would be ideal to have contemporary
forest loss data from before 2000. The most spatially contiguous dataset
for 1990–200019 covered > 80% of the ranges for only 58.7% of the
species in our analyses. However, locations of forest loss were highly
spatiotemporally correlated at the scale of species’ ranges between
1990–2000 and 2000–2014 (Methods, Intermediate-term forest change;
Extended Data Fig. 7).
We also expected that historical deforestation over much longer
temporal scales could influence species vulnerability, a phenomenon
known as ‘extinction debt’
20,21
. We calculated historical forest loss as
the difference between the extents of area within species’ ranges that
historically supported forest cover and the area that remained forested
in the year 2000
6
(Fig. 1). We also calculated the mean ‘human foot-
print’ value22 within each species’ range, because forest loss could be
confounded with other broad-scale anthropogenic pressures (Fig. 1).
Using these data, we fit a spatial autologistic regression model to test
whether forest loss within species’ ranges is associated with the like-
lihood that a species: (i) is listed as threatened; (ii) has qualified for
uplisting to a higher category of extinction risk in recent decades (see
Methods); and (iii) has a declining population trend (as classified by
IUCN Red List assessors).
1Forest Biodiversity Research Network, Department of Forest Ecosystems and Society, Oregon State University, Corvallis, Oregon 97331, USA. 2Global Trophic Cascades Program, Department of
Forest Ecosystems and Society, Oregon State University, Corvallis, Oregon 97331, USA. 3Department of Zoology, University of Cambridge, Downing Street, Cambridge CB2 3EJ, UK. 4Department of
Fisheries and Wildlife, Oregon State University, Corvallis, Oregon 97331, USA. 5Oregon Cooperative Fish and Wildlife Research Unit, Department of Fisheries and Wildlife, Oregon State University,
Corvallis, Oregon 97331, USA. 6BirdLife International, David Attenborough Building, Pembroke Street, Cambridge CB2 3QZ, UK.
* These authors contributed equally to this work.
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter
reSeArCH
442 | NATURE | VOL 547 | 27 JULY 2017
As expected, we found a strong association between rate of recent
forest loss and each response variable. The odds of threatened status,
declining population trends, and uplisting increased by 5.06% (95%
confidence interval: 1.01–9.27), 11.34% (6.45–16.45), and 8.39%
(1.53–15.70), respectively, for each 1% increase in recent forest loss for
forest-exclusive species. This is not surprising, given that estimated or
inferred rates of habitat loss are used to inform IUCN Red List assess-
ments under criterion A2, particularly for species lacking direct data
on population trends
17
. Nevertheless, our results confirm that previous
categorical estimates of habitat decline (based on a mixture of inference,
qualitative and quantitative analysis) match with our global, systematic
analysis of quantitative data on forest loss16.
More importantly, we found strong support for the initial intrusion
hypothesis for both forest-optional and forest-exclusive species. Species
were more likely to be threatened, exhibit declining population trends
and have been uplisted if their ranges contained intact landscapes
(> 90% forest cover) with high rates of recent forest loss (Fig. 2). Evidence
for this lies in the strong positive statistical interaction between forest
loss and cover (that is, forest loss × cover, Figs 2a, 3) on all response
variables for both forest-exclusive and forest-optional species (maximum
false discovery rate (FDR)-adjusted P = 0.025, minimum z = 2.51,
Fig. 2a, Supplementary Table 3). For example, at high proportions of
initial forest cover (90%), the odds of a forest-exclusive species being
uplisted were 15.78% (95% confidence interval: 6.99–25.30) greater
for each 1% increase in deforestation. At average proportions of forest
cover (57%), the equivalent increase in deforestation was much smaller,
with the odds of a forest-exclusive species being uplisted reduced to
3.45% (95% confidence interval: − 3.91 to 11.36) (Fig. 3). These results
were generally similar across vertebrate classes, but amphibians showed
the strongest and most consistent effects across response variables
(Extended Data Fig. 2). Predictably, forest loss and its interaction with
forest cover had little effect on non-forest species (Figs 2, 3).
Historical forest loss also exhibited a strong negative influence on
vertebrate biodiversity (Fig. 2), which may be evidence of an extinc-
tion debt in which some species are capable of persisting in landscapes
long after initial forest loss has occurred, but subsequently decline.
0
25
50
75
100
Cell
mean (%)
aForest cover (2000)
0
20
40
60
Cell
mean (%)
bRecent forest loss (2000−2014)
0
10
20
30
40
Cell
mean (%)
c
Recent forest gain (2000−2012)
0
25
50
75
Cell
mean (%)
dHuman footprint
0
10
20
30
40
50
Cell
mean (%)
e
Forest loss × cover
0
25
50
75
Cell
mean (%)
fHistorical forest loss
Figure 1 | Spatial distribution of the six variables used to predict
species’ IUCN Red List response variables. a, b–d, Forest cover in the
year 2000 (a), forest loss between 2000–2014 (b), forest gain (2000–2012)
(c), and human footprint (d). e, The interaction term ‘forest loss × cover’
tested alternative hypotheses that forest loss exerts the greatest negative
influence on biodiversity at low versus high initial levels of forest cover.
High values of this variable (shown in e) correspond to regions of both
high forest cover and loss. f, Historical forest loss represents long-term
forest loss in years preceding 2000. Values plotted are averages taken over
0.4° grid cells. The maps are derived from current forest change maps6
(a–c, e, f), an intact forest landscapes map32 (f), biomes of the world33
(f),and human footprint22 (d).
Benecial effect
on biodiversity
a
c
Detrimental effect
on biodiversity
b
d
Historical forest loss Human footprint
Forest loss × coverForest gain
−0.5 0.0 0.5 −0.5 0.
00
.5
Fo
rest-exclusive
Forest-optional
Non-forest
Fo
rest-exclusive
Forest-optional
Non-forest
Standardized coefcient
Response Threatened status Declining trend Uplisted in
threatened status
FDR adjusted P value 0 < P ≤ 0.05 0.05 < P ≤ 0.1 P > 0.1
Figure 2 | Effects of four predictors on the status of 19,432 vertebrate
species worldwide. a, Positive ‘forest loss × cover’ terms indicate that
the negative effects of forest loss are amplified in landscapes with greater
initial forest cover. b–d, Forest gain tended to have a positive effect on
forest optional and exclusive species (b), whereas historical forest loss
(c) and human footprint (d) tended to have negative effects. ‘Threatened
status’ refers to IUCN Red List categories of ‘Vulnerable’, ‘Endangered’,
or ‘Critically Endangered’. ‘Uplisted in threatened status’ means that the
most recent genuine Red List category change for a species has been in the
direction of higher endangerment. Forest loss and cover variables were
included as main effects, but coefficient estimates are not shown here as
they are not readily interpretable in the presence of the interaction term.
Error bars represent 95% confidence intervals. Categories for P values
are listed as ranges (that is, 0 < P ≤ 0.05, 0.05 < P ≤ 0.1, P > 0.1), and
sample sizes (also given in Supplementary Table 1) for non-forest/forest-
optional/forest-exclusive are 4,218/3,430/4,218, 10,457/8,827/10,457,
4,757/4,073/4,757 for Threatened status, Declining trend, and Uplisted in
threatened status, respectively.
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter reSeArCH
27 JULY 2017 | VOL 547 | NATURE | 443
Predictably, increased human footprint has had a generally negative
influence on the status of vertebrates associated with forest and non-
forest systems (Fig. 2). We also found recent forest gain decreased the
likelihood of threatened status (forest-exclusive and forest-optional
species) and declining population trend (forest-optional species; Fig. 2).
However, amphibians primarily drove these relationships; bird and
mammal biodiversity did not show statistically significant responses
to forest gain (Extended Data Fig. 2), indicating that young secondary
forest does not appear to be ameliorating biodiversity declines for
these taxa8.
Overall, the global spatial autologistic regression model performed
remarkably well (area under the receiver operating characteristic
curve (AUC) = 0.78; Extended Data Fig. 1), even when we conser-
vatively excluded entire regions one at a time (Africa, Americas,
Asia, Oceania) and evaluated models on these independent data
(AUC = 0.74). Furthermore, results remained consistent when we
statis tically accounted for phylogenetic dependencies, latitude, and time
since each species was initially described Extended Data Fig. 3. We also
applied alternative approaches to account for spatial autocorrelation
and excluded species designated as threatened due to characteristically
small and declining or fragmented ranges (that is, under IUCN Red List
criterion B) (Extended Data Figs 3, 6). Results were also robust to degree
of threat; Critically Endangered, Endangered and Vulnerable species all
showed similar patterns in response to forest loss (Extended Data Fig. 9).
Strong support for the initial intrusion hypothesis may be surprising,
given existing theory
3,23
and that a considerable number of conserva-
tion programs focus on areas that have already lost substantial forest24.
However, such highly deforested landscapes may have already passed
through a substantial local extinction filter, whereby the most sensi-
tive species have been lost25. A recent broad-scale study conducted
in the Brazilian Amazon revealed that landscapes still exceeding 80%
forest cover have lost 46–60% of their conservation value
5
. Our results
suggest that initial forest loss is a potential indicator of other threats to
forest biodiversity that are more challenging to measure at large spatial
extents. Mechanisms for intrusion effects include increased unregulated
hunting
26
(especially near new logging roads
27
), disease and human
disturbance, and invasive species28, as well as the direct effects of habitat
loss for interior forest specialists
29
. Indeed, many of the species with
ranges that were characterized by high initial forest cover (before 2000),
but intensive recent deforestation, tend to be under hunting pressure
(for example, Sira curassow (Pauxi koepckeae)) or are habitat specialists
(Mendolong bubble-nest frog (Philautus aurantium), Mentawi flying
Threatened status Declining trend
Uplisted in
threatened status
Non-forest Forest-optionalForest-exclusive
0% 25%50% 75%100%0%25% 50%75% 100% 0% 25% 50%75% 100%
0%
5%
10%
15%
0%
5%
10%
15%
0%
5%
10%
15%
Forest cover (2000)
Forest loss
0.25
0.50
0.75
Probability
Figure 3 | Predicted probabilities of species status as a function of
recent forest loss and total forest cover within a species range. All other
covariates (forest gain, historical forest loss, and human footprint) were
statistically held at their average values when estimating probabilities.
For forest-optional and forest-exclusive species, the effect of forest loss is
stronger at high levels of initial forest cover; deforestation in intact forests
has the most negative impact, supporting the initial intrusion hypothesis.
0.5 × current
loss rate
Current
loss rate
1.5 × current
loss rate
IUCN
category
Ia Ib II III
IV VVI0204060
Increase in
threatened
richness
2030–2045 2045–2075
Figure 4 | Projected increases in the number of threatened species
under three scenarios of future forest loss. Projections are estimated
using the global model. Increased threatened richness (blue to red colour
scale) is relative to the fitted probabilities of a species being threatened. For
example, a value of 20 would indicate a projected increase of 20 threatened
species in a 0.2° grid cell. Only locations with projected increases in
threatened species are shown and only forest-exclusive species were used
for this projection. Column labels show time spans where the lower limit
assumes the effects of forest loss on status are entirely due to deforestation
from 2000–2014; the upper limit assumes effects could be partly a function
of forest loss in the decades before 2000 (global locations of forest loss are
temporally autocorrelated; see Methods, section ‘Intermediate-term forest
change’). IUCN protected areas (categories I–VI) are shown in greyscale
shading. The maps are derived from the following sources: IUCN Red List
species range maps18, recent forest change6, intact forest landscapes32, human
footprint22, world biomes33, and the World Database of Protected Areas34.
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter
reSeArCH
444 | NATURE | VOL 547 | 27 JULY 2017
squirrel (Iomys sipora)) (Supplementary Table 4). If specialists’ habitat
is targeted in the initial phases of deforestation (for example, accessible
high-economic-value forest (bottomland forest adjacent to rivers)),
habitat will be lost at much greater rates than indicated by the overall
rate of forest loss within a species’ range30.
As a further exploration of the habitat threshold hypothesis, we fit
a model to test whether the strongest negative effects of recent forest
loss occurred in landscapes with both high and low levels of remain-
ing forest cover (a statistical interaction between forest loss and forest
cover squared; Methods). We found no evidence for such an effect for
either threatened status or recent uplisting (Extended Data Figs 4, 5).
Notably, the odds of a declining population trend showed
evidence for this dual effect for forest-optional and -exclusive species;
we speculate that the increased likelihood of a declining trend with
deforestation in landscapes with low levels of forest cover, but no
relationship for threatened status, may constitute early signs of an
extinction debt that remains to be fully paid. Thus, our results do not
imply deforestation effects are benign in regions with low levels of
remaining forest cover. Although species exposed to deforestation in
such landscapes are less likely to be designated as threatened than those
exposed to similar rates of deforestation in more intact areas, their
populations will continue to decline with further habitat loss, which
will in time inevitably lead to increased extinction risk.
The spatially explicit nature of our model enabled quantitative pre
-
dictions of global hotspots where biodiversity is at particularly high
risk given reduced (halving current rates), continued, or accelerated
(1.5× ) future rates of forest loss (each assumes no future forest loss in
protected areas with IUCN categories I–VI; Fig. 4). High-risk hot spots
emerged in southeast Asia (particularly Borneo), the central-western
Amazon and the Congo Basin where the numbers of threatened
forest-exclusive species are predicted to increase by 82–134, 34–74,
and 5–11, respectively, over the next 30 years under current rates of
deforestation. Together, the number of threatened species for these
three regions is predicted to increase by 121–219. Currently, only
17.9% of these areas are formally protected (IUCN classes I–VI;
Supplementary Table 5) and only 8.9% have strict protection (IUCN
classes I–III). These results, alongside evidence of ongoing erosion of
intact forest landscapes31, highlight that areas until recently considered
to be of “low vulnerability”
9
are in fact where anthropogenic distur-
bance is increasingly putting species at most risk of extinction. New
large-scale efforts to reduce both degradation and loss of intact forest
landscapes
7
are needed to protect against an intensified wave of extinc-
tions in the world’s last wildernesses.
Online Content Methods, along with any additional Extended Data display items and
Source Data, are available in the online version of the paper; references unique to
these sections appear only in the online paper.
Received 14 December 2016; accepted 13 June 2017.
Published online 19 July 2017.
1. Joppa, L. N. et al. Filling in biodiversity threat gaps. Science 352, 416–418
(2016).
2. Newbold, T. et al. Has land use pushed terrestrial biodiversity beyond
the planetary boundary? A global assessment. Science 353, 288–291
(2016).
3. Andrén, H. Eects of habitat fragmentation of birds and mammals in
landscapes with dierent proportions of suitable habitat: a review. Oikos 71,
355–366 (1994).
4. Betts, M. G., Forbes, G. J. & Diamond, A. W. Thresholds in songbird
occurrence in relation to landscape structure. Conserv. Biol. 21, 1046–1058
(2007).
5. Barlow, J. et al. Anthropogenic disturbance in tropical forests can double
biodiversity loss from deforestation. Nature 535, 144–147 (2016).
6. Hansen, M. C. et al. High-resolution global maps of 21st-century forest cover
change. Science 342, 850–853 (2013).
7. Peres, C. A. Why we need megareserves in Amazonia. Conserv. Biol. 19,
728–733 (2005).
8. Gibson, L. et al. Primary forests are irreplaceable for sustaining tropical
biodiversity. Nature 478, 378–381 (2011).
9. Brooks, T. M. et al. Global biodiversity conservation priorities. Science 313,
58–61 (2006).
10. Kareiva, P., Watts, S., McDonald, R. & Boucher, T. Domesticated nature: shaping
landscapes and ecosystems for human welfare. Science 316, 1866–1869
(2007).
11. Mendenhall, C. D., Karp, D. S., Meyer, C. F., Hadly, E. A. & Daily, G. C. Predicting
biodiversity change and averting collapse in agricultural landscapes. Nature
509, 213–217 (2014).
12. Fahrig, L. When does fragmentation of breeding habitat aect population
survival? Ecol. Modell. 105, 273–292 (1998).
13. Phalan, B., Onial, M., Balmford, A. & Green, R. E. Reconciling food production
and biodiversity conservation: land sharing and land sparing compared.
Science 333, 1289–1291 (2011).
14. Wilson, E. O. Half-Earth: Our Planet’s Fight For Life. (W. W. Norton & Company,
2016).
15. International Union for Conservation of Nature World Congress. Motion 48:
Protection of primary forests, including intact forest landscapes. (2016).
16. Tracewski, Ł. et al. Toward quantication of the impact of 21st-century
deforestation on the extinction risk of terrestrial vertebrates. Conserv. Biol. 30,
1070–1079 (2016).
17. International Union for Conservation of Nature. IUCN red list of threatened
species. Version 2016.3 http://www.iucnredlist.org (2017).
18. BirdLife International and NatureServe. Bird Species Distribution Maps of the
World Version 5.0 (BirdLife International, 2015).
19. Kim, D.-H. et al. Global, Landsat-based forest-cover change from 1990 to 2000.
Remote Sens. Environ. 155, 178–193 (2014).
20. Tilman, D., May, R. M., Lehman, C. L. & Nowak, M. A. Habitat destruction and
the extinction debt. Nature 371, 65–66 (1994).
21. Wearn, O. R., Reuman, D. C. & Ewers, R. M. Extinction debt and windows of
conservation opportunity in the Brazilian Amazon. Science 337, 228–232
(2012).
22. Sanderson, E. W. et al. The human footprint and the last of the wild. Bioscience
52, 891–904 (2002).
23. Hanski, I. Metapopulation dynamics. Nature 396, 41–49 (1998).
24. Brooks, T. M. et al. Habitat loss and extinction in the hotspots of biodiversity.
Conserv. Biol. 16, 909–923 (2002).
25. Balmford, A. Extinction lters and current resilience: the signicance of past
selection pressures for conservation biology. Trends Ecol. Evol. 11, 193–196
(1996).
26. Ripple, W. J. et al. Bushmeat hunting and extinction risk to the world’s
mammals. R. Soc. Open Sci. 3, 160498 (2016).
27. Benítez-López, A. et al. The impact of hunting on tropical mammal and bird
populations. Science 356, 180–183 (2017).
28. Bellard, C., Genovesi, P. & Jeschke, J. M. Global patterns in threats to vertebrates
by biological invasions. Proc. R. Soc. Lond. B 283, 20152454 (2016).
29. Clavel, J., Julliar, R. & Devictor, V. Worldwide decline of specialist species:
toward a global functional homogenization. Front. Ecol. Environ. 9, 222–228
(2011).
30. Betts, M. G. et al. A species-centered approach for uncovering generalities in
organism responses to habitat loss and fragmentation. Ecography 37, 517–527
(2014).
31. Potapov, P. et al. The last frontiers of wilderness: tracking loss of intact forest
landscapes from 2000 to 2013. Sci. Adv. 3, e1600821 (2017).
32. Potapov, P. et al. Mapping the world’s intact forest landscapes by remote
sensing. Ecol. Soc. 13, 51 (2008).
33. Olson, D. M. et al. Terrestrial ecoregions of the world: a new map of life on
Earth: a new global map of terrestrial ecoregions provides an innovative tool
for conserving biodiversity. Bioscience 51, 933–938 (2001).
34. IUCN and UNEP-WCMC. The World Database on Protected Areas (WDPA)
http://www.protectedplanet.net/terms (2015).
Supplementary Information is available in the online version of the paper.
Acknowledgements Funding from the National Science Foundation (NSF-
DEB-1457837) and the College of Forestry IWFL Professorship in Forest
Biodiversity Research to M.G.B. supported this research. We are grateful for
comments from A. Hadley, U. Kormann, J. Bowman, C. Epps and C. Mendenhall
on earlier versions of this manuscript.
Author Contributions M.G.B., C.W., S.H.M.B., W.J.R. and T.L. conceived the
study, C.W., M.G.B. and T.L. analysed the data, and M.G.B. and C.W. wrote
the first draft of the paper with subsequent editorial input from C.W., B.P.,
S.H.M.B., K.A.M. and A.D.
Author Information Reprints and permissions information is available at
www.nature.com/reprints. The authors declare no competing financial
interests. Readers are welcome to comment on the online version of the
paper. Publisher’s note: Springer Nature remains neutral with regard
to jurisdictional claims in published maps and institutional affiliations.
Correspondence and requests for materials should be addressed to
M.G.B. (matt.betts@oregonstate.edu) or C.W. (wolfch@science.oregonstate.edu).
Reviewer Information Nature thanks J. Barlow, L. Gibson and the other
anonymous reviewer(s) for their contribution to the peer review of this work.
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter reSeArCH
METHODS
No statistical methods were used to predetermine sample size. The experiments
were not randomized and the investigators were not blinded to allocation during
experiments and outcome assessment.
Species data. We obtained data on three classes of terrestrial vertebrates (mam-
mals, amphibians, and birds) from the IUCN Red List
17
. We defined threatened
species as those classified as Vulnerable, Endangered, or Critically Endangered on
the Red List. We also obtained population trends (‘Increasing’, ‘Stable’, ‘Decreasing’,
or ‘Unknown’) from the Red List. We excluded ‘Data deficient’ and ‘Extinct in the
wild’ species from our analysis with threatened status as the response variable.
Similarly, for the decreasing population trend response, we excluded species with
unknown population trends.
For analyses in which we examined change in Red List category, it is necessary
to compare time points in which all species in the taxonomic group were assessed,
and to consider only those Red List category changes between such assessments
that resulted from genuine improvement or deterioration in status (that is, exclud-
ing changes owing to improved knowledge or revised taxonomy). These genuine
changes underpin the Red List Index35,36. We considered species to have been
uplisted if their most recent genuine Red List category change was in the direction
of increasing endangerment (Least Concern < Near Threatened < Vulnerable
< Endangered < Critically Endangered). These data were obtained from Hoffmann
et al.
37
, and were updated to match the taxonomy on the 2016 IUCN Red List; the set
of genuine changes for birds was also updated using data in BirdLife International
38
.
The relevant periods of our primary uplisting dataset are 1980–2004 for amphibi-
ans, 1996–2008 for mammals, and 1988–1994, 1994–2000, 2000–2004, 2004–2008,
2008–2012, and 2012–2016 for birds. Additionally, we used all available genuine
category change data from 2008–2016 for mammals and 2006–2016 for amphibi-
ans. Although these more recent category change data (approximately 100 category
changes) are not yet comprehensive (that is, not all species in these taxa have been
checked for genuine category changes over these times), they cover a wide range of
species and are likely to be reflective of recent changes in forest cover for these species.
Genuine category change data are currently unavailable for other time periods.
We classified non-avian species according to habitat usage (forest-exclusive,
forest-optional, and non-forest) using the IUCN Red List data coding species
against the IUCN habitats classification scheme (http://www.iucnredlist.org/
technical-documents/classification-schemes/habitats-classification-scheme-ver3).
We treated species using only forest habitat as forest-exclusive, those using forest
habitat and at least one other habitat type as forest-optional, and those not using
forest at all as non-forest. To categorize bird species, we used higher-quality data on
forest dependency from BirdLife International
38
, treating species with high forest
dependency as forest-exclusive, medium and low forest dependency as forest-
optional, and not normally using forest as non-forest.
The species range maps used in the analysis were derived from the IUCN Red List
for mammals and amphibians, and from BirdLife International and NatureServe18
for birds. For each species, we used only range polygons where presence was classi-
fied as ‘Extant’ or ‘Probably extant’. Vertebrates without range maps available were
omitted from the analyses (108 mammals, 39 amphibians, and 30 birds). Reptiles
were excluded from the analysis as IUCN reptile data are relatively limited39.
After screening for data availability using the steps above, the dataset consisted
of 19,432 species (19,615 including data-deficient species), 4,396 (22.6%) of which
are listed as threatened. The entire dataset represents 58.2% of the terrestrial
vertebrate species globally (98.9% birds, 84.9% mammals, 63.1% amphibians)
(based on described species totals from IUCN Red List summary table 1).
Predictor variables. We used six predictor variables in our primary analysis
(Fig. 1). Here, we describe these variables in detail.
We used the forest change maps (version 1.2) given in Hansen et al.
6
for our
analyses. The forest cover map indicates the percentage forest cover in each 30 m
pixel in the year 2000. The forest loss and gain maps are both binary and indicate
whether forest loss or gain occurred in each pixel. Following Hansen et al., we
considered forest to have been ‘lost’ if a stand-replacing disturbance (that is, complete
removal of tree cover canopy at the Landsat pixel scale) had occurred between 2000
and 2014, and ‘gained’ if establishment of tree canopy from a non-forest state had
occurred between 2000 and 2012. In addition, we included a forest loss × forest cover
interaction term to test the hypothesis that the effects of forest loss are dependent
upon the total amount of forest within a species’ range. A positive coefficient for such
a term would indicate that the effect of recent forest loss on our response variables
was amplified at when initial forest cover was high (support for the initial intrusion
hypothesis). Conversely, a negative coefficient for this interaction term would indi-
cate that the effect of recent forest loss on our response variables was greatest at low
forest cover (support for the habitat threshold hypothesis; see main text).
The human footprint map that we used (Global Human Footprint v.2,
1995–2004) measures the extent of human impacts on the environment and is
created from nine global data layers covering biome type and biogeographic realm,
human population density, human land use and infrastructure (that is, built-up
areas, night-time lights, land use/land cover), and human access (coastlines, roads,
railroads, navigable rivers)40. Among land cover types, built-up environments
increase the human influence index the most, followed by agricultural land cover,
and mixed-use land cover (other types do not contribute to the index)
22
. Thus,
loss of forest to these land cover types could cause human footprint to be partially
confounded with our forest loss variable, potentially causing our analysis to under-
estimate the effects of forest loss. A more recent version of this map (1993–2009)
was recently released41,42 but the original and updated human footprint maps are
highly correlated (r = 0.935 at 2° resolution), so our choice of human footprint map
is unlikely to have influenced the results.
In our analysis, ‘historical forest loss’ is an estimate of long-term patterns in
forest loss that is not captured by contemporary forest change. To construct this
variable, we took the following steps. First, we used a random forest regression
model to develop a historical (or potential) forest cover map. We modelled the
continuous variable ‘percentage forest cover’ in the year 2000 (from Hansen et al.
7
)
as a function of x and y coordinates, 19 bioclimatic variables (derived from monthly
temperature/precipitation) from the WorldClim database13 along with a categor-
ical variable representing forest biomes
33
. Importantly, to exclude the effects of
contemporary anthropogenic disturbance on percentage forest cover we only used
data from within ‘intact forest landscapes’ (IFLs) in the regression model. An IFL
is defined as “an unbroken expanse of natural ecosystems within areas of current
forest extent, without signs of significant human activity, and having an area of at
least 500 km
2
”
32
. We assumed that forest cover in intact forest landscapes (IFLs) is
representative of the degree of canopy cover that could be historically supported in
across the globe. We then extrapolated the fitted values of this model to the areas
for a map of potential or historical forest cover (Extended Data Fig. 10a). Second,
we subtracted recent forest cover from historical cover to estimate historical loss
(Extended Data Fig. 10b) to yield a map of historical forest loss (Extended Data
Fig. 10c). We restricted our modelling to within forest biomes, excluding non-forest
biomes and the boreal forest/taiga. Although some forest cover may be present out-
side forest biomes (for example, in savannahs), limitations in available IFL data for
these cover types and the taiga make reconstructing historic forest cover in these
biomes impractical. Moreover, forest obligate species—our primary focus—seldom
occur outside forest biomes. Modelling was conducted at 5-km resolution using
rasters in Behrmann cylindrical equal-area projection. We used ArcGIS 10.1 and
R for the geospatial analyses
43,44
. The random forest model was fit using the Rborist
R package with the default settings45. We acknowledge that the period of time since
historical deforestation can vary widely across locations globally. Nevertheless, in
the absence of globally available forest loss data before 2000, this variable is the best
available test of whether long-term reductions in forest cover within a species range
affects Red List category and overall direction of population trend.
Statistical analysis. We used a 2-decimal degree equivalent equal-area grid
(constructed using the Behrmann cylindrical equal-area projection). This resolu-
tion is considered appropriate for macroecological analyses that involve species’
range maps
46
. We rescaled covariates to this resolution by taking their average
values across each grid cell (ignoring regions over water). We rescaled species’
ranges to the grid by treating a species as present in a grid cell if any part of its
range overlapped that cell. We then averaged covariates across species ranges
using the averages of their cell values weighted by the proportion of land in
each grid cell.
We modelled the probability of species being threatened, having a declining
population trend, or having been uplisted (three separate binary responses) using
autologistic regression to account for potential spatial autocorrelation47. The
spatial autocovariate was calculated for each species using a symmetric spatial
weights matrix as:
∑
=
∈
Awy
i
jk
ij j
i
where i is the ith species, k
i
is the set of its neighbours, y
j
is the response for the jth
species, and w
ij
= 1 corresponding to the (i, j) entry of the binary spatial weights
matrix
48
. Geographic distance was calculated using species’ range centroids. The
spatial weights matrix and spatial autocovariate were calculated using the spdep
package for R44,49.
We used the generalized linear model (GLM) function glm in R to fit the logistic
regression model, including the covariates described above, the spatial autocovariate,
and taxonomic class (as a fixed effect). We estimated standardized coefficients
and 95% confidence intervals for all predictor variables (each was standardized
(z-transformed) before analysis). Our hypothesis tests were conducted across all three
vertebrate classes with six predictor variables, which risks inflating Type I error rate.
Sequential Bonferroni-type multiple comparisons are sometimes used to account
for such error inflation, but are highly conservative50. Therefore, we used a FDR
procedure (the ‘graphically sharpened method’
50
) which does not suffer from the
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter
reSeArCH
same loss of power but corrects for multiple comparisons. FDR-adjusted P values
were calculated with p.adjust in R44,51.
Projecting future status changes. We used our model for threatened status of for-
est-exclusive species to map the predicted increase in threatened species richness at
multiple forest loss rates over time. We did this by simulating continued forest loss
at rates of 50%, 100% and 150% the current rates for the time spans 2030–2045 and
2045–2070. For example, at the current loss rate the area of forest lost would double
in 15 years (by 2030). We modified these forest loss projections by setting predicted
future loss to zero within IUCN category I–VI protected areas using the polygon
type protected area maps in the World Database of Protected Areas (WDPA)
34
.
Assuming that there are no substantial time lags between forest loss and species
being listed as threatened, the resulting predictions (probabilities of species being
threatened) correspond to 2030. In the event that intermediate term (approximately
1950–2000) forest loss is also closely linked to threatened status (that is, there are
time lags between forest loss and status decisions), we included conservative upper
time limits corresponding to half the stated forest loss rates. In all cases, predicted
current probabilities of being threatened (from the fitted model) were subtracted
from the estimated future probabilities; we then mapped the result by summing
probabilities for all species in each raster grid cell. As the maps show qualitatively
similar patterns, they can conservatively be interpreted as showing ‘relative hot
spots’—an interpretation that is valid even if the true intermediate-term forest rate
of loss is substantially higher than in our scenario.
To assess overlap between existing protected areas and hot spots (at high risk
of increases to the Red List), we used the ‘predicted increase in threatened spe-
cies richness’ map for 2030–2045 at the current loss rate. Within each regional
panel of this map set in Fig. 4, we considered hot spot areas to be those with at
least one quarter of the maximum predicted increase in threatened richness for
that region. We estimated the percentage of these areas that is protected using the
World Database of Protected Areas (WDPA)
34
. For this analysis, we report both
strictly protected areas: IUCN categories Ia (Strict Nature Reserve), Ib (Wilderness
Area), II (National Park), and III (Natural Monument or Feature) and all other
IUCN categories (IV, Habiat/Species Management Area; V, Protected Landscape;
VI, Protected area with sustainable use of natural resources). In addition, we only
consider protected areas with polygon data in the WDPA, which results in a con-
servative estimate of the percentage of high-risk area that is protected.
Assessing model performance. We used the area under the receiver operating
characteristic curve (AUC) to assess model performance for our primary model
(predicting threatened status for forest exclusive species). The AUC reflects the true
versus false positive rates for a binary classifier with continuous output as a function
of the threshold used to determine which outputs correspond to which categories
52
.
We calculated AUC both for the ‘A ll species’ model (with ‘class’ as a fixed effect)
and separately for each class using models fit to individual classes. We did this
with and without the spatial autocovariate term. In each case, we also quanti-
fied model performance using fourfold cross-validation by regions of the world
(Supplementary Table 2). We used a regional grouping (Africa, Americas, Asia,
Oceania) based on the United Nations Statistics Division classification system
53
.
Using entire regions as hold-out test datasets further reduces the positive effects
of dependency (spatial, taxonomic, and so on) on model performance metrics
54
.
The raw and cross-validated AUCs (0.784 without cross-validation, 0.743 with
cross-validation) for ‘All species’ together (with the autocovariate) indicate that our
models perform well (Extended Data Fig. 1). For each model, we also calculated
P values from a Wilcoxon rank-sum test55 to quantify whether the AUCs were
significantly greater than 0.5 (a baseline at which the model is performing no better
than random chance). All P values except the one for mammals with cross validation
and no auto-covariate term were highly significant (< 0.001) (Extended Data Fig. 1).
Alternative statistical methods to account for spatial autocorrelation. We tested
the residuals of our global autologistic regression model for spatial autocorrelation;
for all response variables, Moran’s I was < 0.15 across all distance classes, indicating
that the autocovariate had removed spatial autocorrelation. Further more, to ensure
that our results were robust to the sort of spatial model applied, we fit other spatial
logistic regression models (that is, Moran eigenvector filtering, simultaneous
spatial autoregressive models (SAR), and Bayesian conditional autoregressive models
(CAR)) to assess sensitivity to the procedure used for modelling or accounting
for spatial autocorrelation. We also fit a non-spatial generalized linear model for
reference along with our primary spatial autologistic regression model using the
50 nearest neighbours of each species (instead of 5). In each case, the models were
fit using forest-exclusive species with threatened status as the response. We fit
models for each taxonomic class separately, as not of all of the procedures could
readily incorporate the hierarchical structure of the data. Our results were robust
to the spatial autocorrelation modelling method (Extended Data Fig. 6). Details
on other spatial models applied are given below.
We fit a Moran eigenvector GLM filtering model by adding covariates to the
generalized linear model that were computed using the ME function in the spdep R
package49,56. This spatial filtering model involves augmenting the predictor matrix
with eigenvectors computed from the spatial weights matrix so as to reduce the
spatial autocorrelation of the residuals (as estimated using the Moran’s I statistic).
The smallest subset of eigenvectors that causes the permutation-based Moran’s I
test P value to exceed a threshold α is chosen for inclusion (we used α = 0.2, which
is a common default value).
We fit CAR and SAR models using the binary spatial weights matrix described
above. The conditional autoregressive model was fit using the CARBayes R
package
57
. Markov chain Monte Carlo sampling errors were encountered when
fitting a few of the CAR models. In such cases, the model results are not available.
The simultaneous autoregressive model was fit using the splogit function in the
MCSpatial package58. It is based on an approximation (linearization), which allows
the model to be fit to large datasets59.
Estimates within taxonomic classes. While the primary results presented in the
main text (Fig. 2) are for all classes together (with class included as a fixed effect),
we also fit models using data from each class separately (Extended Data Fig. 2).
We did this to assess the extent to which our results, particularly for the forest
loss × cover interaction, are consistent between classes.
Accounting for the effects of latitude. We fit models including latitude as a main
effect (Extended Data Fig. 3a). We did this to test whether our results were robust
to this potential confounding variable, which is correlated with numerous variables
that may be linked to endangerment such as net primary productivity (NPP) and
per capita gross domestic product (GDP). The estimated forest loss × cover inter-
action term did not change substantially when accounting for (absolute) latitude
(Extended Data Fig. 3a).
Quadratic models (loss × cover squared interaction). We fit models with quad-
ratic interaction terms corresponding to forest loss × cover
2
to test whether the
models with only the linear forest loss × cover terms were adequate for forest
exclusive and optional species. Support for a quadratic interaction term would
provide evidence for both the initial intrusion hypothesis and the threshold
hypothesis; in other words, the effects of forest loss on species status and trends
are most substantial both at very high and very low initial forest amounts (see main
text). These quadratic terms were generally non-significant (Extended Data Fig. 4)
supporting the hypothesis that the effect of forest loss on the odds of species being
threatened, declining, or uplisted varies linearly with forest cover. However, in the
overall (all species) models, we found strong evidence that the forest loss × cover2
term was positive when declining trend was the response variable (Extended Data
Fig. 4). This suggests that the effect of loss on population trends may be most
negative at both low and high levels of forest cover, and smallest (near zero) at
intermediate levels of forest cover (Extended Data Fig. 5).
Tropical forest species. As the ecology of tropical forests often responds differently
to non-tropical forests, we also examined model results for species found exclusively
in tropical forests (Extended Data Fig. 3b). We did this by restricting the species set
to those with ranges containing only grid cells that overlap tropical forests. We deter-
mined tropical forest regions using a map of biomes
33
and treating the following
biomes as tropical forest: ‘Tropical & Subtropical Moist Broadleaf Forests’, ‘Tropical
& Subtropical Dry Broadleaf Forests’, ‘Tropical & Subtropical Coniferous Forests’
and ‘Mangroves’. The restriction of our dataset to tropical forest species did not
substantially alter our primary results, although it did weaken the forest loss × cover
effect on the likelihood of declining population trends (Extended Data Fig. 3b).
Range area. Species’ geographic range area is a key predictor of extinction risk,
and extent of occurrence and area of occupancy are two parameters used to assess
species under criterion B of the IUCN Red List. This can pose a circularity issue
for comparative extinction risk analyses, particularly those that attempt to assess
the effect of geographic range area relative to the effects of other predictors on
species endangerment60. A common remedy is to run the analysis on species
classified as Least Concern and those that are listed as Near Threatened or threat-
ened for reasons not directly linked to small geographic range area (that is, not
under criterion B)
60
. We followed this procedure as part of our sensitivity analysis.
Specifically, we excluded species listed as threatened under criterion B. Such
species made up 2,529 (approximately 58%) of the 4,396 threatened species in our
full dataset. The results (Extended Data Fig. 3c) show that our overall conclusions
are robust to the exclusion of these species.
Forest loss and cover threshold. In our primary analysis, we used the forest loss
and cover variables directly as given in Hansen et al.
6
. Forest cover is a continuous
variable ranging from 0% to 100% cover within each pixel and forest loss is a binary
variable indicating whether or not tree cover canopy had been completely removed
between 2000 and 2014. Since the effects of forest loss and cover on endangerment
(status/trends/uplisting) probably vary depending on the initial amount of forest
cover, we replicated our analyses, but truncated forest loss and cover at the 75%
threshold (Extended Data Fig. 3d). That is, we treated cover and loss as zero in
pixels that had less than 75% initial forest cover. This change did not influence our
results substantially (Extended Data Fig. 3d).
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter reSeArCH
Forest loss and gain standardization. The forest loss and gain variables in our
analysis can be thought of in terms of percentages of species’ ranges since they are
averages of spatial variables across species’ ranges. An alternative way to compute
the forest loss and gain variables is as percentages of forest cover within species’
ranges. We used these standardized loss and gain variables (that is, loss divided by
cover and gain divided by cover) as part of our sensitivity analysis (we similarly
standardized historical loss by dividing by potential cover), and found that their
use had little effect on our results (Extended Data Fig. 3e). This provides another
way of quantifying forest loss and gain, which may be particularly appropriate for
species that have little forest cover within their ranges. This was uncommon in our
core dataset as we focused on forest-optional and -exclusive species, that tend to
have high forest cover across their ranges.
Accounting for phylogeny. The models that we fit assume that the dependence
structure of the observations is purely spatial. However, this may not be valid as
species that are phylogenetically similar may be more likely to have the same status,
trend, or uplisting variable values, even after accounting for the covariates in the
models. To explore this issue of potential phylogenetic dependence and its effect
on our results, we fit generalized linear mixed models using glmer in the lme4
R package61, including random effects by taxonomic order (Extended Data
Fig. 3f). We were unable to fit more complex phylogenetic models that use full trees
(for example, phylogenetic logistic regression) because detailed phylogenetic data
are not available for many of the species in our analysis
62
. However, the addition of
taxonomic-based random effects did not substantially alter our results, suggesting
that the effects of phylogenetic dependence are weak after accounting for spatial
autocorrelation and the other predictors (Extended Data Fig. 3f).
Assessing sensitivity to resolution. We tested the sensitivity of the results to the
spatial resolution used in our analysis (2 decimal degree equivalent equal-area) by
re-computing the covariates (averages across species’ ranges) at a finer resolution
of approximately 5 km. In this analysis, we refined the species’ ranges by clipping
them using the species’ altitude limits coded on the IUCN Red List, when available
(6,047 of 19,615 species). We also excluded forest loss, gain, and cover inside of
known tree plantations using a map of plantations for seven tropical countries
63
.
Covariate averages at high resolution were calculated using Google Earth Engine.
Coefficient estimates show relatively low sensitivity to our choice of resolution,
clipping ranges by altitudinal limits, and masking out forest variables within known
plantati ons (Extended Data Fig. 3g).
Intermediate-term forest change. Our primary forest change variables are from
2000 to 2014. We also included a derived ‘historical forest cover’ variable to account
for long-term forest change. However, given that many species were listed in the
early period of our recent forest-loss data (or even before this), it would be ideal to
have contemporary forest loss data from before 2000. Unfortunately, no spatially
contiguous datasets exist for this period. Nevertheless, to extend the time span
for the more recent forest change variables, we added 1990–2000 forest loss and
gain estimates to the 2000–2014 estimates, producing estimates of loss and gain
for the period 1990–2014
19
. This summed dataset covered > 80% of the ranges for
only 58.7% of the species in our analyses. Using these data, the forest loss × cover
interaction term was weaker. However, consistent with our primary analyses, esti-
mates still tended to be positive for forest-optional and -exclusive species (Extended
Data Fig. 3h). It is likely that the smaller effect size estimates are related to uncer-
tainty in the 1990–2000 dataset caused by missing data (Extended Data Fig. 7).
Importantly, we found a high correlation between 1990–2000 and 2000–2014
forest loss at low levels of missing data, which suggests that locations of interme-
diate-term and recent forest loss are correlated at the scale of species’ ranges (there
is temporal autocorrelation in forest loss; Extended Data Fig. 7). This correlation
is further supported by the country-level correlations between 1990–2000 and
2000–2015 net forest loss (that is, change in percentage cover) obtained using the
Food and Agriculture Organization’s (FAO) Global Forest Resources Assessment
country-level data
64
(unweighted correlation 0.705, country land-area-weighted
correlation 0.805; Extended Data Fig. 8). This explains strong effects of forest loss
during the 2000–2014 period even though some species may not yet have fully
felt the effects of this most recent loss (or had their status updated accordingly).
Year of discover y. Newly described species are often from remote areas (that is,
with initial high forest cover) where development is starting to take place (dis-
covery was facilitated by access); such species are highly likely to be classed as
threatened
65
. To explore how time since initial species description influenced our
results, we conducted a sensitivity analysis including ‘year of species description’ as
a predictor. We gleaned year of description from the taxonomic authority sections
of Red List fact sheet accounts. For 18 of the species in our analysis, two adjacent
years were reported (for example, “Highton, 1971 (1972)”). In these cases, we used
the average of the two years. In addition to a main effect for year, we included
the three-way forest loss × forest cover × year interaction. This directly tests the
hypothesis that the initial intrusion effect (the statistical interaction between forest
loss and cover) is mediated by the time when a species was initially described,
with the expectation that most recently described species are more likely to show
such effects. However, there was little statistical support for this hypothesis; the
strength of the forest loss × forest cover interaction (our primary focus) was largely
unchanged (Extended Data Fig. 3i).
Threshold for threatened species. It is possible that species in different threat cat-
egories could respond in contrasting ways to forest loss. For instance, we expected
species listed as Endangered and Critically Endangered to be more likely to support
the habitat threshold hypothesis; these species only become extremely threatened
when forest continues to be lost at high rates after most original habitat has been lost.
Therefore, we tested effects of forest loss, forest amount and their interaction on suc-
cessive levels of IUCN threat categories (Extended Data Fig. 9). We compared model
results to those obtained when threatened species were taken to be Endangered or
Critically Endangered species and Critically Endangered species alone. Our overall
conclusions were consistent across threat categories (Extended Data Fig. 9).
Data availability. Data that support the findings of this study have been deposited
with figshare at: https://doi.org/10.6084/m9.figshare.4955465.v4.
35. Butchart, S. H. et al. Improvements to the red list index. PLoS ONE 2, e140 (2007).
36. Butchart, S. H. et al. Measuring global trends in the status of biodiversity: red
list indices for birds. PLoS Biol. 2, e383 (2004).
37. Homann, M. et al. The changing fates of the world’s mammals. Phil. Trans. R.
Soc. Lond. B 366, 2598–2610 (2011).
38. BirdLife International. IUCN Red List for birds. http://www.birdlife.org (2017).
39. Böhm, M. et al. The conservation status of the world’s reptiles. Biol. Conserv.
157, 372–385 (2013).
40. Wildlife Conservation Society and Center for International Earth Science
Information Network, Columbia University. Last of the Wild, v2: Global Human
Footprint Dataset (Geographic). http://dx.doi.org/10.7927/H4M61H5F (2005).
41. Venter, O. et al. Sixteen years of change in the global terrestrial human
footprint and implications for biodiversity conservation. Nat. Commun. 7,
12558 (2016).
42. Venter, O. et al. Data from: Global terrestrial Human Footprint maps for 1993
and 2009. http://dx.doi.org/10.5061/dryad.052q5 (2016).
43. ESRI. ArcGIS Desktop: Release 10.1 (Environmental Systems Research
Institute, 2012).
44. R Core Team. R: A Language and Environment for Statistical Computing.
http://www.R-project.org/ (2013).
45. Seligman, M. Rborist: extensible, parallelizable implementation of the random
forest algorithm. https://cran.r-project.org/web/packages/Rborist/index.html
(2015).
46. Hurlbert, A. H. & Jetz, W. Species richness, hotspots, and the scale dependence
of range maps in ecology and conservation. Proc. Natl Acad. Sci. USA 104,
13384–13389 (2007).
47. Augustin, N. H., Mugglestone, M. A. & Buckland, S. T. An autologistic model for
the spatial distribution of wildlife. J. Appl. Ecol. 33, 339–347 (1996).
48. Dormann, C. F. Assessing the validity of autologistic regression. Ecol. Modell.
207, 234–242 (2007).
49. Bivand, R. et al. spdep: spatial dependence: weighting schemes, statistics and
models. https://cran.r-project.org/web/packages/spdep/index.html (2017).
50. Pike, N. Using false discovery rates for multiple comparisons in ecology and
evolution. Methods Ecol. Evol. 2, 278–282 (2011).
51. Benjamini, Y. & Hochberg, Y. Controlling the false discovery rate: a practical and
powerful approach to multiple testing. J. R. Stat. Soc. B 57, 289–300 (1995).
52. Swets, J. A. Signal Detection Theory and ROC Analysis in Psychology and
Diagnostics: Collected Papers (Psychology Press, 2014).
53. United Nations. United Nations Statistics Division. Standard Country and Area
Codes Classications (M49). http://unstats.un.org/unsd/methods/m49/
m49regin.htm (2013).
54. Bahn, V. & McGill, B. J. Testing the predictive performance of distribution
models. Oikos 122, 321–331 (2013).
55. Gilleland, M. E. verication: weather forecast verication utilities. https://
cran.r-project.org/web/packages/verication/index.html (2015).
56. Dray, S., Legendre, P. & Peres-Neto, P. R. Spatial modelling: a comprehensive
framework for principal coordinate analysis of neighbour matrices (PCNM).
Ecol. Modell. 196, 483–493 (2006).
57. Lee, D. CARBayes: an R package for Bayesian spatial modelling with
conditional autoregressive priors. J. Stat. Softw. 55, 13 (2013).
58. McMillen, D. McSpatial: nonparametric spatial data analysis. https://
cran.r-project.org/web/packages/McSpatial/index.html (2013).
59. Klier, T. & McMillen, D. P. Clustering of auto supplier plants in the United States.
J. Bus. Econ. Stat. 26, 460–471 (2012).
60. Cardillo, M. et al. Multiple causes of high extinction risk in large mammal
species. Science 309, 1239–1241 (2005).
61. Bates, D. M. lme4: Mixed-eects modeling with R. http://lme4.r-forge.r-project.
org/book (2010).
62. Ives, A. R. & Garland, T. Jr. Phylogenetic logistic regression for binary
dependent variables. Syst. Biol. 59, 9–26 (2010).
63. Petersen, R. et al. Mapping Tree Plantations with Multispectral Imagery: Preliminary
Results for Seven Tropical Countries. (World Resources Institute, 2016).
64. MacDicken, K. G. Global Forest Resources Assessment 2015: what, why and
how? For. Ecol. Manage. 352, 3–8 (2015).
65. Giam, X. et al. Reservoirs of richness: least disturbed tropical forests are centres
of undescribed species diversity. Proc. R. Soc. Lond. B 279, 67–76 (2012).
66. Hijmans, R. J., Cameron, S. E., Parra, J. L., Jones, P. G. & Jarvis, A. Very high
resolution interpolated climate surfaces for global land areas. Int. J. Climatol.
25, 1965–1978 (2005).
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter
reSeArCH
Extended Data Figure 1 | Receiver operating characteristic (ROC)
curves for the models predicting status of forest exclusive species. Class
was included as a fixed effect (as in our main results) for the ‘All species’
group. The other results (by class) are based on models fit to each class
separately. The left column is based on results where the model was fit to
the entire dataset. The right column shows ROC curves for predictions
using a fourfold cross-validation scheme where the probability of species
being threatened was predicted for each of four regions with the model
fit using data from all other regions. P values are based on the Mann–
Whitney U statistic and test whether the population AUC is greater than
0.5 (that is, better than random predictions). Results are presented both
with (bottom row) and without (top row) the spatial autocovariate.
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter reSeArCH
Extended Data Figure 2 | Model results for models fit by class (mammals, amphibians, birds) and for all classes together (All). Each row shows
standardized coefficient estimates and 95% confidence intervals (as error bars) for each single model. All covariates are shown in this figure.
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter
reSeArCH
Extended Data Figure 3 | Sensitivity analysis results. The plotted
variable is the estimated standardized coefficient for the forest loss × cover
term with 95% confidence interval (as error bars). Each column
corresponds to a different sensitivity analysis (other covariates are not
shown). a–i, In general, we found that our primary results were robust to
the inclusion of absolute latitude as a predictor variable (a), the restriction
of the dataset to tropical species only (b), the exclusion of species listed as
threatened based on small geographic range (c), using a 75% pixel-scale
threshold for the forest loss and forest cover variables (d), standardizing
forest loss and gain by forest cover (that is, dividing forest loss and gain
by forest cover so that these variables can be interpreted as approximate
percentages of species’ forested range) (e), accounting for potential
phylogenetic dependence using generalized linear mixed models with
random intercepts by taxonomic order (and by class for the ‘all species’
model) (f), using high-resolution species’ range maps and covariate
maps (approximately 5 km), clipping species ranges based on altitudinal
limits, and setting forest loss and cover to zero in regions of known tree
plantations (g), including forest loss and gain from 1990–2000 by adding
1990–2000 and 2000–2014 forest change variables (h), and the inclusion
of year of initial species description as a main effect and in a three-way
interaction term with forest loss × cover (i).
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter reSeArCH
Extended Data Figure 4 | Estimated standardized coefficients for
each model term (with 95% confidence intervals as error bars) when
a quadratic forest loss × cover2 interaction (forest loss × cover2)
is included in the model. This allows for the effect of loss to vary
quadratically with cover. A significant and positive forest loss × cover2
interaction term would suggest that the (negative) effects of forest loss
are greatest in areas with both high and low proportions of forest cover.
However, this term was non-significant for most taxa and response
variables, indicating that the linear model for the interaction is more
parsimonious.
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter
reSeArCH
Extended Data Figure 5 | The effect of forest loss (for 2% additional
loss) in relation to total forest cover using quadratic models. These
models allow the effect of forest loss to vary nonlinearly as a function
of forest cover, allowing us to test the hypothesis that forest loss is
detrimental to species at both high and low levels of forest cover. However,
the quadratic model reveals very similar results to the linear model.
The exception is when ‘declining trend’ is used as the response; species’
populations were more likely to be in decline when forest amount is very
low (the habitat threshold hypothesis), and upon initial intrusion into
intact forests (the initial intrusion hypothesis). For statistical significance
of the quadratic models, see confidence intervals in Extended Data
Fig. 4, far right panel. For context, the histograms (grey bars) show the
(normalized to maximum 100%) distributions of forest cover across
species. For example, if one bar in a panel is twice as high as another, then
twice as many species have average forest cover of this percentage in their
ranges. The black lines show the cumulative percentages of species with
at most x per cent forest cover. For example, approximately half of forest-
optional species have 50% forest cover or less.
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter reSeArCH
Extended Data Figure 6 | Results of multiple spatial models (estimates
and 95% confidence intervals as error bars) for forest exclusive species
when status (that is, whether or not a species is threatened) is used as
the response. Coefficients across multiple models that account for spatial
autocorrelation were very similar. ‘Method’ indicates the procedure
(if any) used to account for spatial autocorrelation: non-spatial ordinary
GLM (non_spatial), autologistic model with spatial autocovariate (AL_b),
autologistic model using 50 nearest neighbours in the spatial weights
matrix (AL_b_50), Moran eigenvector filtering (filtering), spatial
autoregressive model (SAR_approx), or Bayesian condition autoregressive
model (CAR_Bayes). Details on each method are given in the sensitivity
analyses section of the Methods.
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter
reSeArCH
Extended Data Figure 7 | Relationship between forest loss 1990–2000
(from ref. 34) and 2000–2014 (from ref. 7). Overall, rates of forest loss
are temporally autocorrelated; species ranges with high forest loss in
the 1990s also show high forest loss in 2000s. However, this relationship
is strongly affected by data availability; approximately 12.1% of forest
loss data are missing across the globe and as we expected, the more data
missing from a species range, the weaker the relationship between 1990s
and 2000s rates of forest loss. The plots show correlations (in red; top right
of each panel) between forest loss across the two time periods for various
levels of missing data. Each point corresponds to a single species and the x
and y axis values indicate average values of each variable across its range.
Panel titles show the proportion of missing 1990–2000 forest loss data in
species ranges. For example, the top left panel contains results for species
with between 0% and 4% of their ranges missing 1990 forest data (owing
to clouds, lack of satellite coverage, and so on). The correlation between
1990–2000 and 2000–2014 forest loss is highest for species with the least
missing data.
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter reSeArCH
Extended Data Figure 8 | Country-level forest net loss (that is, change
in percentage forest cover) for the 1990–2000 and 2000–2015 periods
according to the Food and Agriculture Organization’s (FAO) Global
Forest Resources Assessment. Based on these data, the correlation between
1990–2000 and 2000–2015 forest loss is 0.705. Weighting by country area
increases the correlation to 0.805. The relatively high correlation suggests
that the spatially explicit recent (2000–2014) forest loss data that we used is
closely related to less recent (1990–2000) forest loss.
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter
reSeArCH
Extended Data Figure 9 | Sensitivity of our results to alternative
categories of threat. In the main text we considered a species to be
‘threatened’ if it fell into the IUCN Red List category Vulnerable,
Endangered or Critically Endangered. We conducted further analysis
considering as threatened only species that are Endangered and Critically
Endangered, and again for only species that are Critically Endangered.
Dots show estimated standardized coefficients for each model term (with
95% confidence intervals as error bars) for all main effects and the forest
loss × cover interaction term. Our overall conclusions were consistent
across these different definitions of threat.
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.

letter reSeArCH
Extended Data Figure 10 | Maps showing the methods used to quantify
historical forest loss. First, we used random forests (a machine-learning
method) to estimate potential forest cover globally (within forest
biomes33). a–c, This model was fit using current forest cover within intact
forest landscapes36 and bioclimatic and other predictor variables66 (a; see
Methods). We then subtracted current forest cover (b; Hansen et al.6) from
this map to obtain estimated historical forest loss (c). The map of land is
taken from http://thematicmapping.org/downloads/world_borders.php.
© 2017 Macmillan Publishers Limited, part of Springer Nature. All rights reserved.