1 Spending limited resources on de-extinction could lead to net biodiversity loss
2 Joseph R. Bennett1, Richard F. Maloney2, Tammy E. Steeves3, James Brazill-Boast4, Hugh P. 3 Possingham5,6, Phillip J. Seddon7
4 1. Department of Biology 5 Carleton University 6 1125 Colonel By Dr 7 Ottawa, ON K1S 5B6 8 Canada 9 [email protected] 10 ph: 613-520-2600 x 3124 11 fax: 613-520-3539 12 13 2. Science and Policy Group 14 Department of Conservation 15 70 Moorhouse Ave 16 Addington 17 Christchurch 8011 18 New Zealand 19 20 3. School of Biological Sciences 21 University of Canterbury 22 Private Bag 4800 23 Christchurch 8140 24 New Zealand 25 26 4. NSW Office of Environment and Heritage 27 59 Goulburn St 28 Sydney, Australia 29 30 5. University of Queensland 31 ARC Centre of Excellence for Environmental Decisions 32 School of Biological Sciences 33 St Lucia, QLD, Australia 34 35 6. Conservation Science 36 The Nature Conservancy 37 245 Riverside Drive 38 West End, QLD 4101 39 Australia 40 41 7. University of Otago 42 Department of Zoology 43 340 Great King Street 44 Dunedin 9016 45 New Zealand
46
1
47 There is contentious debate surrounding the merits of de-extinction as a biodiversity
48 conservation tool. Here, we use extant analogues to predict conservation actions for
49 potential de-extinction candidate species from New Zealand and New South Wales, and
50 use a prioritization protocol to predict the impacts of reintroducing and maintaining
51 populations of these species on conservation of extant threatened species. Even using the
52 optimistic assumptions that resurrection of species is externally sponsored, and that
53 actions for resurrected species can share costs with extant analogue species, public
54 funding for conservation of resurrected species would lead to fewer extant species that
55 could be conserved, suggesting net biodiversity loss. If full costs of establishment and
56 maintenance for resurrected species populations were publicly funded, there could be
57 substantial sacrifices in extant species conservation. If conservation of resurrected
58 species populations could be fully externally sponsored, there could be benefits to extant
59 threatened species. However, such benefits would be outweighed by opportunity costs,
60 assuming such discretionary money could directly fund conservation of extant species.
61 Potential sacrifices in conservation of extant species should be a crucial consideration in
62 deciding whether to invest in de-extinction or focus our efforts on extant species.
63 Technological advances are reducing the barriers to resurrecting extinct species or their close
64 genetic proxies, allowing de-extinction to be considered as a biodiversity conservation tool1,2.
65 Arguments in favour of de-extinction include necessity, driven by the rapid rate of species
66 and habitat loss3,4, an ethical duty to redress past mistakes5, as well as potential technological
67 and ecological knowledge that could stem from de-extinction programs4. Counter-arguments
68 include high risk of failure due to difficulties of cloning for some species6, technical risks
69 inherent in re-introductions7,8,9, loss of culture in resurrected animal species8, and lack of
70 remaining habitat for some species10,11, as well as negative consequences for extant species
2
71 including reduced incentive for traditional conservation12, and ecological impacts of
72 introducing long-absent or genetically-modified species12.
73 The relative cost versus benefit for biodiversity is fundamental to the debate surrounding de-
74 extinction. Assuming species are resurrected to be released into former habitats, the cost of
75 de-extinction includes the process of producing initial founder populations, translocating
76 individuals, then monitoring and managing new wild populations. If conservation funds are
77 re-directed from extant to resurrected species, there is risk of perverse outcomes whereby net
78 biodiversity might decrease as a result of de-extinction12,13. Although private agencies might
79 fund the resurrection of extinct species out of technical or philanthropic interest, the
80 subsequent ongoing management of such species (many of which would face the same threats
81 that made them extinct) would fall upon government agencies as commonly occurs with
82 extant threatened species. Alternatively, if private agencies are willing to provide new
83 funding for post de-extinction management, there could be additional benefits to species
84 sharing habitats or threats.
85 Here, we test the potential impact of establishing and sustaining wild populations of
86 resurrected extinct species (or proxies of such species) on the conservation of extant species.
87 Specifically, we use long-term conservation programs for extant analogue species in New
88 Zealand (NZ) and the Australian state of New South Wales (NSW), to infer potential
89 conservation actions for resurrected species, and predict the impact of resurrected species
90 programs on conservation of extant species. We use these datasets because they contain
91 detailed prescriptions and costs of actions designed to achieve population recovery for most
92 of the extant threatened species requiring specific management actions in either jurisdiction.
93 We estimate the net number of extant species that can be conserved, using the following
94 scenarios: 1) resurrected species become the burden of government conservation programs;
95 and 2) establishment and maintenance of resurrected species populations are funded
3
96 externally using non-public resources. In Scenario 1 the use of government resources on
97 resurrected species results in less funding for extant species programs, but provides potential
98 benefits for species that share actions with resurrected species. In Scenario 2 there are also
99 potential benefits to extant species conservation programs through shared conservation
100 actions. However, there are potential opportunity costs, if private agencies use resources they
101 could otherwise have used on conservation of extant species. Our analysis assumes that
102 species would be resurrected to be re-introduced into their former habitats, rather than for
103 other potential reasons such as research or public display.
104 Because little is known about the costs of producing viable initial populations of resurrected
105 species, we do not consider this in our analysis, and assume it is covered by a private agency.
106 Instead, we focus on the long-term cost of conservation for resurrected species, assuming that
107 such species would have small founder populations that require conservation actions similar
108 to those required for extant threatened species.
109 Although there is considerable uncertainty regarding necessary conservation actions for many
110 extinct species should they be resurrected7, we assume that such species would share many
111 actions with closely related extant species that share habitats, threats, and ecological roles.
112 We therefore chose focal extinct species from among the endemic, fully extinct species from
113 our study areas, based on similarity in taxonomy, range, habitat, life history, and threats with
114 an extant threatened analogue species. Among 70 recently extinct (1000 AD to present)
115 species in NZ, we found 11 for which we could assign reasonable analogues (Supplementary
116 Information Table 1). For NSW, we considered 29 recently extinct species, and found 5 with
117 reasonable extant analogues. Our inferred conservation programs for the extinct species
118 (assuming they were resurrected), were the same as for their analogue extant species, with the
119 addition of captive breeding and translocation costs, based on average costs of captive
120 breeding and translocation from extant species of the same taxonomic group (e.g. bird,
4
121 amphibian). Although cost and shared actions were not criteria for choosing our focal
122 species, our chosen group represented a broad range of estimated costs and number of extant
123 species with shared actions (Supplementary Information Table 1). We note that using
124 analogues in this way likely underestimates both cost and risk of failure, and that we assume
125 actions could be completely shared between resurrected and extant analogue species. It is
126 somewhat unlikely that an effective conservation program for a resurrected species would
127 completely share actions with that of an extant species. We also note that our analysis makes
128 the largely untested assumption that technical barriers to creating initial populations of these
129 species can be overcome6,14. Thus, our results should be regarded as being optimistic in
130 favour of net benefits of resurrected species conservation programs.
131 To assess the potential influence of resurrected species on extant species conservation, we
132 incorporated the proposed programs for resurrected species into threatened species project
133 prioritization protocols developed for the New Zealand Department of Conservation (NZ
134 DOC) and NSW Office of Environment and Heritage (see Methods for details). Costs of
135 shared actions (e.g. predator control that benefits several species sharing a site) were shared
136 among prioritized species recovery projects. Thus, if private funding covers the cost of
137 actions for a resurrected species, the cost of the same actions for any other species (including
138 the resurrected species’ analogue) would also be covered, potentially allowing more species
139 to be conserved within a given budget.
140 In Scenario 1 where resurrected species become the burden of governments, we subtracted
141 the budgets for resurrected species conservation programs from realistic baseline budgets for
142 NZ ($30M NZD15,16) and NSW ($4.65M AUD17), and set the cost of any specific actions that
143 were shared in location and time with other species (e.g. predator fence on a shared habitat
144 patch) to zero. We compared the number of extant species that could be prioritized for
145 funding in this scenario with the number of extant species that would normally be prioritized
5
146 with the same baseline budget. We did this for each resurrected species considered
147 individually (cost of only one focal species subtracted), as well as all resurrected species (11
148 for NZ and five for NSW) considered together.
149 For Scenario 2, where resurrected species programs are entirely externally sponsored, we
150 determined potential benefits for conservation of extant species by setting the cost of any
151 shared actions between resurrected and extant species to zero, and re-ran the prioritization
152 algorithms with our baseline budgets. To determine the opportunity cost associated with this
153 scenario, we added the cost of actions for resurrected species’ sponsorship programs to the
154 baseline budget for extant species prioritization, then determined the number of species that
155 would be prioritized for conservation if such funding could be used on extant instead of
156 resurrected species.
157 For both NZ and NSW species, the number of extant species that could be prioritized for
158 conservation is generally lower in Scenario 1, where resurrected species become the burden
159 of the government (Fig. 1, red bars). This suggests a potential long-term net loss in
160 biodiversity if conservation efforts are shifted towards resurrected species. For NZ, there
161 were potential net gains associated with a single resurrected species, Coenocorypha
162 chathamica. This is because the conservation prescription for this species contained many
163 shared actions with 39 extant species that inhabit its former habitat on Chatham Island.
164 Shared costs for some of these species allowed more to be prioritized than in the baseline
165 scenario. However, for NSW the estimated conservation costs for two extinct species are
166 greater than the most recent baseline budget estimate, suggesting that the government budget
167 would have to be drastically increased if conservation of either species were publically
168 funded. Given that the NZ and NSW algorithms are designed to efficiently conserve species
169 using limited resources, and to account for shared costs, it is possible that the impact of
6
170 government payment for resurrected species programs could be greater in other jurisdictions
171 where spending is less efficient.
172 In Scenario 2 where the conservation costs of resurrected species are covered by an external
173 agency, lowered costs of shared actions would allow more extant species to be prioritized for
174 conservation (Fig. 1, yellow bars). However, the potential biodiversity benefits are
175 outweighed by the opportunity costs of not applying the same funding to extant species (Fig.
176 1, blue bars).
177 For both scenarios, including the costs of all resurrected species together amplified results
178 (Supplementary Information Table 2). For example, government-funded conservation for all
179 11 focal extinct species in NZ would sacrifice conservation for nearly three times as many
180 (31) extant species. External funding for conservation of the five focal extinct NSW species
181 could instead be used to conserve over eight times as many (42) extant species.
182 Including the costs of producing viable initial populations of these species would likely have
183 greatly increased our estimates of sacrifices in extant species conservation. The cost of such
184 programs is difficult to project, but likely to be substantial. A program to use stem cell
185 technology and surrogates to prevent the extinction of the northern white rhinoceros (which
186 would have fewer technical hurdles than resurrecting species from only preserved materials)
187 has been estimated to cost several million dollars2. In addition, our analyses make the
188 generous assumption that we would have perfect analogues for the extinct species, such that
189 conservation programs for resurrected and extant species would share costs. Breeding and
190 husbandry of resurrected species before and after reintroduction could well be more
191 expensive and more prone to failure than for extant species, because we typically know less
192 about behaviour and physiology of extinct species. Reintroduction of locally extirpated
7
193 species would likely have considerably lower risk and costs, given better knowledge of the
194 ecology and physiology of such species.
195 Debates regarding the merits of de-extinction tend to centre on either ethical or biological
196 arguments. Ethical arguments often focus on the potential of de-extinction to right past
197 wrongs, versus the ‘moral hazard’ arising from diminished motivation to conserve extant
198 species, if it is assumed that extinctions can be reversed sometime in the (potentially distant)
199 future12. Biological arguments often focus on the relative benefit to biodiversity. For
200 example, there may be conservation gains through applying technical lessons learned in the
201 process of attempting de-extinctions4. There could also be gains through restoration of
202 ecosystem processes that were provided by the extinct species18. For example, extinct
203 ‘ecosystem engineers’ such as woolly mammoths or passenger pigeons could potentially be
204 resurrected in attempts to restore their lost functional roles19. In addition, resurrected species
205 could act as ‘flagships’ to promote conservation5, and potentially increase resources
206 management of extant threatened species.
207 However, there is considerable risk in assuming that resurrected species would fill these
208 intended roles. Resurrected ecosystem engineers would be introduced into environments that
209 have been much altered by humans, and they could fail to thrive in these new
210 circumstances7,19. Resurrecting populations large enough for such species to fill their former
211 roles could also prove very challenging19. Conversely, there may be biodiversity losses if
212 resurrected species become invasive or spread disease12,20,21. Experience with extant iconic
213 species also suggests a high risk that iconic species resurrected as ‘flagships’ could draw
214 resources away from programs for extant species22, or even create self-reinforcing biases
215 whereby the public profile of resurrected species and resources spent on them would
216 synergistically increase, at the expense of non-iconic extant species23.
8
217 More fundamentally, de-extinction could lead to either biodiversity gains via resurrection of
218 the extinct species themselves and shared conservation actions with extant species, or to
219 losses through missed opportunities to allocate resources to extant species. Conservation
220 resources are scarce24, necessitating careful allocation of funds25. Our analysis strongly
221 suggests that resources expended on long-term conservation of resurrected species could
222 easily lead to net biodiversity loss, compared to spending the same resources on extant
223 species. If costs and risks of failure associated with establishing viable populations could also
224 be calculated, estimates of potential net losses or missed opportunities would likely be
225 considerably higher. Given this considerable potential for missed opportunity, as well as the
226 risks inherent in assuming a resurrected species would fulfil its role as an ecosystem engineer
227 or flagship species, it is unlikely that de-extinction could be justified on grounds of
228 biodiversity conservation.
229 Methods:
230 Focal extinct species were chosen based on taxonomic relatedness as well as having ranges,
231 habitats, threats, and life-history strategies shared with extant analogue species. We chose
232 fully extinct species (i.e., not locally extirpated) that went extinct after 1000 AD, assuming
233 that feasibility of resurrection (e.g. availability of genetic material, knowledge of life history
234 and physiology) would be prohibitively low for species that went extinct before this time.
235 However, we did not assess the availability of genetic material in our focal species, nor
236 consider the feasibility or cost of producing viable initial populations.
237 Species Conservation Projects and Prioritization Algorithms
238 Species conservation projects for all species in the NZ and NSW datasets (including the
239 analogue species) were determined using information gathered from threatened species
240 experts (>100 experts for NZ, ~250 for NSW). The projects include the specific actions
9
241 (including specific location, timing and cost) considered necessary to ensure ~95%
242 probability of each species’ persistence over 50 years.
243 The prioritization algorithms rank species by the cost-effectiveness of their conservation
244 projects, using the following equation:
× 245 = ,
246 where Ei is the cost effectiveness of the conservation project for species i; Bi is the benefit of
247 the project to the species, defined as the difference between estimated probabilities that a
248 species will be secure in 50 years with and without the project; Si is the estimated probability
249 of success for the conservation project; and Ci is the total cost of all actions across all sites for
250 a species’ project. The NZ algorithm uses an additional parameter that estimates a species’
251 evolutionary distinctiveness (see ref. 15 for details). Costs of actions are shared among
252 prioritized species recovery projects. For example, the cost of predator control at a site that
253 benefits two prioritized species sharing the site is reduced by 50% for each of the two
254 species.
255 The algorithm begins with all species ranked, then eliminates the lowest-ranked species
256 sequentially until the set first year budget is reached. As species are removed from the ranks,
257 cost sharing is updated for remaining species. Species that are no longer prioritized no longer
258 share costs with remaining species. Additional details regarding the algorithm are found in
259 refs. 15,26.
260 Data Availability The data and code for the NZ Prioritization Protocol have been deposited
261 in the Dryad Digital Repository (http://dx.doi.org/10.5061/dryad.3qn55). The data and code
262 for the NSW Prioritization protocol have been deposited in the Dryad Digital Repository
263 (http://dx.doi.org/10.5061/dryad.p86t5).
10
264 References
265 1. Folch, J. et al. First birth of an animal from an extinct subspecies (Capra pyrenaica
266 pyrenaica) by cloning. Theriogenology 71, 1026-1034 (2009).
267 2. Callaway, E. Stem-cell plan aims to bring rhino back from brink of extinction.
268 Nature 533, 20 (2016).
269 3. Kumar, S. Extinction need not be forever. Nature 492, 9 (2012).
270 4. Sherkow, J. S. & Greely, H. T. What if extinction is not forever? Science 340, 32-33
271 (2013).
272 5. Cohen, S. The ethics of de-extinction. NanoEthics 8, 165-178 (2014).
273 6. Shapiro, B. Mammoth 2.0: will genome engineering resurrect extinct species? Genome
274 Biol. 16, 228 (2015).
275 7. Seddon, P. J., Moehrenschlager, A. & Ewen, J. Reintroducing resurrected species:
276 selecting DeExtinction candidates. Trends Ecol. Evol. 29, 140-147 (2014).
277 8. Jones, K. E. From dinosaurs to dodos: who could and should we de-extinct? Front.
278 Biogeog. 6, 20-24 (2014).
279 9. Weeks, A. R. et al. Assessing the benefits and risks of translocations in changing
280 environments: a genetic perspective. Evol. Appl. 4, 709-725 (2011).
281 10. Jenkins, C. N., Van Houtan, K. S., Pimm, S. L. & Sexton, J. O. US protected lands
282 mismatch biodiversity priorities. Proc. Natl. Acad. Sci. USA 112, 5081-5086 (2015).
283 11. Peers, Michael J. L. et al. De-extinction potential under climate change: Extensive
284 mismatch between historic and future habitat suitability for three candidate birds. Biol.
285 Cons. 197, 164-170 (2016).
286 12. IUCN Species Survival Commission. IUCN/SSC Guiding Principles on Creating Proxies
287 of Extinct Species for Conservation Benefit, V 1.0. (International Union for Conservation
288 of Nature (IUCN), 2016).
11
289 13. Camacho, A. E. Going the Way of the Dodo: De-Extinction, Dualisms, and Reframing
290 Conservation. Wash. Univ. Law Rev. 92, 849-906 (2015).
291 14. Shapiro, B. Pathways to de-extinction: How close can we get to resurrection of an extinct
292 species? Funct. Ecol., in press. DOI: 10.1111/1365-2435.12705
293 15. Bennett, J. R. et al. Balancing phylogenetic diversity and species numbers in conservation
294 prioritization, using a case study of threatened species in New Zealand. Biol. Cons. 174,
295 47-54 (2014).
296 16. Tulloch, A. I. T. et al. Effect of risk aversion on prioritizing conservation projects. Cons.
297 Biol. 29, 513-524 (2015).
298 17. New South Wales Office of Environment and Heritage. More plants and animals to be
299 saved from extinction: Saving our Species 2016-2021. (Government of New South Wales,
300 2016).
301 18. Wood, J. R. et al. Resolving lost herbivore community structure using coprolites of four
302 sympatric moa species (Aves: Dinornithiformes). Proc. Natl. Acad. Sci. USA 110, 16910-
303 16915 (2013).
304 19. McCauley, D. J., Hardesty‐Moore, M., Halpern, B. S., & Young, H. S. A mammoth
305 undertaking: harnessing insight from functional ecology to shape de‐extinction priority
306 setting. Funct. Ecol. (2016). DOI: 10.1111/1365-2435.12728.
307 20. Minteer, B. A. Is it right to reverse extinction? Nature 509, 261 (2014).
308 21. Nogués-Bravo, D., Simberloff, D., Rahbek, C. & Sanders, N. J. Rewilding is the new
309 Pandora’s box in conservation. Current Biol. 26, R87-R91 (2016).
310 22. Restani, M., & Marzluff, J. M. Funding Extinction? Biological Needs and Political
311 Realities in the Allocation of Resources to Endangered Species Recovery. BioSci., 52,
312 169-177. (2002).
12
313 23. Martín-López, B., Montes, C., Ramírez, L., & Benayas, J. What drives policy decision-
314 making related to species conservation? Biol. Cons., 142, 1370-1380. (2009).
315 24. McCarthy, D. P. et al. Financial costs of meeting global biodiversity conservation targets:
316 current spending and unmet needs. Science 338, 946-949 (2012).
317 25. Wilson, K. A., Carwardine, J. & Possingham, H. P. Setting conservation priorities. Ann.
318 NY Acad. Sci. 1162, 237-264 (2009).
319 26. Joseph, L. N., Maloney, R. F. & Possingham, H. P. Optimal allocation of resources
320 among threatened species: a project prioritization protocol. Cons. Biol. 23, 328-338
321 (2009).
322
323 Supplementary Information is available in the online version of the paper.
324 Acknowledgements JRB was supported the Natural Science and Engineering Research
325 Council of Canada (NSERC) and the Australian Research Council (ARC) Centre of
326 Excellence for Environmental Decisions (CEED); HPP was funded by an ARC Laureate
327 Fellowship and CEED.
328 Author Contributions JRB, RFM and PJS designed the study. JRB, RFM and JBB analyzed
329 the data. JRB wrote the paper, with input from all other authors.
330 Author Information The authors declare no competing financial interests. Readers are
331 welcome to comment on the online version of the paper. Correspondence and requests for
332 materials should be addressed to JRB ([email protected]).
13
333 Figure Caption:
334 Fig. 1 – Mean differences in number of extant species prioritized for conservation from a
335 baseline prioritization of extant species (vertical dashed line), for New Zealand (a) and New
336 South Wales (b). Red bars represent Scenario 1, where conservation of resurrected species
337 becomes the responsibility of government. Yellow bars represent Scenario 2, where
338 conservation of resurrected species is externally sponsored. Blue bars represent the number of
339 extant species that could be prioritized for conservation if funding for conservation costs of
340 resurrected species could instead be applied to extant species. Thus, they represent the
341 opportunity costs associated with Scenario 2. Error bars represent standard errors. Note that
342 Scenario 1 costs for two species in NSW are higher than the set government budget, so mean
343 differences could not be measured.
344
14