Philippa Kaur

Possingham, H. P., Wilson, K. A., Andelman, S. J. & Vynne, C. H. Protected areas. Goals, limitations, and design. In Principles of Conservation Biology (eds Groom, M. J. et al.) 507–549 (Sinauer Associates Inc, 2006).
Google Scholar 
Marquet, P. A., Lessmann, J. & Shaw, M. R. Protected-area management and climate change. In Biodiversity and Climate Change: Transforming the Biosphere (eds Lovejoy, T. E. & Hannah, L.) 283–293 (Yale University Press, 2019).Chapter 

Google Scholar 
Geldmann, J., Manica, A., Burgess, N. D., Coad, L. & Balmford, A. A global-level assessment of the effectiveness of protected areas at resisting anthropogenic pressures. PNAS https://doi.org/10.1073/pnas.1908221116 (2019).Article 

Google Scholar 
Potapov, P. et al. The last frontiers of wilderness: Tracking loss of intact forest landscapes from 2000 to 2013. Sci. Adv. 3, e1600821 (2017).Article 
ADS 

Google Scholar 
Cazalis, V. et al. Effectiveness of protected areas in conserving tropical forest birds. Nat. Commun. 11, 4461 (2020).Article 
ADS 
CAS 

Google Scholar 
Dudley, N., Mansourian, S., Stolton, S. & Suksuwan, S. Do protected areas contribute to poverty reduction?. Biodiversity 11, 5–7 (2010).Article 

Google Scholar 
Dudley, N. & Stolton, S. Arguments for Protected Areas (Earthscan, 2010).Book 

Google Scholar 
CBD. Strategic Plan for Biodiversity 2011–2020, Including Aichi Biodiversity Targets. http://www.cbd.int/sp/ and http://www.cbd.int/decision/cop/?id=12268 (2010).UNEP-WCMC & IUCN. Protected Planet: The World Database on Protected Areas (WDPA). www.protectedplanet.net. Accessed October 2022 (2022).Watson, J. E. M. et al. Persistent disparities between recent rates of habitat conversion and protection and implications for future global conservation targets. Conserv. Lett. 9, 413–421 (2016).Article 

Google Scholar 
Díaz, S. et al. Summary for Policymakers of the IPBES Global Assessment Report on Biodiversity and Ecosystem Services. (2019).Barnes, M. D., Glew, L., Wyborn, C. & Craigie, I. D. Prevent perverse outcomes from global protected area policy. Nat. Ecol. Evol. 2, 759–762 (2018).Article 

Google Scholar 
Visconti, P. et al. Protected area targets post-2020. Science 364, 239–241 (2019).Article 
ADS 
CAS 

Google Scholar 
Kukkala, A. S. & Moilanen, A. Core concepts of spatial prioritisation in systematic conservation planning. Biol. Rev. 88, 443–464 (2013).Article 

Google Scholar 
Joppa, L. N. & Pfaff, A. High and Far: Biases in the location of protected areas. PLoS ONE 4, e8273 (2009).Article 
ADS 

Google Scholar 
Maxwell, S. L. et al. Area-based conservation in the twenty-first century. Nature 586, 217–227 (2020).Article 
ADS 
CAS 

Google Scholar 
CBD. CoP 7 Decision VII/30. Strategic Plan: Future Evaluation of progress. 12 https://www.cbd.int/doc/decisions/cop-07/cop-07-dec-30-en.pdf (2004).Venter, O. et al. Bias in protected-area location and its effects on long-term aspirations of biodiversity conventions. Conserv. Biol. 32, 127–134 (2017).Article 

Google Scholar 
Kuempel, C. D., Chauvenet, A. L. M. & Possingham, H. P. Equitable representation of ecoregions is slowly improving despite strategic planning shortfalls. Conserv. Lett. 9, 422–428 (2016).Article 

Google Scholar 
Barr, L. M., Watson, J. E. M., Possingham, H. P., Iwamura, T. & Fuller, R. A. Progress in improving the protection of species and habitats in Australia. Biol. Conserv. 200, 184–191 (2016).Article 

Google Scholar 
Hoffmann, S., Irl, S. D. H. & Beierkuhnlein, C. Predicted climate shifts within terrestrial protected areas worldwide. Nat. Commun. 10, 1–10 (2019).Article 

Google Scholar 
Hannah, L. Protected areas and climate change. Ann. N. Y. Acad. Sci. 1134, 201–212 (2008).Article 
ADS 

Google Scholar 
Thomas, C. D. & Gillingham, P. K. The performance of protected areas for biodiversity under climate change. Biol. J. Lin. Soc. 115, 718–730 (2015).Article 

Google Scholar 
Ramirez-Villegas, J. et al. Using species distributions models for designing conservation strategies of Tropical Andean biodiversity under climate change. J. Nat. Conserv. 22, 391–404 (2014).Article 

Google Scholar 
Bax, V. & Francesconi, W. Conservation gaps and priorities in the Tropical Andes biodiversity hotspot: Implications for the expansion of protected areas. J. Environ. Manage. 232, 387–396 (2019).Article 

Google Scholar 
Jenkins, C. N., Pimm, S. L. & Joppa, L. N. Global patterns of terrestrial vertebrate diversity and conservation. PNAS 110, E2602–E2610 (2013).Article 
ADS 
CAS 

Google Scholar 
Rodrigues, A. S. L. et al. Global gap analysis: Priority regions for expanding the global protected-area network. Bioscience 54, 1092–1100 (2004).Article 

Google Scholar 
Thuiller, W., Georges, D., Engler, R. & Breiner, F. biomod2: Ensemble Platform for Species Distribution Modeling. (2015).Breiman, L. Random forests. Mach. Learn. 45, 5–32 (2001).Article 
MATH 

Google Scholar 
Friedman, J. H. Greedy function approximation: A gradient boosting machine. Ann. Stat. 29, 1189–1232 (2001).Article 
MATH 

Google Scholar 
Phillips, S. J., Anderson, R. P., Dudík, M., Schapire, R. E. & Blair, M. E. Opening the black box: An open-source release of Maxent. Ecography 40, 887–893 (2017).Article 

Google Scholar 
IUCN. The IUCN Red List of Threatened Species. (2017).Gotelli, N. J. & Graves, G. R. Null Models in Ecology. (1996).Araújo, M. B. & Pearson, R. G. Equilibrium of species’ distributions with climate. Ecography 28, 693–695 (2005).Article 

Google Scholar 
Watson, J. E. M., Grantham, H. S., Wilson, K. A. & Possingham, H. P. Systematic conservation planning: Past, present and future. In Conservation Biogeography (eds Ladle, R. J. & Whittaker, R. J.) (Wiley, 2011).
Google Scholar 
Bevilacqua, M. Áreas protegidas y conservación de la diversidad biológica. Biodivers. Venezuela 2, 922–943 (2003).
Google Scholar 
Franco, P., Saavedra-Rodríguez, C. A. & Kattan, G. H. Bird species diversity captured by protected areas in the Andes of Colombia: A gap analysis. Oryx 41, 57–63 (2007).Article 

Google Scholar 
Barzetti, V. Parks and Progress: Protected Areas and Economic Development in Latin America and the Caribbean. (1993).Schulman, L. et al. Amazonian biodiversity and protected areas: Do they meet?. Biodivers. Conserv. 16, 3011–3051 (2007).Article 

Google Scholar 
Dourojeanni, M. J. Áreas naturales protegidas e investigación científica en el Perú. Rev. For. Perú 33, 91–101 (2018).
Google Scholar 
Rodriguez, L. & Young, K. Biological diversity of peru: Determining priority areas for conservation. Ambio 29, 329–337 (2000).Article 

Google Scholar 
Ministerio del Ambiente & SERNANP. Plan Director de las Áreas Naturales Protegidas (Estrategia Nacional) (2009).Cuesta-Camacho, F. et al. Identificación de Vacíos y Prioridades de Conservación Para la Biodiversidad Terrestre en el Ecuador Continental. http://protectedareas.info/upload/document/ecuador_terrestrial_gap_analysis.pdf (2006).Naveda, J. A. Evaluación del grado de representatividad ecológica y geográfica del sistema de parques nacionales de Venezuela al norte del Orinoco: Anteproyecto. Rev. Geog. Venez. 38, 193–208 (1997).
Google Scholar 
Araujo, N., Müller, R., Nowicki, C. & Ibisch, P. L. Prioridades de conservación de la biodiversidad de Bolivia (editorial FAN, 2010)Arango, N. et al. Vacíos de Conservación del Sistema de Parques Nacionales Naturales de Colombia desde una Perspectiva Ecorregional. https://wwflac.awsassets.panda.org/downloads/vacios_de_conservacion.pdf (2003).Margules, C. R. & Pressey, R. L. Systematic conservation planning. Nature 405, 243–253 (2000).Article 
CAS 

Google Scholar 
Sarkar, S., Sánchez-Cordero, V., Londoño, M. C. & Fuller, T. Systematic conservation assessment for the Mesoamerica, Chocó, and Tropical Andes biodiversity hotspots: A preliminary analysis. Biodivers. Conserv. 18, 1793–1828 (2009).Article 

Google Scholar 
Lessmann, J., Muñoz, J. & Bonaccorso, E. Maximizing species conservation in continental Ecuador: A case of systematic conservation planning for biodiverse regions. Ecol. Evol. 4, 2410–2422 (2014).Article 

Google Scholar 
Young, B. E. et al. Using spatial models to predict areas of endemism and gaps in the protection of Andean slope birds. Auk 126, 554–565 (2009).Article 

Google Scholar 
Fajardo, J., Lessmann, J., Bonaccorso, E., Devenish, C. & Muñoz, J. Combined use of systematic conservation planning, species distribution modelling, and connectivity analysis reveals severe conservation gaps in a megadiverse country (Peru). PLoS ONE 9, 1–23 (2014).Article 

Google Scholar 
Butchart, S. H. M. et al. Shortfalls and solutions for meeting national and global conservation area targets. Conserv. Lett. 8, 329–337 (2015).Article 

Google Scholar 
Pimm, S. L. et al. The biodiversity of species and their rates of extinction, distribution, and protection. Science 344, 6187 (2014).Article 

Google Scholar 
Swenson, J. J. et al. Plant and animal endemism in the eastern Andean slope: Challenges to conservation. BMC Ecol. 12, 1 (2012).Article 

Google Scholar 
Lessmann, J., Fajardo, J., Bonaccorso, E. & Bruner, A. Cost-effective protection of biodiversity in the western Amazon. Biol. Conserv. 235, 250–259 (2019).Article 

Google Scholar 
Rodrigues, A. S. L. & Gaston, K. J. How large do reserve networks need to be?. Ecol. Lett. 4, 602–609 (2001).Article 

Google Scholar 
Reyes-Puig, C. Diversity, threat, and conservation of reptiles from continental Ecuador. Amphib. Reptile Conserv. 11, 8 (2017).
Google Scholar 
Shanee, S. et al. Protected area coverage of threatened vertebrates and ecoregions in Peru: Comparison of communal, private and state reserves. J. Environ. Manage. 202, 12–20 (2017).Article 

Google Scholar 
Kujala, H., Moilanen, A., Araújo, M. B. & Cabeza, M. Conservation planning with uncertain climate change projections. PLoS ONE 8, e53315 (2013).Article 
ADS 
CAS 

Google Scholar 
Hannah, L. et al. 30% land conservation and climate action reduces tropical extinction risk by more than 50%. Ecography 43, 1–11 (2020).Article 

Google Scholar 
Velásquez-Tibatá, J., Salaman, P. & Graham, C. H. Effects of climate change on species distribution, community structure, and conservation of birds in protected areas in Colombia. Reg. Environ. Change 13, 235–248 (2013).Article 

Google Scholar 
del Avalos, V. R. & Hernández, J. Projected distribution shifts and protected area coverage of range-restricted Andean birds under climate change. Glob. Ecol. Conserv. 4, 459–469 (2015).Article 

Google Scholar 
Warren, R. et al. Quantifying the benefit of early climate change mitigation in avoiding biodiversity loss. Nat. Clim. Change 3, 678–682 (2013).Article 
ADS 

Google Scholar 
Golden Kroner, R. et al. COVID-era policies and economic recovery plans: Are governments building back better for protected and conserved areas?. PARKS 27, 135–148 (2021).Article 

Google Scholar 
IPCC Summary for Policymakers. In Climate Change 2013: The Physical Science Basis. Contribution of Working Group I to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change (eds Stocker, T. F. et al.) (Cambridge University Press, 2013).
Google Scholar 
Chevalier, M., Zarzo-Arias, A., Guélat, J., Mateo, R. G. & Guisan, A. Accounting for niche truncation to improve spatial and temporal predictions of species distributions. Front. Ecol. Evol. https://doi.org/10.3389/fevo.2022.944116 (2022).Article 

Google Scholar 
Watson, J. E. M. et al. Bolder science needed now for protected areas. Conserv. Biol. 30, 243–248 (2016).Article 

Google Scholar 
CBD. Kunming-Montreal Global Biodiversity Framework, Draft Decision Submitted by the PRESIDENT. (2022). CBD/COP/15/L.25. https://www.cbd.int/doc/c/e6d3/cd1d/daf663719a03902a9b116c34/cop-15-l-25-en.pdfCBD. Report of the Expert Workshop on the Monitoring Framework for the Post-2020 Global Biodiversity Framework (CBD, 2022).
Google Scholar 
Chaplin-Kramer, R. et al. Mapping the planet’s critical natural assets. Nat. Ecol. Evol. https://doi.org/10.1038/s41559-022-01934-5 (2022).Article 

Google Scholar 
Watson, J. E. M. et al. The exceptional value of intact forest ecosystems. Nat. Ecol. Evol. 2, 599–610 (2018).Article 

Google Scholar 
Elbers, J. Las Áreas Protegidas de América Latina: Situación Actual y Perspectivas PARA el Futuro (2011).Miller, D. C. & Nakamura, K. S. Protected areas and the sustainable governance of forest resources. Curr. Opin. Environ. Sustain. 32, 96–103 (2018).Article 

Google Scholar 
Guisan, A. & Zimmermann, N. E. Predictive habitat distribution models in ecology. Ecol. Model. 135, 147–186 (2000).Article 

Google Scholar 
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).Article 

Google Scholar 
van Proosdij, A. S. J., Sosef, M. S. M., Wieringa, J. J. & Raes, N. Minimum required number of specimen records to develop accurate species distribution models. Ecography 39, 542–552 (2016).Article 

Google Scholar 
Breiner, F. T., Guisan, A., Bergamini, A. & Nobis, M. P. Overcoming limitations of modelling rare species by using ensembles of small models. Methods Ecol. Evol. 6, 1210–1218 (2015).Article 

Google Scholar 
Phillips, S. J. et al. Sample selection bias and presence-only distribution models: Implications for background and pseudo-absence data. Ecol. Appl. 19, 181–197 (2009).Article 

Google Scholar 
Thornhill, A. H. et al. Spatial phylogenetics of the native California flora. BMC Biol 15, 96 (2017).Article 

Google Scholar 
Radosavljevic, A. & Anderson, R. P. Making better Maxent models of species distributions: Complexity, overfitting and evaluation. J. Biogeogr. 41, 629–643 (2014).Article 

Google Scholar 
Kershaw, F. et al. Informing conservation units: Barriers to dispersal for the yellow anaconda. Divers. Distrib. 19, 1164–1174 (2013).Article 

Google Scholar 
Venter, O. et al. Targeting global protected area expansion for imperiled biodiversity. PLoS Biol. 12, e1001891 (2014).Article 

Google Scholar 
Gaston, K. J. The Structure and Dynamics of Geographic Ranges (Oxford University Press, 2003).
Google Scholar 
Yin, L., Fu, R., Shevliakova, E. & Dickinson, R. E. How well can CMIP5 simulate precipitation and its controlling processes over tropical South America?. Clim. Dyn. 41, 3127–3143 (2013).Article 

Google Scholar  More

Fischer, E. M. & Knutti, R. Nature Clim. Change 5, 560–564 (2015).Article 

Google Scholar 
Román-Palacios, C. & Wiens, J. J. Proc. Natl Acad. Sci. USA 117, 4211–4217 (2020).Article 
PubMed 

Google Scholar 
Ma, G., Hoffmann, A. A. & Ma, C.-S. J. Exp. Biol. 218, 2289–2296 (2015).PubMed 

Google Scholar 
Dillon, M. E., Wang, G. & Huey, R. B. Nature 467, 704–706 (2010).Article 
PubMed 

Google Scholar 
Garcia, R. A., Cabeza, M., Rahbek, C. & Araújo, M. B. Science 344, 1247579 (2014).Article 
PubMed 

Google Scholar  More

With human movement and globalization, invasive container breeding vectors responsible for dengue, Zika, chikungunya and now malaria, with An. stephensi, are being introduced and establishing populations in new locations. They are bringing with them the threat of increasing or novel cases of vector-borne diseases to new locations where health systems may not be prepared.Anopheles stephensi was first detected on the African continent in Djibouti in 2012 and has since been confirmed in Ethiopia, Somalia, and Sudan. Unlike most malaria vectors, An. stephensi is often found in artificial containers and in urban settings. This unique ecology combined with its initial detection in seaports in Djibouti, Somalia, and Sudan has led scientists to believe that the movement of this vector is likely facilitated through maritime trade.By modeling inter- and intra-continental maritime connectivity in Africa we identified countries with higher likelihood of An. stephensi introduction if facilitated through maritime movement and ranked them based on this data. Anopheles stephensi was not detected in Africa (Djibouti) until 2012. To determine whether historical maritime data would have identified the first sites of introduction, 2011 maritime data were analyzed to determine whether the sites with confirmed An. stephensi would rank highly in connectivity to An. stephensi endemic countries. Using 2011 data on maritime connectivity alone, Djibouti and Sudan were identified as the top two countries at risk of An. stephensi introduction if it is facilitated by marine cargo shipments. In 2021, these are two of the three African coastal nations where An. stephensi is confirmed to be established.When 2011 maritime data were combined with the HSI for An. stephensi establishment, the top five countries remain the same as with maritime data alone: Sudan, Djibouti, Egypt, Kenya and Tanzania, in that order. The maritime data show likelihood of introduction and HSI shows likelihood of establishment. When combined, the analyses show a likelihood of being able to establish and survive once introduced. Interestingly, the results of the combined analyses align with the detection data being reported in the Horn of Africa. The 2011 maritime data reinforces the validity of the model as it points to Sudan and Djibouti, where An. stephensi established in the following years. Similarly, the HSI data for Ethiopia has aligned closely with detections of the species to date15. Interestingly, around this time of initial detection in Djibouti, Djibouti City port underwent development and organizational change. The government of Djibouti took back administrative control of the port as early as 201230.Following this method, maritime trade data from 2020 could point to countries at risk of An. stephensi introduction from endemic countries as well as from the coastal African countries with newly introduced populations. Here we provide a prioritization list and heat map of countries for the early detection, rapid response, and targeted surveillance of An. stephensi in Africa based on this data and the HSI (Fig. 4). Further invasion of An. stephensi on the African continent has the potential to reverse progress made on malaria control in the last century. Anopheles stephensi thrives in urban settings and in containers, in contrast to the rural settings and natural habitats where most Anopheles spp. are found20. The situation in Djibouti may be a harbinger for what is to come if immediate surveillance and control strategies are not initiated18.Figure 4Prioritization Heat Map of African Countries. These 2020 heat maps rank African countries using (A) the Likelihood of An. stephensi through Maritime Trade Index (LASIMTI) data alone and (B) LASIMTI and HSI combined, based on maritime connectivity to countries where An. stephensi is endemic. Higher ranking countries which are at greater risk of An. stephensi introduction are darker in red color than those that are lower ranking (lighter red). Countries which are shaded grey are inland countries that do not have a coast and therefore no data on maritime movement into ports. Countries which are grey and checkered have established or endemic An. stephensi populations and are considered source locations for potential An. stephensi introduction in this analysis. Map was generated using MapChart (mapchart.net).Full size imageMaritime data from 2020, with Djibouti and Sudan considered as potential source populations for intracontinental introduction of An. stephensi, indicate the top five countries at risk for maritime introduction are Egypt, Kenya, Mauritius, Tanzania, and Morocco, suggesting that targeted larval surveillance in these countries near seaports may provide a better understanding of whether there are maritime introductions. When the data from 2020 data is combined with HSI for An. stephensi, the top five countries are instead Egypt, Kenya, Tanzania, Morocco, and Libya. Interestingly, historical reports of An. stephensi in Egypt exist; however, following further identification these specimens were determined to be An. ainshamsi31. With several suitable habitats both along the coast and inland of Egypt, revisiting surveillance efforts there would provide insight into how countries that are highly connected to An. stephensi locations through maritime traffic may experience introductions.Further field validation of this prioritization list is necessary, because it is possible that An. stephensi is being introduced through other transportation routes, such as dry ports or airports32, or may even be dispersed through wind facilitation33. However, countries highlighted here with high levels of connectivity to known An. stephensi locations should be considered seriously at risk and surveillance urgently established to determine whether An. stephensi introduction has already occurred or to enable early detection. Primary vector surveillance for both Ae. aegypti and An. stephensi are through larval surveys, and the two mosquitoes are commonly detected in the same breeding habitats. It could therefore be beneficial to coordinate with existing Aedes surveillance efforts to be able to simultaneously gather data on medically relevant Aedes vectors while seeking to determine whether An. stephensi is present. Similarly, in locations with known An. stephensi and not well established Aedes programs, coordinating surveillance efforts provides an opportunity to conduct malaria and arboviral surveillance by container breeding mosquitoes simultaneously.Efforts to map pinch points or key points of introduction based on the movement of goods and populations could provide high specificity for targeted surveillance and control efforts. For example, participatory mapping or population mobility data collection methods, such as those used to determine routes of human movement for malaria elimination, may simultaneously provide information on where targeted An. stephensi surveillance efforts should focus. Several methods have been proposed in the literature for modeling human movement and one in particular, PopCAB, which is often used for communicable diseases, combined quantitative and qualitative data with geospatial information to identify points of control34.Data on invasive mosquito species has shown that introduction events are rarely a one-time occurrence. Population genetics data on Aedes species indicate that reintroductions are very common and can facilitate the movement of genes between geographically distinct populations, raising the potential for introduction of insecticide resistance, thermotolerance, and other phenotypic and even behavioral traits which may be facilitated by gene flow and introgression35. Djibouti, Sudan, Somalia, and Ethiopia, countries with established invasive populations of An. stephensi, should continue to monitor invasive populations and points of introduction to control and limit further expansion and adaptation of An. stephensi. Work by Carter et al. has shown that An. stephensi populations in Ethiopia in the north and central regions can be differentiated genetically, potentially indicating that these populations are a result of more than one introduction into Ethiopia from South Asia, further emphasizing the potential role of anthropogenic movement on the introduction of the species17.One major limitation of this work is that Somalia is the third coastal nation where An. stephensi has been confirmed; however, marine traffic data were not available for Somalia so it could not be included in this analysis. The potential impact of Somalia on maritime trade is unknown and it should not be excluded as a potential source population. Additionally, this model does not account for the possibility of other countries with An. stephensi populations that have not been detected yet. As new data on An. stephensi expansion becomes available, more countries will be at higher risk. Other countries with An. stephensi populations, such as Iran, Myanmar, and Iraq, constitute lower relative percentages of trade with these countries so were not included in the analysis. However, genetic similarities were noted from An. stephensi in Pakistan, so this nation was included10.Due to the nature of maritime traffic, inland countries were also not included in this prioritization ranking. Countries which are inland but share borders with high-risk countries according to the LASTIMI index should also be considered with high priority. For example, the ranking from 2011 highlights Sudan and Djibouti, both which border Ethiopia, and efforts to examine key land transportation routes between bordering nations where humans and goods travel may provide additional insight into the expansion routes of this invasive species.In Ethiopia, An. stephensi was detected in 2016. It has largely been detected along major transportation routes although further data is needed to understand the association between movement and An. stephensi introductions and expansion since most sampling sites have also been located along transport routes. Importantly, Ethiopia relies heavily on the ports of Djibouti and Somalia for maritime imports and exports. Surveillance efforts have revealed that the species is also frequently associated with livestock shelters and An. stephensi are frequently found with livestock bloodmeals15. Interestingly, the original detection of An. stephensi was found in a livestock quarantine station in the port of Djibouti. Additionally, livestock constitutes one of the largest exports of maritime trade from this region. For countries with high maritime connectivity to An. stephensi locations, surveillance efforts near seaports, in particular those with livestock trade, may be targeted locations for countries without confirmed An. stephensi to begin larval surveillance.As Ae. aegypti and Culex coronator were detected in tires or Ae. albopictus through tire and bamboo (Dracaena sanderiana) trade, An. stephensi could be carried through maritime trade of a specific good36,37,38. Future examination of the movement of specific goods would be beneficial in interpreting potential An. stephensi invasion pathways. Additionally, the various types of vessels used to transport certain cargo such as container, bulk, and livestock ships could affect An. stephensi survivability during transit. Sugar and grain are often shipped in bulk or break bulk vessels which store cargo in large unpackaged containers. Container ships transport products stored in containers sized for land transportation via trucks and carry goods such as tires. Livestock vessels are often multilevel, open-air ships which require more hands working on deck and water management39.Using LSBCI index data from 2020, we developed a network to highlight how coastal African nations are connected through maritime trade (Fig. 4). The role of this network analysis is two-fold, (1) it demonstrates an understanding of intracontinental maritime connectivity; and (2) it highlights the top three countries connected via maritime trade through an interactive html model (Supplemental File). For example, if An. stephensi is detected and established in a specific coastal African nation such as Djibouti, selecting the Djibouti node reveals the top three locations at risk of introduction from that source country (Djibouti links to Sudan, Egypt and Kenya). This can be used as an actionable prioritization list for surveillance if An. stephensi is detected in any given country and highlights major maritime hubs in Africa which could be targeted for surveillance and control. For example, since the development of this model, An. stephensi has been detected in Nigeria. Through the use of this interactive model, Ghana, Cote d’Ivoire, and Benin have been identified as countries most connected to Nigeria through maritime trade and therefore surveillance prioritization activities could consider these locations.The network analysis reveals the significance of South African trade to the rest of the continent. Due to the distance, South Africa did not appear to be high in risk of An. stephensi introduction. However, this analysis does reveal that if An. stephensi were to enter nearby countries, it could very easily be introduced because of its high centrality. Western African countries such as Ghana, Togo, and Morocco are also heavily connected to other parts of Africa. Interestingly, Mauritius appears to be highly significant to this network of African maritime trade. Based on 2020 maritime data, Mauritius is ranked as the country with the third greatest likelihood of introduction of An. stephensi and has the second highest centrality rank value of 0.159. Considering these factors, Mauritius could serve as an important port of call connecting larger ports throughout Africa or other continents. With long standing regular larval surveillance efforts across the island for Aedes spp., this island nation is well suited to look for Anopheles larvae as part of Aedes surveillance efforts for early detection and rapid response to prevent the establishment of An. stephensi. If An. stephensi were to become established in countries with high centrality ranks, further expansion on the continent could be accelerated drastically. These ports could serve as important watchpoints and indicators of An. stephensi’s incursion into Africa. Anopheles stephensi is often found in shared habitats with Aedes spp. and a great opportunity exists to leverage Aedes arboviral surveillance efforts to initiate the search for An. stephensi, especially in countries that have high potential of introduction through maritime trade. More

Timothy Searchinger and his colleagues raise concerns that the European Union’s plan to produce energy from biomass could compromise forest carbon stocks and biodiversity (Nature 612, 27–30; 2022). However, it is possible for improved forest management to reconcile increased bioenergy production by maintaining and restoring forest ecosystems.
Competing Interests
The authors declare no competing interests. More

Study siteThe field experiment was conducted at Namco Station (30°47’N, 90°58’E, altitude 4730 m) of the Institute of Tibetan Plateau Research, Chinese Academy of Sciences (ITPCAS), which is located in the alpine steppes of TP in China. The experiment was permitted by ITPCAS, complied with local and national guidelines and regulations. From 2006 to 2017, the mean annual temperature (MAT) and mean annual precipitation (MAP) was about − 0.6 °C and 406 mm, respectively. Monthly mean temperature varied from − 10.8 °C in January to 9.1 °C in July and most of the precipitation occurred from May to October37,38. During our six-year observations (2010, 2011, 2012, 2013, 2015 and 2017), climate change during the growing season from May to September varied differently, with the annual precipitation ranged from 255.9 mm to 493.8 mm and the MAT from 6.7 to 7.4 °C. Androsace tapete, Kobresia pygmaea, Stipa purpurea and Leontopodium pusillum were the dominant plant species at the alpine steppe.Experimental design and treatmentsThe long-term experiment began in May, 2010. Three homogenous plots were randomly arranged as replicates at the alpine steppe and six subplots (~ 13 m2) were distributed in each plot by a cycle, with a 2 m buffer zone between each adjacent subplot (Appendix S1: Fig. S1). In this experiment, six treatments of N fertilization rate (0, 1, 2, 4, 8, and 16 g N m−2 yr−1) were clockwise applied in each subplot. The subplots of 0 g N m−2 yr−1 were control group. We sprayed NH4NO3 solution on the first day of each month in the growing season (from May to September) each year. After fertilizing, we rinsed plant residual fertilizer with a little deionized water (no more than 2 mm rainfall). For the control groups, we added equivalent amount of water. The experiment was conducted from 2010 to 2017 (it should be pointed out that there was no fertilization in 2014 and 2016).Sampling and measurementsThe samples were collected with the training and permission of ITPCAS and involved plants that are common species and not endangered or protected. The identification of the plants was done by referring to a book of Chen and Yang39. Pictures of the corresponding specimens can be seen on the website of ITPCAS (http://itpcas.cas.cn/kxcb/kxtp/nmc_normal_plant/).Vegetation samples were collected in August in 2011 and repeated at the same time in 2012, 2013, 2015 and 2017. We established one 50 × 50 cm quadrat in each subplot, clipped aboveground biomass (AGB) and sorted species by families. The biomass was used to measure ANPP (g m−2 yr−1). Following aboveground portion collected, we used three soil cores (5 cm diameter) to collect the belowground roots at 0–30 cm depth and mixed into one sample, which were used to assess belowground net primary productivity (BNPP, g m−2 yr−1). The roots were cleaned with running water to remove sand and stones.Both plant and root samples were dried at 75 °C for 48 h and then ground into powder (particle size ~ 5 μm) by a laboratory mixer mill (MM400, Retsch). To determine N and C content of plants, we weighed the samples into tin capsules and measured with the elemental analyzer (MAT253, Finnigan MAT GmbH, Germany).Estimation of the critical N rate (Ncr), N retention fraction (NRF), N retention capacity and N-induced C gainAccording to the N saturation hypothesis, plant productivity increases gradually during N addition, reaches a maximum at the Ncr, and eventually declines16,17. We considered the Ncr to be the rate where ANPP no longer remarkably changed with N addition (Fig. 1).We defined plant N retention fraction (NRF, %; Eq. 1) as the aboveground N storage caused by unit N addition rate, and N retention capacity (g N m−2 yr−1; Eq. 2) was the increment of N storage due to exogenous N addition compared to the control40. The equations are as following:$$N;retention;fraction = frac{{ANPP_{tr} times N;content_{tr} – ANPP_{ck} times N;content_{ck} }}{N;rate}$$
(1)
$$N;retention;capacity = ANPP_{tr} times N;content_{tr} – ANPP_{ck} times N;content_{ck}$$
(2)
where ANPPtr and N contenttr (%) refer to those in the treatment (tr) groups, and ANPPck and N contentck refer to those in the control (ck) groups. These expressions are also used in the following equations (Eqs. 3–5).The N-induced C gain (g C m−2 yr−1; Eq. 3) was estimated by the increment of C storage owing to exogenous N addition compared to the control40. Maximum N retention capacity (MNRC, Eq. 4) and maximum N-induced C gain (Eq. 5) mean the maximum N and C storage increment in plant caused by exogenous N input at Ncr, respectively. The formulas are as following:$$N{text{-}}induced;C;gain = ANPP_{tr} times C;content_{tr} – ANPP_{ck} times C;content_{ck}$$
(3)
$$MNRC = ANPP_{max } times N;content_{max } – ANPP_{ck} times N;content_{ck}$$
(4)
$$Maximum;N{text{-}}induced;C;gain = ANPP_{max } times C;content_{max } – ANPP_{ck} times C;content_{ck}$$
(5)
where ANPPmax, N contentmax and C contentmax refer to the value of ANPP, N content and C content at Ncr, respectively.Data synthesisTo evaluate N limitation and saturation on the TP more accurately, we searched papers from the Web of Science (https://www.webofscience.com) and the China National Knowledge Infrastructure (https://www.cnki.net). The keywords used by article searching were: (a) N addition, N deposition or N fertilization, (b) grassland, steppe or meadow. Article selection was based on the following conditions. First, the experimental site must be conducted in a grassland ecosystem. Second, the experiment had at least three N addition levels and a control group. Third, if the experiment lasted for many years, we analyzed data with multi-year average. Based on the above, we collected 89 independent experimental cases. Among these, 27 cases were located on the TP alpine grasslands, 25 in the Inner Mongolia (IM) grasslands and 37 in other terrestrial grasslands (detailed information sees Appendix S2: Table S1).We extracted ANPP data and N addition rate of these cases and estimated Ncr and ANPPmax (Appendix S2: Fig. S2). We then calculated NRF, N retention and C gain of each group of data for further analysis (Appendix S2: Table S2). Most of the 89 cases did not have data on N and C content. To facilitate the calculation, we summarized N and C content from 40 articles in the neighboring areas of the cases and divided the N and C content into seven intervals according to the N addition rate (Appendix S2: Table S3 and Fig. S3). The unit of N addition rate was unified to “g N m−2 yr−1”. All the original data were obtained directly from texts and tables of published papers. If the data were displayed only in graphs, Getdata 2.20 was used to digitize the numerical data. For the estimation of N retention and C gain of the TP at current N deposition rates and future at Ncr, we fitted the exponential relationship to the data from 27 cases on the TP, and then substituted N rates into the fitted equations (Eq. 6):$$y = a times left[ {1 – exp left( { – bx} right)} right].$$
(6)
We also included MAT, MAP, soil C:N ratio, fencing management (fencing or grazing) and grassland type (meadow, steppe and desert steppe) of the experiment sites for exploring the drivers affecting N limitation (Appendix S2: Table S1). When climatic data were missing from the article, MAT and MAP were obtained from the WorldClim (http://www.worldclim.org).Species were usually divided into four functional groups (grasses, sedges, legumes and forbs) to study the response of species composition to N addition in previous study41. We synthesized 13 TP experimental cases (including our field experiment) from the data synthesis and each case included at least three functional groups (detailed references see Appendix S2).Statistical analysisThere were 42 species in our field experiment. We divided them by family into eleven groups: Asteraceae (forbs), Poaceae (grasses), Leguminosae (legumes), Rosaceae (forbs), Boraginaceae (forbs), Caryophyllaceae (forbs), Cyperaceae (sedges), Labiatae (forbs), Primulaceae (forbs), Scrophulariaceae (forbs) and Others. Due to species in the group of Others contributed only 1.22% of AGB, we analyzed AGB and foliar stoichiometry among other ten families (Appendix S1: Table S1). In Namco steppe, forbs, grasses, sedges and legumes accounted for 78.0%, 7.4%, 8.2% and 5.2% of the AGB respectively (Appendix S1: Table S1 and Fig. S2). Such a large number of forbs suggested that our experiment was conducted on a severely degraded grassland.For our field data, two-way ANOVAs were used to analyze the effects of year, N fertilization rate and their interactions on species AGB. One-way ANOVAs were used to test the response of ANPP, BNPP, root:shoot ratio, species foliar C content, N content and C:N ratio to N addition rate. Duncan’s new multiple range test was used to compare the fertilization influences at each rate in these ANOVAs. Prior to the above ANOVAs, we performed homogeneity of variance test and transformed the data logarithmically when necessary. Simple regression was used to estimate the relevance among ANPP, NRF, N retention capacity and C gain with N addition rates.Structural equation modeling (SEM) was used to explore complex relationships among multiple variables. To quantify the contribution of drivers such as climate and soil to Ncr, ANPP, NRF and MNRC, we constructed SEM based on existing ecological knowledge and the possible relationships between variables. We considered environmental factors (MAT, MAP and soil C:N) and ANPPck as explanatory variables, and Ncr, NRF and MNRC as response variables. We included the ANPPck in the SEM rather than the ANPPmax because we wonder whether there was a relationship between ANPP in the absence of exogenous N input and the ecosystem N retention in the presence of N saturation. This has important implications for assessing N input. Before constructing the SEM, we excluded collinearity between the factors. In addition, Student’s t-test and one-way ANOVAs were performed to explain the effect of fencing management and grassland type on above response variables, respectively. The SEM was constructed using the R package “piecewiseSEM”42. Fisher’s C was used to assess the goodness-of-model fit, and AIC was for model comparison.Given the influence of extreme values in the data synthesis, we calculated the geometric mean of Ncr, NRF, N retention and N-induced C gain. All statistical analyses were performed with SPSS 26.0 and RStudio (Version 1.2.1335) based on R version 3.6.2 (R Core Team, 2019). More

Plaisance, L., Caley, M. J., Brainard, R. E. & Knowlton, N. The diversity of coral reefs: what are we missing? PLoS One. 6, e25026 (2011).Article 
ADS 
CAS 

Google Scholar 
Hoegh-Guldberg, O., Poloczanska, E. S., Skirving, W. & Dove, S. Coral reef ecosystems under climate change and ocean acidification. Front. Mar. Sci. 4, 158 (2017).Article 

Google Scholar 
Mora, C. et al. Global human footprint on the linkage between biodiversity and ecosystem functioning in reef fishes. PLoS Biol. 9, e1000606 (2011).Article 
CAS 

Google Scholar 
Burke, L., Reytar, K., Spalding, M. & Perry, A. Reefs at Risk Revisited. 130 pp. (World Resources Institute, Washington, D.C., 2011).Hicks, C. C., Graham, N. A. J., Maire, E. & Robinson, J. P. W. Secure local aquatic food systems in the face of declining coral reefs. One Earth. 4, 1214–1216 (2021).Article 
ADS 

Google Scholar 
Cinner, J. E. et al. Gravity of human impacts mediates coral reef conservation gains. PNAS 115, E6116–E6125 (2018).Article 
CAS 

Google Scholar 
Eddy, T. D. et al. Global decline in capacity of coral reefs to provide ecosystem services. One Earth. 4, 1278–1285 (2021).Article 
ADS 

Google Scholar 
Graham, N. A. J. et al. Human disruption of coral reef trophic structure. Curr. Biol. 27, 231–236 (2017).Article 
CAS 

Google Scholar 
Sherman, C. S., Heupel, M. R., Moore, S. K., Chin, A. & Simpfendorfer, C. A. When sharks are away rays will play: effects of top predator removal in coral reef ecosystems. Mar. Ecol. Prog. Ser. 641, 145–157 (2020).Article 
ADS 

Google Scholar 
Ruppert, J. L. W., Travers, M. J., Smith, L. L., Fortin, M.-J. & Meekan, M. G. Caught in the middle: combined Impacts of shark removal and coral loss on the fish communities of coral reefs. PLoS One. 8, e74648 (2013).Article 
ADS 
CAS 

Google Scholar 
Last, P. R. et al. Rays of the World. (CSIRO Publishing, 2016).Ebert, D. A., Dando, M. & Fowler, S. Sharks of the World. 2nd edn, 608 (Princeton University Press, 2021).Heupel, M. R., Lédée, E. J. I. & Simpfendorer, C. A. Telemetry reveals spatial separation of co-occurring reef sharks. Mar. Ecol. Prog. Ser. 589, 179–192 (2018).Article 
ADS 

Google Scholar 
Heupel, M. R., Papastamatiou, Y. P., Espinoza, M., Green, M. E. & Simpfendorfer, C. A. Reef shark science – key questions and future directions. Front. Mar. Sci. 6, 12 (2019).Article 

Google Scholar 
Roff, G., Brown, C. J., Priest, M. A. & Mumby, P. J. Decline of coastal apex shark populations over the past half century. Commun. Biol. 1, 223 (2018).Article 

Google Scholar 
Williams, J. J., Papastamatiou, Y. P., Caselle, J. E., Bradley, D. & Jacoby, D. M. P. Mobile marine predators: an understudied source of nutrients to coral reefs in an unfished atoll. Proc. R. Soc. B. 285, 20172456 (2018).Article 

Google Scholar 
Heithaus, M. R., Wirsing, A. J. & Dill, L. M. The ecological importance of intact top-predator populations: a synthesis of 15 years of research in a seagrass ecosystem. Mar. Freshw. Res. 63, 1039–1050 (2012).Article 

Google Scholar 
Peel, L. R. et al. Stable isotope analyses reveal unique trophic role of reef manta rays (Mobula alfredi) at a remote coral reef. R. Soc. Open Sci. 6, 190599 (2019).Article 
ADS 
CAS 

Google Scholar 
O’Shea, O. R., Thums, M., van Keulen, M. & Meekan, M. Bioturbation by stingrays at Ningaloo Reef, Western Australia. Mar. Freshw. Res. 63, 189–197 (2012).Article 

Google Scholar 
Takeuchi, S. & Tamaki, A. Assessment of benthic disturbance associated with stingray foraging for ghost shrimp by aerial survey over an intertidal sandflat. Continental Shelf Res. 84, 139–157 (2014).Article 
ADS 

Google Scholar 
Burkholder, D. A., Heithaus, M. R., Fourqurean, J. W., Wirsing, A. & Dill, L. M. Patterns of top-down control in a seagrass ecosystem: could a roving apex predator induce a behaviour-mediated trophic cascade? J. Anim. Ecol. 82, 1192–1202 (2013).Article 

Google Scholar 
Creel, S. & Christianson, D. Relationships between direct predation and risk effects. TRENDS Ecol. Evolution. 23, 194–201 (2008).Article 

Google Scholar 
Ward-Paige, C. A. et al. Large-scale absence of sharks on reefs in the greater-Caribbean: a footprint of human presence. PLoS One. 5, e11968 (2010).Article 
ADS 

Google Scholar 
Espinoza, M., Cappo, M., Heupel, M. R., Tobin, A. J. & Simpfendorfer, C. A. Quantifying shark distribution patterns and species-habitat associations: implications of marine park zoning. PLoS One. 9, e106885 (2014).Article 
ADS 

Google Scholar 
Graham, N. A., Spalding, M. D. & Sheppard, C. R. Reef shark declines in remote atolls highlight the need for multi-faceted conservation action. Aquat. Conserv.: Mar. Freshw. Ecosyst. 20, 543–548 (2010).Article 

Google Scholar 
Nadon, M. O. et al. Re-creating missing population baselines for Pacific reef sharks. Conserv. Biol. 26, 493–503 (2012).Article 

Google Scholar 
MacNeil, M. A. et al. Global status and conservation potential of reef sharks. Nature 583, 801–806 (2020).Article 
ADS 
CAS 

Google Scholar 
Dulvy, N. K. et al. Overfishing drives over one-third of all sharks and rays toward a global extinction crisis. Curr. Biol. 31, 1–15 (2021).Article 

Google Scholar 
Walls, R. H. L. & Dulvy, N. K. Eliminating the dark matter of data deficiency by predicting the conservation status of Northeast Atlantic and Mediterranean Sea sharks and rays. Biol. Conserv. 246, 108459 (2020).Article 

Google Scholar 
Yan, H. F. et al. Overfishing and habitat loss drives range contraction of iconic marine fishes to near extinction. Science Adv. 7, eabb6026, (2021).Butchart, S. H. M. et al. Using Red List Indices to measure progress towards the 2010 target and beyond. Philos. Trans. R. Soc. B 360, 255–268 (2005).Article 
CAS 

Google Scholar 
Sherman, C. S. et al. Taeniura lymma. The IUCN Red List of Threatened Species, eT116850766A116851089 (2021). 10.2305/IUCN.UK.2021-1.RLTS.T116850766A116851089.enPacoureau, N. et al. Half a century of global decline in oceanic sharks and rays. Nature 589, 567–571 (2021).Article 
ADS 
CAS 

Google Scholar 
Cardeñosa, D. et al. Small fins, large trade: a snapshot of the species composition of low-value shark fins in the Hong Kong markets. Anim. Conserv. 23, 203–211 (2019).Article 

Google Scholar 
Haque, A. B. & Spaet, J. L. Y. Trade in threatened elasmobranchs in the Bay of Bengal, Bangladesh. Fish. Res. 243, 106059 (2021).Article 

Google Scholar 
Alcala, A. C. & Russ, G. R. A direct test of the effects of protective management on abundance and yield of tropical marine resources. ICES J. Mar. Sci. 47, 40–47 (1990).Article 

Google Scholar 
Serrano, A. et al. Effects of anti-trawling artificial reefs on ecological indicators of inner shelf fish and invertebrate communities in the Cantabrian Sea (southern Bay of Biscay). J. Mar. Biol. Assoc. U. Kingd. 91, 623–633 (2011).Article 

Google Scholar 
Cortés, E. Perspectives on the intrinsic rate of population growth. Methods Ecol. Evolution. 7, 1136–1145 (2016).Article 

Google Scholar 
McClenachan, L., Cooper, A. B. & Dulvy, N. K. Rethinking trade-driven extinction risk in marine and terrestrial megafauna. Curr. Biol. 26, 1–7 (2016).Article 

Google Scholar 
Tamburello, N., Cote, I. M. & Dulvy, N. K. Energy and the scaling of animal space use. Am. Naturalist 186, 196–211 (2015).Article 

Google Scholar 
Dulvy, N. K. et al. Challenges and priorities in shark and ray conservation. Curr. Biol. 27, R565–R572 (2017).Article 
CAS 

Google Scholar 
Davidson, L. N. K. & Dulvy, N. K. Global marine protected areas to prevent extinctions. Ecol. Evolution. 1, 1–6 (2017).
Google Scholar 
Pauly, D., Zeller, D. & Palomares, M. L. D. Sea Around Us Concepts, Design and Data, (2021).Simpfendorfer, C. A. & Dulvy, N. K. Bright spots of sustainable shark fishing. Curr. Biol. 27, R83–R102 (2017).Article 

Google Scholar 
Booth, H., Squires, D. & Milner-Gulland, E. J. The mitigation hierarchy for sharks: a risk-based framework for reconciling trade-offs between shark conservation and fisheries objectives. Fish. Fish. 21, 269–289 (2019).Article 

Google Scholar 
Grorud-Colvert, K. et al. The MPA Guide: A framework to achieve global goals for the ocean. Science 373, eabf0861 (2021).Article 
CAS 

Google Scholar 
Enright, S. R., Meneses-Orellana, R. & Keith, I. The Eastern Tropical Pacific Marine Corridor (CMAR): The emergence of a voluntary regional cooperation mechanism for the conservation and sustainable use of marine biodiversity within a fragmented regional ocean governance landscape. Front. Mar. Sci. 8, 674825 (2021).Article 

Google Scholar 
Chin, A., Kyne, P. M., Walker, T. I. & McAuley, R. B. An integrated risk assessment for climate change: analysing the vulnerability of sharks and rays on Australia’s Great Barrier Reef. Glob. Change Biol. 16, 1936–1953 (2010).Article 
ADS 

Google Scholar 
Dwyer, R. G. et al. Individual and population benefits of marine reserves for reef sharks. Curr. Biol. 30, 480–489 (2020).Article 
CAS 

Google Scholar 
Speed, C. W., Cappo, M. & Meekan, M. G. Evidence for rapid recovery of shark populations within a coral reef marine protected area. Biol. Conserv. 220, 308–319 (2018).Article 

Google Scholar 
Mizrahi, M. I., Diedrich, A., Weeks, R. & Pressey, R. L. A systematic review of the socioeconomic factors that influence how marine protected areas impact on ecosystems and livelihoods. Soc. Nat. Resour. 32, 4–20 (2019).Article 

Google Scholar 
IPBES. Global assessment report on biodiversity and ecosystem services of the Intergovernmental Science-Policy Platform on Biodiversity and Ecosystem Services. 1148 (Bonn, Germany, 2019).Butchart, S. H. M. et al. Global biodiversity: indicators of recent declines. Science 328, 1164–1168 (2010).Article 
ADS 
CAS 

Google Scholar 
Hanh, T. T. H. & Boonstra, W. J. What prevents small-scale fishing and aquaculture households from engaging in alternative livelihoods? A case study in the Tam Giang lagoon, Viet Nam. Ocean Coast. Manag. 182, 104943 (2019).Article 

Google Scholar 
Ahmed, N., Troell, M., Allison, E. H. & Muir, J. F. Prawn postlarvae fishing in coastal Bangladesh: challenges for sustainable livelihoods. Mar. Policy. 34, 218–227 (2010).Article 

Google Scholar 
Prasetyo, A. P. et al. Shark and ray trade in and out of Indonesia: addressing knowledge gaps on the path to sustainability. Mar. Policy. 133, 104714 (2021).Article 

Google Scholar 
McClanahan, T., Polunin, N. & Done, T. Ecological states and the resilience of coral reefs. Conserv. Ecol. 6, 18 (2002).
Google Scholar 
Bellwood, D. R., Hughes, T. P. & Hoey, A. S. Sleeping functional group drives coral-reef recovery. Curr. Biol. 16, 2434–2439 (2006).Article 
CAS 

Google Scholar 
Cinner, J. E. et al. Vulnerability of coastal communities to key impacts of climate change on coral reef fisheries. Glob. Environ. Change. 22, 12–20 (2012).Article 
ADS 

Google Scholar 
Víe, J.-C., Hilton-Taylor, C. & Stuart, S. N. Wildlife in a Changing World – An analysis of the 2008 IUCN Red List of Threatened Species. 180 (Gland, Switzerland, 2009).Mace, G. M. et al. Quantification of extinction risk: IUCN’s system for classifying threatened species. Conserv. Biol. 22, 1424–1442 (2008).Article 

Google Scholar 
Sherley, R. B. et al. Estimating IUCN Red List population reduction: JARA – A decision-support tool applied to pelagic sharks. Conserv. Lett. 13, e12688 (2019).
Google Scholar 
IUCN Red List. Threats Classification Scheme (Version 3.2), (2021).Salafsky, N. et al. A standard lexicon for biodiversity conservation: unified classifications of threats and actions. Conserv. Biol. 22, 897–911 (2008).Article 

Google Scholar 
Moore, A. Chiloscyllium arabicum. The IUCN Red List of Threatened Species 2017, e.T161426A109902537 (2017). 10.2305/IUCN.UK.2017-2.RLTS.T161426A109902537.enSadovy de Mitcheson, Y. J. et al. Valuable but vulnerable: Over-fishing and under-management continue to threaten groupers so what now? Mar. Policy. 116, 103909 (2020).Article 

Google Scholar 
R: A language and environment for statistical computing (R Foundation for Statistical Computing, Vienna, Austria, 2021).Regression Models for Ordinal Data v. 2019.12.10 (CRAN, 2019).Econometric Tools for Performance and Risk Analysis v. 2.0.4 (2020).Dormann, C. F. et al. Collinearity: a review of methods to deal with it and a simulation study evaluating their performance. Ecography 36, 27–46 (2013).Article 

Google Scholar 
Akinwande, M. O., Dikko, H. G. & Samson, A. Variance inflation factor: As a condition for the inclusion of suppressor variable(s) in regression analysis. Open J. Stat. 5, 754–767 (2015).Article 

Google Scholar 
Burnham, K. P. & Anderson, D. R. Multimodel inference: understanding AIC and BIC in model selection. Sociological Methods Res. 33, 261–304 (2004).Article 
MathSciNet 

Google Scholar 
Plots Coefficients from Fitted Models v. 1.2.8 (2022).Fisheries and Aquaculture Software. FishStatJ – Software for Fishery and Aquaculture Statistical Time Series., (2020).Reynolds, R. W., Rayner, N. A., Smith, T. M., Stokes, D. C. & Wang, W. An improved in situ and satellite SST analysis for climate. J. Clim. 15, 1609–1625 (2002).Article 
ADS 

Google Scholar 
NASA Ocean Biology (OB.DAAC). Mean annual sea surface chlorophyll-a concentration for the period 2009-2013 (composite dataset created by UNEP-WCMC). Data obtained from the Moderate Resolution Imaging Spectroradiometer (MODIS) Aqua Ocean Colour website (NASA OB.DAAC, Greenbelt, MD, USA), (2014).General Bathymetric Chart of the Oceans. GEBCO_2014 Grid. version 20150318. www.gebco.net (2015).XGBoost: A Scalable Tree Boosting System v. 1.4.1.1 (In Proceedings of the 22nd ACM SIGKDD International Conference on Knowledge Discovery and Data Mining (pp. 785–794). New York, NY, USA: ACM, 2016).Elith, J., Leathwick, J. R. & Hastie, T. A working guide to boosted regression trees. J. Anim. Ecol. 77, 802–813 (2008).Article 
CAS 

Google Scholar 
ArcGIS Pro 2.7.0 (Environmental Systems Research Institute) (2020).Ferreira, L. C. & Simpfendorer, C. Galeocerdo cuvier. The IUCN Red List of Threatened Species 2019, e.T39378A2913541 (2019).Beta Regression v. 3.1-4 (2021).Butchart, S. H. et al. Improvements to the Red List Index. PLoS ONE. 2, e140 (2007).Article 
ADS 

Google Scholar 
Sherman, C. S. et al. Half a century of rising extinction risk of coral reef sharks and rays, sammsherman27/CoralReefSharkRayIUCN: Data and Code Used in Sherman et al. Half a century of rising extinction risk of coral reef sharks and rays v1.0.0. https://doi.org/10.5281/zenodo.7267904 (2022). More

D’Costa, V. M. et al. Antibiotic resistance is ancient. Nature 477, 457–461. https://doi.org/10.1038/nature10388 (2011).Article 
ADS 

Google Scholar 
Cytryn, E. The soil resistome: The anthropogenic, the native, and the unknown. Soil Biol. Biochem. 63, 18–23. https://doi.org/10.1016/j.soilbio.2013.03.017 (2013).Article 

Google Scholar 
Holmes, A. H. et al. Understanding the mechanisms and drivers of antimicrobial resistance. Lancet 387, 176–187. https://doi.org/10.1016/S0140-6736(15)00473-0 (2016).Article 

Google Scholar 
Regulation (EC) No 1831/2003 of the European parliament and of the council of 22 September 2003 on additives for use in animal nutrition.European Commission. Communication from the commission to the European parliament, the council, the European economic and social committee and the committee of the regions: A farm to fork strategy for a fair, healthy and environmentally-friendly food system COM/2020/381 Final, (2020).Kumar, K. C., Gupta, S. C., Chander, Y. & Singh, A. K. Antibiotic use in agriculture and its impact on the terrestrial environment. Adv. Agron. 87, 1–54. https://doi.org/10.1016/S0065-2113(05)87001-4 (2005).Article 

Google Scholar 
Chee-Sanford, J. C. et al. Fate and transport of antibiotic residues and antibiotic resistance genes following land application of manure waste. J. Environ. Qual. 38, 1086–1108. https://doi.org/10.2134/jeq2008.0128 (2009).Article 

Google Scholar 
Heuer, H., Schmitt, H. & Smalla, K. Antibiotic resistance gene spread due to manure application on agricultural fields. Curr. Opin. Microbiol. 14, 236–243. https://doi.org/10.1016/j.mib.2011.04.009 (2011).Article 

Google Scholar 
Epelde, L. et al. Characterization of composted organic amendments for agricultural use. Front. Sustain. Food Syst. 2, 44. https://doi.org/10.3389/fsufs.2018.00044 (2018).Article 

Google Scholar 
Youngquist, C. P., Mitchell, S. M. & Cogger, C. G. Fate of antibiotics and antibiotic resistance during digestion and composting: A review. J. Environ. Qual. 45, 537–545. https://doi.org/10.2134/jeq2015.05.0256 (2016).Article 

Google Scholar 
Ma, X., Xue, X., González-Mejía, A., Garland, J. & Cashdollar, J. Sustainable water systems for the city of tomorrow: A conceptual framework. Sustainability 7, 12071–12105. https://doi.org/10.3390/su70912071 (2015).Article 

Google Scholar 
Wang, Y. et al. Degradation of antibiotic resistance genes and mobile gene elements in dairy manure anerobic digestion. PLoS ONE 16, e0254836. https://doi.org/10.1371/journal.pone.0254836 (2021).Article 

Google Scholar 
Thanomsub, B. et al. Effects of ozone treatment on cell growth and ultrastructural changes in bacteria. J. Gen. Appl. Microbiol. 48, 193–199. https://doi.org/10.2323/jgam.48.193 (2002).Article 

Google Scholar 
Sousa, J. M. et al. Ozonation and UV254nm radiation for the removal of microorganisms and antibiotic resistance genes from urban wastewater. J. Hazard. Mater. 323, 434–441. https://doi.org/10.1016/j.jhazmat.2016.03.096 (2017).Article 

Google Scholar 
Park, J. H., Choppala, G. K., Bolan, N. S., Chung, J. W. & Chuasavathi, T. Biochar reduces the bioavailability and phytotoxicity of heavy metals. Plant Soil 348, 439–451. https://doi.org/10.1007/s11104-011-0948-y (2011).Article 

Google Scholar 
Jeffery, S. et al. The way forward in biochar research: targeting trade-offs between the potential wins. GCB Bioenergy 7, 1–13. https://doi.org/10.1111/gcbb.12132 (2015).Article 

Google Scholar 
Krasucka, P. et al. Engineered biochar: A sustainable solution for the removal of antibiotics from water. Chem. Eng. J. 405, 126926. https://doi.org/10.1016/j.cej.2020.126926 (2021).Article 

Google Scholar 
Ken, D. S. & Sinha, A. Recent developments in surface modification of Nano zero-valent iron (nZVI): remediation, toxicity and environmental impacts. Environ. Nanotechnol. Monit. Manag. 14, 100344. https://doi.org/10.1016/j.enmm.2020.100344 (2020).Article 

Google Scholar 
Zhao, X. et al. An overview of preparation and applications of stabilized zero-valent iron nanoparticles for soil and groundwater remediation. Water Res. 100, 245–266. https://doi.org/10.1016/j.watres.2016.05.019 (2016).Article 

Google Scholar 
Diao, M. & Yao, M. Use of zero-valent iron nanoparticles in inactivating microbes. Water Res. 43, 5243–5251. https://doi.org/10.1016/j.watres.2009.08.051 (2009).Article 

Google Scholar 
Shi, C. J., Wei, J., Jin, Y., Kniel, K. E. & Chiu, P. C. Removal of viruses and bacteriophages from drinking water using zero-valent iron. Sep. Purif. Technol. 84, 72–78. https://doi.org/10.1016/j.seppur.2011.06.036 (2012).Article 

Google Scholar 
Anza, M., Salazar, O., Epelde, L., Alkorta, I. & Garbisu, C. The application of nanoscale zero-valent iron promotes soil remediation while negatively affecting soil microbial biomass and activity. Front. Environ. Sci. 7, 19. https://doi.org/10.3389/fenvs.2019.00019 (2019).Article 

Google Scholar 
FAOSTAT. Mushrooms and truffles, production quantity (tons). https://www.tilasto.com/en/topic/geography-and-agriculture/crop/mushrooms-and-truffles/mushrooms-and-truffles-production-quantity/spain, (2020).Polat, E., Uzun, H., Topçuo, B., Önal, K. & Onus, A. N. Effects of spent mushroom compost on quality and productivity of cucumber (Cucumis sativus L.) grown in greenhouses. Afr. J. Biotechnol. 8, 176–180 (2009).
Google Scholar 
Fazaeli, H. & Masoodi, A. R. T. Spent wheat straw compost of Agaricus bisporus mushroom as ruminant feed. Asian-Australas. J. Anim. Sci. 19, 845–851. https://doi.org/10.5713/ajas.2006.845 (2006).Article 

Google Scholar 
Phan, C. W. & Sabaratnam, V. Potential uses of spent mushroom substrate and its associated lignocellulosic enzymes. Appl. Microbiol. Biotechnol. 96, 863–873. https://doi.org/10.1007/s00253-012-4446-9 (2012).Article 

Google Scholar 
Lau, K. L., Tsang, Y. Y. & Chiu, S. W. Use of spent mushroom compost to bioremediate PAH-contaminated samples. Chemosphere 52, 1539–1546. https://doi.org/10.1016/S0045-6535(03)00493-4 (2003).Article 
ADS 

Google Scholar 
Mayans, B. et al. An assessment of Pleurotus ostreatus to remove sulfonamides, and its role as a biofilter based on its own spent mushroom substrate. Environ. Sci. Pollut. Res. Int. 28, 7032–7042. https://doi.org/10.1007/s11356-020-11078-3 (2021).Article 

Google Scholar 
Congilosi, J. L. & Aga, D. S. Review on the fate of antimicrobials, antimicrobial resistance genes, and other micropollutants in manure during enhanced anaerobic digestion and composting. J. Hazard. Mater. 405, 123634. https://doi.org/10.1016/j.jhazmat.2020.123634 (2021).Article 

Google Scholar 
Oliver, J. P. et al. Invited review: fate of antibiotic residues, antibiotic-resistant bacteria, and antibiotic resistance genes in US dairy manure management systems. J. Dairy Sci. 103, 1051–1071. https://doi.org/10.3168/jds.2019-16778 (2020).Article 

Google Scholar 
Beneragama, N. et al. Survival of multidrug-resistant bacteria in thermophilic and mesophilic anaerobic co-digestion of dairy manure and waste milk. Anim. Sci. J. 84, 426–433. https://doi.org/10.1111/asj.12017 (2013).Article 

Google Scholar 
Sun, W., Qian, X., Gu, J., Wang, X. J. & Duan, M. L. Mechanism and effect of temperature on variations in antibiotic resistance genes during anaerobic digestion of dairy manure. Sci. Rep. 6, 30237. https://doi.org/10.1038/srep30237 (2016).Article 
ADS 

Google Scholar 
Sun, W., Gu, J., Wang, X., Qian, X. & Peng, H. Solid-state anaerobic digestion facilitates the removal of antibiotic resistance genes and mobile genetic elements from cattle manure. Bioresour. Technol. 274, 287–295. https://doi.org/10.1016/j.biortech.2018.09.013 (2019).Article 

Google Scholar 
Zou, Y., Xiao, Y., Wang, H., Fang, T. & Dong, P. New insight into fates of sulfonamide and tetracycline resistance genes and resistant bacteria during anaerobic digestion of manure at thermophilic and mesophilic temperatures. J. Hazard. Mater. 384, 121433. https://doi.org/10.1016/j.jhazmat.2019.121433 (2020).Article 

Google Scholar 
Agga, G. E., Kasumba, J., Loughrin, J. H. & Conte, E. D. Anaerobic digestion of tetracycline spiked livestock manure and poultry litter increased the abundances of antibiotic and heavy metal resistance genes. Front Microbiol. 11, 614424. https://doi.org/10.3389/fmicb.2020.614424 (2020).Article 

Google Scholar 
Jauregi, L., Epelde, L., González, A., Lavín, J. L. & Garbisu, C. Reduction of the resistome risk from cow slurry and manure microbiomes to soil and vegetable microbiomes. Environ. Microbiol. 23, 7643–7660. https://doi.org/10.1111/1462-2920.15842 (2021).Article 

Google Scholar 
Zhang, Z. et al. Assessment of global health risk of antibiotic resistance genes. Nat Commun 13, 1553. https://doi.org/10.1038/s41467-022-29283-8 (2022).Article 
ADS 

Google Scholar 
He, Y. et al. Antibiotic resistance genes from livestock waste: occurrence, dissemination, and treatment. npj Clean Water 3, 4. https://doi.org/10.1038/s41545-020-0051-0 (2020).Article 

Google Scholar 
Cui, E., Wu, Y., Zuo, Y. & Chen, H. Effect of different biochars on antibiotic resistance genes and bacterial community during chicken manure composting. Bioresour. Technol. 203, 11–17. https://doi.org/10.1016/j.biortech.2015.12.030 (2016).Article 

Google Scholar 
Fu, Y., Zhang, A., Guo, T., Zhu, Y. & Shao, Y. Biochar and hyperthermophiles as additives accelerate the removal of antibiotic resistance genes and mobile genetic elements during composting. Materials (Basel) 14, 5428. https://doi.org/10.3390/ma14185428 (2021).Article 
ADS 

Google Scholar 
Forsberg, K. J. et al. Bacterial phylogeny structures soil resistomes across habitats. Nature 509, 612–616. https://doi.org/10.1038/nature13377 (2014).Article 
ADS 

Google Scholar 
Li, H. et al. Effects of bamboo charcoal on antibiotic resistance genes during chicken manure composting. Ecotoxicol. Environ. Saf. 140, 1–6. https://doi.org/10.1016/j.ecoenv.2017.01.007 (2017).Article 
ADS 

Google Scholar 
Bondarenko, O., Ivask, A., Käkinen, A. & Kahru, A. Sub-toxic effects of CuO nanoparticles on bacteria: Kinetics, role of Cu ions and possible mechanisms of action. Environ. Pollut. 169, 81–89. https://doi.org/10.1016/j.envpol.2012.05.009 (2012).Article 

Google Scholar 
Wang, X. et al. Bacterial exposure to ZnO nanoparticles facilitates horizontal transfer of antibiotic resistance genes. NanoImpact 10, 61–67. https://doi.org/10.1016/j.impact.2017.11.006 (2018).Article 
ADS 

Google Scholar 
Qiu, X., Zhou, G. & Wang, H. Nanoscale zero-valent iron inhibits the horizontal gene transfer of antibiotic resistance genes in chicken manure compost. J. Hazard. Mater. 422, 126883. https://doi.org/10.1016/j.jhazmat.2021.126883 (2022).Article 

Google Scholar 
Zeng, T., Wilson, C. J. & Mitch, W. A. Effect of chemical oxidation on the sorption tendency of dissolved organic matter to a model hydrophobic surface. Environ. Sci. Technol. 48, 5118–5126. https://doi.org/10.1021/es405257b (2014).Article 
ADS 

Google Scholar 
Pak, G. et al. Comparison of antibiotic resistance removal efficiencies using ozone disinfection under different pH and suspended solids and humic substance concentrations. Environ. Sci. Technol. 50, 7590–7600. https://doi.org/10.1021/acs.est.6b01340 (2016).Article 
ADS 

Google Scholar 
Zhuang, Y. et al. Inactivation of antibiotic resistance genes in municipal wastewater by chlorination, ultraviolet, and ozonation disinfection. Environ. Sci. Pollut. Res. Int. 22, 7037–7044. https://doi.org/10.1007/s11356-014-3919-z (2015).Article 

Google Scholar 
Park, S., Rana, A., Sung, W. & Munir, M. Competitiveness of quantitative polymerase chain reaction (qPCR) and droplet digital polymerase chain reaction (ddPCR) technologies, with a particular focus on detection of antibiotic resistance genes (ARGs). Appl. Microbiol. 1, 426–444. https://doi.org/10.3390/applmicrobiol1030028 (2021).Article 

Google Scholar 
European Medicines Agency. European surveillance of veterinary antimicrobial consumption, (2020). Sales of Veterinary Antimicrobial Agents in 31 European Countries in 2018 (EMA/24309/2020).Heuer, H. et al. The complete sequences of plasmids pB2 and pB3 provide evidence for a recent ancestor of the IncP-1β group without any accessory genes. Microbiology (Reading) 150, 3591–3599. https://doi.org/10.1099/mic.0.27304-0 (2004).Article 

Google Scholar 
World Health Organization. Critically Important Antimicrobials for Human Medicine, 6th Revision (WHO, Geneva, Switzerland, 2019).Zhu, Y. G. et al. Diverse and abundant antibiotic resistance genes in Chinese swine farms. Proc. Natl Acad. Sci. U. S. A. 110, 3435–3440. https://doi.org/10.1073/pnas.1222743110 (2013).Article 
ADS 

Google Scholar 
Guo, T. et al. Increased occurrence of heavy metals, antibiotics and resistance genes in surface soil after long-term application of manure. Sci. Total Environ. 635, 995–1003. https://doi.org/10.1016/j.scitotenv.2018.04.194 (2018).Article 
ADS 

Google Scholar 
Nõlvak, H. et al. Inorganic and organic fertilizers impact the abundance and proportion of antibiotic resistance and integron-integrase genes in agricultural grassland soil. Sci. Total Environ. 562, 678–689. https://doi.org/10.1016/j.scitotenv.2016.04.035 (2016).Article 
ADS 

Google Scholar 
Chen, Q. L. et al. Effect of biochar amendment on the alleviation of antibiotic resistance in soil and phyllosphere of Brassica chinensis L.. Soil Biol. Biochem. 119, 74–82. https://doi.org/10.1016/j.soilbio.2018.01.015 (2018).Article 

Google Scholar 
Zhu, B., Chen, Q., Chen, S. & Zhu, Y. G. Does organically produced lettuce harbor higher abundance of antibiotic resistance genes than conventionally produced?. Environ. Int. 98, 152–159. https://doi.org/10.1016/j.envint.2016.11.001 (2017) .Article 

Google Scholar 
Métodos, M. A. P. A. Oficiales de análisis de suelos y Aguas Para riego. Minist. Agric. Pesca Aliment. Métodos Oficiales Anal. III (1994).Muziasari, W. I. et al. Aquaculture changes the profile of antibiotic resistance and mobile genetic element associated genes in Baltic Sea sediments. FEMS Microbiol. Ecol. 92, fiw052. https://doi.org/10.1093/femsec/fiw052 (2016).Article 

Google Scholar 
Muurinen, J. et al. Influence of manure application on the environmental resistome under Finnish agricultural practice with restricted antibiotic use. Environ. Sci. Technol. 51, 5989–5999. https://doi.org/10.1021/acs.est.7b00551 (2017).Article 
ADS 

Google Scholar 
Muziasari, W. I. et al. The resistome of farmed fish feces contributes to the enrichment of antibiotic resistance genes in sediments below Baltic Sea fish farms. Front. Microbiol. 7, 2137. https://doi.org/10.3389/fmicb.2016.02137 (2017).Article 

Google Scholar 
Livak, K. J. & Schmittgen, T. D. Analysis of relative gene expression data using real-time quantitative PCR and the 2−ΔΔCT method. Methods 25, 402–408. https://doi.org/10.1006/meth.2001.1262 (2001).Article 

Google Scholar 
Ovreås, L., Forney, L., Daae, F. L. & Torsvik, V. Distribution of bacterioplankton in meromictic Lake Saelenvannet, as determined by denaturing gradient gel electrophoresis of PCR-amplified gene fragments coding for 16S rRNA. Appl. Environ. Microbiol. 63, 3367–3373. https://doi.org/10.1128/aem.63.9.3367-3373.1997 (1997) .Article 
ADS 

Google Scholar 
Caporaso, J. G. et al. Ultra-high-throughput microbial community analysis on the Illumina HiSeq and MiSeq platforms. ISME J. 6, 1621–1624. https://doi.org/10.1038/ismej.2012.8 (2012).Article 

Google Scholar 
Lanzén, A. et al. Multi-targeted metagenetic analysis of the influence of climate and environmental parameters on soil microbial communities along an elevational gradient. Sci. Rep. 6, 28257. https://doi.org/10.1038/srep28257 (2016).Article 
ADS 

Google Scholar 
Pinna, N. K., Dutta, A., Monzoorul, H. M. & Mande, S. S. Can targeting non-contiguous V-regions with paired-end sequencing improve 16S rRNA-based taxonomic resolution of microbiomes?: An in silico evaluation. Front. Genet. 10, 653. https://doi.org/10.3389/fgene.2019.00653 (2019).Article 

Google Scholar 
Andrews, S. FastQC: A quality control tool for high throughput sequence data. http://www.bioinformatics.babraham.ac.uk/projects/fastqc/ (2010).Martin, M. Cutadapt removes adapter sequences from high-throughput sequencing reads. EMBnet J. 17, 10–12. https://doi.org/10.14806/ej.17.1.200 (2011).Article 

Google Scholar 
Bolyen, E. et al. Reproducible, interactive, scalable and extensible microbiome data science using QIIME 2. Nat. Biotechnol. 37, 852–857. https://doi.org/10.1038/s41587-019-0209-9 (2019).Article 

Google Scholar 
Amir, A. et al. Deblur rapidly resolves single-nucleotide community sequence patterns. mSystems 2, e00191-e216. https://doi.org/10.1128/mSystems.00191-16 (2017).Article 

Google Scholar 
Yang, Y., Li, B., Zou, S., Fang, H. H. P. & Zhang, T. Fate of antibiotic resistance genes in sewage treatment plant revealed by metagenomic approach. Water Res. 62, 97–106. https://doi.org/10.1016/j.watres.2014.05.019 (2014).Article 

Google Scholar 
Wickham, H. ggplot2: Elegant Graphics for Data Analysis (Springer-Verlag, 2016).Book 
MATH 

Google Scholar 
de Mendiburu, F. Agricolae: Statistical procedures for agricultural research. R package version 1.3-3. https://CRAN.R-project.org/package=agricolae (2020).Paradis, E. & Schliep, K. ape 5.0: An environment for modern phylogenetics and evolutionary analyses in R. Bioinformatics 35, 526–528. https://doi.org/10.1093/bioinformatics/bty633 (2019).Article 

Google Scholar 
Oksanen, J. et al. Vegan: Community ecology package. R Package Version 2.3-1. (2015). More

Data collectionWe obtained trade data from two different sources: the United States Fish and Wildlife Service (USFWS) Law Enforcement Management Information System (LEMIS)31 and the International Union for Conservation of Nature (IUCN) Red List32. We used the former to obtain data on the live wildlife trade in general and the latter for data on the pet trade specifically. We then matched trade data with our previously compiled global scale datasets of life history traits and introductions in mammals, reptiles and amphibians25,26.We obtained data on the US live wildlife trade from LEMIS by a Freedom of Information Act Request on 12/08/2019. We requested summary data on all US imports and exports of wildlife across all available years (1999-2019) and all trade purposes, including information on species identities and shipment contents (e.g. live individuals, meat, skins, etc.). For each species, we summed the total number of recorded shipments of live individuals (including individuals that died in transit, and live eggs) as a measure of trade frequency. We classified species as in trade if there was at least one shipment of live individuals recorded in the LEMIS database, and as not traded otherwise. The LEMIS dataset is geographically limited to trade by the US, and therefore may not capture the full diversity of species involved in the wildlife trade. For example, the LEMIS database may be missing some species involved in the substantial trade in live wildlife between South–East Asian countries50. However, the US represents one of the most dominant players in the global market for live wildlife16, and by summing both imports and exports we capture demand for species in countries beyond the US to some extent. Supplementary Fig. 2 illustrates the frequency of trade between the US and countries represented in the US LEMIS dataset. LEMIS data should be considered a minimum estimate of the diversity of species involved in the wildlife trade since they mostly record only legal trade (although confiscated shipments are recorded), and shipments are sometimes not identified to the species level16,51,53,53. The LEMIS database also contains some mis-spelled and incorrectly identified species due to human input errors52. To minimise the effect of misidentified shipments on our species level classifications of US trade status, we discarded all LEMIS records that were not identified to the species level (i.e. those identified using genus, common or generic names only), and manually checked the LEMIS data for synonyms and alternate spellings when we could not automatically match any records in LEMIS with species in our life history datasets. Species classified as traded on the basis of a single recorded live shipment in LEMIS are most vulnerable to species level misclassification due to misidentified shipments. The vast majority of traded species have multiple shipments recorded in LEMIS (259/312 [83%] of traded mammals, 265/285 [93%] of traded reptiles and 72/75 [96%] of traded amphibians), reducing the potential impact of shipment level misidentification over the reliability of species level trade classifications. However, to investigate the robustness of our findings to possible errors in species identification in LEMIS, we re-ran our key analyses excluding species classified as traded on the basis of a single live shipment. We found qualitatively the same effects of life history traits on the probability of trade when removing these species as in our full sample (Supplementary Tables 25–27). Despite its limitations, LEMIS is an invaluable resource for identifying broad scale trends in the wildlife trade since few other countries maintain such detailed records, and it is the only large-scale international trade dataset that includes both CITES- and non-CITES-listed species16,41. Including non-CITES listed species in our analyses is important because CITES-listed species represent only a small minority of those in trade14 and are likely to be a biased sample in terms of life history traits, since species vulnerable to extinction typically have slower life histories40.We obtained separate data on the pet trade from the IUCN Red List. The IUCN has assessed the vast majority of mammal, reptile and amphibian species (91%, 79% and 86% respectively54). Here, we classified a species as involved in the pet trade if the IUCN species account included at least one clear description of involvement in the pet trade. Otherwise, we considered a species as not involved in the pet trade. Although LEMIS records the purpose of trade, it uses broad categories (e.g. ‘Commercial’, ‘Personal’, ‘Breeding in captivity’), none of which refers specifically to nor necessarily equates to trade for pets. Therefore, we sought this additional data on the pet trade from the IUCN Red List instead of following the approach of some previous studies which have used LEMIS data as a proxy for the pet trade (e.g. Refs. 15,19). In contrast, the IUCN Red List contains clear textual descriptions of use and trade for many species, allowing us to identify which species are traded specifically for pets32. The IUCN data has further complementary strengths compared with LEMIS in that it is global in scope and includes both legal and illegal trade. We obtained data from the IUCN Red List by manually searching the binomial name of each species in our samples and consulting the ‘Threats’ and ‘Use and Trade’ sections of the species accounts. We classified species as in the pet trade if the information clearly stated this was the case (e.g. “It has been recorded in the pet trade”, “This species appears in the international pet trade”). We discounted descriptions where the information was uncertain (e.g. the species is described as “probably” or “possibly” traded for pets). We did not count as pets those species that the IUCN categorises as used for “Pets/display animals, horticulture” but which are used only for zoos or captive display, such as beluga whales (Delphinapterus leucas). All species described as pets by the IUCN are ‘exotic’, i.e. those without a long history of domestication14, since the IUCN does not list domesticated species.We matched trade data with our previously published global scale datasets on life history traits and introductions25,26. Internationally traded species may or not be released in the wild outside their native range: some may remain in the confines of captivity (e.g. in zoos or kept by private owners). We defined a species as introduced if there was at least one reliable record of its release, by humans, into the wild outside of its native range, either accidentally or intentionally25,26. We included only species with complete data for the same life history traits as used in our prior analyses (mammals: body mass, gestation period, weaning age, neonatal body mass, litter size, litters per year, age at first reproduction and reproductive lifespan; reptiles: body mass, hatchling mass, clutch size, clutches per year, age of sexual maturity, reproductive lifespan and parity; amphibians: snout-vent length, egg size, clutch size, age of sexual maturity and reproductive lifespan) to facilitate direct comparisons with previous results and to allow us to account for covariation between life history traits55. Species with complete life history data represent 7.8%, 3.5% and 1.6% of the total estimated number of species of mammals, reptiles and amphibians respectively56,57,58. These samples are not random as they over-represent orders containing many species of interest and utility to humans (e.g. ungulates, primates, crocodilians) (Supplementary Tables 28–30). However, these biases are unlikely to undermine our results since we examine life history effects on trade and introduction within these samples. Trade and introduction data do not necessarily cover the same time periods: the US dataset covers only the years 1999-present and the IUCN descriptions also typically refer to recent trade. In contrast, our introduction dataset includes both historical and recent introductions25,26. Therefore, the goal of our analyses is not to test causal hypotheses on the direct relationship between trade and introduction but rather to investigate whether the same life history traits predispose species towards both trade and introduction across diverse taxa, locations and circumstances. When combining the datasets and phylogenies59,60,61,62,63, we resolved species name mis-matches by referring to taxonomic information from the IUCN Red List32, the Global Biodiversity Information Facility (GBIF33) and the Integrated Taxonomic Information System (ITIS64). Table 1 summarises final sample sizes and Supplementary Table 1 the degree of overlap between the trade datasets. Most species in the pet trade are also in the general live wildlife trade, but many more species are traded by the US for general purposes than are involved in the pet trade specifically.Finally, we obtained data for a proxy measure of species detectability in order to control for a potential confounding effect on relationships between life history traits and introduction: larger bodied and longer-lived species may be more likely to be recorded by human observers when introduced compared with smaller and shorter-lived species. We obtained data on species occurrence records, geographic range size and population density, assuming that highly detectable species will have a disproportionately large number of recorded observations than expected based on the size of their geographic ranges and average population densities, following similar approaches by e.g. Refs. 65,66. We obtained occurrence records from the Global Biodiversity Information Facility (GBIF33) via the R package rgbif67 selecting only records resulting from human observation. We obtained range sizes (in decimal degrees squared) from the IUCN Red List32 and processed them for analysis using functions from the rgdal package68, excluding areas of uncertain presence (i.e. limiting range to presence code 1, ‘extant’). We obtained population density estimates from the TetraDENSITY database (version 134), a global database of population density estimates for terrestrial vertebrates. Most species in the TetraDENSITY dataset are represented by estimates from multiple different studies (median = 3, range 1–408). We collapsed density estimates to the species level by taking the median value across studies, including all estimates regardless of sampling method to maximise sample size, and converting all units to individuals/km2 to ensure comparability.Statistical analysesTo investigate relationships between life history traits and trade, we run models treating US or pet trade as the outcome variable and life history traits as the predictors. For all analyses, all life history variables were included in the same models to account for covariation among life history traits55. For US trade, where data on trade frequency are available, we run models both in which trade is treated as a binary variable (traded vs. not traded) and as a count variable (frequency of live shipments, including zero values), while for the pet trade, we have no data on trade frequency and so we treat pet trade as a binary variable only. To investigate the effects of life history traits on introduction, we run models in which introduction is the outcome variable and life history traits are the predictors. In introduction models, we only include traded species (running separate models for the set of species in US trade and the set of species in the pet trade). This approach allows us to disentangle effects associated with trade and introduction and thus identify at which stage(s) life history biases emerge. We also run introduction models in which frequency of US trade is included as an additional predictor alongside life history traits, anticipating that highly traded species are more likely to be introduced. Finally, to investigate possible confounding effects of species detectability on relationships between life history traits and introduction, we investigate effects of number of observations, geographic range size and, where sample sizes allowed, population density on the probability of introduction. If highly detectable species are more likely to be recorded as introduced, we expect to find a positive effect of the number of observations (while accounting for geographic range size and population density) on the probability of introduction. If this effect confounds relationships between body mass/lifespan and introduction, the effect of these life history traits on the probability of introduction should disappear when detectability measures are included in the models alongside life history traits. All analyses were conducted using the R statistical programming environment (Version 4.2.069). Plots were coloured using palettes from the viridis package70.To estimate effects of predictor variables, we fit generalized linear mixed models (GLMMs) using Markov chain Monte-Carlo (MCMC) estimation, implemented in the MCMCglmm package35,36. For analyses with binary outcome variables (traded vs. not traded, introduced vs. not introduced) we fit probit models, while for analyses with US trade frequency as the outcome variable we fit hurdle models. Hurdle models estimate two latent variables: the probability that the outcome is zero (on the logit scale), and the probability of the outcome modelled as a Poisson distribution for non-zero values71. This method therefore allows us to estimate effects of life history traits on the probability and frequency of trade in the same model. While the binary component of a hurdle model estimates the probability that outcomes are zero, when reporting results we reverse the sign of coefficients from the binary model for ease of interpretation, so that effects can be interpreted as the probability that the outcome is not zero. Therefore, here predictors with consistent effects on the probability and frequency of trade in hurdle models will have the same sign (so that if, for example, litter size has a positive effect on both the probability and frequency of trade, both coefficients for litter size from the hurdle model will be positive).Datasets comprising biological measures from multiple related species violate the fundamental statistical assumption that observations are independent of one another, since closely related species are more phenotypically similar than expected by chance due to their shared evolutionary history72. To account for the non-independence of species due to shared ancestry, we included a phylogenetic random effect in all models, represented by a variance-covariance (VCV) matrix derived from the phylogeny. The off-diagonal elements of the VCV matrix contain the amount of shared evolutionary history for each pair of species35,37,38 based on the branch lengths of the phylogeny (here proportional to time)59,61,62,63,63. This approach allows us to estimate phylogenetic signal using the heritability (H2) parameter, which measures the proportion of total variance in the latent variable attributable to the phylogeny35,37,38. Heritability is interpreted in the same way as Pagel’s λ in phylogenetic generalized least squares regression35,37,38,72. Specifically, phylogenetic signal is constrained between 0, indicating no phylogenetic effect so that species can be treated as independent, and 1, indicating that similarity between species is directly proportional to their amount of shared evolutionary history35,38,72. As hurdle models estimate two latent variables, for each hurdle model we report two heritability estimates, one for the binary and one for the Poisson component. All continuous independent variables were log-10 transformed due to positively skewed distributions. Although GLMMs do not require normally distributed predictor variables, log-transforming positively skewed life history predictors in phylogenetic comparative analyses allows us to model life history evolution on proportional rather than absolute scales. This is important as it facilitates biologically meaningful comparisons between species across large scales of life history variation73. Further, log-transforming positively skewed predictors helps to meet assumptions of the underlying Brownian motion model of evolutionary change, which assumes that phenotypic change along branches of the phylogeny is normally distributed74.We calculated variance inflation factors (VIFs) using functions from the car R package75 to check for multicollinearity between predictor variables. Where any model reported a variance inflation factor of 5 or above, indicating potentially problematic levels of collinearity76, we re-ran the model removing the variable with the highest VIF iteratively until all the remaining variables had VIFs of More