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58.4: An Evolutionary Perspective on the Biodiversity Crisis

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    This discussion is excerpted from Santamaría, L., & Mendez, P. F. (2012). Evolution in biodiversity policy–current gaps and future needs. Evolutionary applications5(2), 202-218.  Blackwell Publishing Ltd under CC-BY-NC license. Accessed from  

    https://www.ncbi.nlm.nih.gov/pmc/articles/PMC3353340/ on 14 November 2023

    To date, biodiversity policy largely rested in the assumption that evolutionary processes take place at a temporal scale that largely exceeds that of most human operations. However, evidence indicating that detectable evolutionary changes commonly occur over ecological time scales has mounted over the last decade (e.g. Thompson 1998). Rapid evolutionary changes often arise in response to new forms of selection caused (directly or indirectly) by human action – termed ‘anthropogenic selection’, as opposed to natural selection (Palumbi 2001Stockwell et al. 2003). Anthropogenic evolution is widespread in nature, and numerous recent examples show that anthropogenic trait change in the wild is a global phenomenon, documented in marine, freshwater and terrestrial ecosystems worldwide (Palkovacs 2011). Moreover, and because eco-evolutionary dynamics are inherently bi-directional, contemporary evolution can have important effects on the dynamics of populations, communities and ecosystems; these effects may occur over large spatial scales and impact system-wide processes, such as trophic cascades (Carroll et al. 2007; Palkovacs 2011).

    Anthropogenic trait changes take place, in the wild, in two primary contexts: anthropogenic disturbance (especially harvest, but also habitat loss and fragmentation, pollution/acidification, and climate change) and biotic exchange (Hendry et al. 2008). Harvest is, probably, the most potent agent of anthropogenic trait change. Trait changes associated with the harvest of wild populations are, on average, three times faster than those caused by non-anthropogenic selection (Darimont et al. 2009). Fisheries, for example, drive the evolution of earlier age and smaller size at maturation in target populations, which affects their population persistence and sustainable yield (Hutchings and Fraser 2008) and results in considerable impacts on food-web interactions, trophic cascades and nutrient cycling in aquatic ecosystems (Palkovacs 2011). For example, fisheries of northwest Atlantic cod resulted, between the mid 1950s and the early 1990s, in estimated declines in age at 50% maturity from 6.5–7.0 to 5.0–5.5 years, resulting in an estimated reduction of population growth by 25–30% (Hutchings and Fraser 2008). These evolutionary consequences are often difficult to reverse; in some cases, the reduction or even cessation of fishing does not lead to rapid population recovery, particularly if directional selection continues over protracted periods as fisheries continue to harvest the largest available individuals (Conover 2000Stockwell et al. 2003). Indeed, one prediction common to all studies of fisheries-induced evolution is that genetic change effected by exploitation will be slow to reverse (Hutchings and Fraser 2008) – as confirmed by the persistence of small size-at-age in some populations of Atlantic cod, for at least 15 years after the cessation of heavy fishing and despite favourable environmental conditions for growth (Swain et al. 2007).

    Sport hunting, the main cause of death for prime-aged adults in many populations of ungulates, may result in selective effects that affect their morphological and life-history traits, favoring an earlier reproduction and increased reproductive investment in young adults – particularly when combined with regulations prohibiting the killing of lactating females, which enhance the survival of early-reproducing ones (Festa-Bianchet 2003Fenberg and Roy 2007). Trophy hunting selects for smaller horn/antler size, delayed horn/antler development and earlier reproduction – influencing male reproductive success and the economic profitability of harvested populations (Festa-Bianchet 2003). Comparable trends may be expected in game-bird species, especially those subjected to trophy hunting (such as capercaillie and black grouse). Even poaching may result in evolutionary pressures that compromise the long-term viability of poached populations – for example, poaching of African elephants for the illegal ivory trade may select for tusklessness (Jachmann et al. 1995).

    Habitat loss and fragmentation, the main global driver of biodiversity loss, often result in reduced population size and increased isolation of the affected biota, which in turns erodes its genetic variation and reduces its evolutionary potential. Small, isolated populations are subject to genetic drift and inbreeding; these processes tend to cause decreased fitness, decreased tolerance to environmental stress, and impeded adaptive responses to changing environmental conditions (K. Bijlsma in Grant et al. 2010). Habitat fragmentation can also impact traits related to migration, movement and habitat selection, with dramatic ecological consequences – such as the reduction of marine-derived subsidies to continental waters, or changed food-web interactions that in turn trigger new evolutionary responses in prey species (see examples in the study by Palkovacs 2011). While the destructive aspects of fragmentation may be accompanied by evolutionary opportunities (e.g. genetic drift may promote evolutionary processes, isolation may promote local adaptation and the rise of evolutionary novelties; F. Boero and F. Bonhomme in Grant et al. 2010), genetic erosion in permanently small, fragmented populations will generally result in decreased adaptive potential, impaired evolutionary processes and local extinctions (K. Bijlsma in Grant et al. 2010). Some species are, however, able to broaden their functional niche and make use of the anthropogenic matrix – a process that may involve evolutionary changes in traits related to both perception and dispersal (Van Dijk 2011).

    Climate change is expected to result in shifts in the geographical distribution and phenology of natural populations, as well as in the (local or global) extinction of species unable to counter its speed and magnitude. Rapid evolutionary adaptation can help species counter stressful conditions or realize ecological opportunities arising from climate change, influencing the resulting patterns of colonization, extinction and distribution shifts (Hoffmann and Sgrò 2011). The ‘evolving metacommunity’ framework (Urban 2011) emphasizes that interactions between ecological and evolutionary mechanisms, taking place at both local and regional scales, will drive community dynamics during climate change. In particular, ecological and evolutionary dynamics are likely to interact to produce outcomes different from those predicted based on either mechanism alone. While some of these dynamics have received recent attention (e.g. species interactions may prevent adaptation of other species to new niches, and resident species may adapt to changing climates and thereby prevent colonization by other species; Urban 2011), the realization that we know much more about how climates will change across the globe than about the likely responses of species to these changes and their effects on global biological diversity is profoundly worrisome.

    Biotic exchange, by which species are moved beyond the limits of their normal geographical ranges by human actions – often to produce biological invasions (Blackburn et al. 2011), can also affect the rates and the trajectories of evolutionary change (Shine 2011). Evolutionary responses are important to both predict the likelihood of biological invasions and manage the spread and impact of already-established invaders. These responses are double-sided: invasive species can induce rapid evolutionary responses on native taxa, which may reduce their ecological impact or exploit the opportunities provided by them, but the invasion process itself can cause substantial evolutionary shifts in invader's traits (Cox 2004Carroll 2007; Shine 2011). Many of these changes are adaptive, but others may result from nonadaptive evolutionary processes (e.g. spatial sorting; Shine 2011). From an applied point of view, evolutionary changes influencing the invader's dispersal rate and establishment ability are particularly important.

    Readers are also referred to Myers, N., & Knoll, A. H. (2001). The biotic crisis and the future of evolution. Proceedings of the National Academy of Sciences98(10), 5389-5392. Accessed from  https://www.pnas.org/doi/abs/10.1073/pnas.091092498 on 14 November 2024

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    Carroll SP. Facing change: forms and foundations of contemporary adaptation to biotic invasions. Molecular Ecology. 2007;17:361–372.

    Conover D. Darwinian fishery science. Marine Ecology Progress Series. 2000;208:299–313.

    Cox GW. Alien Species and Evolution. Washington: Island Press; 2004.

    Darimont CT, Carlson SM, Kinnison MT, Paquet PC, Reimchen TE, Wilmers CC. Human predators outpace other agents of trait change in the wild. Proceedings of the National Academy of Sciences. 2009;106:952–954. 

    Fenberg PB, Roy K. Ecological and evolutionary consequences of size-selective harvesting: how much do we know? Molecular Ecology. 2007;17:209–220.

    Festa-Bianchet M. Exploitative wildlife management as a selective pressure for life-history evolution of large mammals. In: Festa-Bianchet M, Apollonio M, editors. Animal Behavior and Wildlife Conservation. Washington, DC: Island Press; 2003. pp. 191–207. 

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    58.4: An Evolutionary Perspective on the Biodiversity Crisis is shared under a CC BY-NC 3.0 license and was authored, remixed, and/or curated by LibreTexts.

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