Threat-based laws

Our review of the literature indicated that best practice principles require the law to address threats to biodiversity (Scheele et al., Reference Scheele, Legge, Armstrong, Copley, Robinson, Southwell, Westgate and Lindenmayer2018) with the identification of threats and the necessary actions to address them based on scientific input (Evans et al., Reference Evans, Che-Castaldo, Crouse, Davis, Epanchin-Niell, Flather, Frohlich, Goble, Li, Male, Master, Moskwik, Neel, Noon, Parmesan, Shwartz, Scott and Williams2016) and Traditional and local knowledges (Artelle et al., Reference Artelle, Zurba, Bhattacharyya, Chan, Brown, Housty and Moola2019). Best practice also means that threat abatement should also occur simultaneously with protection policies and laws, with processes in place to mitigate the threat source (Hutchings et al., Reference Hutchings, Stephens and VanderZwaag2016).

Our review also found that laws should adopt a scaled approach, and biodiversity conservation law and relevant policies should accommodate different species’ threat levels and incorporate varying responses depending on that threat. For example, species with a small population and narrow habitat are specifically vulnerable to land use change (Sykes et al., Reference Sykes, Santini, Etard and Newbold2020) which enhances extinction risk (Staude et al., Reference Staude, Navarro and Pereira2020). The law should also provide for decision-makers and authorities to undertake a science-based assessment of the level of threat to a species (e.g., vulnerable, endangered) and the specific source of the threat (e.g., land clearing), with species that face a higher risk of extinction to be the focus of immediate and concerted conservation actions (Kraus et al., Reference Kraus, Murphy and Armitage2021).

Law and policy should also allow for threat-based laws to adapt to changing circumstances and threats that arise in the future, adequately abating the drivers of species extinction as they evolve. For example, incidental catching of seabirds during oceanic longline fishing operations was listed as a key threatening process under the Australian Environmental Protection and Biodiversity Conservation Act, but successful threat abatement plans have reduced the deaths of albatross and petrels in Australia (Baker et al., Reference Baker, Gales, Hamilton and Wilkinson2002). Elements recognised as contributing to the success of this threat abatement plan included semi-regular updates, implementation through regulations and management plans, wide support from stakeholders (especially government departments), criteria against which outcomes are measured, specific enforcement mechanisms, and a strong NGO advocate (Invasive Species Council and Bush Heritage Australia, 2020). The funding available to support the instrument is also likely to be critical to success (Bottrill et al., Reference Bottrill, Walsh, Watson, Joseph, Ortega-Argueta and Possingham2011).

Target-based laws

Our review identified that one of the most important features of best practice biodiversity conservation is the inclusion of goals, targets and indicators within biodiversity conservation laws (Maseyk et al., Reference Maseyk, Dominati and Mackay2019). Appropriate and explicit targets and indicators allow the effectiveness and efficiency of the law or policy to be measured, as well as ensure accountability for the success of prevention measures (Scheele et al., Reference Scheele, Legge, Armstrong, Copley, Robinson, Southwell, Westgate and Lindenmayer2018). Targets should communicate clear benchmarks for assessing whether objectives, principles or rules around a legal biodiversity mechanism have been met, which leads to further evaluation and adaptation (Evans et al., Reference Evans, Che-Castaldo, Crouse, Davis, Epanchin-Niell, Flather, Frohlich, Goble, Li, Male, Master, Moskwik, Neel, Noon, Parmesan, Shwartz, Scott and Williams2016). While specific targets and indicators may vary depending on a range of circumstances, biodiversity laws can still legislate avenues for these targets and indicators to be established. For example, the Australian state of Victoria has legislated a requirement that a Biodiversity strategy is prepared, which must include targets to measure achievements (Flora and Fauna Guarantee Act 1988 section 17), without legislating the specific targets in the law itself.

Broad targets and indicators have been proposed or implemented on an international scale, which can be reinforced in domestic or national law. The proposed Post-2020 Global Biodiversity Framework is an example of this, providing a framework for the development of national and regional goals and targets that further the objectives of the Convention on Biological Diversity (‘CBD’) (CBD Working Group, 2021). A biodiversity framework that incorporates an integrated system of goals, targets, and indicators is a best practice approach to biodiversity conservation. For example, Canada’s government released the 2020 Biodiversity Goals and Targets for Canada in response to global Aichi Biodiversity Targets. These included a range of quantitative and qualitative targets to be met by 2020, including percentages of land to be conserved through protected areas (17% of terrestrial areas and inland water and 10% of coastal marine areas protected by 2020), implementing priority adaption measures, reduction of pollution levels, and Aboriginal traditional knowledge regularly, meaningfully, and effectively informing biodiversity conservation and decision-making processes (Environment and Climate Change Canada, 2016).

Transparent and independent decision-making and review

The decision-making process in laws that aim to prevent species extinction should reflect transparency, accountability, and participation (van Doeveren, Reference van Doeveren2011). Laws should ensure that government and ministerial decision-making processes are scrutinised by independent review, with triggers in place for intervention in ineffective management processes. Independent or quasi-independent scientific authorities (such as an Environmental Protection Authority) may minimise ministerial discretion and prioritise evidence-based policy over political influences (Department of Environment and Science (Qld), 2022). Decision processes prescribed by law should be based on empirical scientific data and understanding, and be informed by independent scientific authorities (Evans et al., Reference Evans, Che-Castaldo, Crouse, Davis, Epanchin-Niell, Flather, Frohlich, Goble, Li, Male, Master, Moskwik, Neel, Noon, Parmesan, Shwartz, Scott and Williams2016; Dorey and Walker, Reference Dorey and Walker2018; Bethlenfalvy and Olive, Reference Bethlenfalvy and Olive2021). For this best practice principle to be reflected in law and policy, provisions should limit the discretion of decision-makers to carry out their functions based on social and economic factors that do not prioritise species extinction prevention.

The literature also suggests that biodiversity conservation law and policy should be transparent and made publicly available (Richardson, Reference Richardson2016; Scheele et al., Reference Scheele, Legge, Armstrong, Copley, Robinson, Southwell, Westgate and Lindenmayer2018; Hilty et al., Reference Hilty, Worboys, Keeley, Woodley, Lausche, Locke, Carr, Pulsford, Pittock, White, Theobald, Levine, Reuling, Watson, Ament, Groves and Tabor2020). This includes any law or decision-making process, as well as any monitoring, evaluation, triggers for intervention, costs and burdens, and terminology. Publication of these elements should also be accompanied by public consultation and review processes. Public consultation ensures participation in the decision-making process, with open invitations to the public to make submissions on legal mechanisms that aim to prevent species extinction (with accompanying provisions that require the decision-maker to properly consider these submissions). Best practice is also indicated by law that enables avenues for the review of these decisions, not only for those immediately affected by a decision (e.g., those prevented from performing an activity) but also instilling broader provisions for reviews for a wide range of parties who have an interest in preventing species extinction (Tsioumani, Reference Tsioumani and Faure2018).

Effective biodiversity conservation law should also be backwards-looking, and it has been argued that where a species extinction does occur there should be a public inquiry to examine the causes and actions that contributed to it (Woinarski et al., Reference Woinarski, Garnet, Legge and Lindenmayer2017; see also McCormack in this special issue). This can ensure accountability and review of previous decisions, and ensure that future decisions are informed by any identified shortcomings (Smith et al., Reference Smith, Allan, McGowan, Szymanski, Rogers-Oetker and Bell2018).

Inclusive decision-making

Best practice biodiversity conservation law should incorporate the principle of inclusive decision-making. Whilst this overlaps to an extent with public consultation in transparent decision-making as discussed above, inclusive decision-making as a distinct principle expands upon the former by reinforcing the need for laws to include all relevant members of the public or specific groups involved in decisions that affect the environment. Public participation is reflected in human rights law, for example, the inclusion of the right to effective participation of a minority population as part of the right to culture (Human Rights Committee, 1994). Public participation promotes democratic values within a governance system, with benefits including stakeholder empowerment, as they can influence matters that affect them, which leads to greater community cooperation and an increase in behavioural change, including a greater conservation ethic (Cattino and Reckien, Reference Cattino and Reckien2021; Hao et al., Reference Hao, Nyaranga and Hongo2022).

In the specific context of biodiversity conservation law, the literature suggests that best practice biodiversity conservation laws and policies should develop and maintain Indigenous collaborations and partnerships. These partnerships extend beyond mere participation and require the recognition of Indigenous authority in traditional ecological knowledge systems (Artelle et al., Reference Artelle, Zurba, Bhattacharyya, Chan, Brown, Housty and Moola2019). Literature suggests that biodiversity law and policy should not limit Indigenous rights and access to land and should aim to enhance the position of Indigenous peoples through the preservation of language and culture (Berkes and Davidson-Hunt, Reference Berkes and Davidson-Hunt2006; Reimerson, Reference Reimerson2013; Satterfield et al., Reference Satterfield, Gregory, Klain, Roberts and Chan2013). This principle is further discussed below.

The literature also notes that best practice inclusive decision-making in biodiversity law is also reflected in effective efforts to develop and maintain partnerships between private landholders and regulators (Langpap et al., Reference Langpap, Kerkvliet and Shogren2018; Lee and Wakefield-Rann, Reference Lee and Wakefield-Rann2021; Pannell et al., Reference Pannell, Buckley, Case and Norton2021).

Effective evaluation and review

Best practice biodiversity law and policy should include provisions that allow for effective evaluation and review of the law itself, as well as any function, activity, or program carried out under the law (Verschuuren et al., Reference Verschuuren, Mallarach, Bernbaum, Spoon, Brown, Borde, Brown, Calamia, Mitchell, Infield and Lee2021). Measuring the effectiveness of biodiversity mechanisms, through monitoring and evaluation, results in laws being adapted to changing circumstances, new scientific findings, and any ineffective management processes identified through the evaluation process (Hutchings et al., Reference Hutchings, Stephens and VanderZwaag2016). Monitoring and evaluation are integral to the effectiveness of international biodiversity frameworks. This is essential to provide baseline data against which to determine ‘effective’ implementation, promote accountability, track the progress of the conservation targets and promote policy and decision-making that is informed by evidence (Mascia et al., Reference Mascia, Pailler, Thieme, Rowe, Bottrill, Danielsen, Geldmann, Naidoo, Pullin and Burgess2014; Cardesa-Salzmann, Reference Cardesa-Salzmann and Faure2016). The literature also suggests that laws must also provide timely and public systems of review to the law and any programs carried out under the law, with monitoring and evaluation informing this review process. Operating in conjunction with established targets and indicators, monitoring and review programs are outlined in the proposed Post-2020 Global Biodiversity Framework (CBD Working Group, 2021). Legislated dates for the review of frameworks, as well as provisions for adaptive management principles, can assist with provisioning for this principle (Evans et al., Reference Evans, Che-Castaldo, Crouse, Davis, Epanchin-Niell, Flather, Frohlich, Goble, Li, Male, Master, Moskwik, Neel, Noon, Parmesan, Shwartz, Scott and Williams2016; Dimitropoulou, Reference Dimitropoulou2018; Scheele et al., Reference Scheele, Legge, Armstrong, Copley, Robinson, Southwell, Westgate and Lindenmayer2018).

Climate and biodiversity mainstreaming

One prominent theme that emerged from the literature was that biodiversity conservation requires mainstreaming in law and policy. Mainstreaming can be achieved through different means but in general, it is supported when a particular issue, in this case, climate change and biodiversity conservation, becomes a core consideration of the government, and government processes and systems are redesigned and re-organised from the perspective of addressing those issues (Forester and Bleby, Reference Forester and Bleby2022). Mainstreaming in biodiversity law – also referred to as horizontal policy integration (e.g., Lafferty and Hovden, Reference Lafferty and Hovden2003) – mainly involves placing one issue more centrally on the agenda of another domain. For example, species extinction prevention becomes a core concern of the legislation governing industries that typically affect biodiversity (i.e., forestry, agriculture, fisheries, water management, land management and planning) (Karlsson-Vinkhuyzen et al., Reference Karlsson-Vinkhuyzen, Kok, Visseren-Hamakers and Termeer2017). The literature diverges on the extent to which mainstreaming must be implemented in law and policy to constitute ‘best practice’. The argument that environmental factors must take priority over other concerns to constitute mainstreaming (Lafferty and Hovden, Reference Lafferty and Hovden2003) is contrasted with the notion that mainstreaming involves the process of merging the concerns of two domains (Jordan and Lenschow, Reference Jordan and Lenschow2008). As a best practice approach to mainstreaming, legislation should include an intention to make species extinction prevention a core priority of the law, forming a primary and compulsory directive or guiding principle, rather than only one factor to be considered. Social, cultural or economic considerations should be ancillary to the main goal of preventing species extinction to constitute biodiversity conservation mainstreaming (Tallis et al., Reference Tallis, Kennedy, Ruckelhaus, Goldstein and Kiesecker2015; Maseyk et al., Reference Maseyk, Dominati and Mackay2019). Threat and target-based laws, with objective criteria that are underpinned by scientific evidence and scrutiny, can ensure species extinction is prioritised over social or economic influences, with only very limited exceptions where environmental or social justice requires it. This means biodiversity and environmental concerns must not only be present in the broad objectives and principles of law, but also in the specific functions of authorities under the act.

Any function or activity which further prioritises species extinction prevention may constitute ‘mainstreaming’, with no exhaustive list of actions that may evince ‘mainstreaming’ and further the above principles. For example, mainstreaming may include the establishment of a central climate or biodiversity conservation agency that other departments will report to, reflecting structural integrity principles through enhancing coordination between government agencies for decision-making and reducing the cost of conservation bureaucracy (see, e.g., the Climate Change Act 2017 in the Australian state of Victoria). Other examples range from introducing compliance reporting, and developing Ministerial guidelines that direct proper consideration of species status in decision-making, to the inclusion of key performance indicators in executive contracts, performance plans and training modules – all of which reflect effective evaluation and review principles (Forester and Bleby, Reference Forester and Bleby2022). Mainstreaming can be furthered through a range of other, specific mechanisms, and as a result of this scope, reflect and further a range of best practice principles.

Indigenous collaborations and partnerships

The literature was clear that biodiversity law and policy that provides avenues for Indigenous collaborations and partnerships reflect inclusive decision-making best practice principles. However, biodiversity law and policy that promotes these collaborations should aim to balance conservation goals and human rights goals (i.e., the rights of recognition, and well-being of Indigenous peoples) to ensure mutual benefit for both parties (Artelle et al., Reference Artelle, Zurba, Bhattacharyya, Chan, Brown, Housty and Moola2019). Recognising Indigenous ecological knowledge can empower Indigenous peoples through joint decision-making (Berkes and Davidson-Hunt, Reference Berkes and Davidson-Hunt2006), as well as contribute to recovery from colonial assimilation, state violence, and ecological degradation (Satterfield et al., Reference Satterfield, Gregory, Klain, Roberts and Chan2013). In furthering conservation efforts, Indigenous peoples have an in-depth knowledge of land or place and are often well-positioned to provide effective monitoring and evaluation in biodiverse areas, with traditional knowledge informing biodiversity assessment and conservation priorities (Ban et al., Reference Ban, McDougall, Beck, Salomon and Cripps2014; Artelle et al., Reference Artelle, Zurba, Bhattacharyya, Chan, Brown, Housty and Moola2019).

These collaborations and partnerships, as a best practice mechanism, must recognise the authority and rights that Indigenous people can have about preventing species extinction while ensuring these partnerships benefit the position of Indigenous peoples (Goolmeer et al., Reference Goolmeer, Skroblin and Wintle2022b). The Indigenous Land and Sea Management Programs funded by the Australian Federal Government are examples of these goals being balanced. Such programs aim to advance biodiversity conservation and protect natural resources while also creating substantial employment and economic opportunities for Indigenous people, as well as improving the well-being of these communities through a wide range of social and cultural benefits (Larson et al., Reference Larson, Stoeckl, Jarvis, Addison, Grainger and Watkin Lui2020). Best practice Indigenous collaborations should also extend beyond rights to natural resources and land management in traditionally owned areas, but also incorporate traditional ecological knowledge in assessment processes and combine with western sciences to underpin decision-making processes.

Protected areas

Empirical evidence has suggested that identifying an area as protected can provide additional protection for biodiversity conservation. However, legislative provisions that establish and provide for the management of these protected areas must incorporate certain aspects to constitute best practice, including broad representation, size and connectivity and co-existence with other laws including species listing (Mitchell et al., Reference Mitchell, Stolton, Bezaury-Creel, Bingham, Cumming, Dudley, Fitzsimons, Malleret-King, Redford, Solano and Groves2018).

Reflective of the inclusive and transparent decision-making principles, input into the nomination of areas should extend beyond government authorities and scientific bodies to NGOs and community groups (Scott, Reference Scott2016). The New Zealand Marine Reserves Act demonstrates a particularly wide approach, with applications open to universities, private bodies, scientific research groups, Indigenous people who have tangata whenua status over the area, as well as the Director-General of Conservation (Marine Reserves Act 1971 s 5; Scott, Reference Scott2016). This could be broadened further by legislating that applications may be open to anyone with a special interest in nominating a protected area.

Law and policy must also recognise a need for larger and interconnected protected areas, specifically providing connectivity with existing protected areas, ecological corridors and Other Effective Area-based Conservation Measures (OECMs) (Hilty et al., Reference Hilty, Worboys, Keeley, Woodley, Lausche, Locke, Carr, Pulsford, Pittock, White, Theobald, Levine, Reuling, Watson, Ament, Groves and Tabor2020). Law and policy surrounding protected areas should provide avenues for interim protection measures, so that areas and the inhabiting species not yet legally listed may still be protected. For example, in New Zealand, land not designated as protected but has also not been fully assessed for its conservation value can be designated as ‘stewardship land’, allowing for greater or lesser protection measures once the assessment is complete (Koolen-Bourke and Peart, Reference Koolen-Bourke and Peart2021).

The principle that requires law and policy to be designed to address threats is particularly relevant for protected area laws. Threatened species can only benefit from the establishment of protected areas if the threat abatement of pervasive threats beyond land use, such as fire patterns, exotic plants, animals and pathogens, and visitors, occurs (Taylor et al., Reference Taylor, Sattler, Evans, Fuller, Watson and Possingham2011). Protected areas cannot exist in isolation to constitute best practice, but rather work in conjunction with other mechanisms to ensure these areas remain completely protected.

Listing and recovery plans

The listing of threatened species, and associated recovery plans for these species, were identified in the literature as best practice mechanisms when law and policy incorporate specific markers. The IUCN Red List categories and criteria are an example of a best practice for the listing process. This process provides an objective criterion as well as a standardised and transparent approach to evaluating a species’ status, with constant updates and improvements being implemented (Braby, Reference Braby2018). Species being associated with a distinct category (i.e., extinct, extinct in the wild, critically endangered) is reflective of the threat-based law principle, recognising the level of threat of each species based on population size reduction, geographic range, and other science-based criteria (Moir and Nand Brennan, Reference Moir and Nand Brennan2020). Using empirical scientific data as the basis for categorising a species was a prominent aspect of best practice, limiting considerations for socio-economic considerations and political influence (e.g., not listing a species as it may impact economic growth for a certain primary industry) (Dorey and Walker, Reference Dorey and Walker2018).

Other markers for best practice in listing include the need to list ecological communities, as well as individual species, broadening the impact listing processes can have (Kraus et al., Reference Kraus, Murphy and Armitage2021). Law that provides for the listing processes to identify and protect critical habitats is also considered an element of best practice, as is evident in the United States’ Endangered Species Act requiring that a listed species have a designated habitat (Henson et al., Reference Henson, White and Thompson2018). Law that considers listing of critical habitat and ecological communities allows for recovery to address threats to an area rather than a specific species (Walsh et al., Reference Walsh, Watson, Bottrill, Joseph and Possingham2013; Braby, Reference Braby2018; Kraus et al., Reference Kraus, Murphy and Armitage2021), which is reflective of law that is threat-based and serves to further this principle in the best practice mechanism. Walsh et al., argued that the concept of a national threatened species list, rather than specific state or provincial lists improves the listing processes’ efficiency, knowledge and data-sharing capabilities and effectively utilises limited funding (Walsh et al., Reference Walsh, Watson, Bottrill, Joseph and Possingham2013), although we acknowledge that this may also be contingent on this national list being accompanied by implementing legislation. Implementing this would further the structural integrity of the mechanism, with consolidated legislation ensuring more efficient coordination of resources and reducing any overlap or unnecessary bureaucracy that delays urgent action (although there may be discrepancies in a species’ level of risk in one area vs. another).

The literature described a range of best practice markers for the recovery plan mechanism. These included (Dee Boersma et al., Reference Dee Boersma, Kareiva, Fagan, Alan Clark and Hoekstra2001; Li et al., Reference Li, Malcom, Che-Castaldo and Neel2020):

• • legal avenues for adaptative plans that adjust to new information or changed conditions, with changes based on systematic monitoring and review;

• • legal requirements for implementation of monitoring; target-based laws that allow for timely and appropriate goals to be set and defined;

• • single-species focus to be balanced with multi-species and ecosystem-based recovery to ensure effective analysis is not sacrificed for efficiency;

• • and laws that strengthen inclusivity through non-government participation in the recovery of threatened species.

Recovery planning should also incorporate threat-based laws as a best practice principle, as recovery is likely to be most effective when aiming to ameliorate primary threats driving decline (Woinarski et al., Reference Woinarski, Garnet, Legge and Lindenmayer2017). In the past, recovery measures have failed to prevent species extinction in instances where primary threats to a species were not identified and not subject to threat-abatement measures, rendering recovery ineffective (Woinarski et al., Reference Woinarski, Garnet, Legge and Lindenmayer2017). The US Endangered Species Act (‘ESA’) implements some of these factors, including recovery plans that estimate the time and cost of recovery, descriptions of location-specific recovery actions, and criteria for recovering a species. However, recovery measures under the US System do not require an explanation of the causes behind a species’ decline, hindering the ability to mitigate the drivers of species extinction (Li et al., Reference Li, Malcom, Che-Castaldo and Neel2020).

Other issues with listing and recovery processes have also been raised, and these can indicate what best practice should incorporate. Some literature refers to a bias in these listing mechanisms, with less charismatic species receiving little attention regardless of their status (Dorey and Walker, Reference Dorey and Walker2018). Walsh et al. also identified the problem of bias in the listing of more charismatic species and noted that this may overlook some that are nevertheless vital for ecosystem function such as invertebrates, plants and fungi (Walsh et al., Reference Walsh, Watson, Bottrill, Joseph and Possingham2013). Some literature also identified a lack of accountability for or allocation of responsibility for the implementation of recovery measures (Woinarski et al., Reference Woinarski, Garnet, Legge and Lindenmayer2017).

Stewardship and PPAs

Stewardship, while varying in its formality (ranging from ‘hand-shake agreements’, to formalised programs) and PPAs, can generally be characterised by collaboration with private landowners, with efficient inclusivity a key principle in determining the best practice of this mechanism. Lee and Wakefield-Rann argue that this collaboration mechanism is a high priority due to the significant portion of natural resources located on private land, with the accessibility of this land and permissible management practices largely determined by landowners (Lee and Wakefield-Rann, Reference Lee and Wakefield-Rann2021). For example, the effectiveness of the US ESA is largely attributed to provisions that foster partnerships between government and private landowners, implementing conservation on private land through ‘Habitat Conservation Plans’ (Langpap et al., Reference Langpap, Kerkvliet and Shogren2018; Kraus et al., Reference Kraus, Murphy and Armitage2021).

The literature revealed that engaging landowners in species extinction prevention on private land is dependent upon close-working relationships as well as clear and accessible information on the benefits and costs of conservation (Pannell et al., Reference Pannell, Buckley, Case and Norton2021). Issues in the design and implementation of PPAs and stewardship arise due to the voluntary nature of this mechanism, relying on the willingness of landowners to participate in any conservation programs (Pannell et al., Reference Pannell, Buckley, Case and Norton2021). Embedding these ‘soft law’ programs in legislation can allow for greater institutional support. The mutual benefits of these mechanisms must be clearly communicated. Permits under the ESA enable landowners to develop lands inhabited by endangered species contingent on an approved Habitat Conservation Plan (HCP), which aims to mitigate harm to that species, while the ‘Conservation Banking’ programs allow landowners to protect habitat for endangered species and then sell conservation credits to developers to mitigate habitat alteration (Evans et al., Reference Evans, Che-Castaldo, Crouse, Davis, Epanchin-Niell, Flather, Frohlich, Goble, Li, Male, Master, Moskwik, Neel, Noon, Parmesan, Shwartz, Scott and Williams2016). The ‘Safe Harbour’ and ‘Candidate Conservation Agreements with Assurances’ programs also allow landowners to gain assurance that their voluntary actions to improve habitat or increase species numbers will not result in additional regulation (Evans et al., Reference Evans, Che-Castaldo, Crouse, Davis, Epanchin-Niell, Flather, Frohlich, Goble, Li, Male, Master, Moskwik, Neel, Noon, Parmesan, Shwartz, Scott and Williams2016). While limited in its effect, best practice of this mechanism is dependent on transparency and appropriate engagement with stakeholders, balancing compensation payments and regulatory assurances with conservation benefits (Evans et al., Reference Evans, Che-Castaldo, Crouse, Davis, Epanchin-Niell, Flather, Frohlich, Goble, Li, Male, Master, Moskwik, Neel, Noon, Parmesan, Shwartz, Scott and Williams2016). The ‘voluntary’ nature of these mechanisms means that ‘best practice’ generally requires that other actions, such as direct acqusition and protection of private land by conservation managers, are encouraged (Kearney et al., Reference Kearney, Carwardine, Reside, Adams, Nelson, Coggan, Spindler and Watson2022).

Ecosystem restoration

While the term ‘ecosystem restoration’ may cover a broad range of practices occurring in terrestrial spaces, there were general markers of best practice identified in the literature. This includes establishing legislative frameworks with clear and consistent terminology, including statutory definitions, to ensure greater public accessibility that enhances accountability (Nzyoka et al., Reference Nzyoka, Minang, Wainaina, Duguma, Manda and Temu2021). Clear and appropriate statutory goals should also be evident in the legislation, to ensure efforts are focused, targeted, and can be monitored to measure success (which can lead to adaption if deemed to be unsuccessful) (Gann et al., Reference Gann, McDonald, Waler, Aronson, Nelson, Jonson, Hallett, Eisenberg, Guariguata, Liu, Hua, Echeverria, Gonzales, Shaw and Dixon2019; Nzyoka et al., Reference Nzyoka, Minang, Wainaina, Duguma, Manda and Temu2021). Ecosystem restoration as a best practice mechanism also requires avenues for public participation and stakeholder support (Richardson, Reference Richardson2016; Campbell et al., Reference Campbell, Alexandra and Curtis2017; Gann et al., Reference Gann, McDonald, Waler, Aronson, Nelson, Jonson, Hallett, Eisenberg, Guariguata, Liu, Hua, Echeverria, Gonzales, Shaw and Dixon2019). This is a result of ecosystem restoration programs displacing people, which risks generating opposition that may hinder success (Richardson, Reference Richardson2016). Negating this effect relies on increasing social acceptability, which may involve legal avenues for offsetting disproportionate costs for a community or financial compensation, as well as sharing of experience and expertise (Richardson, Reference Richardson2016; Gann et al., Reference Gann, McDonald, Waler, Aronson, Nelson, Jonson, Hallett, Eisenberg, Guariguata, Liu, Hua, Echeverria, Gonzales, Shaw and Dixon2019). Implementing clear and consistent target-based laws, inclusive decision-making, and furthering general social acceptability is essential to garnering support for the demanding workload involved in ecosystem restoration. The literature also raised the dependency ecosystem restoration projects have on financial capital through public funding, with the suggestion that tax law is ‘reworked to stimulate philanthropic commitments while business law may overcome barriers to commercial approaches to eco-restoration’ (Richardson, Reference Richardson2016; Campbell et al., Reference Campbell, Alexandra and Curtis2017).

There are examples of these principles in ecosystem restoration projects. In the US, the Omnibus Public Land Management Act 2009 defines the specific ‘environmental baseline’ that the Collaborative Landscape Restoration Program aims to restore (the structure and composition of old-growth stands according to pre-fire suppression old growth conditions characteristic of the forest type), as well as outlining general ecological outcomes (improving fish and wildlife habitat, water quality, and preventing invasions of exotic species) (Richardson, Reference Richardson2016). In contrast, statutory goals and targets under New Zealand law are broad and unspecified (i.e., ‘restoring and enhancing the area and its heritage features, restoring and enhancing degraded landscapes’ (Waitakere Ranges Heritage Area Act 2008)). However, there are examples of local community participation in New Zealand restoration projects, with Māori and environmental groups establishing individual ‘eco-sanctuaries’, as well as private sectors shouldering the financial burden in the wake of budget cuts in the public sector (Richardson, Reference Richardson2016).

Offsets and no net loss schemes

The literature we reviewed was predominantly critical of offset and no net loss schemes (NNL), suggesting that offsets should only be used as a last resort by developers to compensate for unavoidable damage (Bull et al., Reference Bull, Suttle, Gordon, Singh and Milner-Gulland2013; Simmonds et al., Reference Simmonds, Sonter, Watson, Bennun, Cost, Dutson, Edwards, Grantham, Griffiths, Jones, Kiesecker, Possingham, Puydarrieux, Quetier, Rainer, Rainery, Roe, Savy, Souquet, Kate, Victurine, von Haxe and Maron2020a; Schoukens and Van Hoorick, Reference Schoukens, Van Hoorick, Pozzo and Jacomett2021). This is because offsets have suffered from persistent implementation failures due to weak compliance or regulatory enforcement, inconsistencies with governance arrangements (Zu Ermgassen et al., Reference Zu Ermgassen, Utamiputri, Bennun, Edwards and Bull2019), high costs for participation, and administrative delays (Martin et al., Reference Martin, Hamman, Leuzinger, Qin, Daya-Winterbottom and Du Plessis2022), as well as NNL policies being subject to the influence of uneven power dynamics or vested interests (Zu Ermgassen et al., Reference Zu Ermgassen, Utamiputri, Bennun, Edwards and Bull2019). As a result, environmental markets are largely viewed as not adequately addressing socio-ecological issues (Lockhart and Rea, Reference Lockhart and Rea2019), with minimal regard for their practicality (Zu Ermgassen et al., Reference Zu Ermgassen, Utamiputri, Bennun, Edwards and Bull2019). This suggests that offsets and NNL schemes may not be considered a best practice mechanism in biodiversity law and policy.

However, offsets and NNL policies are becoming increasingly relied upon by countries to mitigate the impact developments have on biodiversity (Gardner et al., Reference Gardner, Hase, Brownlie, Estrom, Pilgrim, Savy, Theo Stephens, Treweek, Ussher, Ward and Kate2013; Zu Ermgassen et al., Reference Zu Ermgassen, Utamiputri, Bennun, Edwards and Bull2019). This means recognising that potential best practice markers are important to develop and improve biodiversity offset outcomes. To understand, define and action these markers more research is needed and should be prioritised in circumstances where regulators wish to introduce offset policies. Examples of these best practice aspects include the need for offsets to always be accompanied by a no-net loss policy to be effective, the adoption of a precautionary approach to offset design, access to adequate data and technical expertise, economic and financial safeguards, and strong monitoring and enforcement, as well as outlining strict and explicit thresholds where impacts cannot be offset (Bull et al., Reference Bull, Suttle, Gordon, Singh and Milner-Gulland2013; Gardner et al., Reference Gardner, Hase, Brownlie, Estrom, Pilgrim, Savy, Theo Stephens, Treweek, Ussher, Ward and Kate2013; Benabou, Reference Benabou2014; Bergman et al., Reference Bergman, Binley, Murphy, Proctor, Ngyuyen, Urness, Vala, Vincent, Fahrig and Bennet2020).

Legal rights of nature

The literature recognised that both property rights and the sustainable development paradigm (or the misunderstanding of it) have resulted in a disproportionate emphasis on positive economic outcomes, leading to widespread environmental degradation (Riley, Reference Riley2019). Many commentators argued that adopting either an environmental personhood approach or recognising legal rights for the environment would be an effective means for overcoming these problems. The differences between the two approaches can be subtle, but important when considering how they can support genuine conservation. A limitation of the rights approach is that the protection offered by providing legal rights will be limited to the specific rights that are granted to either the environment or nature in general or to distinct ecosystems (Riley, Reference Riley2019). For instance, the rights granted could potentially include the right to restoration for any actions that cause environmental damage (Athens, Reference Athens2018). The alternative, which is recognising ‘legal personhood’ for the environment means that ‘an entity exists for its own interests and not for the value it contains for others’ (Athens, Reference Athens2018). There are several practical examples where jurisdictions have extended legal rights to the environment. The recognition of environmental legal personhood has been both sporadic and jurisdiction-specific however, which makes it difficult to evaluate the effectiveness of this as a legal mechanism to support conservation (Putzer et al., Reference Putzer, Lambooy, Jeurissen and Kim2022).

The first constitutional recognition of rights to the environment was in 2008 when Ecuador passed a constitutional amendment acknowledging the rights of nature (Athens, Reference Athens2018). However, arguably a more effective (but less general approach) was taken in New Zealand when the Parliament passed the Te Urewera Act in 2014. This Act embraced the traditional values of the Māori people in support of adopting legal personhood for Te Urewera national park and the Whanganui River. In this instance, the rights were attached to particular ecosystems but gave those ecosystems all the “rights, powers, duties and liabilities of a legal person” (Te Awa Tupua (Whanganui River Claims Settlement) Act 2017 (NZ) s 14(1)) Overall the literature suggests that although this approach aligns with a recognition of the intrinsic value of the environment, it is by no means a panacea for biodiversity conservation (Putzer et al., Reference Putzer, Lambooy, Jeurissen and Kim2022).

Despite recognising this, ecocentric approaches to ownership could be a useful mechanism for environmental protection laws (Hoops, Reference Hoops2022). Commentators considered this (and the limits of western legal systems’ property rights) both for the environment and for fauna impacted by environmental degradation (Best, Reference Best2021). Indeed, some commentators argue that the legal status of animals as property has significantly increased their vulnerability in times of disaster (Best, Reference Best2021).

In terms of best practice markers, these are difficult to articulate across jurisdictions generally as different nations have different requirements for constitutional amendments and likewise may approach rights differently. The New Zealand example demonstrated that common law systems can indeed provide the flexibility to support legal personhood for the environment. Where conservation of biodiversity can sit outside regulations (such as through constitutional recognition), a rights-based approach may indeed lead to long-term benefits.