what part of mining are ecosystems? defining success ...

Chapter 4

WHAT PART OF MINING ARE ECOSYSTEMS? DEFINING SUCCESS FOR THE ‘RESTORATION’ OF HIGHLY DISTURBED LANDSCAPES D. Doley1,* and P. Audet2,† 1

Honorary Research Fellow, Centre for Mined Land Rehabilitation, University of Queensland, Australia 2 Senior Scientist, Terrestrial Ecology & Land Reclamation Group, EDI Environmental Dynamics Inc., Canada

SYNOPSIS Contemporary mining activities are predicated on the expectation that (a) mining is a temporary land-use and (b) proponents should return affected lands to a suitable postdisturbance ecological condition. As such, mine closure criteria are often set with the assumption that any desired ecosystem can be established on the reconstructed landscape and near/natural recovery processes can be fostered through carefully planned environmental management. However, most mining activities are sufficiently disruptive that fundamental characteristics of the post-disturbance environment bear very little semblance to the original landscape, such as its lithology, hydrology, soil biogeochemistry, and biodiversity. Consequently, there are few examples of ‘successful’ mine site recovery demonstrating near/natural restoration. This Chapter addresses some of the primary challenges underlying mine site restoration with emphasis on the reconciliation of ecological and socio-economic expectations for highly disturbed landscapes. It is argued that there is an incongruity between the ecological targets set for mine sites vs. the limitations of ecosystem recovery following such severe and possibly irreversible alterations to natural ecosystems. Given that mining is a deliberate (and highly lucrative) disturbance activity often leading to non-natural landscape attributes, ‘traditional’ restoration approaches may not always be appropriate and altered physical and social circumstances might be expected to lead to alternative ecological outcomes. Notwithstanding these constraints, a critical objective is to achieve the highest standards of biological conservation and ecosystem stewardship, particularly in an era of global environmental change.

* †

E-mail: [email protected]. E-mail: [email protected].

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D. Doley and P. Audet

1. INTRODUCTION During the mid-20th century, large-scale disturbances associated with mining accelerated due to technologies that were able to remove greater depths of overburden for the extraction of near-surface minerals and fossil fuel reserves (Beaver 1944; Evans 1944; Kohnke 1950). Inevitably, the increase in mining’s scope and efficiency resulted in a commensurately large anthropogenic footprint, but also a burgeoning awareness of and requirement for environmental stewardship (Bates 1957; Spitz and Trudinger 2008). Although the earliest objectives of mine rehabilitation were pragmatic (e.g., for industrial or urban expansion; UK Government 1958), in some rural environments mined land was returned to its previous agricultural applications (e.g., cropping or pasture) or converted to forestry purposes where land surface and soil conditions were less favorable (Kohnke 1950; Sisam and White 1944). In time, a more comprehensive view of mine rehabilitation objectives and standards arose from the Surface Mining Control and Reclamation Act (SMCRA) (US Congress 1977) that was followed by similar legislation in many other jurisdictions. In the late-20th century, a collective realization of the limited and declining extent of natural habitats and their dependent species (cf. Wilson 1988; West 1993) resulted in a change of focus for the use of mined land, and disturbed landscapes in general, whereby it was then expected to contribute to the maintenance of biological diversity (Jordan et al. 1988; European Commission 1992). This marked a shift away from re-greening affected landscapes and towards establishing near/natural ecological patterns and processes where much greater emphasis would be given to the conservation of native species and reinstatement of natural (i.e., pre-disturbance) ecosystems (Bell 1996; Bradshaw 1997; Burger 2015). These developments complemented recognition of the need for a conceptual foundation in restoration ecology as well as understanding of the limitations to ecosystem recovery (Botkin 1990; Hobbs and Norton 1996); especially in light of inchoate public expectations (Tarlock 1996) that all reclaimed, rehabilitated or restored lands (depending on jurisdiction) should necessarily be expected to have equal biological form and function to the pre-disturbance condition (European Commission 2011; Humphries 2014). While these latter objectives were and continue to be highly desirable, they would require comprehensive and careful planning as well as the allocation of considerable resources throughout the mining process – steps that were not necessarily formalized or applied during earlier periods of development. Recently, commitment to life-of-mine planning has become a necessary precondition for sustainable development in the mining industry (Warhurst and Noronha 2000b; DITR 2006a,b; Laurence 2006; DMP/EPA 2015). At its core, this approach involves consideration of the physical, financial and social requirements for mine closure at the earliest stages of development in hopes of achieving timely lease relinquishment and adequate post-mining environmental stewardship once mining has ended. Stemming from historical underpinnings, the primary assumptions of mined land restoration are that mining is a temporary land-use and affected lands should be returned to a suitable post-disturbance ecological condition (Jordan et al. 1988; European Commission 1992). Although near/natural restoration is a logical goal presumably representing the highest end-point following any disturbance (Clewell and Aronson 2006; Murcia et al. 2014; Humphries and Tibbett 2015), it should be noted that our knowledge of ecological restoration is drawn from various investigative contexts (often compliance-driven and financially constrained) and wide ranging disturbance

What Part of Mining Are Ecosystems?

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factors (slight and extensive vs. severe and localized). Likewise, our expectations for ‘success’ are under continual revision as we both expand and refine our knowledge of ecosystem recovery (Bell 2001; Clewell and Aronson 2006, 2013; Tongway and Ludwig 2011; Hobbs et al. 2013), which applies directly to our understanding of what can be achieved in terms of ecological development on a mine site (Warhurst and Noronha 2000a; Bell 2001; Doley et. al 2012; Audet et al. 2015; Humphries and Tibbett 2015). Albeit instructive, broader ecological knowledge and associated theories do not always reflect or account for the conditions and limitations typically associated with mining, such as the alteration of higher order landscape components (landform, lithology and soil) and development of various nonnatural landscape attributes (overburden, waste material and tailings). Notwithstanding the major advances in mine site restoration that have occurred over recent decades, it is the authors’ view that we have yet to establish how best to address mining legacies along with the challenges that they may pose to the application of ecological restoration within life-of-mine planning in resource-limited environments and financially competitive economies. Drawing from the mining experience in Australia, in this Chapter we argue that most contemporary forms of mining (especially open-pit strip mining operations) cause permanent rather than temporary changes to natural landscapes and these changes may represent considerable technical and financial obstacles for the establishment or re-instatement of a desired ecological outcome (i.e., over the time-frame of human appreciation and within the context of sustainable development). For better or worse, we should expect that altered physical and social circumstances may lead to alternative ecological outcomes. Clearly, relaxed environmental responsibility and/or grossly oversimplified rehabilitation ‘success’ criteria are unacceptable approaches to this dilemma in the 21st Century (Clewell and Aronson 2013; Perring et al. 2013). Still, a major challenge lies in achieving the highest standards of biological conservation and ecosystem stewardship when ecological outcomes may not be readily predicted, particularly in an era of global environmental change, heightened environmental awareness and diminishing profitability of mining operations. Likewise, how do we prioritize ecological outcomes under these conditions?

2. DEFINING RESTORATION ‘SUCCESS’ FOR MINE SITES Mine closure criteria are often based on the assumption that any desired ecosystem can be established (or re-instated) on the reconstructed landscape. For a given location and climate, ecosystems can be characterized by abiotic (geology and landscape) and biotic (vegetation, soil biota and fauna) properties (cf. Table 1; Tongway & Ludwig; 2011) that can be used both to define pre-disturbance ecological boundaries and to guide restoration practices. This ecological evaluation for restoration or its nomenclatural variants (Yeldell 2015) involves consideration of bioregional climate, with post-disturbance geology, topography and predominant soil types, to identify vegetation and habitat types that could be accepted as a final target ecosystem.

Table 4.1. Categories and hierarchies of site information included in landscape and ecosystem functional analyses. From Doley et al. (2012), adapted from Tongway and Ludwig (2011). Factor

Attribute

Geology*

1. 2. 3. 4.

Consolidation Origin Mineralogy Texture

Landscape

1.

Form

2.

Function

Physical property

Stability Toxicity Persistence

a. b. c. d.

Landform class Landform pattern Landform element Quantitative attributes

a.

Surface

b.

Profile

I. Relief II. Slope III. Texture I. Erosion II. Deposition I. II. III. IV.

pH EC Nutrient status Water status

Ecosite


Factor

Attribute

Physical property

1.

Structure

2.

Composition

3.

Condition

Soil biota

1. 2.

Abundance Diversity

Beneficial vs. Non-Beneficial

Fauna

1. 2.

Abundance Diversity

Predominant Rare (Endemic) Susceptible

Vegetation

Note:

a. b. c. d. e. f.

Height Cover Biomass Species richness Diversity Structural class

Natural vs. Non-Natural

*Geology and biochemistry of the parent material (including, e.g., tailings and spoils).

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With this information, an appropriate schedule for ecological development can be negotiated through engagement of relevant stakeholders and rights holders, referring to the proponents, regulators, and both local and traditional landowners. As such, successful mine rehabilitation with appropriate landscape reconstruction and reinstatement of biodiversity has been described for: • •

•

•

Lignite mining in Germany (Hüttl and Webber 2001; Simon et al. 2014; Tischew et al. 2014), Coal mining in the UK and USA (Humphries 2013, 2014; Holl 2002; Simmons et al. 2008; Zipper et al. 2013; Wilson-Kokes et al. 2013; Burger 2015), Mineral sand mining in Australia (Enright and Lamont 1992; Rokich et al. 2000; Cummings et al. 2005; Ross et al. 2004; Herath et al. 2009a; Millar et al. 2010; Smith and Nichols 2011; Gravina et al. 2011, 2012; Williams et al. 2011; Cristescu et al. 2012, 2013; McCullough and van Etten 2012; Audet et al. 2013b), and Bauxite mining in Amazonia and Australia (refer to Box 4.1) (Parrotta et al. 1997; Parrotta and Knowles 2001; Schwenke et al. 2000a,b; Spain et al. 2006; Gould 2011, 2012; Koch 2015; Spain et al. 2015).

Common among these examples – besides commitment to adaptive management and active industry research programs – is that a sufficient combination of higher order landscape features could be conserved or reinstated throughout the mining process and/or reasonably improved over time. This usually refers to the reconstruction of the essential landform, the conservation and careful replacement of topsoil (including its fertility and indigenous microbial attributes) (Koch 2007; Evans et al. 2013; Humphries 2014), and the maintenance of native vegetation as a basis for native wildlife and habitat development (Rokich et al. 2000; Koch 2007b; Herath et al. 2009b; Cristescu et al. 2013; Majer et al. 2013; Zipper et al. 2013; Humphries and Tibbett 2015). As a case in point, mineral sand mining on North Stradbroke Island in Queensland (Australia) for zircon, rutile and ilmenite did not significantly alter the parent material characteristics. Consequently, the tailings (representing ‘cleaned’ residual sand) were generally non-toxic and non-polluting and could be used directly and immediately in the reconstruction of post-mining landform features. With careful topsoil and vegetation management as well as adequate climatic and landscape conditions (Smith and Nichols 2011), forest communities resembled reference communities and provided suitable fauna habitats within ~15 to 20 years following mining (Audet et al. 2013b; Cristescu et al. 2012, 2013). For broadly similar reasons, another extensively documented example is the relinquishment of a mining lease following bauxite mining in Western Australia’s jarrah forests (cf. Box 4.1).


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a) Mining as a Special Case of Environmental Disturbance Despite some notable examples of successful ecological recovery following mining, ecological restoration inevitably becomes more difficult with increasing size and/or severity of the disturbance (Figure 4.1) and most mining activities are sufficiently severe, extensive and long-lasting that critical physical and biological characteristics of the original landscape are drastically altered (Mulligan 1996; Tibbett 2010). A logical first objective is to minimise the impact of mining by restricting its extent in sensitive landscapes, for example those containing rare and threatened species (Gibson et al. 2015; Howard 2015). Where open-pits, extensive overburden or waste rock, and tailings persist within the landscape, very few reconstructed mined landscapes may ever resemble the pre-disturbance condition (Ludwig et al. 2003). Hence, veritable restoration can seldom be claimed unless environments are amenable to plant growth and restoration work is intensive (Humphries 2013, 2014; Humphries and Tibbett 2015). Here, it should be acknowledged that the number of abandoned, orphaned and/or highly degraded mine sites in Australia still far exceeds the number of successfully restored sites (Unger et al. 2012), and the numerous reports of unacceptable changes to environments resulting from mining activities do not enhance the reputation of restoration practice (Worrall et al. 2009). And yet, regulators in Australia frequently require that the post-disturbance mining environment bears a close semblance to its pre-disturbance (often pre-European settlement) level of composition, structure and functionality. We do not regard ecosystems that have become established on long-abandoned mine sites as acceptable reference conditions for restoration of existing mine sites in Australia because they often have undesirable 1 soil (old and weathered) and/or water quality conditions (semi/arid climate). The lack of clarity over restoration goals has sometimes resulted in disjunction between the social and biological expectations for mine sites and the limitation(s) of ecosystem recovery (Worall et al. 2009; Soltanmohammadi et al. 2010; Van Kooten 2011; Burton et al. 2012; Glenn et al. 2014; Hodge 2014; Smith 2015). Mining is a deliberate activity that, unlike more gradual agents of change (such as climatic patterns or dispersal of invasive species), is engineered and optimized to extract the desired commodity by the most efficient means possible. For these reasons, we view mining as a special case of environmental disturbance causing more rapid ecosystem degradation than can be offset by the rate of ecosystem building, and potentially resulting in the formation of artificial and sometimes toxic landscape features. Accordingly, the realization of ecological restoration has been questioned from the perspectives of philosophy (Katz 1996), ecological theory (Hobbs and Norton 1996; Hobbs et al. 2009, 2013) and field practice (Bell 1996; Hobbs 2007), which has led ecologists to question ‘what are natural ecosystems’ and ‘what are their likely anthropogenic antitheses’.

b) Mine Sites as Novel (Not Natural) Ecosystems Representing a major step forward in the depiction of managed ecosystem recovery, the definitions of natural vs. novel ecosystems are based on the principle of ecosystem state and transition and the likelihood that disturbance factors can cause (ir)reversible changes to 1

I.e., in terms of production agricultural capability.

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various attributes of the natural landscape. Against a backdrop of alternative stable-states theory and state-and-transition modelling (cf. Grant 2006; Hobbs and Suding 2009), Hobbs et al. (2013) proposed that novel ecosystems are an outcome of irreversible transfer across abiotic and (or) biotic thresholds. Whereas some authors have focused on semantic underpinnings and consequences for regulatory criteria (Murcia et al. 2014; Aronson et al. 2014; Humphries and Tibbett 2015), the depiction of ecosystems in terms of their transitional stabilities (or state-and-transition characteristics) in relation to disturbance is a very valuable concept for assessing the inherent complexity of natural vs. reconstructed ecosystems and the requirements for or limitations to ecological restoration (Figure 4.2). As with any method of ecological classification (Austin 2002), boundaries between natural, hybrid and novel ecosystem categories are difficult to define precisely, so the concept must be applied with suitable care. Here, it should be noted that the novel ecosystem concept is most definitely not a rigid framework that seeks to justify reduced restoration expectations or excuse poor performance; but, it does assist land managers to identify the need for and the magnitude of intervention required to achieve a desired ecological state. Rather than simply criticize existing practices, we suggest it is preferable to consider the attributes of mine sites that must be managed in order to achieve acceptable ecological restoration. With due deference to empirical evidence, we consider that alternative ecosystems can provide safe, stable and manageable environments with acceptable ecological functions (Perring et al. 2013) that ultimately contribute to the delivery of ecological stewardship and biological conservation benefits (Bullock et al. 2011; Chapin et al. 2010). Yet, the question remains as to how the most responsible, effective and economical outcome may be achieved for the restoration of an extensively disturbed landscape without compromising social or environmental standards. As a starting point for discussion of improved mine stewardship – especially highly degraded sites in resource-limited environments – Doley and Audet (2013, 2014) suggested moving away from strict adoption of ecological reference sites and applying the natural-novel ecosystem paradigm in order to identify limitations to ecological recovery and obstacles to achievement of a given ecological outcome. However, this conceptual approach requires clear descriptions of the necessary conditions for restoration, along with a willingness to consider alternative ecological trajectories as valued outcomes for the post-disturbance environment.


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Figure 4.1. (a) Summary of contemporary life-of-mine procedures associated with openpit and/or strip mining, and (b) relative rehabilitation cost and management effort for each phase or process. Note: During the pre-mining phase, vegetation is removed and cover soils (~20–50 cm containing native plant propagules and indigenous soil microbial) are stripped for later use in the landscape reconstruction phase of reclamation. After mining, subsoils consisting of recovered overburden and tailings materials are configured to reconstruct the post-disturbance landform and facilitate revegetation. During the monitoring and maintenance phase of reclamation, site maintenance may range from soil amendment, vegetation biocontrol and/or management of habitat to facilitate desirable land-uses and self-sustaining ecological processes.

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Figure 4.2. Schematic relationship between the effects on landscape integrity, costs of environmental stewardship and severity of disturbance, as it is applied to mining environments.


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Box 4.1. Bauxite mine restoration – A benchmark for environmental stewardship Bauxite mines have been operated by Alcoa World Alumina Australia (Alcoa) in the south west of Western Australia since 1963, clearing and rehabilitating approximately 550 ha of jarrah (Eucalyptus marginata) forest (Bell and Hobbs 2007) to produce in excess of 30 million tonnes of bauxite per year. The forest is valuable for its biodiversity values and commercial timber production and it occurs in important surface water catchments and recreational areas for the Perth metropolitan area (Dell et al. 1988). Consequently, there were many reasons for engagement in mined land restoration, which has been sustained through 50 years of close collaboration between managers, regulators and researchers (Koch 2015). Bauxite mining has greater similarity with quarrying than with the mining of coal or other metals because the major consequence is a lowering of the land surface following the removal of a layer of pisolitic or massive bauxite (a mixture of aluminium oxide and iron oxide) that may vary from 3 to 15 m in thickness (Sadleir and Gilkes 1976). Topsoils and overburden layers, which are commonly shallow in comparison with many other surface mining operations, are stripped and stored separately; the bauxite is removed and the new land surface shaped and ripped to a depth of at least 1 m before overburden and topsoil are replaced in sequence after minimum storage times (Koch 2007a). Fresh topsoil may be moved up to 10 km between origin and placement, frequently across quarantine barriers for Phytophthora cinnamomi, a soil-borne pathogen of jarrah (Colquhoun and Kemp 2007). Revegetation relies on establishment of favourable soil microbiological conditions (Gardner and Malajczuk 1998; Banning et al. 2008; Glen et al. 2008) and the presence of both seeds and vegetative propagules in the topsoil, augmented by carefully managed broadcast seed mixtures and selected planting of species that are difficult to establish by seed (Koch 2007b). These activities have resulted in successful establishment of jarrah (Koch and Samsa 2007), the development of vegetation succession (Norman et al. 2006) and species richness, including terrestrial orchids (Collins and Brundrett 2015) on restored sites at least equal to that on undisturbed sites (Koch and Hobbs 2007; Koch 2015). Post-establishment ecosystem management is a topic of continuing concern, involving thinning of jarrah stems, hazard reduction burning (Smith et al. 2000; Grant et al. 2007; Craig et al. 2010) and activities to enhance faunal habitat development (Nichols and Grant 2007; Christie et al. 2013). Reassessment of management priorities has been necessary in order to realise the final goal of habitat restoration (Grant and Koch 2007; Majer et al. 2013). This work has contributed almost half of all the published research on fauna recolonization of mine sites in Australia (Cristescu et al. 2012). Importantly, there has been little change to the quantity (Hughes et al. 2012) or quality (Croton and Reed 2007) of water draining from mined catchments, and the restored areas appear to have satisfied the desire by members of the public for the reinstatement of plant species diversity and vertebrate habitats in a socially costeffective manner (Koch and Hobbs 2007; Burton et al. 2012). Bell and Hobbs (2007) summed up the international-award-winning environmental achievement as: The ability to reconstruct or restore a piece of damaged land following disturbance relies heavily on strong ecological knowledge, and the basis of this acquired knowledge is research related to the particular problem sites and the objectives presented for the restoration of that site. Of specific importance is that native organisms have inherent capacities to tolerate local climatic and edaphic conditions and are generally preferable to introduced species in revegetation programs. Second, amelioration of the post-disturbance habitat can ensure establishment of plants under difficult circumstances. And, third, the long-term functional aspects of the post-mining areas may be of great importance to the future well-being of the region and must be a major consideration of restoration programs.

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Critically, the performance of Alcoa has been largely driven by local management initiatives within the company to go beyond regulatory compliance in the setting and pursuit of restoration goals (Gardner 2001; Koch 2015). There is no substitute for vision and determination on the part of mine owners, managers and operators, and vigilance and participation on the part of the adjacent community.

3. PLANNING FOR THE FUTURE – OPPORTUNITIES FOR NOVEL ECOSYSTEM DESIGN Mining does not represent insurmountable obstacles to restoration across all geographic locations, as described earlier. However, most major mining operations in Australia are located in resource-limited sub-humid or semi-arid regions that offer many challenges to postmining land restoration (Groves 1994; Mulligan 1996; Bell 2001; Tongway and Ludwig 2011; Dunkerley 2012; Arnold et al. 2013; Audet et al. 2013a; Erskine and Fletcher 2013), due to: • • •

Climate (especially low or highly erratic precipitation and high evaporation rates); Landscape (i.e., commonly flat or rocky) and edaphic factors (i.e., soils that are commonly infertile, shallow or toxic), and Indigenous vegetation associations (i.e., containing species that only regenerate naturally at long and irregular intervals, and are prone to invasion by exotic species − particularly if nutrients/fertilizers are added).

In order to overcome these obstacles (which are by no means exclusive to Australasia), it is necessary to consider the elements of landscape structure and function that may enable existing (reference) ecosystems to be emulated or new ecosystems to be established. Table 2 provides a checklist of restoration objectives and tests, followed by design responses and activity or process responses. Some of these issues will be discussed briefly below. Notably, these observations apply particularly to the more disruptive forms of mining and thereby provide a framework for the description of stable landscape elements in different environments. It is argued that incorporation of these landscape objectives in mine planning and operations can make the mining operations more efficient and rehabilitation outcomes less uncertain and less costly by focusing on (but not being limited to) the influence of landform and local aridity, the amelioration of soil conditions (especially toxic materials), and the selection of suitable plant species.

Table 4.2. Objectives, tests, design and implementation checklist for mined land restoration Objective Secure the landscape

Secure the soil

Secure the biota

Test • Can the pre-disturbance landform be reinstated? • Can the new landscape conditions meet predisturbance functional criteria? • Is there enough soil? • Is substrate remediation required? • Are agents of pedogenesis intact and functioning? • Do restored soils support pre-disturbance vegetation types? • Do soils support rare, endemic, environmentally sensitive flora? • Does climatic seasonality seriously limit revegetation?

Design response • Design the disturbance footprint and operating schedule • Design the reconstructed landform: safe, stable, non-polluting • Design landscape elements: best practicable ecosystem substrate • Align soil characteristics with broad bioregional ecosystems • Design storage and replacement footprint and procedures • Design substrate remediation procedures • Select best practicable target ecosystem for each landscape element • Develop management plans for rare, endemic, environmentally sensitive flora and fauna • Develop management plans for climatic and seasonality obstacles to revegetation

Activity/process response • Build to be safe, stable, non-polluting • Steep slopes of bare competent rock can minimise footprint and erosion • Restrict dispersive materials to low slope areas • Physically isolate potentially hazardous materials (wherever possible, avoid store and release covers) • Shape land surface to optimise conditions for plant establishment (e.g. mounding, contour banks, drains) • Apply soil only where it can support acceptable vegetation • Apply soil amendments to facilitate plant establishment (chemical treatments, organic matter, microbes, fertiliser) • Allow time for soil improvement • Select or test species that may best fit the new environment (not only local ecotypes) • Allow time for plant establishment • Repeat plant establishment if necessary • Uniform or continuous high plant cover and species diversity are not the only valid goals • Manage key plant species that provide landscape stability • Manage key plant species that provide critical fauna habitat

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Objective Secure the ecosystem


Test • Do restored landscapes provide stable and manageable habitats? • Are agreed ecological functions achieved?

Design response • Identify and review progressive targets using State-and-Transition model • Monitor to identify and apply management inputs to maintain stable habitats • Monitor to identify and apply management inputs to maintain essential ecological functions and service

Activity/process response • Assess condition against agreed targets using state-andtransition model • Assess biological development in stages o Soil conditions and major soil fauna species o Major structural and functional plant species o Key fauna indicator species • Make prompt physical repairs (e.g. erosion, nutrient and water status) • Apply species management (re-introduction, removal) to achieve habitat goals • Apply necessary site treatments (fire, grazing management) • Accept a dynamic equilibrium for target ecosystem (not a constant state)

a) Securing the Landscape Natural environmental stresses at a mine site may be exacerbated by the alteration of lithology and landform through mining activities, with consequent changes to the geochemistry and hydrology of landscapes and the release of high concentrations of toxic soluble metals and radionuclides into the environment (Mulligan 1996; Howard et al. 2011). For example, where vegetative cover may be neither continuous nor permanent, erosion of bare surfaces must be strictly minimised by physical means such as the installation of competent rock covers (Waggitt 2004). Under these circumstances, we consider that it is impractical to reinstate the pre-disturbance ecosystems (i.e., vegetation communities plus associated fauna) on landscapes that may bear very little resemblance to their pre-disturbance analogues. The rates of weathering and erosion vary with climate, and are generally slower in cold or arid environments (Wainwright et al. 2011) than in warm and humid environments (Parsons 1988); but, eventually, they lead to changes in surface shape (van Beek et al. 2008) that can be described quantitatively by erosion models (Hollingsworth et al. 2010; Hancock et al. 2003, 2008). Natural surfaces of arid landscapes are commonly covered by stone pavements (Laity 2011) or crusts (Nash 2011) that contribute substantially to their stability whereas, in more humid environments, this surface stability is more often a function of the vegetation that occurs naturally (Breckle 1999) or can be established and maintained (Norris et al. 2008). In a reconstructed landscape, both the landform and soil may be substantially different from the pre-disturbance historical condition (Willgoose and Riley 1989; Doley and Audet 2014). Therefore, it is necessary to consider the combinations of conditions that may influence the amount of vegetation that could be expected to develop on a particular site. Ideally, a constructed slope should be stable in the absence of vegetation. Parsons (1988) identified elements of a hillslope, the physical characteristics of which often result in a concave natural surface that can be predicted by erosion modelling (Hancock et al. 2003). A clear case has been made that landform design can and should be modelled to take account of the quantities and types of material available, their physical and chemical properties and their consequences for erosion, contaminant discharges from the site and vegetation development, and these shapes should be copied during landform reconstruction (Hancock et al. 2003, 2006; Hollingsworth et al. 2010; Howard et al. 2011). According to the geo-biosphere concept (Breckle 1999), a critical component of the environment for vegetation development on rehabilitated sites is the plant available water (PAW), which is determined by climate (rainfall, evapotranspiration, soil properties (depth and water release characteristics), slope of the reconstructed landform and position in the landscape (runoff or runon) (Tromp-van Meerveld and McDonnell 2006; Tongway and Ludwig 2011). Over the long term, PAW will be a function of the ratio of precipitation to evaporation at a point on the landscape where there is no net drainage, as used in regional vegetation analyses by Specht (1972). Critically for reconstructed landscapes, sites of runoff will have lower PAW and sites of runon will have higher PAW than the climatic median, and therefore the potential to support less or more vegetation respectively. Climate is a global determinant of vegetation development (Breckle 1999; Harris et al. 2006; Reich et al. 2014), including reconstructed environments such as post-mining sites (Audet et al. 2013a). Yet, the climate is not subject to manipulation in the

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course of land rehabilitation, therefore it is necessary to examine the extent to which other factors can or should be managed to achieve stable post-disturbance landscapes.

b) Securing the Soil Before vegetation can be introduced to a reconstructed site, it may be necessary to address the toxicity of the landscape, its physical or chemical amelioration, and the biotic components (i.e., vegetation assemblages) that may survive/persist given these very challenging circumstances. Metalliferous geological formations commonly have elevated concentrations of plant available metals or metalloids in their surface soils and it is not surprising that mine sites may also have high concentrations of available metals in waste materials, such as waste rock or tailings (Lottermoser 2010). Contaminant concentrations may be phytotoxic and totally inhibit the development of vegetation on waste rock heaps or on tailings storage facilities. This creates serious problems for the attainment of mine lease extinguishment, which usually depends on the establishment of some type of vegetative cover. Capping waste rock or tailings with soil is a common practice, and this often relies on the maintenance of vegetation on the benign material of the cap (Bell 2001). Where there is ample capping material (Zipper et al. 2013; Tongway and Ludwig 2011) or the climate is conducive to development of a stable vegetation cover, this approach can lead to satisfactory vegetation establishment (Smith and Nichols 2011; Perring et al. 2013; Shackelford et al. 2013) and reduction of erosion (Carroll and Tucker 2000; Stokes et al. 2008). However, in some situations, capping materials may be scarce or totally lacking (Schneider et al. 2010) or the climate may be very unfavourable (Audet et al. 2013a), and sites may remain unvegetated for many years (Mulligan 1996). In such cases, it is necessary to decide whether or not revegetation should be attempted. This is not a frivolous question. If, for whatever reason, the desired vegetation cannot be maintained on a reconstructed landscape, it may be preferable to design the post-mine landscape to not be dependent on vegetation cover, whether this results in a rocky hillside or a salt lake. Both of these landscape features occur naturally, especially in the sub-humid to arid environments of Australia. While very rocky landforms may support endemic species of great ecological value, they are often distributed sparsely and confined to more favourable environmental niches (Gibson et al. 2015; Howard 2015). Therefore, revegetation activities should seek to identify the critical combination of resources that allow plants to survive in these extreme landscapes and to provide those conditions where it is feasible to do so. The overriding requirement for any restoration would be that the landform was safe, stable and non-polluting, and it would be the responsibility of the mining engineers to ensure that these conditions were always attained. Management of surface leaching and evaporation of salts, subsurface drainage and aeration of the substrate may be necessary, followed by the addition of microbial inoculants may further ameliorate the plant root environment to the point where at first tolerant but then more pollutant sensitive plant species may be introduced (Mendez and Maier 2008; Meier et al. 2012). This approach requires careful attention to both the physical and biological properties of the substrate. For example, alumina refinery residue (red mud) may be either highly alkaline (if discharged directly to a disposal site) or saline (if sea water is used to neutralise the alkalinity) (Menzies et al. 2004; Wehr et al. 2006). Similarly, tailings from metalliferous ore treatment may have physical characteristics that are unfavourable to plant


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growth, together with high salinity or sulfide contents that result in acid formation on wet oxidation (Lottermoser 2010). Redistribution of solutes within the profile or landscape may assist in the biological amelioration, so the generation of small-scale topographic relief may produce mounds that are more favourable and hollows that are less favourable for plant establishment. Discontinuous vegetation cover is considered to be preferable to no vegetation cover. Where limited supplies of topsoil are available in an environment that is not conducive to continuous plant growth, there is a temptation to spread the topsoil thinly over a substrate (Dragovich and Patterson 1995) in the hope that this will be equivalent to reconstruction of the original soil profile. The number of revegetation failures attending this technique is large and does not need to be increased any further. If topsoil is scarce, we consider that it is far better to place it in a few selected sites from which it will not be removed by erosion or washing into substrate voids and in sufficient quantity that it could be expected to maintain a few plants through a sequence of unfavourable seasons. Particularly in semi-arid areas, vegetation distribution is naturally patchy, and it is not reasonable to expect that a restored mine site should have a vegetative cover that is greater or more uniform than that of a native community on a truly analogous landform. The limited cover that does develop may have little effect on the water balance of the site (Arnold et al. 2014) and should not be relied upon to do so. Fresh topsoil has been shown repeatedly to be the most favourable medium for any revegetation work (Bell 2001; cf. Box 4.1) Therefore, it is critical not to waste this resource by storing it for any longer than absolutely necessary. Immediate relocation of topsoil may be far preferable to storage for projects with a long life, even though it may compromise the quality of restoration in the first and final areas to be disturbed. This process requires careful planning and meticulous implementation (Koch 2007a) but it can assure the most rapid ecosystem recovery (Gardner and Bell 2007; Waterhouse et al. 2014) In situations where topsoil is not available but where the post-mining substrate theoretically may have sufficient plant available water to maintain vegetation, an alternative to capping a reconstructed landform with sufficient topsoil to support the desired vegetation is to modify the land surface (e.g., waste rock or tailings) so that some biological activity may become possible. The simplest method to ameliorate a toxic surface is to dilute the pollutant materials with benign bulk additives, such as sand, soil, organic matter (straw, mulch, compost, wood chips or biochar) or to add small quantities of specific synthetic materials that bind pollutants irreversibly, to the point where the growth of selected pollutant-tolerant species may be possible immediately. The required quantities and costs of application of these different additives vary greatly, and the selection of a method will depend on economic as well as technical evaluations at each remediation site. Diverse soil microbial taxa are known to enhance plant growth under conditions of nutrient and toxic metal stress (Audet 2012, 2013), leading to a progressive increase in the diversity of organisms and the complexity of biological associations in metalliferous mine tailings (Mendez and Maier 2008; Courtney et al. 2009; Huang et al. 2012) and alkaline alumina refinery residue (Courtney et al. 2009; Schmalenberger et al. 2013; Santini et al. 2015). One promising approach to soil amelioration is to apply synthetic metal binding polymers that sequester sufficient toxic soluble (i.e., plant available) metals to enable the germination and early seedling establishment of metal tolerant grasses in metalliferous mine tailings (Rossato et al. 2011; Bigot et al. 2013). Root growth is likely to be limited to the depth of incorporation of the particles, but the polymers may also increase the volume of plant available water (Bigot et al. 2013), improving the chances of

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plant survival under conditions that would otherwise be lethal to young plants. Diverse soil microbial taxa are also known to enhance plant growth under conditions of nutrient and toxic metal stress (Audet 2012, 2013), so it is prudent to consider all options for improving substrate conditions. The progressive improvement in environmental conditions following establishment of first metal-tolerant and second more sensitive plant species on contaminated sites is well described (Baker et al. 1994; Bradshaw 1997; Cooke and Johnson 2002; Parraga-Aguado et al. 2013; Hansi et al. 2014). An important contribution of novel site amelioration techniques is that they allow this process of improvement to begin under otherwise prohibitive environmental conditions (Rajikumar et al. 2012; Ali et al. 2013; Miransari et al. 2011; Rascio and Navari-Izzo 2011). For various reasons, optimising phytoremediation strategies (and all of its sub-genres) (Pilon-Smits 2005) means that compromises must be made regarding types of species used at a given location and that local biodiversity would have to be momentarily overlooked in favour of more tolerant species (e.g., metallophytes). This is an unavoidable dilemma faced by mine rehabilitation practitioners having to reconcile considerable biological and ecological obstacles (i.e., referring to the reality that reinstating ecosystems takes time, if it is at all possible) (Botkin 2012; Allen et al. 2014) against public expectations that justifiably require highly lucrative industries to return disturbed landscapes to suitable standards of ecological integrity. Such a dilemma is far from being resolved. Then again, recognition of hybrid and novel ecosystems (sensu lato) and the reality that rehabilitation outcomes may ultimately differ from any historical or pre-disturbance antecedent offers alternative perspectives that potentially lead to greater flexibility when selecting final post-mining landscapes, land uses, and species for establishment (Guisan and Zimmermann 2000; Soltanmohammadi et al. 2010; Maczkowiack et al. 2012; Masoumi 2014).

c) Establishing the Vegetation The first stage of habit construction is establishment of the required vegetation composition and structure. For regional and nominally flat landscapes, there is an approximately linear relationship between the relative wetness (precipitation/evaporation) or annual rainfall of a site and the vegetative biomass it can support (Breckle 1999; Audet et al. 2013a; Ngugi et al. 2014). However, in semi-arid and arid locations, the unreliability of water availability (Audet et al. 2013a) may also constrain the frequency and success of regeneration, so that successful establishment events for woody species may be rare, occurring only once in several decades (Winkworth 1973). For a particular climate and geological situation, soil depth is associated with slope angle and length, and with soil water holding capacity and consequent potential of the site to support biomass (Tromp-van Meerveld and Mc Donnell 2006). The influence of surface slope on this relationship is complex, but it can be indicated by the variation in vegetation communities associated with different landforms within a bioregion (Queensland Herbarium 2014). While relatively high vegetative cover and complex structure can be maintained on steep slopes (20 to 30 degrees) under humid conditions, more arid conditions are associated with reduced vegetation cover and stature. These differences are particularly important in reconstructed landscapes because local variations in plant water availability can result from the placement of coarse textured


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materials, the amount of topsoil that is available for distribution over the site and the patterns of surface and subsurface water movement and consequent erosion (Hancock et al. 2003; Howard et al. 2011). That being said, if the soil characteristics of the reconstructed site cannot be specified closely, then it follows that the specification of the vegetation association cannot be any more precise. An example of the difficulty of such prediction − recently drawn-upon by Doley and Audet (2013) − is seen in the detailed soil and vegetation analysis carried out over 300 000 ha in sub-humid tropical Queensland, Australia, by Burgess (2003). Sixty-nine soil profile types were identified on geological formations that ranged from igneous (granite and basalt), through sedimentary to colluvial and unconsolidated, and landscapes that ranged from steep slopes to flat clay pans and flood plains. A total of 50 vegetation associations were identified, but only 42 of them occurred over sufficient areas to be analysed. Still, for all of the 69 soil profile types, there was no unique association with a single vegetation type, but between two and five vegetation associations could be found on any soil (Figure 4.3a). Where more than one soil profile location was examined, the dominant vegetation association sometimes differed between these locations and the minor vegetation associations often differed. This could indicate that there is a stochastic component to the occurrence of vegetation types on soil profile types. Therefore, the plant species list for introduction to a nominal soil type on a reconstructed landscape should be more extensive than the dominant species list for a single Regional Ecosystem. This approach would permit the sorting of species according to their fitness for the range of environments represented in the new landscape (Whalley et al. 2013) and may well lead to site-species combinations that were not previously recorded, and were therefore novel. The occurrences of vegetation types on different soil classes showed greater variation. Five vegetation associations were restricted to only one soil profile type (Figure 4.3b), but 60% of vegetation associations were found on between two and five different soil profile types (Burgess 2003). This phenomenon is consistent with the widespread occurrence of mixed vegetation associations and hence mixed Regional Ecosystem categories at 1:100 000 mapping scales (Neldner et al. 2012). One vegetation association (brigalow, Acacia harpophylla) occurred on 25 different soil profile types. This may be an artefact of the soil survey technique as vegetation associations dominated by or containing A. harpophylla occurred in a variety of land zones and variants of soil profile types, but always where fine textured soils were associated with relatively high electrical conductivity (Burgess 2003). A consequence of this lack of unique soil-vegetation associations for setting rehabilitation goals is that specification of a particular Regional Ecosystem for a reconstructed landscape may not be appropriate. If specific vegetation associations are identified for replacement in particular landform elements such as, e.g., plateau tops or riparian areas, it must be assumed that the necessary physical conditions, especially of plant available water, nutrition and toxicity can be provided. Where these assumptions cannot be met, different vegetation associations might be expected to occur, regardless of the nature of the topsoil that is spread over the substrate. Then again, mixed-vegetation associations and overlapping bioregional ecosystems necessarily provide some flexibility in the range of possible ecological outcomes for a given mine site within a geographic location. Provided the species introduced to a rehabilitated area do not develop weedy characteristics (Bullied et al. 2012; Kenkel et al. 2002), there does not appear to be a functional reason why hybrid or novel ecosystems containing species or

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ecotypes suited to these new conditions should not develop and be assisted to develop on these novel landscapes (Whalley et al. 2013).

Figure 4.3. Frequency distributions of: (a) number of vegetation associations occurring on one soil type and (b) number of soil profile types occupied by one vegetation class in central Queensland, Australia. Data from Burgess (2003) redrawn from Doley and Audet (2014).

d) Securing the Ecosystem A common perception in mine rehabilitation work is that initial vegetation establishment is the critical process and there are widely accepted prescriptions for (e.g., DITR 2006a; SERI/IUCN 2004; Tongway and Ludwig 2011) and many accounts of such associated activities. Pasture systems are common land use objectives for reconstructed landscapes (Bell 1996; Maczkowiack et al. 2012). In Australia, pastures may be established quickly on rehabilitated coal mines, but they may not achieve the biomass productivity of the pre-


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disturbance pastures without careful fertilizer and grazing management in order to reinstate their nutrient status (Brown and Grant 2000). While vegetation cannot be established without due care in these early stages, it does not necessarily follow that the desired ecological outcome (e.g., habitat development) will emerge without clear articulation of the goals and further resource and management inputs (Miller and Hobbs 2007). The importance of this phase of mined land restoration is shown by the wide range of activities required for the development of desired conditions in jarrah forests re-established on bauxite mine sites in Western Australia. Carefully managed tree thinning (Grigg and Grant 2009) and burning (Craig et al. 2010) treatments have been shown to be necessary to maintain the desired patterns of development, while uncontrolled fires may impede plant community development (Ross et al. 2004). Even this unusual intensity of restoration effort for mined land outside Europe (cf. Hüttl and Weber 2001; Humphries 2013) may not result in the attainment of suitable habitat conditions for certain fauna after 37 years (Majer et al. 2013). On the other hand, certain critical habitat for invertebrate fauna (Bisevac and Majer 1999; Williams et al. 2011) and mammals (Craig et al. 2012; Cristescu et al. 2013) may be reinstated within 15 years after of rehabilitation of mined forest areas. Decades-long persistence with restoration is rare amongst mine managers, but for ecosystems dominated by trees, it should be the norm rather than the exception. Allen et al. (2014) recognised that dynamic feedback between biological and physical processes was critical for ecosystem development, but they concluded that there would be an asymptotic approach to a stable condition. This long-term stability may be a mirage, as even long-lived vegetation associations are subject to periodic autochthonous change (Botkin 2012; Botkin et al. 2014). Perhaps more importantly are the steps (as outlined above) toward securing the desired landscape, soils and broader vegetation assemblies that could provide predictable attributes of the post-disturbance environment. Therefore, it is critical that any post-mining land use allows for continuing management, whether the objective is to return to area to its pre-disturbance condition or to develop a novel ecosystem.

4. CONFIRMING THE LICENCE TO OPERATE IN AN ERA OF GLOBAL CHANGE – A NEW ENVIRONMENTAL ZEITGEIST? An increasing world population and increasing material well-being in countries undergoing industrial development means an increasing demand for resources of all descriptions (van Zanden et al. 2014). At the same time, technological improvements in mining and the ability to apply increased energy and financial resources to mining activities have vastly increased the scale of mining activities (Mudd 2010; Batterham 2013). In recent times, much of this activity has occurred in developing countries (ICMM 2012) and this trend is likely to continue. These mining developments have heightened the intensity of debate over the balance between material wellbeing and environmental quality (Giljum et al. 2014), specifically in relation to mining (Mulligan 1996; UNEP et al. 2005; Yeldell and Squires, this volume), with the result that many mining companies have embraced environmental sustainability as a corporate goal (Hilson and Murck 2000; Tuazon et al. 2012; Barkemeyer et al. 2015; Smith 2015). However, even for technical matters of mining and mineral processing, there may be a disjunction between corporate aspiration and operational achievement (Mudd

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2010; Tuazon et al. 2012). The same disjunction occurs with respect to mine land restoration (Jenkins and Yakovleva 2006). This need not be so if mine owners and operators could be convinced that a socially acceptable mine closure process is both technically feasible and commercially affordable (Van Kooten 2011; Maczkowiack et al. 2012; Hodge 2014; Masoumi 2014). When attempting to address What part of mining are ecosystems, the close connections between geophysical and ecological processes (Allen et al. 2014; Fu et al. 2013) make it both logical and necessary to consider the contributions of all phases of landscape and ecosystem management to the overall goal of sustainable land management in the mining sector (Laurence 2006, 2011; Barkemeyer et al. 2015). Consequently, it is critical that land management practices such as mine restoration are designed carefully, approved by all stakeholders, resourced adequately, executed diligently and monitored appropriately (Watson et al. 2015). The cost of these activities will be greater than the initial cost of neglect, but the final cost to society of abandoned mines is daunting and unfairly placed (Unger et al. 2015). Some jurisdictions require the establishment of self-sustaining ecosystems on restored mine sites. Natural colonisation of mined land can occur (Bradshaw 1997), so the notion that a selfsustaining ecosystem will develop inevitably after sowing or planting the specified mixture of plant species is an attractive option for mine operators who wish to minimise their current expenditure. However, agricultural and horticultural practice recognises that constant management is required in order to achieve desired outcomes from those ecosystems. Even restored bauxite mines in Western Australia require continuing study and intervention in order to ensure that desired ecosystem conditions can be maintained into the future. Because mining operations are a temporary activity that often incur permanent environmental modifications (i.e., over the timeframe of human appreciation), the physical, biological, social and financial requirements for ecosystem restoration should be incorporated in all stages of mine planning and operation (Yeldell and Squires, this volume). With suitable care at each stage of the mining operation, mined land can be returned to a useful function, but that function may not be the one that occurred at that precise point prior to mining disturbance. So long as the net benefit to society from the mining, restoration and subsequent activities is greater than their costs, then the mining disturbance can be justified.


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