Grassy Community Restoration

Abstract

Each case study presented shows how grassy community restoration can be undertaken to various degrees. While there are some differences in the approaches used, or in how they are structured, resourced and timed, all demonstrate that restoration can be used to recover or reintroduce species and grassy communities to lands where they are absent or degraded. This is a very important message given the threat these communities face and should give hope that there is the knowledge and the tools available to halt and even reverse grassy community loss should peoples choose to do so and are properly resourced.

Aust Case Study: Paul Gibson-Roy (1). US Case Study, Chris Heltzer (2); Belgian Case Study: Sandrine Godefroid (3); Thibaut Goret (3); Maïké Dellicour (3). Brazilian Case Study: Fernando A. O. Silveira (4)

FormalPara Summary and Key Lessons

Each case study presented shows how grassy community restoration can be undertaken to various degrees. While there are some differences in the approaches used, or in how they are structured, resourced and timed, all demonstrate that restoration can be used to recover or reintroduce species and grassy communities to lands where they are absent or degraded. This is a very important message given the threat these communities face and should give hope that there is the knowledge and the tools available to halt and even reverse grassy community loss should peoples choose to do so and are properly resourced.

They also give insights into various issues that make grassy community restoration difficult to undertake or undermine its chances of success. Several such as seed supply (e.g., it’s accessing, quantity, quality and cost) or unsuitable site conditions (e.g., high weed or nutrient loads) are consistent across case studies, while others such as being able to secure land for restoration, creating opportunities for local communities (e.g., via seed networks), providing incentives for landholders or others to undertake restoration (e.g., funded programs) and improving sector capacity (e.g., skills, training, technology) are more pertinent to specific situations or settings.

While these Case Studies show that we have the knowledge to undertake grassy community restoration, for this to occur at landscape and global scales will require well-tailored support from governments to create the legislative structures and opportunities that support human communities to achieve these outcomes. Indeed, restoration at the scale needed to repair anthropogenic damage done over millennia will take a commensurate effort in terms of time, resourcing and commitment from countries, jurisdictions and their peoples. Some countries are more advanced in this than others; however, this should create the opportunity for knowledge sharing and even of cross-jurisdictional support. If native grassy ecosystems can be better integrated into the fabric of our landscapes, be they farm-scapes, urban regions, transport corridors or others, then humans and a rich natural biota will benefit. For this to occur, purposeful decisions and goals, clear pathways and concrete actions must occur so that at times ad-hoc and intermittent successes of the past are turned into purposeful strategies and widespread global advances of the future.

FormalPara Management Implications

Grassy communities can be restored to agricultural, forested, urban and other landscapes using regenerative and reintroduction approaches.

• Grassy community restoration can achieve high levels of species and functional diversity as well as temporal resilience.

• Restored grassy communities create a myriad of biodiversity, ecological and ecosystem service benefits.

• Restored communities must be purposefully managed and maintained over time to preserve their structural and compositional integrity.

• Effective seed supply chains and delivering seed in quantity, quality, price and ethically are critical to successful restoration.

• Creating better employment opportunities for individuals and communities, increasing technical skills and training and improving infrastructure and technology are all crucial to overcoming barriers to increasing the effectiveness and scale of restoration.

• Landscape-scale grassy community restoration will rely on the formulation of insightful and finely crafted government strategies and policies that create the settings, frameworks and coordination required to build markets, improve sector capacity and meet ambitious grassy community restoration targets.

Introduction

Grassy communities, including grasslands, prairies, steppes, meadows, grassy woodlands, savannahs and grassy-forest complexes, are present on all continents except Antarctica and cover an estimated 40.5% of these landmasses (White et al., 2000). They occur across many disparate regions, from tropical to tundra and alpine areas and from arid to temperate zones (Gibson, 2009; Squires et al., 2018; Wilsey, 2018). Grassy communities are dominated by ground layer vegetation, primarily of grasses and forbs, whilst within grassy woodlands and savannahs, trees and shrubs are key functional components giving a sparse open upper stratum (Wilsey, 2018; Raghurama & Sankaran, 2021). A common feature of all these different community types is that they support high plant diversity, even within relatively small-scale pockets (Partel et al., 2005; Morgan & Williams, 2015).

Continual vegetation disturbance due to grazing herbivores has been instrumental in the formation and maintenance of grassy community structure, where conditions such as soils and climate might have otherwise been suited to the development of forests (Nerlekar & Veldman, 2020). Other factors such as fire, aridity and cold are also known to aid in the restriction of woody dominance and the consequent retention of grassy communities (Gibson, 2009; White et al., 2000). These areas, maintained by such factors over long periods, are referred to as ‘old-growth’ grasslands (Buisson et al., 2019, 2021a, b; Nerlekar & Veldman, 2020; Silveira et al., 2020), whilst ‘new-growth’, ‘derived’ or ‘semi-natural’ grasslands are terms given to those areas formed more recently through human-mediated disturbance to other vegetation types, such as through forest clearing.

The rise and spread of modern humans (Homo sapiens sapiens) saw humans also become intimately connected to the formation and maintenance of grassy communities. This was done through their use of fire primarily to meet food, cultural, spiritual and other needs (Bird et al., 2008; Gibson, 2009; Gammage, 2010). Grassy communities provided direct food sources for hunter-gatherers, such as seeds and tubers, as well as being a source of animal fodder, which attracted game. Over time, this also provided reliable feed for domesticated livestock (Gott et al., 2015; Gammage, 2010, White et al., 2000). However, the transition by many human societies from nomadic to sedentary agricultural lifestyles resulted in increasing degradation and loss of native grassy communities, especially after the domestication of a narrow group of plant species, including corn, millet, rice, rye, sorghum and wheat, which led to the conversion of native grassy communities to agricultural land supporting annual cultivated crops.

As human populations grew, areas of land transformed to croplands to feed societies increased and the areas occupied by native grasslands consequently decreased, even noting that in some areas forest clearing had led to localised increases of semi-natural grasslands. It is thought that approximately 40% of temperate grassy communities have now been converted to cropland or other forms of intensive agriculture. In some countries and regions, the degree of loss is much higher. Indeed, in Australia, temperate grasslands have been reduced to less than 1% of their once extensive range (Kirkpatrick et al., 1995), while in the United States, tall-grass prairies have been reduced to only 3% of their previous cover (Samson et al., 1998); likewise, in Europe, Schutyser and Condé (2009) reported continuing and substantial decreases in grasslands of ~260,000 ha between 1990 and 2000. And while in some parts of Europe grasslands still occupy substantial areas of the landscape, overall their quality has declined and more than 75% have been classified in unfavourable conservation status’ condition (Silva et al., 2008). This situation of loss, degradation and fragmentation of grassland communities has been further exacerbated by the cessation of traditional cultural management practices, ongoing clearing for the development of towns and cities, agricultural expansion, conversion to forestry, invasive species, climate change and other anthropogenic factors (White et al., 2000; Gibson, 2009; Valko et al., 2016; Torok et al., 2021).

Grassy community loss is a tragic occurrence on many levels. First, such communities represent a vast number of endemic plant species, and loss will mean many unique forms of biodiversity will be lost (White et al., 2000). Not only are they floristically and functionally diverse, but they also provide habitat and resources for a vast number of organisms from different trophic levels, including animals, birds, fungi and bacteria. Together, these form intricate webs of existence that should be valued and preserved for their own sake. Beyond these attributes, grassy communities provide an array of ecosystem services which continue to benefit humans. These include the provision of food and raw materials (Valko et al., 2016), regulating erosion and soil loss (Cain & Lovejoy, 2004), improving water quality by reducing nitrogen and phosphorus run off (Vilsack, 2016), reducing flood risks (Johnson et al., 2016), sequestering of soil carbon (Gebhart et al., 1994, Yang et al., 2019), improving pollinator services for agriculture (Kremen & M’Gonigle, 2015; White et al., 2017, McMinn-Sauder et al., 2020) and improving air quality (Johnson et al., 2016). Yet, despite these many attributes, the formal protection of grassy communities across the globe is extremely poor, with percentages of protected areas ranging from as little as 1% to 3% in grassland regions (Henwood, 2010), highlighting again the urgent need for their restoration.

Grassy Community Restoration

Given the degree of loss of grassland across the world, it is important that where these native grassy communities are still present, their conservation, protection and maintenance should be among key environmental goals for societies and governments. However, given that in many countries there are few areas of native grassy community left to retain, their restoration through active or passive means must become a key environmental goal. Over recent times, ambitious global restoration targets to remediate the impacts of human-induced land degradation (in the order of 350 million hectares by 2030) have been set under programs such as the Bonn Challenge (https://www.bonnchallenge.org) and the United Nations Decade on Ecosystem Restoration (About the UN Decade| UN Decade on Restoration). These goals should naturally include targets related to the restoration of grassy communities since these always represent an affected vegetation type (Dudley et al., 2020; Tolgyesi et al., 2022).

Common goals for grassy community restoration are to increase their extent, range, quality and connectedness and to reinstate as high a proportion as possible of native species and functional diversity representing a desired or reference community (Prober & Thiele, 2005; Gibson-Roy & Delpratt, 2015; Valko et al., 2016; Barr et al., 2017). In this respect, the extent of species diversity attained in restorations often can depend upon the degree to which appropriate species are either available in the broader landscape as collectable seed, retained in soil seed or bud banks, or available as colonizing propagules (Price et al., 2021; Gibson-Roy, 2022). The relative contribution from each may depend on the restoration strategy used, such as restoration through reintroduction (seeding or planting), or restoration through assisted natural regeneration (spontaneous recovery), or combinations of both these approaches.

Where diverse and plentiful native seed/bud banks exist and/or there are local colonizing sources remaining near a restoration site, the probability of spontaneous natural recovery ranges from likely in temperate grasslands (Valko et al., 2016) to remaining unlikely in tropical edaphic grasslands due to dispersal limitation, low seed quality and slow seedling growth (Nerlekar & Veldman, 2020). And while the process of natural recovery is typically slow and long term (Partel et al., 2005), it can be promoted and accelerated by management interventions that mimic natural disturbances such as stock grazing, controlled burns, hay baling or the careful use of herbicides (even in tropical settings). These treatments can limit the buildup of biomass and/or litter, thereby creating suitable niches for seedling recruitment and/or restricting competition from mature plants (natives and exotics) for emerging or colonizing natives (McDonald, 2000; Prober et al., 2005). Conversely, in highly disturbed locations such as agricultural landscapes or along transport corridors, where land clearing, long-term cultivation, ongoing herbicide application or other forms of disturbance mean that native vegetation has been absent for long periods and soil-borne or colonizing propagule sources are largely depleted, here restoration through species reintroduction by direct seeding, seed hay or plantings is typically required (Barr et al., 2017; Gibson-Roy & Delpratt, 2015; Kiss et al., 2021). Encouragingly, both regenerative and reintroduction approaches have been shown to be effective in restoring grassy communities when used to their best effect (Valko et al., 2016).

There are several key factors that act as constraints to restoration. Foremost among them is gaining access to native seed in the quantity, quality, diversity and timing required (Gibson-Roy & Delpratt 2013; Delpratt & Gibson-Roy, 2015; Ladouceur et al., 2018; León-Lobos et al. 2018; Pedrini et al., 2020; Torok et al., 2021; Zinnen et al., 2021). In many countries and regions seed is limited due to the rarity of grassy communities themselves, whilst in other situations, grassy communities are still present to some degrees, but there are poor levels of training and workforce capacity that limit the effectiveness, quality and quantity of wild collections (Peppin et al., 2010; de Urzedo et al., 2019; Tangren & Toth, 2020; Gibson-Roy et al., 2021a, b). In some countries, such as North America and parts of Europe, well developed and structured markets for restoration mean that while rarity and access to remaining wild communities may still constrain supply of this vital resource, the value of the market has meant that it has been supported and supplemented by cultivated seed production. (De Vitis et al., 2017; Gibson-Roy, 2018 – see Seed Chapter X). In many other countries, the size of restoration markets remains small and so does not support well-developed seed production capacity, leaving seed supply as a major threat to project success (Hancock et al., 2020; Schmidt et al., 2019a).

Apart from the restrictions imposed by seed supply issues, other factors can and do constrain outcomes. Many can change spatially and temporally depending on the site, region or country and include excessive weed loads (soil stored and standing vegetation), unsuitable nutrient settings (because of prior agricultural practices), limitations of training and/or workforce capacity and poorly developed restoration policies or markets (De Vitis et al., 2017; Gibson Roy et al., 2021a, b; Cortina-Segarra et al., 2021). However, despite this, continued advances in knowledge gained from research and practice give a clearer understanding of the pathways towards successful grassy community restoration (Torok et al., 2021).

The following Case Studies give examples of grassy community restoration and allow us to consider work completed in recent decades. Each gives a picture of how restoration has developed and what it has been achieved over this period - including successes and persisting challenges. These projects highlight work done to restore grassy community types in different parts of the world and so allow us to expand on the approaches, philosophies and techniques used by those involved.

Case Study 1: Grassy Community Restoration in the United States¹

Project Rationale(s) and Strategy(ies)

The Platte River Prairie Restoration project (PRPRP) is in the region between Grand Island and Kearney in Nebraska and was overseen by the Nature Conservancy USA. The main rationale for this project’s instigation was based on the Conservancy’s desire to assist in the conservation of America’s once extensive native prairies (Fig. 2.1). To help achieve this aim, the PRPRP worked to restore native prairies located in ex-agricultural landscapes where remnant prairies are today severely fragmented and degraded. The project’s goal was to convert several thousand acres of Conservancy-owned former crop fields back to species-rich native prairie habitat (and in some areas, to wetland) and in doing so to also reconnect various small and isolated remnant areas of native prairie. It was anticipated that successfully achieving this aim would have the effect of increasing the vigour and extent of these areas and thereby provide higher quality habitat for faunal species impacted by the continued loss of native prairies. A secondary aim of the project was to promote and disseminate the techniques and approaches which were specifically developed to achieve these outcomes so that they might be taken up and used by others to achieve comparable results across various regions and states.

Fig. 2.1

Restored Sandhill Prairie with Large Beardtongue (Penstemon grandiflorus) in foreground

Major Project Concerns and Barriers

One of the prime constraints faced by those undertaking this type of work in the United States, related to strong cultural resistance from people living and working in rural regions to the belief that restoration takes land out of ‘productive use’. Many rural Americans feel a ‘moral obligation’ of land ownership that land should be made useful to its full productive extent, and for the farming community, this meant the land should produce an economic or agronomic output, such as crops or livestock. Restoring productive farmland to native prairie was at odds with this objective, because there was no visible economic or agronomic output from this action. Because of this widespread view, there is limited interest in restoring prairie outside of land owned and managed by conservation organizations and ‘recreational land’ owned by people interested in outdoor recreation more than agriculture.

A second local issue, in addition to widespread cultural resistance, was the issue of farmland taxes. In the State of Nebraska, these taxes are based on the government-assessed value of the land’s highest productive potential. If, under this system, a landholder is not able to reach reasonable productive or economic potential from their land, taxes are fixed to its ‘assessed’ potential. Beyond the obvious implications to farm livelihoods of such a situation, this land tax model creates serious negative outcomes for anyone wanting to turn farmland back to native vegetation, given there is no productive output from doing so and that land would thus continue to attract land tax at its assessed productive potential.

Beyond the cultural resistance and tax implications, there were other obstacles to be faced. For example, at the time of the initiation of this project, many among those in the conservation sector did not believe that resilient species-rich prairies could be recreated by direct seeding ex-crop fields with wild native seed mixtures. Indeed, these views were in some ways well-founded. For instance, securing the seed resources required to undertake this type of high diversity restoration, with up to 200 species being involved was a major challenge because in the region (and the State of Nebraska more broadly), there was not a well-developed native seed industry (Oldfield, 2019). Even if that were not the case, the cost of buying native seed (which is much higher than comparable pasture species) in the quantities required to restore thousands of acres of land would have been prohibitive for an environmental NGO. This meant all seeds had to be sourced locally from wild remnants, which created difficulties such as where the seeds and extant species could be located, in addition to the task of locating appropriate people and training them in the techniques required to harvest and process these seeds efficiently and effectively.

Accessing funds to undertake such an ambitious restoration program was also a considerable and ongoing project challenge. Beyond whatever resources, the Conservancy itself could garner through donations; project managers worked with various government agencies that ran conservation-focussed programs to identify any potential funding sources. Interestingly, the Conservation Reserve Program (CRP), which is the most established operational US farm environmental support program, was not a prime funding source for the PRPRP. This was because the CRP set a limit as to how much land a single landholder could hold to be eligible to enter their program, and because the Conservancy owned several properties across the country, it had exceeded the limit. However, project managers were able to identify other government programs that were able to provide suitable pathways that subsidized project costs.

Another seed-related issue arose once seeding itself came into focus. This centred on the nature of actual restoration seeding rates of native grasses. Whilst this would seem to be straightforward, because much of the project’s work was linked to government-funded programs, this meant having to align project seeding rates with their established program rules and stipulations. Experience with on-farm seeding programs showed that most projects sought to rapidly establish native grasses as grazing pasture or for weed control, not, as in our case, for recreating diverse communities. This meant that government regulations were framed on the assumption that grasses would be drill seeded; therefore, they set high seeding rates. However, as the PRPRP’s goal was to establish a complex species mix of grasses and forbs and that they intended to surface broadcast the seed because many species were not suitable for use with drill seeders, the regulations stipulated even higher seeding rates since it was assumed that this technique would be less effective. This regulatory approach would have created serious issues for seed supply, in terms of both quantities and costs. To manage this situation, the project coordinators worked closely with funding authorities to explain the approach and to show why relatively low seeding rates and the broadcasting of diverse seed mixes would be appropriate and successful. This friction between rapid grass introduction versus high diversity restoration remains an issue in the sector, and while the PRPRP has made progress with changing attitudes toward restoration practice in their area of influence, in many other regions and states, it is still the case that seeding grasses remains the main goal of native species establishment and programs continue to resist requests to support high diversity restoration.

The project also faced technical issues, especially relating to equipment, given there was almost no off-the-shelf, purpose-built machinery for restoring native prairies. This meant that important implements such as harvesters, seed cleaners and seeders all had to be adapted or modified from agricultural machinery. This also presented issues with high costs for equipment purchase and around finding people suitably trained in their use and maintenance. On this point, tight budgets meant that project managers had limited staff numbers. Very few were full-time, and most were students employed as seasonal technicians. This meant working with an enthusiastic but transient and untrained volunteer work force with high annual turnover. Volunteers certainly gained a unique set of experiences, opportunities, and knowledge that they took away when they left, but it also meant it was difficult for project managers to develop a long-term local work force with skills and knowledge around plant identification, seed harvesting and handling, seeding or site management. This is still an on-going problem in this area.

Key Project Features

Limitations and constraints aside, there were several features that defined the essence of this project. Foremost among them was its ongoing and faithful focus on restoring species-rich prairie. Regarding this point, the project received critical early support from Bill Whitney and his Prairie Plains Resource Institute, which had pioneered early prairie restoration methods in Nebraska during 1980s. Whitney helped to guide the PRPRP during its formative planning stages in the mid and late 1990s and gave valuable advice and perspectives to the project managers.

As indicated earlier, the long-term goal for restoring large-scale areas of prairie habitat required significant quantities and varieties of native seeds, and this needed to be sourced from the local region. Knowledge of local remnant areas was built up over time, and seed and propagule collections were planned and undertaken on an annual basis. Most of the seed harvesting was done by hand, with mechanical strippers or combines used to harvest key grass and forb species where possible (Fig. 2.2a–d). Collectors aimed to source and harvest as many species as possible, and estimates showed that there were over 200 species eventually used in restoration seed mixes. Even with limited staff and predominantly hand harvest, the project was able to secure enough seed to sow up to two hundred acres per year. In terms of seed handling and processing, a simple and straightforward approach was adopted by project managers. Seed was only cleaned to a basic level, focusing mainly on breaking seeds apart from each other and removing them from pods/stems so that they could be effectively spread when broadcast. Collections were also run through a hammermill cleaner where various screens separated the bulk of seed from chaff, after which it was stored in sacks or open buckets until sowings.

Fig. 2.2

(a) Hand harvest (Top left); (b) combine harvest (Top right); (c) brush harvest (Bottom left); (d) seed mixture (Bottom right)

Each year, different farm locations were identified and prepared for seeding. Ideally, these fields had been under crops for up to a decade, which meant long-term herbicide programs controlling crop weeds had depleted invasive species’ seed banks. Stubble from the most recent crop was mechanically hoed back into the soil, ensuring there was a levelled and workable bed. Because soils in the project area were primarily sandy loams, after stubble incorporation, no further harrowing or soil preparation was required. Seeding was then undertaken in the fall or winter. Seed was sown as a high diversity mixture onto the bare surface or onto a lightly snow-covered soil (which helped draw the seed back into the soil as the snow thawed and melted), by an EZ-Flow drop spreader or by hand (Fig. 2.3a).

Fig. 2.3

(a) Broadcast seeding (Right); (b) early successional weeds (Left)

Following sowings and during the early years of establishment, little was done to the sown fields other than to monitor and control any problematic weeds that might become issues. Of main concern were tree regrowth or tree colonisation, particularly of cottonwood (Populus deltoides) (Fig. 2.4a) and siberian elm (Ulmus pumila), and/or the establishment of dominant exotic perennial grasses, such as smooth brome grass (Bromus inermis), reed canary grass (Phalaris arundinacea) and Kentucky bluegrass (Poa pratensis). Fortunately, annual weeds were not a large issue over extended time periods, as the sown perennial prairies became established and competed more strongly for resources. Once the prairie species were established, other than for periodic weed control (Fig. 2.4b), management primarily focussed on restricting grass biomass by prescribed burning and strategic grazing to preserve forb diversity. Importantly, these older restored prairies became important seed resources for future restoration.

Fig. 2.4

(a) Cotton Wood seedlings (Left); (b) Herbaceous weeds (Right)

Major Project Outcomes

The PRPRP has been a great success. Since the 1990s it has restored more than 1500 acres of rare native prairie in the Platte River region of Nebraska, where it was once extensive (Fig. 2.5a, b). Importantly, this has been done using large numbers of native species, ensuring that the restorations are functionally complex and resilient. Together they have created a chain of prairies and wetlands representing a corridor of native habitat in an otherwise biologically depleted agricultural landscape. These restored prairies have also become ‘working laboratories’ for continued development of innovative techniques and knowledge related to prairie restoration and management. Most importantly, they are full of life and include birds, mammals, reptiles and amphibians, invertebrates and plants with animals moving from local fragmented remnants into and through these restored prairie lands.

Fig. 2.5

(a and b) Mature native-dominated restorations

Remarkably, ongoing monitoring and assessment have shown that almost every species used in sowings has become established, although experience has shown that some have proved more difficult to introduce or maintain than others. Perhaps even more importantly, monitoring has shown that these restored plant communities, established and managed over as long as 20 years, have maintained their ecological integrity and resilience. This is an important outcome given the unpredictable future faced given the likelihood of major climate change. Using a combination of approaches, including field days, site visitations, tours for government program managers and social media, project staff have tried to communicate to the broader conservation sector that it is possible to restore prairie landscapes using relatively simple techniques around seed harvest, handling, seeding and management. In doing so they hope to influence prevailing beliefs about the feasibility of prairie restoration locally, across Nebraska and at a national level.

What About the Project Worked, What Did Not Work and Why?

Despite these many successes, adoption of high diversity prairie restoration is still relatively uncommon across the United States. There are likely to be many contributing factors, but prime among them remain cultural resistance (i.e., productive land) and high financial costs (i.e., expense). It remains too expensive for most landholders or farmers to undertake such complex restoration on a larges cale, and government programs to date are not yet able to provide sufficient funding to prompt broader uptake. This is not to diminish the importance of agri-environmental programs such as the Conservation Reserve Program the Grassland Reserve Program, the Wetland Reserve Program or the Monarch Butterfly Program, each of which go some way to encouraging and incentivizing landholders to restore native vegetation on parts of their lands and thereby promoting growth and capacity of the seed and restoration sector more broadly. However, most of these initiatives do not result in permanent protection of restored landscapes, leading some to question the investment of public funds in these transient projects.

In the United States, there has been a long history of farm support programs, even in the case of very marginal lands. This could mean that governments and society now need to seriously consider alternative decisions about how and where farm support monies could or should be spent. It is possible that instead of programs trying to support increased farm productivity in very marginal landscapes, they might be better utilized by financially supporting these farmers to restore and manage, into perpetuity those marginal lands back to their native habitat. This will increase the ecological integrity and biological diversity of such fringe landscapes. Studies of restored Conservation Reserve Program lands have shown that there are many flow-on benefits to farmers from restoring native vegetation. These include improved water quality, reduced soil loss, carbon capture and potential alternative income streams. For these practices to be seriously and widely adopted (in marginal or other landscapes), governments would also have to significantly revise their approaches and metrics to collecting farmland tax.

It is also important to recognise that the goal of the PRPRP was not to reverse historical agricultural progress or to convert a sizeable percentage of currently cropped lands back to native prairie. Rather, it aimed to demonstrate that prairies can be returned to strategic parts of agricultural landscapes to increase native biodiversity and natural beauty, whilst at the same time providing other important ecosystem services. For over two decades now, the PRPRP has worked towards restoring up to 200 acres of prairie per year, engaging with local communities, training future generations of conservationists and spreading a message of hope that these activities can be replicated elsewhere. To that end, the Conservancy has initiated similar projects in other states and hopes that one day works of this nature are commonplace across the United States.

Case Study 2: Grassy Community Restoration in Australia²

Project Rationale(s) and Strategy(ies)

Prior to European settlement in the continent of Australia, temperate native grassy landscapes were managed and maintained by indigenous peoples (Gott et al., 2015). With the arrival of European settlers, there was a sharp cessation of indigenous cultural and management practices and the introduction of Northern hemisphere-based agricultural approaches, which, together, had a dramatic and disruptive impact on native grassy communities (Williams & Morgan, 2015). During the 1970s, 1980s and 1990s, there was much focus turned towards grassland conservation, and this was supported through legislative protection and regulation from Federal and State governments. However, despite this signal, agriculturally linked factors (in addition to others associated with human activities) continued to degrade native grasslands and grassy woodlands, leaving them among Australia’s most threatened communities (Fig. 2.6a, b, Kirkpatrick et al., 1995).

Fig. 2.6

(a & b) High-quality remnant grassy communities

The Grassy Groundcover Restoration Project (GGRP) was initiated to respond to this alarming situation at a time when there was little confidence from conservationists, ecologists or researchers that native grasslands and grassy woodlands could be reinstated by restoration. The project began as a collaboration between the University of Melbourne and Greening Australia (an environmental non-government organisation), and its underpinning centred on promising recently completed doctoral studies focused on the feasibility of grassland restoration (Gibson-Roy, 2004). The project aimed to further explore and develop findings from these and other early studies under ‘real world’ conditions (Gibson-Roy, 2005). The project began in 2004 and continued in various forms until 2019, eventually leaving a legacy of grassy restoration sites across Southeastern Australia and inspiring other groups to take up these learnings and practices (Gibson-Roy, 2022).

Major Project Concerns and Barriers

Sadly, the early 2000s saw little appetite or formal support from government or their agencies for undertaking grassy restoration. Indeed, following the Federal government’s 1989 commitment to replant one billion trees nationally, most restoration programs were almost totally focused on the woody strata. Also, the extreme degradation faced by grassy communities meant that many conservationists questioned activities that might further negatively impact on them, including seed-based restoration, which was viewed as a well-intentioned but inappropriate use of critically rare native seed. Under these settings, grassy community research and endeavour tended to focus on their ecology and management rather than on their restoration. Assessments of the few small-scale restorations undertaken, which were typically by hand plantings, showed only limited success due to harsh climatic conditions and herbivore impacts, but most commonly due to weed competition (Berkeley & Cross, 1986; Scarlett & Parsons, 1992; Shears, 1998; Delpratt, 1999; Morgan, 1999; Gibson-Roy, 2000; Smallbone et al., 2007) (Fig. 2.7).

Fig. 2.7

(a & b) Woody tree and shrub plantings typical of government-funded programs

Key Project Features

Experimentation

Over its life, the GGRP maintained a focus on the restoration of native grassy communities in the context of disturbed landscapes such as ex-agricultural land, urban development areas and transport corridors. Core elements included (i) a focus on high diversity restoration; (ii) the management of elevated nutrient levels; (iii) the manipulation of weed-dominated soil seed banks; (iv) the refinement of seed production, seed harvesting, seed processing and direct seeding technologies and techniques; and (v) post-establishment management.

The early years of the project were strongly focused on experimentation and capacity development (Gibson-Roy, 2005, 2010, 2012, 2013; Gibson-Roy et al., 2010a, b; Taylor et al., 2013; Gibson-Roy, 2014a; Gibson-Roy et al., 2014), whilst in the latter years, this moved to applying or refining initial learnings on a larger scale or under different conditions and settings (Gibson-Roy, 2014b; Gibson-Roy & Denham, 2014; Gibson-Roy & McDonald, 2014; Delpratt & Gibson-Roy, 2015; Gibson-Roy & Delpratt, 2015; White et al., 2017; Morris & Gibson-Roy, 2018; Cuneo et al., 2018; Morris & Gibson-Roy, 2019a, b; Schmidt et al. 2020).

For the early experimental phase of the project, a steering committee was established to facilitate good governance and a technical panel to advise on the design and undertaking of experiments. To promote the project and its aims and to garner interest from landholders prepared to host one-hectare experimental sites, public presentations were held across the central and southwestern parts of Victoria. Landholder interest was overwhelming, and 11 locations (from over 50 offered) representing a wide range of land tenures, including farms, roadsides and public reserves, were chosen. These landholders agreed to preserve and manage restored sites in consultation with the project managers and to provide access for long-term monitoring, which was estimated to be 10 plus years.

Experimental treatments were developed under the guidance of the technical panel and applied across these 11 sites. These aimed to address the issues of excessive nutrification and weed-dominated seed banks by either exhausting or physically removing weeds and nutrient loads. ‘Exhaustion’ plots were treated with either 1, 2 or 3 years of fallowing by herbicide treatment (four per year), with each application preceded by shallow cultivation to stimulate weed banks before spraying. These were compared to ‘removal’ plots, which were treated by topsoil removal through mechanical stripping to a depth of 10 cm (Fig. 2.8).

Fig. 2.8

Topsoil manipulation. (a) small-scale excavator (Left); (b) large-scale grader and scraper (Right)

Seed Resourcing – Collection

Due to a lack of markets for their restoration, commercial seed supplies were largely unavailable. To meet this fundamental requirement, seeds were initially sourced through field collections from remnant areas and later supplemented through cultivated seed production techniques. Not surprisingly, locating remnant populations in highly fragmented landscapes was a great challenge, and collection zones were defined in relation to each sowing site, taking into consideration their current and historical distributions and elements of past connectivity between species and local populations. The aim of collection boundaries was to minimize the risk of creating inbreeding populations (in both restored and seed production sites), increase the diversity of species and the amounts of seed available for restoration and improve the adaptive potential within the restored communities and to preserve regional identities (Broadhurst et al., 2008; Bischoff et al., 2010).

Within seed zones, collectors aimed to match source population and sowing site conditions based on soil type and topography (Cole et al., 1999). Most of the seeds were collected within 50 km of each seeding site, with collectors targeting all relevant species which had been located and which were producing seeds. Hand collections aimed to take seeds from 50+ plants per species per population and to avoid conscious or unconscious selection to reduce the potential for relatedness. For mechanical collections (e.g., brush harvesters) population size routinely exceeded 10,000 individuals. Seeds were harvested from multiple source populations within a collection zone over the entire span of a ripening season. Project staff worked closely with seed collectors and seed production area (SPA) growers across all regions to improve their species recognition, and harvesting and processing skills and processing skills using regular project forums, workshops and technical newsletters. Ultimately the project established a highly proficient and dedicated group of collectors and growers across the regions in which it operated (Fig. 2.9).

Fig. 2.9

(a) Roadside grassy remnant (Top left); (b) brush harvested seed (Top right); (c & d) teams hand harvesting (Bottom left and right)

Seed Resourcing – Seed Production

Wild collections only provided enough seed for a limited range of species (primarily grasses). Therefore, the project began to cultivate species in seed production areas (SPAs) to supplement wild sources. SPAs established early in the project were each linked to a growing region and provided seeds for one or more restoration sites (up to three). SPAs were usually set up in association with local native plant nurseries and grew plants in simple containerized settings (typically foam boxes) as high-density irrigated crops where competition from weeds and herbivores was minimized. Over time, in the latter stages of the project, these approaches were refined, modified and expanded and a smaller number of SPAs were developed at larger scales to service the seed needs of bigger restorations using more advanced growing systems that included weed-mat-covered in-ground beds, open field beds, raised covered beds and vertical trellis beds (Fig. 2.10).

Fig. 2.10

(a & b) Containerised production systems (Top left and right); (c & d) weed mat production systems (Bottom left and right)

SPAs proved critical to the project’s success. Through these facilities, large quantities of high-quality seed from numerous species indigenous to the restoration region were produced. Furthermore, these SPAs relied on only small amounts of wild-collected seeds to establish crops, which reduced impacts on remnant communities. Importantly, while collection protocols for remnant populations aimed to capture a broad range of genetic traits, similar protocols were established to increase the likelihood that these traits would be preserved through the seed production phase. In practice, this meant appropriate mixing and sub-sampling of wild seed-lots when propagating production crops, avoidance of selection bias when pricking-out seedlings for bed plantings and harvesting seed from production crops over the whole fruiting period. SPA populations for a given species also contained as many individuals as possible (given space and resource considerations) but were typically composed of between hundreds and several thousands of individuals. In many cases, this meant that SPA crops were actually much larger than the source population/s. To further lessen the potential for genetic bottlenecks, crops were maintained for only two harvest seasons before new genetic material was introduced from wild populations.

Many hundreds of species were grown in seed production areas by the project over its life. Most species were readily propagated from seed and suited to some form of cultivated production system. Seed production enabled the project to grow sowable quantities of seed from many threatened species that would otherwise not have been available for use in restoration (Gibson-Roy & Carland, 2023; Gibson-Roy, 2010). These ex-situ populations of rare species also afforded them some protection from localized extinction where in situ populations may have been further impacted or even destroyed by some form of human disturbance. Another key feature of SPAs was that they represented large collections of species growing in centralised locations as weed-free monocultures, and this dramatically simplified collection in comparison to wild harvest. No longer did collectors have to spend long months travelling large distances, often in harsh conditions to locate and harvest seeds. SPA crops were maintained for ease of harvest and produced more reliable quantities of seed at times when source populations were often severely impacted by harsh climatic conditions such as drought, storms and other events, including uncontrolled fires, grazing and predation. In addition, most species cultivated in SPAs produced seeds over much extended periods in comparison to those in the wild.

Seed Resourcing – Seed Quality Characterisation

Seeds used in sowings (wild or production) were assessed for quality characteristics using purity and germination tests. Seed mixtures were also sampled at the time of sowing and germinated under nursery conditions, and where possible, in germination cabinets to gauge germination and emergence potential at the time of sowing (Gibson-Roy et al., 2010a). These approaches enabled important understanding of the seed’s characteristics, both post-harvest and at the time of sowing, making it possible to reliably consider seed quality in the post-analysis of field emergence patterns rather than attributing good or poor field outcomes solely to post-sowing factors such as soils, rainfall, temperature or predation (Fig. 2.11).

Fig. 2.11

(a) SPA seed drying (Top left); (b) seed mixing (Top right); (c) seed testing under temperature- and light-controlled conditions (Bottom left); (d) seed storage under temperature 15 C and 15% relative humidity conditions (Bottom right)

Site Preparation and Seeding

GGRP sites had various land use histories. Therefore, at each, soil testing was undertaken to determine the key soil characteristics, including soil texture, colour, pH, nitrogen, phosphorus and electrical conductivity. Some sites were located on cropping land, and these exhibited elevated nutrient levels and, because of long cultivation histories, had deep weed-dominated soil seed banks. Others had a history of pasture grazing, with lower fertilization and minimal cultivation, which left shallower weed seed banks and less nutrified soils. A small number was located on road verges where vegetation was typically dominated by colonising pasture grasses from nearby farms, together with broadleaf weeds. These had been exposed to regular soil and vegetation disturbance by road managers such as using slashers or graders or herbicide spray machinery, often meaning they were volatile in terms of weed loads and soil health. At all sites, native herbaceous species were largely absent or possibly represented by a few common species in very low numbers.

In the first year of experimentation, treatments were applied to 2 × 2 m plots prior to being seeded by hand. As larger quantities of seed became available from SPAs in the following 2 years, plot size increased to 2000 m⁻². In later years (beyond the experimental phase), sites ranged in size up to 16 ha and almost exclusively used soil scalping, soil inversion or subsoil capping as the core nutrient and weed seed/bud bank treatment methods. All 2000 m⁻² and larger plots were machine seeded using oscillating rotating tines to develop lightly tilled friable soils, while modifications to the seeding box allowed the seed mix (and sand carrier) to fall as a curtain onto the prepared seed bed with a mounted rake and roller lightly covering and pressing seed into the soil. Seed flow and tractor speed could be adjusted to achieve accurate sowing rates. This machine proved very effective on a broad range of soil types and conditions and for all native seeds used (ground layer, shrub or tree) (Fig. 2.12).

Fig. 2.12

(a) Small experimental plots following seeding (Left); (b) large-scale plots being mechanically seeded (Right)

Sowing mixes contained up to 100 species. Grasses represented approximately 60–70% of seed mixes by mass, dominant forbs 10–15%, with sub-dominant forbs (or other functional groups) making up the remainder. These sub-dominant species represented the bulk of species diversity in the mix. Sowing rates varied from site to site and from year to year, being linked to seed availability, seed quality and restoration goals, but, in general, rates of between 40 to 50 kg per ha (representing pure seed and chaff) were used.

Major Project Outcomes

Approximately 230 species were established in the first 3 years of experimental sowings. These included 20 grass genera, 74 forb genera and 10 sub-shrub genera, showing that a wide range of ground layer species could be established by direct seeding (Gibson-Roy & Delpratt, 2015). A great many more species were used in later sowings across different regions of Victoria and into other States over the following decade and a half (e.g., Morris & Gibson-Roy, 2018; Cuneo et al., 2018). In terms of experimental treatments, the most outstanding finding related to the differences in establishment success between soil stimulation plus herbicide-treated plots (n = 96 − 2 × 2 m plots & n = 22 − 2000 m⁻² plots) and soil-removed plots (n = 96 − 2 × 2 m plots & n = 22 − 2000 m⁻² plots). Here, monitoring revealed that species diversity, plant densities and structural composition were significantly higher or better on soil-removed plots compared to the long-term cultivation and herbicide-treated plots (Table 2.1). In many cases, this left them comparable in quality and composition with reference remnant communities, and importantly, these characteristics were largely maintained over the following decade and a half since (Gibson-Roy & Carland, 2023).

Table 2.1 Comparisons of differences in mean values from measurements taken on topsoil removed and soil stimulation plus herbicide-treated plots located at 11 sites in western Victoria. Diversity = species number per plot; plant counts = plant number m⁻²; vegetative cover = percentage cover m⁻²

Underpinning this outcome was the effect of soil removal in restricting key nutrifying elements in topsoil-removed plots compared to non-removed. Soil testing of plots revealed that following soil removal, phosphorus levels declined to an average of 14 mg/kg, making them like those observed in reference communities (<20 mg/kg – Gibson-Roy et al., 2010a, b). Likewise, nitrogen levels were also reduced by half or more in comparison to non-soil removed plots. Analysis of vegetation data revealed a strong relationship between low P levels and higher native diversity and density, while low nitrogen levels corresponded to a reduced dominance of all grasses (exotic and native), which further aided sub-dominant native species’ persistence (Fig. 2.13).

Fig. 2.13

(a & b) Seeded restorations on topsoil-removed sites. (a & b) wildflower rich restorations (Top). (c & d) grassy-dominated restorations (Bottom)

Conversely, diversity and structural complexity on cultivated and herbicide-treated plots was lower, regardless of the duration of treatment application (1–3 years). Whilst this treatment removed all standing vegetation at the time of sowings, monitoring revealed that large numbers of weeds continued to reemerge from soil seed and bud banks to compete with the sown natives, and because soil P and N remained at agronomic levels, these nutrients helped exotic species grow to a larger size or greater percentage vegetative cover than co-occurring natives. This finding clearly showed that weed-dominated agronomic soil seed banks were not exhausted despite long-term cultivation and herbicide treatment and continued to produce large numbers of emergent weeds which would compete vigorously with colonising, sown or planted natives. This provided a stark reminder of the degree of alteration that agriculture has had on soils and plant communities in these landscapes. It also highlighted another positive feature of soil removal which was that, as well as treating nutrient-laden layers, the process also removes weed seed and bud banks.

Among the species established by restorations were those that were locally, regionally and nationally threatened. Indeed, several GGRP restorations represented new populations of endangered species and some featured populations which greatly exceeded the size of wild source populations (Cuneo et al., 2018; Gibson-Roy, 2010; Gibson-Roy & Carland, 2023). This was made possible largely through the combination of nutrient and weed seed bank reductions, the utilization of seed production approaches to increase seed supply from these species for sowing and purpose-designed seeding equipment. Ongoing seedling recruitment from restored species was also verified by seedling emergence close to mature adults and from seedlings appearing in unsown areas such as adjoining walkways and large bare ‘recruitment zones’ left adjoining sowing areas. In later years monitoring highlighted second, third and fourth generations of plants dispersed over considerable distances of up to several hundred metres from the initial planting (Gibson-Roy et al., 2010a; Gibson-Roy & Carland, 2023) (Fig. 2.14).

Fig. 2.14

Seeded listed threatened species established in restorations with recruiting seedlings surrounding mature adults; (a) hoary sunray (Leucochrysum albicans subsp. albicans var. tricolor) (Left); (b) button wrinklewort (Rutidosis leptorrhynchoides) (Right)

Another important feature of GGRP restorations was their high levels of colonization by other native animals and plants, indicating increasing functionality at multiple trophic levels. Insects, birds, mammals, amphibians and reptiles were routinely observed feeding, sheltering or nesting within restorations, where they had not been present before restoration (Gibson-Roy & Delpratt, 2015) and also within seed production areas (White et al., 2017, Schmidt et al., 2020). Native trees (eucalypts and acacias) commonly reappeared within restored areas where a nearby tree canopy provided a seed source. In other situations, unsown native ground layer species emerged from seed or bud banks or from colonisation from outside restored sites (especially where they adjoined remnants) Gibson-Roy & Carland (2023). Importantly, investigations of native plant roots from several restorations (scalped and non-scalped) as well as from reference areas also showed functioning arbuscular mycorrhiza at similar levels across all (Gibson-Roy et al., 2014) (Fig. 2.15).

Fig. 2.15

Examples of fauna colonising restored sites. (a) Native spider on grasses (Top left). (b) growling grass frog (Litoria raniformis) (Top right). (c) chequered copper butterfly (Lucia limbaria) (Bottom left); (d) little whip snake (Suta flagellum) (Bottom right)

What About the Project Worked, What Did Not Work and Why?

Experience from the GGRP over many years has highlighted (i) the need for good planning and goal setting; (ii) the importance of re-establishing complexity and function in restored communities; (iii) the value of the application of horticultural and agricultural principles, as well as ecological understandings in restorations; (iv) the success of seed production in addressing seed and species limitations; (v) the importance of the development and use of specialised restoration technology; (vi) the worth of embedding (where possible) experimentation within projects; (vii) the need to monitor and quantify outcomes to inform future practice and (viii) the value in purposeful involvement of stakeholders, communities and others in projects. The extended period over which the GGRP operated and the packaging of these key points into a single project enabled it to clearly demonstrate the feasibility of grassland and grassy woodland restoration in disturbed landscapes.

The early experimental phase of GGRP was critical to its success. It was during this period that various replicated, field-applied experimental treatments were tested and where outcomes were monitored, evaluated and verified. These findings provided strong evidence for the efficacy of soil removal in treating elevated nutrient levels and weed seed and bud banks. These two factors are typically fundamental constraints to ensuring native ground layer species can establish and persist in restorations (Gibson Roy et al., 2010a, b). Based on this knowledge, the project confidently undertook many other larger restorations across several states (Cuneo et al., 2018; Morris & Gibson-Roy, 2019a, b; Gibson-Roy & Carland, 2023). An important application and expansion of these approaches occurred between 2013 and 2018, when after developing regional-scale seed production capacity, the GGRP undertook a series of restorations totalling nearly 50 ha in the urban matrix of Sydney, which is Australia’s largest city. Mostly undertaken in western Sydney on the Cumberland Plain, the project restored nationally threatened grassy woodlands in parklands, council reserves and in national parks, where before, none had been successfully restored (Cuneo et al., 2018). This showed that after a decade of demonstrating the efficacy of these methods in rural landscapes, these same approaches could be used to restore grassy communities in the urban context. In the intervening years, other groups, informed by these techniques and led by their own goals and motivations, have taken and, in many cases, further refined these approaches to implement successful grassy community restoration projects of their own (Gibson-Roy, 2022) (Fig. 2.16).

Fig. 2.16

Urban restorations. (a) wildflower-rich council reserve (Left); (b) grassy sward on public parkland (Right)

Globally, many people now recognise that full ‘ecological reconstruction’ is crucial to re-integrating native grassy communities into landscapes highly fragmented and degraded by agriculture, and this Case Study has demonstrated examples of pathways to success under Australian conditions. However, this Case Study does not represent the end of this quest. An emerging understanding of the Australian experience is that despite the best efforts of many committed people the prognosis for complex ecological restoration of grassy communities in Australia remains bleak (Gibson-Roy et al., 2021a, b; Gibson-Roy, 2022). Australian governments, and their agencies, together with the academics and researchers to whom they turn to as authorities, continue to overwhelmingly focus on conservation rather than a balance between conservation and restoration. This situation persists even though native grassy communities continue to be lost to agriculture and other forms of urban development. In parallel, the almost total bias of many decades towards tree and shrub plantings (primarily for carbon or functional outcomes) has also been maintained. Indeed, there are effectively no legislative or regulatory incentives for the uptake of complex grassy community restoration across arable landscapes even though the need has been identified (Mappin et al., 2021). Without strong and reliable markets for native seed and restoration services, restoration at a significant landscape-scale has not occurred, leaving the restoration sector small, incapacitated and dysfunctional. Despite clear and long-term evidence from the GGRP and other groups that complex restoration is indeed feasible and there being examples from other parts of the world showing governments can provide the right mix of policy and regulatory mandates to create an environment for seed and restoration markets to develop, for sector capacity to increase and for restoration at landscape scales to occur – in Australia, we remain steadfast in our reluctance to pursue such opportunities (Gibson-Roy, 2022).

On the positive side, Federal government parties of all political hues continue to support a national Environmental Protection Biodiversity Act, which in effect states that Australia’s biodiversity should not be allowed to disappear. It is in this context that the GGRP, and the various small-scale projects that have since followed, have been able to provide a base of evidence, knowledge and experience which can and should be built upon and supported by governments and their agencies. Such a situation could see landholders, land managers and communities joining to restore complex native vegetation across vast areas of arable land and to long distances of road and rail corridors from which it has almost disappeared and where its restoration is most needed. If in the years and decades to come this does not occur, it will not be because it was not possible, but rather it will be because we were timid or too self-centred to ensure it did.

Case Study 3: Grassy Community Restoration in Belgium³

Project Rationale(s) and Strategy(ies)

As a result of a thousand years of mixed agro-pastoral practices, permanent grasslands on the European continent have complex structures and do not have easily identified natural states. For example, the grassland ecosystems of particular interest in this case study, in the southern half of Belgium (called Wallonia), are among the most species-rich vegetation types, sometimes exceeding 60 plant species per m² (Merunková et al., 2012), but these species are liable to represent many phases of invasion assisted by animals, wind, water and human seed transfer.

As a result, notwithstanding this rich vegetation environment, the question of the ‘restoration’ of these grasslands to their natural state is still somewhat contested. In this respect, in more recently colonised countries such as the United States or Australia, there are still remnant natural grassland communities which can act as benchmarks for restoration in many climatic and regional areas. In these instances, the term ‘restoration’ has a specific and criterion-based meaning. By comparison, in Europe, it has been estimated that about 20% of all European grasslands still have a so-called favourable conservation status within the meaning of Article 1 of the Habitat Directive (European Environment Agency, 2020),

Three conservation status categories can be assigned to European grasslands (A, B, C) (European Environment Agency, 2020). Recent assessments have shown that 76% are now in conditions that meet the ‘unfavourable conservation status’ category, indicating the extreme pressures these landscapes have been under. The conservation status ‘goal’ for our project restorations was to achieve an A (favourable) category. To this end, project managers sought to assist the recovery of degraded, damaged or destroyed grasslands at numerous sites and settings.

Furthermore, in Wallonia, only 5% of regional grasslands are currently in a favourable conservation status, and it is known that between 1955 and 2009, the area of permanent grasslands in Belgium was reduced by a third (Belgian Federal Government – https://statbel.fgov.be/). The major causes of reduction were urbanization, conversion to cropland, abandonment, and plantations of exotic species (particularly to Norway spruce, Picea abies). The remaining grassland areas in Wallonia were highly intensified, especially from 1955, through the use of chemical fertilizers, the earliness and frequency of successive mowing and from increases in livestock density. In addition to the gradual loss of species-rich grasslands, essential elements of the landscape, such as ponds, hedges and orchards, have also disappeared. This represents a major cause of local extinction of many animal species living in these environments, such as insects, birds, bats and amphibians (Fig. 2.17).

Fig. 2.17

(a) Low diversity degraded grassland (Right); (b) image showing grassland meadow divided into two plots 1 year after restoration actions have taken place (Right). Left section of this image shows restored plot area with forbs establishing while right image section shows unrestored plot area. Photo credits: (a) Maïké Dellicour; (b) Patrick Lighezzolo

Today, ecosystem restoration is recognized as a priority (UNEP & FAO, 2020). A growing number of scientists are convinced that it is necessary to intervene by reintroducing or reinforcing plant communities for the purpose of nature conservation. This option was adopted by the partners of several ongoing LIFE projects co-financed by the European Commission and in particular the ‘LIFE Bocage Meadows’ project (LIFE11NAT/BE/001059), led by Natagora, a nature conservation NGO (Goret et al., 2020).

Major Project Concerns and Barriers

When trying to restore habitats, (i) it is difficult to decide the nature of the technique which will be chosen for implementation; (ii) budgets are often limited, while restoration is expensive; and (iii) it is often not easy to predict species recovery trajectories. We, therefore, wanted to improve the cost-efficiency of our restoration strategies and ensure that appropriate action plans were developed. For grasslands such as those defined by the European Union (see European Commission, 2013), information exists on the results of previous conservation efforts published through technical notes, detailed action plans and scholarly scientific articles. However, this information is somewhat scattered and provides only approximate indications of the target habitat criteria or the environmental conditions under which restoration has taken place. It was, therefore, sometimes difficult for us to know if, in our cases, it was relevant to apply a technique recommended in other (unknown) contexts. It became clear that any sort of tool proposing restoration measures adapted to each type of European grassland was sorely lacking before the implementation of our project. We, therefore, decided to develop a knowledge and literature-based decision tree to facilitate the adoption of the most appropriate restoration techniques for the Wallonia area (Goret et al., 2021).

Key Project Features

Within the framework of the ‘LIFE Bocage Meadows’ project, we restored around 200 ha of lowland hay meadows (Arrhenatherion community) between 2012 and 2020, following the methodology developed in our decision-making tool (Goret et al., 2021). The first step consisted in carrying out exhaustive floristic surveys in the meadows to be restored. These inventories made it possible to identify the phytosociological alliance of the meadows and to verify whether they were already in the Arrhenatherion target habitat. We subsequently determined their conservation status, according to the methodology outlined in the Habitats Notebook of the Department of Natural and Agricultural Environment Studies. This considered the presence and abundance of characteristic species in the area to be restored. Meadows to be restored are classified into three categories. Meadows in good conservation status are classified as A and thus do not require any further intervention. Medium-level conservation status is classified as B; and poor conservation status is classified as C. Finally, classification X refers to areas where target habitat is absent and can evidence bare soil after deforestation or degraded meadow of the Cynosurion type.

Regarding the techniques that were employed in this work, provided that a threshold of 5 mg of available phosphorus per 100 g dry soil (Janssens et al., 1998) was not exceeded and that the main threats to conservation such as fertilization and agricultural management have disappeared, we can summarize our approach as follows:

(i) If the conservation status of the habitat is C or X and it was not A or B less than 5 years ago (based on the average lifespan of the soil seed bank of characteristic species), then we carried out population reinforcement by two seeding techniques after having exposed 50% of the soil:

• Sowing of harvested seeds from source meadows in good conservation status as close as possible to the target meadow.

• Green hay transfer from source meadows in good conservation status as close as possible to the target meadow.

(ii) If the conservation status of the habitat is B (or it was A or B less than 5 years ago), we considered that the seed bank of species characteristic of the habitat was still present in the soil and that it was necessary to promote their germination and the development of seedlings. After total cessation of fertilization, the restoration consisted of mowing twice a year, the first time from June 15 and again in September. This mowing regime allowed nitrogen to begin to be exported (starting a process that takes 10 to 15 years), and it enabled the plant cover to be as low as possible at the time of germination, namely between April and October. Grass that grows back after mowing can be treated again between August and October (Figs. 2.18 and 2.19).

Fig. 2.18

(a) Mechanical seed harvest (Left); (b) collected seed (Right). (Photo credits: (a & b): Xavier Janssens)

Fig. 2.19

Seed hay application. (Photo credit: Thibaut Goret)

In each case, post-restoration management had to be adapted and controlled for a number of years before being readjusted to recurrent management. This was judged to be appropriate when the conservation status of the area had improved significantly. After seeding, it is therefore essential to carry out two to three mowing sessions per year. In the following years, it is possible to move to two mowing per year, maintained until the meadow moves into a good conservation status and can therefore be managed by a single mowing per year in July (late mowing). To ensure the return of the entomofauna, it is essential in all cases to maintain sufficient refuge areas for insects, and we sequestered at least 10% of the surface on which we were working. Our sowing operations were carried out in September, which was the best period for seed germination of characteristic species and to promote the development of the seedlings. We deliberately avoided the summer droughts season, which can cause significant suffering to seedlings.

Major Project Outcomes

To measure the success of the project, we created a transition matrix showing the evolution of the conservation status of treated areas (Table 2.2). While 71% of the grasslands were initially in a poor conservation status (C or X), after our treatment, 87.6% of these grasslands are now in a good or medium-level conservation status (A or B). Eighty-six percent of the treated area has seen its conservation status improved throughout the project. It is noted that the improvement was not always an increase in a single level of conservation status. Rather, over the 6-year monitoring period, 34.2% of the area improved by 2 conservation status levels and 9.2% improved by 3 levels. Inversely, the reason why some treatments did not result in a conservation status improvement is that most of the work was done only 1 or 2 years before the final monitoring, and thus, the time elapsed was too short to observe significant improvement. The second reason is that the seed bank was probably missing and that restoration by only changing the mowing regime was thus not enough to overcome this constraint.

Table 2.2 Transition matrix showing the evolution of the conservation status of restored meadows

Species richness significantly increased with time for the three restoration techniques (Fig. 2.20). The number of species gained per year was equivalent for all techniques with on average 2.4 additional species being found each year. After 6 years, the mean species richness after mowing and sowing reached that of the reference meadows. This was not the case for fresh hay transfer where an increase to 33 species was found compared to 47 species in the reference meadows. This difference is explained by the fact that initial states of meadows, and thus the species richness and composition, differed among the restoration techniques. Fresh hay transfer had a significantly lower initial species richness than mowing and sowing (23 species compared to 27 and 30 species for sowing and mowing respectively). This result indicates that time is a major factor for the botanical restoration of meadows. Indeed, in the case of mowing, 6 years were needed to observe a successful species richness recovery. Five years were necessary for sowing, and 7 years would probably have been necessary to observe complete success of species richness recovery with fresh hay transfer (Fig. 2.21).

Fig. 2.20

(a & b) Species-rich restored grasslands. Photo credits (a & b): Maïké Dellicour

Fig. 2.21

Linear regression models showing significant positive relationships between species richness and time since restoration (mowing: F = 13.94, p < 0.001, sowing: F = 10.49, p = 0.003, hay: F = 19.1, p < 0.001). (Note: boxplots show the 25th percentile, median and 75th percentile)

The effect of these treatments on plant community composition was also evaluated and showed very encouraging results. Similarity between species composition of treated and reference meadows was calculated, and results clearly showed that similarity of restored plant communities with reference meadows significantly increased with time for the three treatments. On average 2.4% of similarity with reference meadows was gained each year (Fig. 2.22). Remarkably, it took only 2 years for the similarity between the treated and the reference meadows to be equivalent to the similarity observed between alternative reference meadows. This result indicates a restoration success since the recovery of species composition attained the level of similarity found within reference meadows.

Fig. 2.22

Linear regression models showing significant positive relationships between similarity with the average species composition of reference meadows and time since restoration (mowing: F = 12.18, p = 0.002, sowing: F = 25.86, p < 0.001, hay: F = 10.17, p = 0.004). Note: boxplots show the 25th percentile, median and 75th percentile

What About the Project Worked, and What Did Not Work?

Our investigations have shown that sowing, hay transfer and mowing, all led to equivalent significant changes each year in species richness and species composition compared to the reference meadows, which were chosen to represent conservation status A. It is noted that the final species richness of areas treated by fresh hay transfer did not reach that of reference meadows because initial states differed between techniques, and fresh hay transfer had a lower initial species richness. Following the decision-making tool of Goret et al. (2021), mowing was favoured for less degraded meadows and active population enhancement (sowing) was reserved for highly degraded meadows. Thus, having started with a lower number of species, meadows restored by fresh hay transfer will necessarily take more time to reach target species richness. All techniques showed successful regeneration of species composition. These results demonstrate that adapting restoration technique depending on the initial degradation state and the direct vicinity of the reference meadow is a relevant factor in successful conservation.

The outcomes of this project highlight the importance of soil preparation and transfer of seed-containing plant material in more impaired sites. This is consistent with the outcomes of several studies that have tested the effectiveness of species introduction to restore lowland hay meadows or alluvial meadows in Europe (Edwards et al., 2007; Schmiede et al., 2012; Baasch et al., 2016; Harvolk-Schöning et al., 2020). Success of mowing also attests to the efficiency of management extensification as a means of restoring slightly degraded meadows, based on the assumption that target species possibly remain in the seed bank. It is congruent with the results of previous studies which reported a positive effect on species richness after cessation of fertilization and implementation of extensive management through mowing or grazing (Pallett et al., 2016; Van Vooren et al., 2018).

In slightly altered landscapes providing seed sources and stopping disturbances are valuable tools for conserving valuable grassland habitats and achieve restoration goals (Ruprecht, 2006). Postponing mowing from spring to summer was demonstrated to be effective in promoting plant and invertebrate diversity in European meadows (Humbert et al., 2012). Similarly, a twice-a-year defoliation frequency was shown to be efficient in enhancing plant and insect richness and increasing export of potassium in agricultural lands (Uchida & Ushimaru, 2014; Piqueray et al., 2019).

Before any type of treatment, it is recommended that the local ecological conditions should be explored to decide which type of restoration action is more likely to succeed (Prach et al., 2020; Goret et al., 2021). Mowing should be favoured in mildly impaired sites, where there is low environmental stress and evidence of intermediate productivity, which are usually located in more well-preserved landscapes (Prach et al., 2020). When selecting a treatment, financial and practical factors must also be considered. Passive recovery naturally requires lower costs than seed transfer, while fresh hay spreading additionally imposes organisational constraints. Fresh hay must be transferred immediately to the receptor site after cutting, as storage would compromise seed viability due to rapid fermentation (Blakesley & Buckley, 2016). The large volume of fresh hay which needs to be transferred also requires close proximity between donor and receptor sites (Blakesley & Buckley, 2016). Notwithstanding these precursory conditions, compared to sowing, green haymaking is less time-consuming, requires commonly available machinery (Blakesley & Buckley, 2016) and produces a very more efficient seed harvest yield (Scotton & Ševčíková, 2017). The residual hay layer left on the receptor site can also favour seedling establishment (Loydi et al., 2013).

In conclusion, the ‘LIFE Bocage Meadows’ project shows that meadow restoration can be a great success if the treatment is adapted to the local conditions. These depend on the ecological context, which includes initial levels of degradation, presence of a seed bank and an adjacent well-preserved meadow. In addition, there are financial and practical factors that must be factored in depending on the environment. Decisions on the appropriate treatment can be difficult to make given the multiple factors that must be carefully considered. To this end, the recently published dichotomous key which was used in this project should assist practitioners to make appropriate choices for a successful restoration process (Goret et al., 2021).

Case Study 4: Grassy Community Restoration in Brazil⁴

Project Rationale(s) and Strategy(ies)

Tropical grasslands are essential global ecospheres, being home to much unique biodiversity, providing key ecosystem services and sustaining the livelihoods of hundreds of millions of people. However, notwithstanding these remarkable attributes, they are amongst some of the most misunderstood, neglected, mismanaged and threatened ecosystems worldwide. In Brazil, the disproportionate focus on forest ecology and its restoration, coupled with the economic interests of colonial legacies, have created widespread misconceptions on the ecology of old-growth grasslands which has had many detrimental, long-standing ramifications for our understanding of grassland restoration (Overbeck et al., 2015; Silveira et al., 2022).

Restoration programs in Brazilian open area biomes, which include grasslands, savannas and shrublands are currently in their infancy, and, in addition, the paucity and geographically limited nature of the studies which have been undertaken have hindered us from learning transferable lessons. However, the last decade has witnessed an upsurge in theoretical and empirical papers addressing grassland restoration which, collectively, have led to an improved conceptual framework tailored for grasslands and savanna restoration (Buisson et al., 2019). In this regard, the major sources of degradation of tropical grasslands include (i) land conversion for crops and pastures, (ii) quarrying, (iii) mining and (iv) afforestation (tree planting in former non-forested sites).

Fortunately, tropical grassland restoration practice and science are now becoming more common, and in this case study attention is given to three independent initiatives: (i) restoration of Cerrado grasslands (the world’s most biodiverse savanna) which has been degraded by conversion to pasture, (ii) restoration of Cerrado grasslands that are degraded by afforestation and (iii) restoration of nutrient-impoverished megadiverse montane, which include the campo rupestre grasslands in south-eastern Brazil (Silveira et al., 2016) which have been degraded by mining and quarrying. These three grassland types are floristically, climatically and edaphically distinct, and each has been impacted by different degrading factors to varying degrees. Consequently, restoration efforts have been necessarily localised. The goal of this case study is to provide a brief overview of restoration programs established in these three areas, then address the specific learned lessons, rather than to attempt a comprehensive review of all grassland restoration initiatives in Brazil.

Major Project Concerns and Barriers

The Cerrado, the largest Neotropical savanna, originally covered more than two million km². This area mostly involved seasonal climates and was found in the nutrient-poor soils of central Brazil. Its original distribution covered 20 degrees of latitude, with elevations ranging from 100 to almost 2000 metres above sea level (Borghetti et al., 2020). Cerrado vegetation has been found to be extremely heterogeneous, and variations are is driven chiefly by local fire regimes, water availability and soil fertility (Bueno et al., 2018). Open, fire-prone formations have a continuous biodiversity layer of herbaceous strata composed of grasses, graminoids, forbs and sub-shrubs and a discontinuous woody layer formed by scattered, small-sized shrubs and trees (Fig. 2.23a). For decades open grasslands in the core area of the Cerrado have been converted to intensive cattle raising using fertilized pastures composed of invasive African grasses. These pastures dominated by African grasses require intense and constant NPK fertilization, and decade-old fertilization regimes have produced negative legacies that have had drastic consequences for natural communities and ecosystems. Consequently, the major challenges for the restoration of open grasslands are (i) returning soil fertility parameters to pre-disturbance conditions and (ii) removing or suppressing the invasive grasses. In contrast to forests, native herbaceous species of savannas are light-demanding, and shade cannot be used to control or restrict the exotic species (Sampaio et al., 2019). Although challenging, returning sites to natural low soil fertility is expected to provide multiple benefits, including the prevention of invasion by African grasses (Giles et al., 2021), restoring the natural soil microbiota (Wolfsdorf et al., 2021) and the shifting of plant communities from fast-growing to a more appropriate slow-growing functional signature (Giles et al., 2021).

Fig. 2.23

Aerial view of the Chapada dos Veadeiros National Park in Central Brazil, where a large-scale restoration experiment has been implemented in a forest-grassland mosaic. (a) the red arrow points an area where multiple seed-based restoration treatments are being conducted, including seed sowing; (b) seedling planting in the Cerrado; (c) pescribed fires have been used in a nearby area to restrict invasive African grasses; (d) seedling planting in campo rupestre (e) and hay transfer have been experimentally tested to restore campo rupestre grasslands in Southeastern Brazil. (Photo credits: (a). Fernanda Barros; (b). Guilherme Mazzochinia; (c). Alessandra Fidelis; (d–e). Soizig Le Stradic)

In the State of São Paulo, representing the southern portion of the Cerrado, grasslands establish in places with more fertile soils, higher precipitation, and are located close to semi-deciduous forests. In addition to conversion to pastures, these grasslands have been degraded by replacement with or because of ingress from, pine tree plantations and general woody encroachment. Afforestation and encroachment pose major threats to the biodiversity and ecosystem services provided by these grasslands (Honda & Durigan, 2016, Haddad et al., 2020). Ironically, when these areas are targeted for restoration, they are commonly ‘restored’ by tree planting, which in fact represents inadequate understanding of the aim of the intervention, and has several negative consequences, leading to their degradation instead of restoration. Clearly, if restoration goals include the recovery of old-growth savanna biodiversity and structure, interventions are required to prevent woody encroachment and reintroduce a broad suite of native grasses, forbs and shrubs (Cava et al., 2018).

In southeastern Brazil, significant areas of the campo rupestre grasslands have been lost due to opencast iron ore mining and quartzite quarrying. Campo rupestre is characterized by extremely impoverished soils, outstanding plant endemis and communities dominated by slow-growing and seed-limited species. This combination results in negligible natural regeneration after soil disturbance (Le Stradic et al., 2018, Onésimo et al., 2021). As a consequence, major concerns include (i) sourcing large amounts of high-quality seeds required for specific local revegetation (Dayrell et al., 2016), (ii) developing appropriate protocols for contributing species’ propagation (Machado et al., 2013; Figueiredo et al., 2018a, b), (iii) finding suitable strategies for species’ reintroduction and undertaking long-term monitoring (Le Stradic et al., 2014; Gomes et al., 2018; Figueiredo et al., 2021) and (iv) recovering ecosystem function in sites where topsoil’s and sometimes subsoils, which have been entirely removed.

Key Project Features

Experimental Focus

Given the pronounced knowledge gaps in basic biological aspects of grassland species involved in these areas, initial studies have focused on an understanding of seed germination requirements, dormancy-breaking mechanisms, optimum conditions for seedling establishment and monitoring diversity after seed sowing or seedling planting (Pellizzaro et al., 2017). Although initially focused on a few woody species, researchers inevitably turned their attention to herbaceous species, which represent the bulk of diversity and which are essential for promoting soil vegetative cover (e.g., Figueiredo et al., 2020; Oliveira et al., 2021). In these early restoration attempts, seed sowing and seedling planting were understandably based on a trial-and-error approach, owing to virtually absent theoretical frameworks on grassland restoration.

More successful restoration of the open grasslands in Central Brazil was implemented through community-based networks that supply native seeds and seedlings for projects in the Cerrado area (Schmidt et al., 2019b). Many seed-sowing experiments were conducted in these projects, and a series of treatments were attempted to control invasive grasses (Fig. 2.23b, c). Direct seed-sowing experiments tested the survival of 75 native herbaceous and woody species for up to 2.5 years (Pellizzaro et al., 2017). Silva & Vieira (2017) evaluated the effects of seed burial, comparing surface exposure to buried seed and mulching (with no-mulch, 5-cm straw mulch and 10-cm straw mulch). The emergence, survival and growth of 16 woody trees of native Cerrado tree species with variable seed sizes and shapes and seedling type were also involved in this trial. In a second investigation, Coutinho et al. (2019) sowed seeds of 54 native grasses, shrubs and trees in order to test the effects of initial functional-group composition on assembly trajectory. Finally, Sampaio et al. (2019) tested whether seeding density affected native plant cover and whether soil ploughing is effective in controlling invasive grasses. These experiments were done in different soil types and with the different plant guilds of grasses, shrubs and forbs.

In one post-afforestation experiment (Haddad et al., 2020), the composition of herbaceous communities was compared across several treatments, which included (i) a burned and abandoned pine plantation, (ii) a burned and pine harvested site and (iii) an area planted with 82 native species which was previously used as a pine plantation. Haddad et al. (2021) later carried out an experiment consisting of a comparison of plant communities and vegetation structure in (i) abandoned pine plantation stands, (ii) areas open to passive restoration (natural regeneration after pine clearcutting) and (iii) native tree planting, where native tree seedlings were planted at high densities after pine clearcutting. The reference site was a Cerrado location which had never been exposed to tree planting. It has been noted that in previously afforested sites, the pine needle layer may prevent native regeneration after abandonment and cutting of the forest. In this respect, an experiment tested topsoil translocation, plant transplantation, direct seeding, topsoil translocation plus direct seeding and needle layer removal in both dry and wet grasslands (Pilon et al., 2018). Topsoil translocation involved the uprooting of plants and then extracting a 5-cm-deep layer of topsoil, where most seeds were concentrated.

For the campo rupestre project example, small-scale restoration experiments included monitoring the outcome of planting native shrub species (Fig. 2.23d), hay transfer (Fig. 2.23e) and direct seeding in experimentally degraded sites where the soil had been entirely removed. In the shrub planting experiments, Gomes et al. (2018) monitored survival, growth and recruitment of 10 shrubs 8.5 years after transplantation. The hay transfer experiment (Le Stradic et al., 2014) consisted of the spreading of seed hay (collected from pristine areas) at degraded sites and estimating seedling emergence after 14 months. The direct seeding experiment (Figueiredo et al., 2021) tested whether the addition of plant material (litter) improved seedling established of 14 native species. Finally, topsoil transfer was tested as a strategy to overcome the physical, chemical and biological filters of degraded ironstone campo rupestre by monitoring natural regeneration after spreading a 30 cm-depth layer of topsoil on bare soil in degraded areas (Onésimo et al., 2021).

Seed Networks

Since 2012, a partnership between the Brazilian Protected Areas agency (ICMBio), the University of Brasilia and Embrapa (the Brazilian Agriculture and Animal Husbandry Research Enterprise) has performed grassland restoration experiments in Central Brazil. These experiments aimed to develop efficient low-cost, seed-based techniques for restoring grasslands and savannas at a landscape scale (Pellizzaro et al., 2017; Coutinho et al., 2018). Direct seeding experiments also considered the effectiveness of the involvement of local communities and the use of agricultural machinery to reduce restoration costs. Three hectares were restored in 2012 and 2013, and a fruitful partnership with the Cerrado Seeds Network allowed the restoration of a further seven hectares in 2014. Seed collection, preparation and storage techniques were adapted or developed by that group using local ecological knowledge and available scientific information. People from local communities were trained and performed all restoration stages, from seed collection to seed sowing. In 2015, a power line company (Norte Brasil) established a pioneer agreement with ICMBio and local seed collectors to restore 95 hectares inside a protected area through mechanized direct seeding. The experiment sponsored by the power line company significantly increased the demand for native seeds and generated more than US$ 60,000 of income for more than 60 families within local rural communities in 2015–2016 (Schmidt et al., 2019b).

Understanding that seed collection for restoration can result in income generation and livelihood improvement, seed collectors funded a community association, named Standing Cerrado (Cerrado de Pé in Portuguese) that, in partnership with the Cerrado Seeds Network, now sells native seed for restoration projects in central Brazil. This cooperative has now become self-sustaining and is currently generating revenue to improve local livelihoods. It has also indirectly led to a decreased rate of local vegetation conversion, such as that involving pasture areas or other non-native use because conserved areas outside the park became important seed sources, generating income for local dwellers through seed collection and sale. The direct seeding restoration methods developed within this initiative cost less than US$ 3500 per ha compared to US$ 7000 per ha for tree seedling planting (Schmidt et al., 2019b). The cost of establishing a 1-year-old seedling by direct seeding has been shown to be cheaper than an equivalent approach through nurse-seedling planting for 56 of 57 species (Raupp et al., 2020).

Seed price for each species is established collaboratively among collectors, this value being based on the species density in natural areas coupled with the labour, time and equipment required for seed collection and seed processing. The Cerrado Seeds Network holds a National Register of Seeds and Seedlings and, in partnership with research institutions and universities, tests collections for seed quality, according to legal requirements. However, it is worth noting that because the use of native species is still incipient in grassland restoration, there are as yet in Brazil no established seed quality regulatory parameters for these species. Therefore, the Cerrado Seeds Network and partner institutions are developing and proposing parameters for seed quality tests for several species (Schmidt et al., 2019b).

Seed sourcing for large-scale restoration in campo rupestre is extremely challenging due to multiple reasons. First, most native and endemics are locally rare, have irregular fruiting seasons and have limited fruit production (Dayrell et al., 2016). Second, germination requirements and dormancy-breaking mechanisms are largely unknown. Third, large percentages of embryo-less and unviable seeds result in low-quality seed lots. Collectively, these factors make seed collection and seed quality unpredictable, consequently hampering extensive species propagation and restricting the establishment of seed production areas. Despite these initial challenges, intensive sampling in natural areas allowed for seed collection in sufficient amounts to make laboratory experiments, possible, together with greenhouse, trial-and-error seedling production (Dayrell et al., 2017).

Site Preparation and Seeding

In Central Brazil, sites were prepared in multiple ways before sowing seed. In the experiment by Pellizzaro et al. (2017), soil was ploughed prior to seeding to decrease dominance by invasive grasses and soil compaction one or two times during the dry season (May–October). In Coutinho et al. (2019), soils were prepared by repeated harrowing aiming for decompaction, uprooting invasive grasses and destruction of invasive grass seedlings that had germinated from the soil seed bank. In 2014, a controlled burn was conducted before soil harrowing to reduce invasive grass biomass. This approach facilitated the effectiveness and ease of soil harrowing and led to the removal of invasive grass seeds held in vegetation above the soil surface. Silva & Vieira (2017) spread seeds on the soil surface or buried at 3–5 cm depth and tested the effect of 5-cm and 10-cm straw mulching.

In Haddad et al. (2021), there was no soil tillage for native tree planting, so the underground structures of previously existing native species were preserved and therefore could resprout. Additionally, exotic grasses were controlled with glyphosate herbicide before native tree planting and for over 2 years afterwards. Pilon et al. (2018) and Haddad et al. (2020) did not perform site preparation in their experiments, except for what is mentioned above.

Most restoration experiments in campo rupestre did not involve any site preparation. This was most probably due to the assumption that native species adapted to low-fertility soils would be outcompeted by invasive species following increased soil fertility. However, in mined and quarried sites, iron ore or quartz extraction removed both soil and topsoil, so reconstructing the physical substrate remained a challenge that needed to be addressed before the seeding and planting stages. To reinstate native ecosystems post-mining, the development of a substrate similar to the iron-rich cap rock is necessary (Levett et al., 2021). To achieve this outcome, accelerating microbial iron cycling, dissolution and recrystallization of goethite catalysed by root exudates and bacteria and slope stabilization using biocrusts (complex association between soil, microorganisms and extracellular polymeric substances), have been tested as solutions to create environmental conditions suitable for the reintroduction of native species from ferrugineous campo rupestre. Nevertheless, a key challenge remains upscaling such biotechnologies to the landscape-level, which would lead to significant advances in mine-site restoration in Brazil (Levett et al., 2021).

A few studies examined the role of litter addition and topsoil transfer in the establishment of target species in post-mined sites. The establishment of native species was evaluated under four different conditions: (i) seeding on the degraded substrate, (ii) seeding covered by 1 cm degraded substrate layer, (iii) seeding on 1 cm topsoil layer and iv) seeding covered by 1 cm topsoil layer (Figueiredo et al., 2021). Another experiment established permanent plots to monitor floristic and life-form spectra in post-mined sites 4 years after topsoil transfer (Onésimo et al., 2021).

Major Project Outcomes

Under field conditions in the first rainy season after planting, Pellizzaro et al. (2017) found that 62 out of 85 species of trees, shrubs and grasses produced seedlings, of which 30 of them reached at least 20% survival rate. After the first year, 41 species had above 60% of survival, some with an astounding 80% survival rate. A separate study found that seed burial did not affect the emergence of species with round seeds, but negatively affected species with flat seeds (Silva & Vieira, 2017). Another found that harrowing and fire failed to eliminate the seed bank of invasive grasses which subsequently were able to re-establish, while short-lived shrubs and annual grasses lost dominance primarily to invasive species or perennial grasses. Most low-coverage plots shifted to invasive grass dominance after 2 years (Coutinho et al., 2019). Silva & Vieira (2017) showed that despite straw mulching reducing the emergence of native species with flat seed shape, it increased soil moisture and strongly reduced emergence of the invasive Urochloa decumbens, resulting in a higher growth rate of tree seedlings up to 1 year for five species. Encouragingly, results of these various seed-sowing experiments do indicate the feasibility of reintroducing a considerable number of native species from different functional groups in Cerrado restoration (Sampaio et al., 2019), but that controlling invasive grasses remains as a major challenge given, they have been shown to eventually regenerate.

Results from Haddad et al. (2020) showed that herbaceous plant communities of all three post-afforestation sites, regardless of management and fire history, were very different from the old-growth savannas that were destroyed to establish pine plantations five decades ago. Consistently, Haddad et al. (2021) showed that both passive restoration and native tree planting restored the structure, richness, and composition of the woody layer, reaching values like the reference ecosystem, but in all treatments, the herbaceous layer lacked the sub-communities of shrubs and herbs typical of undisturbed savannas even 15 years after passive restoration. Thus, these results are consistent with a growing body of evidence that shows that the species-diverse herbaceous communities of tropical savannas are unable to recover rapidly after afforestation and fire exclusion.

Even after 8 years post gravel extraction degraded campo rupestre sites were characterized by altered soil properties, and plant communities with impoverished seed banks. Species composition was still very different from that at reference sites (Le Stradic et al., 2018). Unfortunately, this result suggests that relying on natural regeneration is not a feasible strategy. Even more disappointing were results from the previous hay transfer experiments which showed that few seedlings emerged following the spreading of this material despite the large number of seeds contained in the hay (Le Stradic et al., 2014). This outcome indicates that hay transfer may not be as useful a method for restoring degraded areas of campo rupestre as envisaged. In this respect, a hypothesis that remains to be tested is whether mechanical seeding on lightly cultivated or slotted soils might create seed niches and constitute a viable restoration alternative.

Despite these findings, other positive outcomes have arisen in studies on ferrugineous campo rupestre. The mineralization of the biocrusts have been found to have led to the formation of biocemented aggregates that mechanically stabilized the crushed iron-rich waste material suggesting the potential of synthetic biocement as a long-term stabilization strategy for waste rock stockpiles, engineered slopes and mine remediation requiring the reformation of iron-rich duricrust (Paz et al., 2021). Another promising result is the finding that root exudates in this iron-rich substrate contributed indirectly to iron dissolution, particularly during phosphate solubilization, and the resulting surplus iron not taken up by the plants was redeposited, promoting the cementation of the residual minerals (Paz et al., 2020). Another study found that litter addition to the first 20 cm of the substrate plus seed sowing promoted the establishment of herbaceous and woody species (Figueiredo et al., 2021). In a further topsoil transfer experiment, 105 species were subsequently identified, and community composition and life-form spectra progressively resembled the reference areas (Onésimo et al., 2021). Unexpectedly, weed presence did not prevent the regeneration of native species. Altogether these results indicate that a combination of the appropriate reconstruction method for the physical environment and the correct site preparation appear to be promising restoration strategies in post-mined sites.

What About the Project Worked, What Did Not Work and Why?

These various direct seeding experiments in Central Brazil have suggested promising strategies for some types of grassland restoration. They indicate that when developed in an inclusive social-economic context, direct seeding has the potential to increase biodiversity, overcome the prohibitive costs of seedling planting and generate income for local communities. Nevertheless, seed sowing still appears to have a limited role in controlling invasive non-native grasses, and for the use of slow-growing species which remain less represented in seed-sowing programs because of their relative lower fecundity when compared to fast-growing species. Notwithstanding the innate problems which are apparent, the success of the Cerrado Seeds Network has been now established, and this is likely to increase following increased legal flexibility in terms of a relaxation of seed testing and commercialization rules (Schmidt et al., 2019b).

Grassland restoration after afforestation has been rarely studied but current evidence suggests that tree cutting and managing appropriate fire regimes have positive effects in re-establishing plant communities. However, the development of more effective strategies for the regeneration of the herbaceous communities is needed (Haddad et al., 2020). Already, prescribed fires have been shown to reduce the biomass of invasive species (Damasceno & Fidelis, 2020) and to help restore post-afforestation Cerrado sites (Zanzarini et al., 2019). Also, there is some suggestion that the combination of fire and ploughing may be an effective method to remove or at least restrict invasive grasses, but this treatment may need to be applied several times and/or for years. Fire should be used carefully as a restoration tool, and the fire regime needs to be managed considering natural fire frequency and the management context of specific vegetation types (Schmidt et al., 2019b; Haddad et al., 2020). Ploughing, which was shown to be effective in decompaction, uprooting of invasive grasses and destruction of invasive grass seedlings, may jeopardize regeneration of natives in bud banks by destroying or damaging underground storage organs.

Scepticism towards the feasibility of campo rupestre restoration has arisen owing to the repeated lack of spontaneous natural regeneration (Le Stradic et al., 2018), failures in hay transfer experiments (i.e., Le Stradic et al., 2014), the impoverished nature of native seed banks (Medina & Fernandes, 2007), overall low seed quantity/quality produced by native species (Dayrell et al., 2017) and low germination under field conditions (Figueiredo et al., 2021). These results unambiguously indicate that seed-based techniques would prove unviable as restoration strategies. Nevertheless, despite the low percentage of establishment, other studies suggest (i) a significant biotechnological potential for biocrust reconstruction, (ii) moderate to high seedling survival and growth across a different range of substrates (Machado et al., 2013, Figueiredo et al., 2018b), (iii) substantial success in sexual and vegetative propagation, (iv) moderate persistence and recruitment of planted individuals (Gomes et al., 2018), (v) a positive effect of the incorporation of plant litter and (vi) topsoil transfer in revegetation of post-mined sites with viable cost (Figueiredo et al., 2021). Taken collectively, the current prospect for campo rupestre restoration is more positive than previously thought a decade ago.

Grasslands and savannas that have been subjected to medium or high-intensity disturbance are typically composed of low-resilience grasses, short-lived herbs, and shrubs that have shallow roots and bud bank organs (Schmidt et al., 2019b). Nevertheless, regeneration after endogenous disturbance is chiefly driven by the resprouting of underground storage organs (Buisson et al., 2019). Therefore, transplanting underground storage organs may be a cost-effective strategy to enhance resilience in degraded grassland restoration, by increasing resprouting particularly for slow-growing species which can increase biodiversity. This hypothesis does, however, remain to be tested under field conditions.

The recently growing body of evidence on tropical grassland restoration indicates that multiple strategies may turn out to be feasible as large-scale alternatives to seedling planting (e.g., Buisson et al., 2019). Regarding this issue, a better understanding on the ecology of tropical grassy biomes, improved recognition of their value to mitigate climate change and better resourcing of restoration projects (Silveira et al., 2022), are expected to provide additional support for the improved restoration policy and higher standards of practice. These improvements are much needed so that tropical grassland restoration science can reach maturity. As such, in addition to improvements in knowledge and practice the importance of developing appropriate and multidimensional indicators of grassland restoration success is likely to emerge as an immediate challenge for the restoration sector in the coming years.

Chapter Synthesis

These four case studies focus on different grassy community types, and each shows that grassy community conservation and restoration is difficult, but feasible. Similar examples of success have been shown in other countries (Buisson et al., 2018; Puthod et al., 2020; Wagner et al., 2020, Freitag et al., 2021). While there were specific factors guiding the planning, approaches and goals set, each case study shows that successful outcomes are possible and the process of undertaking ecological restoration provided important learning experiences. Whilst the differences between regions and countries in these case studies are instructive, there are some key areas of similarity suggesting there may be fundamental factors or principles underlying the approaches taken in grassy community restoration which, when adjusted to suit specific local conditions or settings, will broadly lead to success (e.g., Goret et al., 2021). This is a very important message given the threat that these communities face. It should also give hope that in the future increasing knowledge, capacity and technologies and engagement by people and communities will allow us to halt and even reverse grassy community loss at local, landscape and perhaps even global scales.

For grassy community restoration to occur at the global scale required to repair anthropogenic damage over millennia it will take a commensurate effort in terms of time, resourcing and commitment from countries, jurisdictions, and their communities. Whilst some countries are clearly more advanced in this area than others, this can create the opportunity for knowledge sharing and cross-jurisdictional support. We assert that grassy communities can be, and should be, better integrated into the fabric of our landscapes, be they as farm-scapes, abandoned forestry, urban regions, transport corridors or protected remnant areas. For this to occur, purposeful decision-making and goal-setting, guided by clear pathways and concrete actions that meaningfully involve people, communities and practitioners, must be put in place so that ad-hoc actions and intermittent successes of the past are turned into purposeful strategies and widespread global advances of the future.

Ten key implications that have arisen from our reflections on the four Case Studies presented. These are:

Implications

• 1. Grassy communities can be restored using careful regenerative and reintroduction approaches into agricultural, post-forested, urban and other landscapes where they once existed or where they are now the desired community type.

• 2. Grassy communities can be conserved and enhanced through restorative management approaches.

• 3. Grassy community restoration can achieve high levels of species and functional diversity as well as temporal resilience.

• 4. Restored grassy communities create a myriad of biodiversity, ecological and ecosystem service benefits.

• 5. Restored communities must be purposefully managed and maintained over time to preserve their structural and compositional integrity.

• 6. Effective and inclusive (e.g., to local communities) seed supply chains, delivering seed in quantity, quality, price and ethically are critical to successful restoration.

• 7. Seed production approaches are likely to be critical to supplementing seed supply for restoration, especially for rare or uncommon species.

• 8. Developing overall sector capacity can improve training, technical skills and knowledge and lead to greater employment and career opportunities for people and communities involved in restoration.

• 9. Improved infrastructure and technology development will be crucial to increasing the effectiveness and scale of restoration undertaken.

• 10. Landscape-scale grassy community restoration relies on the formulation of insightful and finely crafted government strategies and policy to create the settings, frameworks and coordination required to build markets, improve sector capacity and meet ambitious grassy community restoration targets.