Invasive Plant Ecology and Management: Linking Processes to Practice CABI INVASIVE SPECIES SERIES Invasive species are plants, animals or microorganisms not native to an ecosystem, whose introduction has threatened biodiversity, food security, health or economic development. Many ecosystems are aff ected by invasive species and they pose one of the biggest threats to biodiversity worldwide. Globalization through increased trade, transport, travel and tour- ism will inevitably increase the intentional or accidental introduction of organisms to new environments, and it is widely predicted that climate change will further increase the threat posed by invasive species. To help control and mitigate the eff ects of invasive species, scien- tists need access to information that not only provides an overview of and background to the fi eld, but also keeps them up to date with the latest research fi ndings. Th is series addresses all topics relating to invasive species, including biosecurity surveil- lance, mapping and modeling, economics of invasive species and species interactions in plant invasions. Aimed at researchers, upper-level students and policy makers, titles in the series provide international coverage of topics related to invasive species, including both a synthesis of facts and discussions of future research perspectives and possible solutions.

Titles Available 1. Invasive Alien Plants: An Ecological Appraisal for the Indian Subcontinent Edited by J.R. Bhatt, J.S. Singh, R.S. Tripathi, S.P. Singh and R.K. Kohli 2. Invasive Plant Ecology and Management: Linking Processes to Practice Edited by T.A. Monaco and R.L. Sheley Invasive Plant Ecology and Management: Linking Processes to Practice

Edited by

THOMAS A. MONACO

US Department of Agriculture, Agricultural Research Service, Forage and Range Research Laboratory, Utah State University, Logan, Utah, USA

ROGER L. SHELEY

US Department of Agriculture, Agricultural Research Service, Eastern Oregon Agricultural Research Center, Burns, Oregon, USA CABI is a trading name of CAB International

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A catalogue record for this book is available from the British Library, London, UK.

Library of Congress Cataloging-in-Publication Data Invasive plant ecology and management : linking processes to practice / edited by Th omas A. Monaco, Roger L. Sheley. p. cm. -- (CABI invasive species series ; 2) Includes bibliographical references and index. ISBN 978-1-84593-811-6 (hbk. : alk. paper) 1. Invasive plants--Ecology. 2. Invasive plants--Control. I. Monaco, Th omas A. II. Sheley, Roger L.

SB613.5.I552 2012 581.62--dc23

2011037079

ISBN-13: 978 1 84593 811 6

Commissioning Editor: David Hemming Editorial Assistant: Gwenan Spearing Production Editor: Simon Hill

Typeset by Columns Design XML Limited, Reading, UK. Printed and bound in the UK by MPG Books Group. Contents

Contributors vii Foreword ix Preface xi Acknowledgements xii

Part I Assessing Ecosystem Processes and Invasive Plant Impacts 1 1 Managing Invasive Species in Heterogeneous Ecosystems 3 Joel R. Brown and Brandon T. Bestelmeyer 2 Linking Disturbance Regimes, Vegetation Dynamics, and Plant Strategies Across Complex Landscapes to Mitigate and Manage Plant Invasions 19 Samuel D. Fuhlendorf, Brady W. Allred, R. Dwayne Elmore, and David M. Engle 3 Land-use Legacy Eff ects of Cultivation on Ecological Processes 36 Lesley R. Morris 4 Resource Pool Dynamics: Conditions Th at Regulate Species Interactions and Dominance 57 A. Joshua Leffl er and Ronald J. Ryel 5 Invasive Plant Impacts on Soil Properties, Nutrient Cycling, and Microbial Communities 79 Th omas A. Grant III and Mark W. Paschke

Part II Principles and Practices to Influence Ecosystem Change 105 6 Weather Variability, Ecological Processes, and Optimization of Soil Micro-environment for Rangeland Restoration 107 Stuart P. Hardegree, Jaepil Cho, and Jeanne M. Schneider 7 Th e Eff ects of Plant-Soil Feedbacks on Invasive Plants: Mechanisms and Potential Management Options 122 Valerie T. Eviner and Christine V. Hawkes

v vi Contents

8 Species Performance: the Relationship Between Nutrient Availability, Life History Traits, and Stress 142 Jeremy J. James 9 Reducing Invasive Plant Performance: a Precursor to Restoration 154 Joseph M. DiTomaso and Jacob N. Barney 10 Revegetation: Using Current Technologies and Ecological Knowledge to Manage Site Availability, Species Availability, and Species Performance 176 Jane M. Mangold Index 197 Contributors

Brady W. Allred, Natural Resource Ecology and Management, Oklahoma State University, 008C Agricultural Hall, Stillwater, Oklahoma 74077, USA; Email: brady.allred@okstate. edu Jacob N. Barney, Department of Plant Pathology, Physiology and Weed Science, Virginia Tech, 435 Old Glade Road (0330), Glade Road Research Center, Blacksburg, Virginia 24061, USA; Email: [email protected] Brandon T. Bestelmeyer, US Department of Agriculture, Agricultural Research Service, Jornada Experimental Range, New Mexico State University, Box 30003 MSC 3JER, Las Cruces, New Mexico 88003, USA; Email: [email protected] Joel R. Brown, US Department of Agriculture, Natural Resources Conservation Service and Jornada, Experimental Range, New Mexico State University, Box 30003 MSC 3JER, Las Cruces, New Mexico 88003, USA; Email: [email protected] Jaepil Cho, US Department of Agriculture, Agricultural Research Service, Northwest Watershed Research Center, 800 Park Blvd, Suite 105, Boise, Idaho 83712, USA; Email: [email protected] Joseph M. DiTomaso, Department of Plant Sciences, Mail Stop 4, University of California, Davis, California 95616, USA; Email: [email protected] R. Dwayne Elmore, Natural Resource Ecology and Management, Oklahoma State University, 008C Agricultural Hall, Stillwater, Oklahoma 74077, USA; Email: dwayne. [email protected] David M. Engle, Natural Resource Ecology and Management, Oklahoma State University, 008C Agricultural Hall, Stillwater, Oklahoma 74077, USA; Email: david.engle@okstate. edu Valerie T. Eviner, Department of Plant Sciences, University of California 1210 PES, Mail Stop 1, One Shields Ave, Davis, California 95616, USA; Email: [email protected] Samuel D. Fuhlendorf, Natural Resource Ecology and Management, Oklahoma State University, 008C Agricultural Hall, Stillwater, Oklahoma 74077, USA; Email: sam. [email protected] Th omas A. Grant III, Department of Forest, Rangeland, and Watershed Stewardship, Colorado State University, 230 Forestry Building, Campus Delivery 1472, Fort Collins, Colorado 80523-1472, USA; Email: [email protected] Stuart P. Hardegree, US Department of Agriculture, Agricultural Research Service, Northwest Watershed Research Center, 800 Park Blvd, Suite 105, Boise, Idaho 83712, USA; Email: [email protected]

vii viii Contributors

Christine V. Hawkes, Section of Integrative Biology, Th e University of Texas, 1 University Station C0930, Austin, Texas 78712, USA; Email: [email protected] Jeremy J. James, US Department of Agriculture, Agricultural Research Service, Eastern Oregon Agricultural Research Center, 67826-A Hwy 205, Burns, Oregon 97720, USA; Email: [email protected] A. Joshua Leffl er, US Department of Agriculture, Agricultural Research Service, Utah State University Forage and Range Research Laboratory, 696 N 1100 E, Logan, Utah 84322- 6300, USA; Email: josh.leffl [email protected] Jane M. Mangold, Department of Land Resources and Environmental Sciences, Montana State University PO Box 173120, Bozeman, Montana 59717-3120, USA; Email: jane. [email protected] Lesley R. Morris, US Department of Agriculture, Agricultural Research Service, Utah State University, Forage and Range Research Laboratory, 696 N 1100 E, Logan, Utah 84322- 6300, USA; Email: [email protected] Mark W. Paschke, Department of Forest, Rangeland, and Watershed Stewardship, Colorado State University, 230 Forestry Building, Campus Delivery 1472, Fort Collins, Colorado 80523-1472, USA; Email: [email protected] Ronald J. Ryel, Department of Wildland Resources and the Ecology Center, Utah State University, Logan, Utah 84322-5230, USA; Email: [email protected] Jeanne M. Schneider, US Department of Agriculture, Agricultural Research Service, Great Plains Agroclimate and Natural Resources Research Unit, 7207 W Cheyenne Street, El Reno, Oklahoma 73036, USA; Email: [email protected] Steven G. Whisenant, Department of Ecosystem Science and Management, Texas A&M University 2138 TAMU, College Station, Texas 77843, USA; Email: s-whisenant@tamu. edu

viii Foreword

Invasive plant species can substantially alter ecosystems. Direct economic losses to agri- culture, livestock production, forestry, and recreation are well known consequences of cer- tain invasive species. Less understood, yet probably more important, are changes to community and ecosystem processes that are caused by invasive plant species. Th ese altered ecosystems, sometimes created by large-scale plant invasions, are not simply structural reorganizations with new species. Many are novel ecosystems that look, function, and react quite diff erently than their predecessors. Invasive plant species that diminish biological diversity, alter nutrient and hydrologic processes, or dramatically change fi re regimes have serious consequences. Ecosystem management strategies based on this knowledge should have signifi cant advantages over strategies simply focused on removing invasive species. Th e ultimate test of ecological restoration is our ability to understand ecosystems and apply relevant science to solve real-world environmental problems such as the detrimental impacts of invasive plant species. My experiences in ecological restoration lead me to believe that invasive species are among our most diffi cult and perplexing challenges. We may be very good at killing and removing individual plants. Yet ultimately we lose the battle by fail- ing to understand the underlying processes or even the scales at which they operate. Th is occurs whether we apply traditional chemical or mechanical plant removal tools to relatively small areas or use less direct strategies over large spatial scales. Too often we treat the symp- toms of invasive plant problems rather than identifying and managing the underlying cause of those problems. Not addressing and removing the underlying causes of invasions usually reduces treatment longevity. Clearly, recognizing and removing these underlying causes is more easily stated than done. Th is diffi culty is why I view this book as a seminal contribution to the science and practice of invasive plant management. In Invasive Plant Ecology and Management: Linking Processes to Practice, editors Th omas Monaco and Roger Sheley have assembled an impressive group of authors for the purpose of applying contemporary ecological knowledge to the practice of invasive plant management. Collectively, at all major levels of ecological organization, they address the attributes that make plants invasive as well as those characteristics that make ecosystems receptive to inva- sions. Signifi cantly, the authors then describe how this knowledge can be used to both enhance the competitive ability of native and other desirable species while limiting the suc- cess of invasive species. Th ese are substantial contributions to the science and practice of invasive plant management.

Steven G. Whisenant College Station, Texas July 2011

ix This page intentionally left blank Preface

Th e primary objective of this book is to illustrate how understanding ecological processes will foster scientifi cally based approaches to invasive plant management in semi-arid ecosys- tems. Ecological processes serve as the underpinning and common ground within the scien- tifi c literature that bridges the gap between researchers and land managers. Our focus on ecological processes is also justifi ed based on the overwhelming realization that invasive plant management must move beyond treating symptoms of damaged lands to repairing and infl uencing the processes responsible for plant community change. While our emphasis is clearly slanted towards semi-arid wildlands, we believe the ecological principles outlined by the contributing authors can easily be applied to many other systems impacted by inva- sive plant species. Assessing ecological processes and how they are impacted by invasive plant species is a critical aspect of land management, which can easily be overlooked when the impetus to “do something” overshadows sound decision-making. Part 1 of this book, comprising Chapters 1 through 5, provides a compelling justifi cation to assess ecosystem and landscape heteroge- neity and historical land-use legacy eff ects before a process-based understanding of how ecosystems operate can be realized. In this same vein, Part 1 also showcases how the emerg- ing concept of resource pool dynamics provides a much more adequate mechanism to assess ecological processes associated with plant resource use than traditional emphasis on compe- tition. Concluding Part 1 with a comprehensive assessment of how invasive species impact soils, nutrient cycling, and microbial communities emphasizes that eff ective invasive plant management will also require designing management practices that infl uence processes that operate within soils. Th e idea of embracing successional-ecological processes to initiate and direct plant com- munity change is not new, but the principles and practices to achieve greater success when managing invasive plant species continue to emerge. If the underlying origin for the persist- ence of invasive species lies with damaged ecological processes, intervening with restorative actions to repair these ecological processes is our responsibility and should be emphasized in management. Part 2 of this book illustrates a mechanistic approach at presenting princi- ples and practices to optimize soil microclimate conditions, eradicate adverse plant–soil feedbacks, and systematically alter plant species performance. Ultimately, it is our hope that adopting these ideas will assist you in achieving desired outcomes and greater predictability when applying land management practices.

Th omas A. Monaco and Roger L. Sheley

xi Acknowledgements

Th e conception and preparation of this book was only possible through fi nancial support from the US Department of Agriculture, Agricultural Research Service and funding they provided to the Areawide Pest Management Program for Annual Grasses in the Great Basin Ecosystem. Th e unique collaboration of this program inspired much of the conceptual and scientifi c organization of the chapters. We also recognize the organizations that supported the authors of this book. Finally, we are grateful to all the authors and publishers who have granted permission for us to reproduce material in this book.

Th omas A. Monaco and Roger L. Sheley

xii Part I Assessing Ecosystem Processes and Invasive Plant Impacts This page intentionally left blank Managing Invasive Species in 1 Heterogeneous Ecosystems Joel R. Brown and Brandon T. Bestelmeyer

US Department of Agriculture, New Mexico State University, USA

Introduction environments diff ers from more intensively managed systems, such as croplands, where Ecologically based invasive plant manage- the spatiotemporal relationships are much ment (Sheley et al., 2010) provides a more clearly defi ned. Th en we present and mechanistic framework for diagnosing demonstrate the use of ecological sites causes of plant invasions and selecting (soil-based landscape subunits) and state- manage ment responses. Th is approach and-transition models (graphic and text requires the organization of multiple sources descriptions of soil:vegetation dynamics) as of information, much of which is highly a means of organizing and synthesizing dependent upon spatial and temporal information about plant community and context. Although there have been sub- landscape susceptibility to invasion. Th ese stantial eff orts to identify the characteristics tools, in combination with a basic under- of successful invaders as a means to predict standing of invasive species biology, can be which plants are likely to be successful when used to predict the pattern of invasion, introduced into new environments, the use develop management responses with an of species attributes alone is a poor predictor increased probability of success, and devise of which plants will invade a particular monitoring systems to assess the eff ective- landscape (Mack et al., 2000). In fact, many ness of management actions. Finally, we of the plants currently defi ned as ‘invaders’ provide examples of the application of in ecological terms and ‘noxious weeds’ in ecological sites and state and transition legal terms are native to the regions and models to the diagnosis of the ecological landscapes, if not the plant communities, process drivers of invasions and the they invade. Th is combination of species development of management responses. attributes (invasiveness) and plant commun- We believe that an Ecological Site-based ity or landscape susceptibility (invasibility) approach off ers an improvement in the complicates the development of a universal management of invasive species on wild- set of principles for prediction, and even lands by increasing the spatial and temporal post hoc analysis, of the interactions of precision of monitoring, analysis, and invasive plants and landscapes. Because responses, thereby increasing the chances of the information necessary for successful success of ecologically based invasive plant implementation of management responses management. is so highly variable in both time and space, as well as by invasive species, a systematic approach to organization, analysis, and Managing Invasive Plants in Complex decision-making is essential. Landscapes In this chapter, we fi rst discuss the challenge of managing invasive species in Th e spatial and temporal patterns of invasive com plex landscapes. In particular, how species movement are critical to under- managing invasive species in wildland standing the processes of invasion and to

© CAB International 2012. Invasive Plant Ecology and Management: Linking Processes to Practice (eds T.A. Monaco and R.L. Sheley) 3 4 J.R. Brown and B.T. Bestelmeyer

developing eff ective prediction models and plexities of decision-making and priority decision-support systems for management setting for management actions are responses (Masters and Sheley, 2001). considered, it is easy to see how invasive Such patterns, if properly described and species populations can be well established quantifi ed, can provide insight into the long before there is a clear motivation to causes and drivers of invasions and are the respond (Brown et al., 1999). foundation for predictive models essential While exotic species have been the focus in developing eff ective responses, both for of much research and analysis, many of the reducing the rates and impacts of invasion ‘invasive species’ with measured negative and for restoring invaded landscapes. economic impacts are actually native with Accurate predictive models for the long histories as a part of the landscape, but movement of invasive species are especially changes in land use, land management or important in wildlands (rangeland, forests, climate have resulted in increased popu- and deserts). Although many important lations. Radosevich et al. (2007) enumerated agronomic and economic principles have the broad range of behaviors, characteristics, emerged from research in crop and timber and impacts ascribed to ‘weeds’ or invasive production systems (weed science), they plants, which in many cases, could just as may be of limited value in the management easily apply to native species. Th us, in the of invasive species in wildlands. Agro- context of natural resource management economic models used for decision making on wildlands, the defi nition of invasive in production systems, where the focus is species must go beyond ‘exotic’ to include largely on recognition of economic impact ‘undesirable’ species. Regardless of the thresholds, are generally based on the geographic source of the invasive plant assumption that control technologies are species, the management of undesirable eff ective against target species (Cousens and species requires: (i) detection methodologies Mortimer, 1995). While the relationships and response protocols; (ii) selection of the between weed density and crop yield are appropriate control technology to limit their relatively straightforward and can be used as impact; and (iii) knowledge of the treatment a solid basis for decisions about the cost:benefi t ratio (Radosevich et al., 2007). deployment of weed control technologies, Quantitative information about these these decisions are typically based on an relationships allows land managers to make annual timeframe and considerations are informed decisions and evaluate progress usually limited to an individual fi eld or toward objectives. However, because wild- management unit. land natural resource management typically Invasive species management in exten- encompasses large (>100 km2) spatial scales sively managed lands is much more complex, and long (decades) timeframes, changes in requiring considerations that extend to the composition and arrangement of plant greater spatial and temporal scales (Brown communities on a landscape can greatly et al., 1999). Th ere is often a substantial infl uence vulnerability to invasive species amount of time, in a managerial context, and can also alter the success of management between the presence of invasive species responses. Management tools must be able propagules or juveniles and the ability of to systematically organize these spatio- land managers to detect nascent populations, temporal changes if they are to be of value. and it is relatively common for years to Invasive species detection in wildland elapse between the appearance of detectable environments commonly involves detecting populations and the impacts on economic very low densities of plant populations that yields. Th is time lag between the initial, may reside as seeds or juveniles within largely irreversible stages of invasion (in the communities for years before reaching a form of persistent banks of seeds or detection threshold. While remote sensing juveniles) and an economic threshold when applications have been developed, they yield is reduced may further complicate currently lack the precision needed to detect management responses. When the com- populations of target plants early in the Managing Invasive Species in Heterogeneous Ecosystems 5

invasion process (Mitchell and Glenn, 2009). Diff erential vulnerabilities to invasive Even well-trained managers struggle to species can result from the complex detect invasive species by direct observation interactions of climate, geomorphology, and initiate a timely response. For the landscape position, soils, and vegetation. immediate future, wildland invasive species Th ese factors govern the movement of management will not benefi t from remote invasive species propagules across the sensing applications intended to detect low landscape via a variety of vectors; directly by level populations (Blumenthal et al., 2007). wind and water or indirectly by endo- and However, remote sensing will remain a core ecto-zoochory. Th e availability of resources technology for inventory and for the at diff erent points in the landscape and in development of restoration strategies and diff erent time periods can aff ect the tactics when populations are more easily recruitment and survival of invasive species observed (Washington-Allen et al., 2006). (Davis et al., 2000). Th ese shifting resources Without the benefi t of a reliable, cost- can vary on time scales of years, months, and eff ective early detection system for invasive even days. In addition, the changes in species, wildland managers must rely more vegetation occupying the site can change the heavily on predictive models to identify vulnerabilities of landscapes to invasive which invasive species are threats and which species over time. Varying soil:vegetation plant communities, landscapes, and regions interactions result in functionally diff erent are most likely to be invaded. Such a ecological confi gurations that can change the prediction approach would allow land availability of resources and the vulnerability managers and technical advisors to target of sites to invasion. Th e varying availability locations for detection and treatment. Th is of soil resources (water and nutrients) approach must address both the spatial and governs the rate and success of ecological temporal dynamics of invasions and requires processes such as seed germination, seedling a well-developed understanding of: (i) survival, juvenile growth rates, interspecies dispersal vectors; (ii) species life histories competition, and the eff ects of disturbance and performance attributes; and (iii) a range at diff erent stages of the life cycle. of vulnerabilities in individual landscape Changes in the patterns of vulnerability components (e.g. soil units) that may change at the community scale and the arrangement the pattern of invasion across landscapes of those communities can also aff ect the (Sheley et al., 2010). Managing low-level way a species spreads across the landscape populations of invasive species in hetero- and how the impacts are manifested. Within geneous landscapes requires either precise this complex opportunity matrix, the application of intensive technologies or vulnerability to invasion, the impacts of the widespread application of extensive, low- invasion, and the range of successful cost technology. Selecting and implementing responses and restoration tactics will the tactics that follow from either of these change with time. In a management context, approaches not only requires a thorough property owners must select and prioritize knowledge of both invasive species life specifi c landscape components to target with history and plant community characteristics, management actions. As the vulnerability of but also an understanding of how the landscape components changes over time targeted invasive species and the plant (either seasonally or across seasons), the community will respond to the treatment vulnerability to invasion changes and the and the removal of the species. Managing deployment of techniques to detect invasions invasive species in wildlands clearly requires would also be aff ected. Likewise, the selection knowledge about the landscape matrix, and and use of management treatments to individual plant communities that comprise address the invasion would change. it. Landscapes have diff ering vulnerabilities Developing responses to invasive species and diff ering responses to invasive species also more often than not involves multiple that, in turn, have diff erent vulnerabilities landowners. When landscape-scale re spon- and impacts on landscapes. ses include landowners with diff erent 6 J.R. Brown and B.T. Bestelmeyer

objectives, resources, and skills, the role of reach their goals within those highly variable policy becomes increasingly important. In conditions (spatial and temporal hetero- this case, policy can be considered as geneity) by directing ecological processes encompassing the full range of government with extensive management (Hammitt and responses: regulations, rules, programs, and Cole, 1998). Understanding this hetero- fi nancial and technical assistance. Depend- geneity in a management framework ing upon the spatiotemporal patterns of requires subdividing the landscape into invasion and restoration needs, success in components that behave similarly in responding to invasive species may require response to management. Th ese groupings multi-scale information to implement of similarly responding landscape units are successful responses. Th e ability of policy- especially important in the detection, developers to incorporate relevant spatio- prediction, and treatment of invasive temporal information is thus critical to species. Th ese landscape subdivisions are responding successfully to the challenges of referred to as ‘sites.’ Th e concept of site and managing invasive species. its application has been one of the central Understanding the eff ects of spatio- tenets of modern natural resource manage- temporal heterogeneity on invasion pro- ment on rangelands and forestlands, cesses is a necessary fi rst step in developing incorporating the evolving ideas of a guide for estimating impacts and selecting ecosystem behavior and management and implementing responses to invasive responses as they developed (Brown, 2010). species. While individual species perform- Grouping portions of a landscape based on ance under a range of environmental their climatic, geomorphic, and edaphic conditions is relatively easy to defi ne and similarities to predict the behavior of soil, investigate, the range of possibilities vegetation, and related resources has been associated with the spatiotemporal distri- the basis for organizing information and bution of favorable establishment con- decision-making. Th e site concept has been ditions (site availability), propagule presence valuable to managers, advisors, and policy (dispersal) and species performance is makers in assessing current conditions, staggeringly large. An ability to organize setting objectives for change (or avoiding knowledge of this ‘spatial heterogeneity’ as a change), selecting and implementing source of conditions necessary for invasive management practices to achieve objectives, plant establishment, growth, and repro- allocating scarce resources, and evaluating duction is a key component in design ing the eff ects of practices, programs, and assessment, prediction, and treatment policies (see Briske, 2011). Th e site, as a approaches. means of organizing information, was also an eff ective means of communicating professionally with land managers and with Ecological Sites and State and the public. Transition Models As the science of land management has advanced, the concepts used to defi ne and Wildlands and their management are describe the dynamics of sites have become generally defi ned more by their limitations more sophisticated and complex (e.g. than their uses. Th ese limitations (aridity, Rumpff et al., 2011). Th e availability of shallow soils, steep slopes) generally more precise observational technologies preclude the use of intensive management (e.g. remotely sensed landscape imagery, technologies because of the high cost of monitoring of ecosystem gas exchange) and implementation and the relatively low nearly two centuries of direct observation probability of achieving a favorable economic has led to several refi nements and improve- return. Th us, the successful management of ments. Th e initial concept of site, developed wildlands, regardless of the objectives, in the early 20th century, was focused on the depends on land managers being able to ‘climax’ vegetation, defi ned as the endpoint Managing Invasive Species in Heterogeneous Ecosystems 7

of ecological succession processes, and landscape position within climate-based assumed to be the most stable and regions called land resource units (LRUs) or productive. In this approach, invasive major land resource areas (MLRAs). MLRAs species, and undesirable plants in general, are similar in extent to ecoregion sections were classifi ed as ‘invaders’ (Dykesterhuis, used by the US Department of Agriculture, 1949) that resulted from inappropriate Forest Service (Cleland et al., 1997). For management, but were not regarded as each ecological site, the physical setting, particularly infl uential in altering ecosystem- ecological dynamics, and management driving processes. Th e resulting weed inter pre tations are contained in an manage ment practices and policies had two ecological site description (ESD; Bestel- serious shortcomings: (i) the plant meyer et al., 2010). Th is document serves as community representing the ‘climax’ or the repository for information supporting successional endpoint was supposed to be the conceptual delineation of the landscape the most competitive, and competition was subunit. the dominant ecological process controlling Ecological sites are typically not mapped invasive plant populations (climax com- directly, but rather they are spatially defi ned munities were less invasible); and (ii) once by their correlation to mapped soils as a part an invasive species was established, manage- of the National Cooperative Soil Survey ment responses emphasized the elimination (http://soils.usda.gov/partnerships/ncss/). of the ‘problem’ species rather than an Because of the inherent heterogeneity of evaluation of the altered disturbance regime wildlands and the more extensive nature of (fi re, grazing, drought, fl ood). Th is soil mapping, soil map units often contain agronomic approach was relatively eff ective much more variability than soils mapped in in communities and ecosystems where the croplands. Th ese ‘complexes’ often contain invasive species had not aff ected ecological suffi cient variability that an assumption of processes, but failed to have an eff ect when homogeneity of ecological processes, plant the processes had been altered (i.e. shrub community dynamics, and response to invasion controlled water and nutrient management can lead to fl awed decision- cycles), and a new disturbance regime was in making. Th us, within the boundaries of an place (Brown et al., 1999). An enhanced ecological site, displayed as part of a understanding of the role of disturbance and hardcopy or electronic soil property-based climate variability in structuring plant map, there may be substantial heterogeneity. communities has led to the acceptance of For the purposes of planning invasive species the existence of multiple stable ‘states’ assessment and management, users can (soil:vegetation combinations) for each site. assume that a particular soil map unit Even subtle shifts in the seasonal dis- will respond as defi ned by the spatially- tribution of precipitation (Archer and dominant, ecological site it contains. Predick, 2008) and atmospheric chemistry However, management-practice implement- (Morgan et al., 2008) can result in changes ation and assessment should not be based on in dominant species and create opportunities the assumption of within-map unit homo- for invasive species. geneity. Assessment and management- Th is change in perspective, which practice implementation necessarily require recognizes the overriding infl uence of con- closer inspection to ensure success. Field trolling variables (climate regime, landscape inspections of soil properties (depth, surface position, soil properties) on the dynamic texture) are necessary to insure that behavior of plant communities, has resulted assumptions about applying the appropriate in an increased emphasis on static site description are valid. properties as a means to classify and group In addition to the defi ning spatial factors, subunits of the landscape (Duniway et al., every ecological site has unique temporal 2010). Th ese landscape subunits are termed soil:vegetation dynamics that result in ‘ecological sites’ (USDA NRCS, 2007) and diff erent plant communities, diff erent are unique groupings of soil properties and relationships among the plant communities, 8 J.R. Brown and B.T. Bestelmeyer

and diff erent responses to management predictors of sites prone to invasive species over time (Bestelmeyer et al., 2010). An establishment. important part of the ESD is the graphical Each of the alternative states has a and textual description of soil:vegetation characteristic resistance to change and dynamics called a ‘state and transition should persist under the defi ned disturbance model.’ Th e primary purpose of a state and regime. For each state, there are a set of transition model (STM) is to display and characteristic diagnostic attributes and describe the possible soil:vegetation con- indicators. Th ese are quantifi able parameters fi gurations (states) on a particular site, and that can be used to distinguish states and the relationships (transitions) among them. should refl ect diff erences in ecological STMs describe the range of dynamics of processes. Plant attributes (species com- plant communities and ecological processes position, cover, basal area, canopy height/ possible on the site (Briske et al., 2005). cover) and soil surface attributes (erosion, Figure 1.1 presents an example STM to litter distribution, and surface stability) are illustrate the components and relationships. typically used (Herrick et al., 2006). Th e large boxes represent ecological states. Resistance to invasion and resilience after States are defi ned as diff erent plant disturbance are key state attributes, which communities that may exist on the same may be helpful in conveying an overall sense site. Each state has diff erent ecological of community dynamics. However, it could process rates and functional relationships be misleading to assume that resistance or and possesses suffi cient resistance such resilience somehow imparts low invasibility. that signifi cant management eff ort or Th e likelihood of a particular species disturbance is required to change (Stringham invading a particular state or community et al., 2003). Within each state, embedded can only be evaluated within the context of boxes represent plant community phases both species attributes and community that are similar in their ecological functional characteristics (Heger and Trepl, 2003). Th e attributes and among which change is community phases within a state represent relatively simple and can occur in response a characteristic range of variability in the to minor management changes or climatic processes and structure for that state. In fl uctuations (Beisner et al., 2003). An particular for invasive species management, individual state may contain multiple each state has an inherent resistance to community phases (a range of variability), invasion based on the manner in which or they may have only one phase. plant species use the available resources and A recent innovation in the information in how disturbance is likely to alter those STMs is the inclusion of dynamic soil utilization patterns and create new properties (DSPs). DSPs characterize near opportunities. If a particular disturbance soil surface attributes that are tightly (e.g. fi re, grazing, drought, or insect attack) coupled through feedbacks with plant creates new opportunities within a plant community structure and function. In most community, which allow an invasive species wildland plant communities, surface soils to establish, persist, and, more importantly, and vegetation are so closely linked that alter ecological functions, then a new, novel they must be considered a single unit when state is reached. If the disturbance does not described (Duniway et al., 2010). For result in invasive plant establishment and example, the amount and distribution of altered ecological function, the changes can soil resources described by DSPs are be described merely as another community particularly important in determining phase within the original state. Th e conditions and site availability for potential distinction between a new state and a new invasive species establishment. In particular, phase within an existing state is critical soil moisture and nutrient availability are because of the implications for future important controls on seedling recruitment. management and restoration. In more mesic environments, bare soil patch Transitions are the pathways of change size and distribution are often important among states. Unlike community pathways, Managing Invasive Species in Heterogeneous Ecosystems 9

(a)

2.1 BOER (15-45%) 1.1 BOER (15-60%) T1a PRGL (1-15%) 1.1 1.2 2.1 2.2 1.2 BOER (3-15%) 2.2 PRGL (1-15%) (large bare patches) R1 BOER (3-15%) (large bare patches)

State 1. Black grama grassland T1b T2a State 2. Mesquite invaded T2b 3.1 Other PG (5-35%) BOER (< 3%) 4.1 PRGL (15-25%) PRGL (1-15%) (erosion) 3.1 3.2 T3 Other PG (< 5%) 3.2 PRGL (1-15%) 4.1 4.2 Other PG (<5%) BOER (< 3%) 4.2 PRGL (15-25%) (erosion) Other PG (5-35%)

State 3. Bunchgrass/mesquite State 4. Shrubland

T1a. Mesquite establishment facilitated by seed transport by cattle, bare patches > 50 cm, and relatively wet springs R1. Shrub removal via herbicide or fire followed by black grama recovery to > 15% T1b, T2a. Black grama is reduced below ca. 3% cover by heavy grazing in drought T2b, T3. At perennial grass cover < 5%, wind and storm events, trigger deep, spreading soil erosion

(b)

State 1. Black grama grassland State 2. Mesquite invaded

State 3. Bunchgrass/mesquite State 4. Shrubland

Fig. 1.1. (a) State-and-transition model for the Sandy ecological site (reference # R042XB012NM) of the southern desertic basins, plains, and mountains major land resource area (MLRA 42) in southern New Mexico, USA. BOER=Bouteloua eriopoda (black grama); PRGL=Prosopis glandulosa (honey mesquite); PG=perennial grasses. Arrows labeled T and R are transitions. Dotted lines are community pathways. Redrawn from Bestelmeyer et al. (2011). (b) Plant community scale photos of different states in Chihuahuan desert grassland. 10 J.R. Brown and B.T. Bestelmeyer

transitions represent changes in ecological relatively fertile soils (usually mollisols) in processes, not merely changes in structure comparatively mesic (>25 cm annual that require substantial management input precipitation) regions. Although the result- or changes in disturbance regime to result in ing perennial vegetation communities have a new state. Transitions generally represent value for a variety of ecosystem services changes in plant life-form (e.g. grass to (erosion control, carbon sequestration, shrub dominance) and sometimes near- wildlife habitat), it would indeed be a surface soil properties. Transitions can be mistake to assume that the primary limiting caused both by gradual change associated factors in the restoration of all plant with slow variables and rapid change communities is the availability of seed of associated with distinct triggers (Bestelmeyer adapted plant species and appropriate et al., 2010). Transitions caused by gradual cultural practices to suppress competition. A change may stem from increasing invasive far greater challenge is the restoration of species recruitment rates or loss of partially intact wildland communities that herbaceous species due to chronic over- have experienced signifi cant changes in life- grazing. In contrast, transitions caused by form dominance and the distribution of triggers are relatively rapid, discrete events, soil properties. Bestelmeyer et al. (2003) such as an exceedingly hot fi re followed by a demonstrated the challenges of restoration drought period, or a fl ooding event that in a desert grassland ecosystem (Fig. 1.1). causes massive soil erosion. Th e loss of seed-producing, native tussock A key point in the process of transition grasses from the system and the diffi culty between two states is the threshold. of reestablishing grass populations via Th resholds are specifi c points in time in seeding, the loss and redistribution of fi ne which negative feedback mechanisms switch textured soil resources via wind erosion, and from reinforcing the inertia inherent within the increased competition from invasive a state (resistance) to positive feedbacks shrubs present a substantial barrier to the accelerating that change (Brown et al., 1999). restoration of important native species and When describing thresholds in STMs, it is ecological processes. Th ere are many other important to identify conditions that ecosystems with similar challenges caused immediately precede a change. However, by redistribution of soil nutrients, changes precise descriptions of thresholds are in climatic patterns, and alteration of unlikely to be accurate given limited pre- groundwater availability in addition to the dictive capability for most ecosystems, competitive eff ects of invasive species which could impart a false sense of (Society for Range Management, 2010). confi dence to managers (Brown, 2010). Because of the information they contain Perhaps the most important application of and the relationships they describe, STMs transition and threshold concepts for can be used to assist the development of invasive species management is the appli- management strategies, choose appropriate cation to ecological restoration (Bestelmeyer, assessment and monitoring approaches, 2006). Restoration technologies have and focus research and modeling eff orts on typically been developed based on the key ecosystem properties and processes assumption that important ecosystem pro- (Herrick et al., 2006). Th ese applications cesses remained intact, or in the terminology are especially appropriate when applied of a state and transition model, there has to invasive species management. STMs not been a state change (Young et al., 2001). organize, and put into context, information For example, the Conservation Reserve that allows managers to: (i) assess Program that has been implemented on vulnerability to invasive species with dif- more than 15 million ha in the USA fering life history characteristics and alter establishes perennial vegetation (mostly management to increase the resistance to grasses) on marginal cropland. Th e program invasion; (ii) devise management practices has been successful in reestablishing exotic and systems to limit the impacts of invasive and native perennial grasses and trees on species; and (iii) develop strategies to restore Managing Invasive Species in Heterogeneous Ecosystems 11

invaded plant communities, landscapes, community scale interactions that are and regions. Because they are conceptual, necessary to develop credible predictions, draft versions can be developed quickly and hypotheses, and experiments. As STMs updated as new information is available. In become more frequently used, and imple- addition to the management applications, mentation extends to more ecosystems, the STMs can be used to improve the link quality of the logic, and data that goes into between research and management by them, will become more critical. focusing hypothesis development, experi- ESDs and STMs are valuable tools in mental design and data collection to refi ne applying ecologically based invasive species the models. Th is link to fi eld research in a management. Ecological processes that hypothesis-testing framework allows them aff ect invasive species dynamics and infl u- to guide the development of quantitative ence the success of management responses models, which can then be used to extend are highly variable in both space and time. the understanding gained through experi- Th e information in ESDs and STMs can aid mentation to other similar sites. Th us, managers in subdividing the landscape community dynamics in response to changes according to soil:vegetation behaviors, and in resource availability (e.g. precipitation), develop and test hypotheses about the disturbance, and post-disturbance recovery dynamics of invasive plant populations. can be described (Westoby et al., 1989; Th ey also can integrate information gained Briske et al., 2003). STMs constructed using from experimentation, monitoring, and these principles are especially valuable in treatment case studies. Together, these estimating the eff ects of changes in life- technologies can be com bined into a history traits associated with invasive powerful system to generate, interpret, and species. distribute relevant ecological information Th e application of STMs to invasive for managing invasive species and natural species management is limited primarily resources. by the fact that our understanding of the dynamic interactions among soil, vegetation, climate, and animals is Applying Ecological Site Information incomplete. Unfortunately, existing datasets to Invasive Species Management are frequently inadequate to provide the information necessary to defi ne states and Ecologically based invasive plant manage- community phases, transitions, and ment relies on the application of ‘adaptive thresholds. In many cases information is management’ to guide the process of adequate to construct initial STMs, but identifying problems, selecting appropriate details and subtleties necessary for complex technologies, implementing responses, and management decisions are lacking. Perhaps assessing impacts (Fig. 1.2). At the core of the greatest limitation of STMs is that the adaptive management is the relevant and dynamics represented within them are not timely use of scientifi c information. In applicable beyond the boundaries of the particular, the approach requires a format mapped soil properties that serve as the for organizing information to ‘understand spatial link to specifi c land units (Briske et the linkages among ecological processes, al., 2005). Few invasions or invasive species vegetation dynamics, management practices distributions are limited to one ecological and assessment’ (see Sheley et al., 2010). In site, and the landscape scale patterns of this section, we provide examples of the ecological sites/states may be just as application of ecological site/state and important in determining the rates and transition models to the process of defi ning patterns of invasions as the within-com- problems, selecting and implementing munity attributes. However, the information management responses, and assessing in ESDs and STMs can be used to calibrate impacts of a common invasive species models of invasive plant dynamics and problem: the invasion of arid and semi-arid behaviors at larger scales that quantify the grasslands by shrubs. 12 J.R. Brown and B.T. Bestelmeyer

infl uence plant invasions is incorporated Problem recognition/ into GIS tools, their utility in decision- definition making will greatly improve. In the implementation phase (Fig. 1.2), identifying spatial heterogeneity is a critical Assessment/ Planning/design requirement in achieving the desired redefinition outcomes of a management strategy (Rumpff et al., 2011). Soils vary in their inherent water and nutrient holding Implementation/ capacity, in their susceptibility to wind and adjustment water erosion, and in their ability to tolerate and recover from disturbance; all of which Fig. 1.2. The adaptive management cycle. Some help defi ne the relative susceptibility of an steps have been combined to more accurately ecosystem to invasion. Likewise, changes in refl ect the application of adaptive management to vegetation over time can create a variety of invasive species. diff erent conditions. Periods of drought or heavy livestock grazing can create a patchwork of bare ground or open spaces While invasive species management among vegetation patches that can vary often focuses on the conditions necessary both within and across seasons. for plant establishment and reproduction, Th e conversion of grasslands to shrub- the spatial distribution of these conditions, lands in a variety of locations around the at the scale of seed dispersal and seedling world over the past century is an excellent establishment, has been poorly described model for illustrating the importance of (Brown et al., 2002). For example, while the multi-scale heterogeneity that can lead to climate and soils of a region may be adequate, ecosystem-scale change. Th e importance of on average, for plant recruitment and spatiotemporal heterogeneity in determin- growth, the availability in both space and ing the rates and patterns of invasive species time of those conditions is what ultimately spread and impact is illustrated by desert determines the ability of a species to grassland invasion by the shrub species establish and become abundant in the plant honey mesquite (Prosopis glandulosa). In a community (Davis et al., 2000; Masters Chihuahuan desert grassland, increasing and Sheley, 2001). In essence, spatial distance between patches of relatively heterogeneity describes the patterning of uniform perennial grass ground cover (black environmental conditions in which an grama grass; Bouteloua eriopoda) in response organism resides or through which it moves. to extended drought (a rapid, discrete Th us, spatial heterogeneity is the most trigger) or grazing-initiated (a slow, gradual important factor in determining how a change), grass loss and soil redistribution particular species invades a community, favors shrub seedling establishment, landscape, or region (Th eoharides and recruitment, and growth (Okin et al., 2006). Dukes, 2007). Th ese patterns are critical in Th e transition from a black grama grassland recognizing and defi ning the problem of state to a mesquite invaded state represents plant invasions (Fig. 1.2). In the planning the establishment of processes that favor phase, accounting for spatial heterogeneity continued shrub recruitment, limit grass is a necessary component in setting manage- competition and fi re frequency/intensity ment priorities and allocating resources to (see Bestelmeyer et al., 2006). As individual achieve results. Th e use of geographic grass plants succumb to drought stress, information systems (GIS) has aided excessive defoliation, or burial by locally tremendously in the development of transported sediment, the loss of cover and invasive species management plans. As increase in bare ground leads to black grama more information about the spatiotemporal loss, dominance of shorter-lived bunch- distribution of ecological processes that grasses, and further wind and water Managing Invasive Species in Heterogeneous Ecosystems 13

redistribution of soil resources to shrub nial grass component (Fredrickson et al., patches in the bunchgrass/mesquite state. 2006). Relatively infrequent fi re and In the bare interspaces, invasive shrubs are competition from a continuous warm season more likely to establish and persist than grass layer are thought to be suffi cient to perennial grass plants, garnering below- limit the recruitment and growth of shrubs. ground resources via their extensive root In the mesquite-invaded or bunchgrass/ system, and capturing aboveground mesquite states (Fig. 1.1), a management resources via their shrub structure. As strategy would have to include a disturbance shrubs capture more resources and enhance regime that suppresses existing shrubs, their performance (growth and seed pro- allowing the existing grasses to recolonize duction), the availability of their propagules bare spaces (Havstad and James, 2010). Th is increases within the community and approach would necessarily include both positively reinforces the process of change, intensive measures to target shrubs via enabling the transition to the shrubland more frequent and intense fi res, herbicide state. Ultimately, these processes increase application, and more extensive grazing in scale to encompass the entire plant management, e.g. low stocking rate to community and cascade across larger scales, ensure adequate herbaceous fuel accu- resulting in landscape-scale conversions mulation and exclusion from grazing (Peters and Havstad, 2006). In this case, during critical periods for grass recruitment mesquite invasion restructures not only the and expansion. Th is management regime plant community, but also the soil resources represents a substantial redirection of (Schlesinger et al., 1990). resources and management to a level that Th e redistribution of vegetation pro- precludes decision-making based primarily duction and soil resources creates multi- on economic considerations related to live- scale heterogeneity, enhancing runoff – stock production. Finally, if the shrub:grass run-on relationships and creating increased relationship has shifted entirely in favor of opportunities for shrub seedling recruit- shrubs (Fig. 1.1; shrubland state), there is ment (Rango et al., 2006). Th e redistribution relatively little opportunity to implement of resources at multiple scales acts as a a management-based strategy, and the positive feedback on the processes imparting emphasis of management actions shifts to resilience to the mesquite shrubland state by either intensive mechanical manipulation of concentrating resources around shrubs and the soil surface or alternative land uses limiting resource availability in the bare (Havstad and James, 2010). Th e transition interspaces for grass recruitment. At any away from mesquite dominance, via point in the change from one state to intensive management actions, is certainly another, diff erent disturbances (e.g. grazing not cost eff ective, and may not even be or fi re) can interact with climatic drivers possible. (e.g. drought, high rainfall) in a variety of Th e preceding example illustrates how ways. Assessing the eff ects of management the various soils × vegetation interactions intervention (Fig. 1.2, step 4) can only be on a single site can have a tremendous done within the context of the particular infl uence on the development of manage- site × season interactions. In particular, ment strategies and tactics to control an adapting management responses to chang- invasive species. However, the model of ing conditions requires explicit knowledge community dynamics presented in Fig. 1.1 is of these unique site × season interactions representative of only a relatively small and how they impact invasion patterns. portion of the entire landscape (Fig. 1.3). Each of these states (Fig. 1.1) features a Each of the geomorphic surfaces (limestone, set of attributes that dramatically alter how gravelly, clayey, sandy, loamy) has diff erent management of invasive species is soil properties, vegetation dynamics and accomplished. In the black grama grassland diff erent vulnerabilities to invasion and state (Fig. 1.1), water and nutrient cycles are other forms of degradation. Limestone soils controlled largely by the herbaceous peren- are relatively stable and only moderately 14 J.R. Brown and B.T. Bestelmeyer

Gravelly (shallow, relict piedmont) Limestone Hills Shallow gravelly soils, fissures permit Rocks protect grass, shrub dominance, water limited for patchily vegetated grass with mix of grass and shrubs at potential Sandy (relict basin floor) Erodible surface soils once grasses removed, prone to shrub dominance Basin (15 km) Jornada

Soil mapping units of the

Loamy (silt loam piedmont) Transportational surface, susceptible to water erosion and grass loss

Clayey (basin floor) Receives water and sediment, good water infiltration and retention, grass cover is high and resilient

Fig. 1.3. The soil-geomorphic template of a southern New Mexico, USA landscape. Each of the soil- geomorphic units possesses unique characteristics that impart differing resistance (tolerance to disturbance) to invasive species and differing resilience (ability to recover after disturbance). Reproduced from Bestelmeyer et al. (2006). susceptible to degradation. Gravelly soils are confer the greatest resistance to invasion is shallow and susceptible to erosion and the certainly desirable. When management invasion of the shrub Larrea tridentata favors plant communities that use available (creosote bush). Sandy and loamy soils have resources as effi ciently as possible, the similar levels of vulnerability to shrub availability of those resources to potential invasion by mesquite via the processes invaders will be limited. However, it is described previously. Th us, the distribution possible that a particular invasive species’ of plant communities across the landscape ability to disperse, establish, and thrive can have a profound impact on invasive across the landscape may overwhelm passive species spread and impact. In the above management designed to maximize invasion example, the processes driving the increase resistance. For example, in another semi- in both density and cover of honey mesquite arid grassland, density and cover increases on the sandy and loamy soils are diff erent in the same invasive species, honey than the processes driving the increase of mesquite, result from a much diff erent set of creosote bush on the gravelly soils. Given processes and presents a diff erent set of the diff ering susceptibility to shrub challenges for management. In more mesic invasion and soil patterns within a land- south Texas and east Texas ecosystems, scape, substantially diff erent management USA, the conversion from grasslands to responses are required depending on which shrublands is essentially a case of invasion state the plant community is in (Bestelmeyer driven by dispersal rather than loss of et al., 2011). resistance (Archer et al., 1988). Th e change Managing ecosystems for invasion in plant community dominance from grasses resistance among individual landscape to shrubs is attributed to a change in components and for landscape patterns that dispersal vectors (increase in domestic Managing Invasive Species in Heterogeneous Ecosystems 15

cattle) as the driving cause of the invasion, predictable with knowledge of the distri- not the changes within community bution of features that attracted livestock attributes (competition from grasses) or the (water development). Once viable popu- spatial patterning of plant communities lations were established proximate to water across the landscape (Brown and Archer, facilities, livestock dispersed the seeds into 1999). Competition from surrounding the uplands. Prior to the introduction of grasses was largely ineff ective in limiting cattle as the dominant grazer in the region, honey mesquite ingress and increase because the fi re-, drought-, and grazing regimes that of its ability to germinate and establish a were in place throughout the early 20th taproot that extended beyond the zone of century were suffi cient to limit the competition with grass roots within one expansion of prickly acacia (Radford et al., season (Brown and Archer, 1989). Th e 2001). When the domestic grazing regime availability of soil water and nutrients was changed (cattle to sheep; chronic over- not a limiting factor because of the ability of grazing), the factors dampening or slowing the invasive shrub to escape competition change (fi re, competition, and insect control with grasses within a time period in which of seed production in sparse populations) soil moisture was abundant. Although honey were diminished in their importance and mesquite recruitment in this ecosystem another set of factors (dispersal, lack of fi re, seemed to transcend community scale and seasonal soil moisture abundance) that controls, Wu and Archer (2005) identifi ed favored the change to shrub-increase was relatively fi ne-scale community topoedaphic ascendant. variability that could be used as a reliable predictor of the spatial distribution of mesquite recruitment success. Conclusion Th e same processes can drive exotic species invasion patterns at the regional and Th ese above examples illustrate the necessity continental scales. For example, an exotic of considering both the attributes of shrub, prickly acacia (Acacia nilotica) has potential invasive species as well as the invaded subtropical perennial grassland in a attributes of the plant communities, land- relatively short time in Australia (Brown and scapes, and regions vulnerable to invasion. Carter, 1998). Prickly acacia was introduced Seemingly minor changes in dispersal to the Mitchell Grass Plains as early as the vectors, weather, vegetation composition late 19th century, but was not considered an and distribution, soil properties, and land- invasive species until the mid-1970s. Its scape position can result in completely increase in density within areas where it had diff erent outcomes. While the understanding been planted, and expansion to new areas of plant attributes is critical to predicting was associated with the replacement of invasion processes and designing responses, sheep by cattle as the dominant domestic other authors in this volume address species livestock species. Both grazers consumed performance in detail (see Eviner and acacia seed pods as seasonally important Hawkes, Chapter 7, this volume; James, parts of their diets, but sheep were far more Chapter 8, this volume). However, know- likely to destroy the seed during chewing ledge of how landscape and regional than cattle. Many of the ingested seeds heterogeneity govern patterns of livestock passed through cattle digestive system movement and seed dispersal is essential to intact and were deposited in dung across a predicting species movements and devising much wider area than seed could be management strategies. transported by wind or water (Radford et al., Th e patterns of invasive species 2001). Th e patterns of species movements abundance can be defi ned by processes that through the region and landscapes were occur within communities (competition) or entirely predictable based on the knowledge by processes that link communities of livestock movements. Within-community (dispersal) and control spread. In either patterns of shrub recruitment were also case, spatial structure is a key piece of 16 J.R. Brown and B.T. Bestelmeyer

information that is necessary to make Sanchez, H., Briske, D.D. and Fernandez- predictions, devise responses, and assess Gimenez, M.E. (2010) Practical guidance for impacts (see Masters and Sheley, 2001). In developing state-and-transition models. this chapter, we have presented two tools Rangelands 32, 23–30. (ecological sites, state and transition Bestelmeyer, B.T., Brown, J.R., Fuhlendorf, S., Fults, G. and Wu, X.B. (2011) A landscape models), which appear critical for the approach to rangeland conservation practices. implementation of ecologically based In: Briske, D.D. (ed.) Conservation Benefi ts of invasive plant management. Th e ever- Rangeland Practices: Assessment, Recom- changing spatiotemporal dynamics of mendations, and Knowledge Gaps. Allen Press, invasive plants in wildlands requires a Lawrence, Kansas/USA, (in press). systematic way to organize, display, and Blumenthal, D., Booth, D.T., Cox, S.E. and Ferrier, analyze information that serves as the basis C.E. (2007) Large-scale aerial images capture for decision-making. While a particular details of invasive plant populations. Rangeland species’ characteristics are an important Ecology and Management 60, 523–528. part of that knowledge base, the ability of an Briske, D.D. (2011) Introduction. In: Briske, D.D. individual to germinate, survive, grow, and (ed.) Conservation Benefi ts of Rangeland Practices: Assessment, Recommendations, reproduce can only be determined by the and Knowledge Gaps. Allen Press, Lawrence, resources available to it within the context Kansas/USA, pp. 1–8. of the surrounding community and land- Briske, D.D., Fuhlendorf, S.D. and Smeins, F.E. scape. An organized knowledge of these (2003) Vegetation dynamics on rangelands: a relationships is the basis for reliable critique of the current paradigms. Journal of prediction of a species’ movements and for Applied Ecology 40, 601–614. the development of management responses Briske, D.D., Fuhlendorf, S.D. and Smeins, F.E. to slow or stop that movement. (2005) State-and-transition models, thresholds, and rangeland health: a synthesis of ecological concepts and perspectives. Rangeland Ecology References and Management 58, 1–10. Brown, J.R. (2010) Ecological sites, their history, Archer, S.R. and Predick, K.I. (2008) Climate status and future. Rangelands 32, 5–8. change and ecosystems of the southwestern Brown, J.R. and Archer, S.R. (1989) Woody plant United States. Rangelands 30, 23–28. invasion of grasslands, establishment of honey Archer, S.R., Scifres, C.J., Bassham, C. and Mesquite (Prosopis glandulosa var. glandulosa) Maggio, R. (1988) Autogenic succession in a on sites differing in herbaceous biomass and subtropical savanna, rates, dynamics and grazing history. Oecologia 80, 19–26. processes in the conversion of grassland to Brown, J.R. and Archer, S.R. (1999) Shrub invasion thorn woodland. Ecological Monographs 58, of grassland, recruitment is continuous and not 111–127. regulated by herbaceous biomass or density. Beisner, B.E., Haydon, D.T. and Cuddington, K. Ecology 80, 2385–2396. (2003) Alternative stable states in ecology. Brown, J.R. and Carter, J. (1998) Spatial and Frontiers in Ecology and the Environment 1, temporal patterns of exotic shrub invasion in an 376–382. Australian tropical grassland. Landscape Bestelmeyer, B.T. (2006) Threshold concepts and Ecology 13, 93–102. their use in rangeland management and Brown, J.R., Herrick, J. and Price, D. (1999) restoration: the good, the bad and the insidious. Managing low-output agroecosystems sustain- Restoration Ecology 14, 325–329. ably, the importance of ecological thresholds. Bestelmeyer, B.T., Brown, J.R., Havstad, K.M., Canadian Journal of Forest Research 29, Alexander, R., Chavez, G. and Herrick, J.E. 1112–1119. (2003) Development and use of state-and- Brown, J.R., Svejcar, T., Brunson, M., Dobrowolski, transition models for rangelands. Journal of J., Fredrickson, E., Krueter, U., Launchbaugh, Range Management 56, 114–126. K., Southworth, J. and Thurow, T. (2002) Are Bestelmeyer, B.T., Ward, J.P. and Havstad, K.M. range sites the appropriate spatial unit for (2006) Soil geomorphic heterogeneity governs measuring and monitoring rangelands? patchy vegetation dynamics at an arid ecotone. Rangelands 24, 7–12. Ecology 87, 963–973. Cleland, D.T., Avers, R.E., McNab, W.H., Jensen, Bestelmeyer, B.T., Moseley, K., Shaver, P.L., M.E., Bailey, R.G., King, T. and Russell, W.E. Managing Invasive Species in Heterogeneous Ecosystems 17

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monitoring of rangelands. Rangeland Ecology features on patterns of woody plant and Management 59, 19–29. encroachment in savanna landscapes. Westoby, M., Walker, B.H. and Noy-Meir, I. (1989) Landscape Ecology 20, 733–742. Opportunistic management for rangelands not Young, T.P., Chase, J.M. and Huddleston, R.T. at equilibrium. Journal of Range Management (2001) Community succession and assembly: 42, 266–274. comparing, contrasting, and combining Wu, X.B. and Archer, S.R. (2005) Scale-dependent paradigms in the context of ecological infl uence of topography-based hydrologic restoration. Ecological Restoration 19, 5–18. Linking Disturbance Regimes, Vegetation Dynamics, and Plant 2 Strategies Across Complex Landscapes to Mitigate and Manage Plant Invasions

Samuel D. Fuhlendorf, Brady W. Allred, R. Dwayne Elmore, and David M. Engle

Natural Resource Ecology and Management, Oklahoma State University, USA

Introduction livestock, altered fi re regimes, short- and long-term fl uctuations in climate, and Composition and structure of ecosystems increased atmospheric trace gases. Rates have long been described as dynamic, but a and patterns of species invasions are: (i) recent focus on invasive species has often non-linear and characterized by revitalized discussion of assembly rules and gradual initial changes followed by rapid dynamics of species composition. For increases; (ii) accentuated by drought with rangelands, our understanding of species variable rates over time; (iii) infl uenced invasions focused fi rst on the eff ects of by soil and topographic features; and (iv) woody plant encroachment on livestock often non-reversible over management production. Recent research emphasizes timeframes (Archer, 1994). Research and intentional and accidental introduction of observations on rangelands leave little exotic species that are capable of dramatic doubt that species invasions have caused changes to ecosystem structure and major changes over the past 100–200 years, function. Increased attention to invasive frequently converting ecosystem types (e.g. species, including woody plants, has altered converting open grasslands to closed canopy our understanding and methods of woodlands and converting shrublands to describing rangeland dynamics. In early annual grasslands). 20th-century descriptions of rangeland Species invasions are best understood dynamics, grazing was seen as the primary through studies at multiple scales. Th e driver that produced dynamics with invasion process is a landscape or regional predictable, linear, and reversible patterns. process that has been analogously described Species invasions that can greatly alter as a glacier moving across a region converting dynamics led to the development of such the landscape as it progresses. An agronomic rangeland concepts as thresholds, multiple approach to research focused on small plots steady states, and our modern perception of may be useful in understanding small- rangeland change. scale mechanisms of invasion or potential Following initial introduction and management approaches, but is not eff ective establishment, invasion of both exotic and at understanding broad-scale invasions or native species has been associated with landscape/regional management schemes. many factors, including introduction of Most research eff orts understandably rely

© CAB International 2012. Invasive Plant Ecology and Management: Linking Processes to Practice (eds T.A. Monaco and R.L. Sheley) 19 20 S.D. Fuhlendorf et al.

on small plots to focus on the immediate species life history traits determines its problem of controlling invasive species, but establishment and abundance in the novel these studies potentially neglect landscape ecosystem. Using this framework we will and regional patterns that might be specifi cally examine the following questions: important in understanding the causes and (i) What traits allow some species to gain implications of invasions. Understanding dominance within some ecosystems and not and eff ectively managing invasion processes others? (ii) What environmental factors requires an appreciation of: (i) the ecological have contributed to the number and extent relationships that contribute to invasive of invasive species? and (iii) What can we species success; (ii) the ecological factors take from our understanding of succession that promote effi cacy of specifi c management and rangeland dynamics that will contribute strategies; and (iii) the most appropriate life to our management of invasive species? stages and techniques for management. Even though causal factors associated with Management of these species also requires species invasions are variable across species an ecological approach that incorporates the and ecosystems, most invasions can be life history of the invasive species and the explained by a few plant adaptations that ecological structure and function of the relate to current and historical environ- ecosystem that is being invaded. mental conditions. Our goal is to present an overview of relevant ecological considerations for invasive species in general and to suggest Evolutionary Disturbance Processes some ecological factors to consider before on Rangelands initiating management eff orts. We will focus much of our discussion on rangelands, which Rangeland ecosystems are characterized by are disturbance-dependent ecosystems with disturbances, and most species within a long history of fi re and grazing, as well as these landscapes vary in abundance highly variable topo-edaphic and weather depending on their adaptation to dis- patterns. We conceptualize species invasion turbance patterns (Fig. 2.2). Generally by focusing on pre-invasion conditions and disturbance refers to a discrete, punctuated trying to understand the interplay between killing, displacement, or damaging of one changes in disturbance processes and traits or more individuals that directly or of invasive species that would best prohibit indirectly creates an opportunity for new or inhibit invasions (Fig. 2.1). From this individuals to become established (Sousa, perspective, the portfolio of factors infl uenc- 1984). By this defi nition, it is somewhat ing species assemblages can be viewed as a circular to argue that diversity is dependent fi lter (climate, soils, grazing, fi re, etc.) and on disturbance because the defi nition

How do species become invasive?

Plant Changing Adaptations Environment • Seed dispersal • Introduced grazing • Seed bank and production • Altered fire regimes

• Seedling establishment • Increased CO2 • Growth rates • Changes in climate • Community influences • Land fragmentation • Plant longevity • Land use legacy • Stress tolerance • Resistance to disturbance

Fig. 2.1. The potential for a species to invade is dependent upon the interaction among the plant traits of the invasive species and the changing environment of the ecosystem. The environment acts as a fi lter and the traits allow certain plants to be more or less successful. Linking Disturbance Regimes, Vegetation Dynamics and Plant Strategies 21

includes opportunity for new species. sometimes considered a disturbance that is However, consideration of the pattern of central to understanding rangeland dynamics this relationship and analysis of variable resulting from evolutionary pressures that responses across diff erent disturbances is shaped rangeland communities. Most pro- warranted. Studies of rangelands suggest cesses that are described as disturbances can the dominant disturbances prior to also be considered as a critical part of the European settlement were fi re and grazing, portfolio of factors that contribute to and their interaction with each other ecosystem structure and function (e.g. (Fuhlendorf et al., 2009a). Rangeland biotic climate, soils, fi re, grazing, etc.). communities also evolved with many Based on the simple defi nition of secondary disturbances (e.g. prairie dogs, disturbance processes, it is circular to predict insect outbreaks) and their interaction with changes in species diversity and species dominant disturbances. Drought also is invasion but it is useful to discuss predicted

(a)

(b)

Fig. 2.2. Fire and grazing are common disturbances on rangelands throughout the world. (a) Fire in a sand sagebrush/mixed prairie ecosystem being invaded by eastern redcedar. (b) Bison grazing on a recently burned patch at the Tallgrass Prairie Preserve in Oklahoma. Both photos by Stephen Winter. 22 S.D. Fuhlendorf et al.

patterns. Th e diversity/disturbance relation- For at least the 10,000–15,000 years ship relates to invasive species because by before Europeans discovered the Americas, reducing the dominance of a few species or herbivory in North America was largely by widening niche space, exotic or invasive controlled by the interaction between fi re species can capitalize on the opportunity and grazing. Disturbance by grazing animals (Elton, 1958; Tilman, 1997; Gurvich et al., on prehistorical landscapes was dependent 2005) depending on the type, duration, upon selection of grazing locations and the timing, frequency, etc. of the disturbance interaction among landscapes, people, and and other processes (e.g. change in disturbance processes. For example, recent atmospheric deposition of nitrogen). Th e evaluations of grazing animal behavior processes that maintain species diversity suggest that grazing animals are strongly can potentially be the dominant contribution responsive to fi res, which were dominant to invasion. While we might assume that forces ignited by lightning or aboriginal evolutionary disturbances benefi t native peoples. Th e net eff ect was a disturbance- species over exotic species, there is no dependent landscape where the interaction certainty that invasive species will not of fi re and grazing was a dominant feature, increase from these disturbances as well. hereafter termed pyric herbivory (i.e. Moreover, invasion can ultimately alter the herbivory shaped by fi re) (Fuhlendorf et al., disturbance regime, increase positive 2009a). From this perspective, grazing and feedbacks at the expense of negative feed- fi re may best be viewed as a single backs, and shift the ecosystem across a disturbance (pyric herbivory) that creates a threshold to a new system state that may shifting mosaic of disturbance patches contain reduced species diversity (Fig. 2.3). across a complex landscape (e.g. Fuhlendorf Finally, it is possible that more species-rich and Engle, 2001; Salvatori et al., 2001; plant communities are more susceptible to Hassan et al., 2008). Th is results from invasion than less species-rich plant com- grazing animals freely selecting between munities (Ganguli et al., 2008) at least at burned and unburned portions of the some scales (Stohlgren et al., 2002). landscape, and the dependence of fi re

Probability of selection by grazing animals (+) (–)

Recently burned, Transitional No fire for 3 years currently grazed state minimal grazing

High production, <1 year High bare ground 2–3 years Accumulated litter quality and and forbs and low and standing availability of forage litter and standing biomass of mostly biomass grasses

(–) (+) Probability of fire

Fig. 2.3. A conceptual model of path dynamics within a shifting mosaic landscape where each patch is experiencing similar but out-of-phase dynamics driven by fi re and grazing (Fuhlendorf and Engle, 2004). Ovals represent the primary drivers (fi re and grazing) while squares represent the ecosystem states within a single patch as a function of time since focal disturbance. All states have the potential for fi re or grazing. Solid arrows indicate positive (+) and negative (–) feedbacks in which plant community structure is infl uencing the probability of fi re and grazing. Linking Disturbance Regimes, Vegetation Dynamics and Plant Strategies 23

occurrence on the removal of fuel by changes in plant communities to traits or herbivores (e.g. Norton-Griffi ths, 1979; ‘vital attributes’ of plants has long been Fuhlendorf and Engle, 2001) (Fig. 2.3). Th at proposed (Noble and Slatyer, 1980; Glenn- rangeland biodiversity depends largely on Lewin, 1992). Research has often focused on the mosaic pattern that results from pyric important life history traits that can herbivory refl ects the notion that fi re and facilitate the establishment of species grazing were historically coupled. Over following disturbance, including propagule the past century or more these processes persistence or dispersal, age at fi rst have been decoupled, largely to facilitate reproduction, and longevity or life span. Th e utilitarian eff orts to maximize livestock importance of these traits, as well as the production (Fuhlendorf and Engle, 2001). ability to persist or reproduce vegetatively, While other disturbance processes were also has been demonstrated for forest community important, the interaction of multiple response to fi re, and they have been used to herbivores with fi re across large and complex predict the species that would most likely landscapes was central to the development increase or decrease immediately following of the Great Plains of North America prior the disturbance. to settlement. Restoration of these processes Other approaches to predicting species and patterns is frequently the focus of invasions from life history traits have conservation of rangeland ecosystems but focused on a species’ ability to persist in conversion and fragmentation of these resource-rich conditions (competitors), landscapes are serious barriers. resource-poor conditions (stress tolerators), and with frequent/intense disturbances (ruderals) (Grime, 1977; Blumenthal et al., Life Histories and Plant Strategies 2009; Seastedt, 2009). Th e use of life history that Promote Invasiveness traits of functional categories of species is easily transferred to invasive species Although life history strategies of invasive invasion following disturbance (Horvitz et species and the conditions of the com- al., 1998), but considerable debate persists munities they invade are highly variable, about whether it is productive to categorize some generalizations are useful in under- species as ‘native’ or ‘exotic’ rather than standing invasion processes and developing simply focus on the species traits. A recent ecological management strategies. Some analysis of species traits (competitor, stress- plant traits can enable a species to invade into tolerator, ruderal) highlighted an important a relatively open environment while other consideration when discussing species traits traits may allow a species to invade under a and potential invasive species. Blumenthal closed canopy. Likewise, some species are well et al. (2009) describe a paradox in which adapted to specifi c dis turbances, such as fi re resource availability can have a strong and grazing, while others may be extremely infl uence on plant success but that would sensitive to only one or all disturbances. hold equally true for native and exotic Effi cient ecological management requires an species. Th ey concluded that life history understanding of plant traits that best traits, resource availability, and the describe the potential strengths and structure of the pathogen community were weaknesses species have toward invasion, as predictive of the potential invasiveness of well as the conditions of the ecosystem that exotic species, suggesting that invasive make it susceptible to invasion. species are unique due to rapid growth and By identifying these traits and under- limited pathogen burden (Blumenthal et al., standing some basic ecological general- 2009; Seastedt, 2009). Pressures from izations, a land manager can more effi ciently herbivory can be evaluated in similar light evaluate the need for management and to the pathogen burden perspective. If a accurately focus on critical phases of species is capable of tolerating or avoiding vegetation change. Linking successional herbivory, it would be similar to low 24 S.D. Fuhlendorf et al.

pathogen burdens in the novel ecosystems open areas for grazing and crops was that they invade. commonplace, possibly resulting in a short- lived increase in fi res (Stambaugh and Guyette, 2006; DeSantis et al., 2010). Not What has Changed? Altered long after settlement, however, natural- Disturbance Regimes and Global occurring and anthropogenic fi res were Change reduced in most places because of fear of fi re, which made wildfi re suppression a high While plant adaptations are useful in priority. Also, heavy stocking of rangelands explaining species invasions on rangelands, with livestock reduced the probability of patterns and rates of invasion have been natural-occurring fi res by reducing the highly variable in space and time. Th is standing biomass available for fuel (Belsky suggests that the biotic and abiotic structure and Blumenthal, 1997; Fuhlendorf et al., and function of these ecosystems, as well as 2008). Th e result is an invasion of species the dynamics over time, may be useful in (both native and exotic) that are intolerant explaining the invasion process. Since we of fi re. are interested in the invasion of species into In some regions, disturbances such as a matrix that was dominated by historic grazing greatly alter the response of disturbances and other processes that acted ecosystems to fi re, resulting in both increased as assembly factors for historical plant and decreased invasion following fi res communities, it is useful to think about (Cummings et al., 2007; Davies et al., 2009). how structure and function of these eco- Grazing alters the fuel load and ultimately systems have been altered since European fi re intensity, which can strongly infl uence settlement. Major alterations of landscapes fi re severity and the rate and pattern of that contribute to species invasions include: invasion (Twidwell et al., 2009). Also, long- (i) altered fi re regimes; (ii) introduction of term grazing can greatly alter composition domestic livestock; (iii) increases in atmos- and seed bank of plant communities limiting pheric trace gases; (iv) climatic infl uences; the rate and pattern of recovery following and (v) land use legacies (Archer, 1994; fi re (Fuhlendorf and Smeins, 1997), leaving Fuhlendorf et al., 1996, 2002; Polley et al., sites available for invasive species establish- 1997). ment. Alternatively, patch fi res can result in focused intense grazing pressure limiting the stature of all species and reducing the Altered fi re regimes importance of traits that may promote tolerance or avoidance of grazing (Cummings Prior to European settlement, periodic fi res et al., 2007). burned across many landscapes maintaining Invasion of species to novel ecosystems fairly open grasslands with woody plants can be associated with a feedback where limited to discrete landscape units where fi re initial invasion can alter fi re regimes, leading frequency and intensity were insuffi cient to to further invasion. Th e common example control their dominance (Axelrod, 1985). of this is the invasion of sagebrush eco- Th ese patterns were highly variable depend- systems in the western USA by cheatgrass, ing on regional diff erences in climate, soils, which promotes fi re, reducing sagebrush vegetation, and people. For centuries, the dominance, and promoting further invasion majority of fi res were set by aboriginal people and additional fi res. Th is occurs due to life for myriad of reasons, but the resulting fi re history, as cheatgrass grows early in the regimes were as critical to the evolution of growing season, rapidly maturing and species and the development of ecosystems providing fuel for subsequent fi res through- as climate and other abiotic factors. At the out the summer (the pre dominant fi re time of settlement, the use of fi re to clear the season for this region; Brooks et al., 2004). land of woody vegetation and provide more Th ese frequent fi res remove non-sprouting Linking Disturbance Regimes, Vegetation Dynamics and Plant Strategies 25

shrubs, such as Artemsia spp., shifting the resulting in high mortality to livestock plant community from shrub to annual during droughts before 1900 (Lehmann, grass. An alternative example that is less well 1969). Livestock mortality rates were described is the invasion of tall fescue (Lolium reported as high as 85% in the southwestern arundinaceum (Schreb.) S.J. Darbyshire) in USA before the turn of the century, central tallgrass prairies, the most abundant suggesting that the resource had been over- plant in the eastern USA (Rudgers and Clay, used soon after settlement. Th ese severe 2007). Th is is a cool-season grass that is stocking rates likely resulted in complete invading warm-season grasslands. Once removal of vegetation cover and may have invaded, the species is active during much of led to signifi cant erosion (Smeins et al., the season when fi re can be used, in some 1997). Th ese infl uences should be considered cases limiting the ability to conduct pre- when evaluating vegetation change over the scribed burns, which can further promote past 100–200 years, particularly when we invasion of woody plants and shift the plant base our ecological discussion and manage- community to woodland. Th ese two examples ment decisions on plant community demonstrate that while we have greatly descriptions of pre-settlement conditions. altered fi re regimes, invasion by some species Th e abiotic environment associated with can result in positive feedbacks that further those descriptions frequently no longer alter fi re regimes and these changes can exists (Fuhlendorf and Smeins, 1997). either increase or decrease fi re-return Invasion of native and exotic woody interval and fi re intensity. plants has been presumed to be facilitated by the introduction of livestock (Walker et al., 1981; Belsky and Blumenthal, 1997). In Introduction of domestic livestock some cases, similar relationships could be drawn to herbaceous invasive species, Over the past 20,000 years, animal although such relationships can be merely infl uences on North American rangelands correlative with many changes (livestock, have gone from a high diversity of now reduced fi re, species introductions, etc.) that extinct large herbivores (mammoths, have occurred for the past century or more. sloths, giant bison, camels, etc.) to vast However, several studies have suggested that herds of smaller herbivores (primarily the presence of dense grassland vegetation, bison) that were largely interactive with fi re as in ungrazed or moderately grazed and large complex landscapes at the time of conditions, may not restrict the establish- settlement (Burkhardt, 1996; Fuhlendorf et ment of woody species or more recently al., 2009a). Many rangelands developed for invasive species, but instead may actually many years with high numbers of herbivores improve conditions for seedling survival and while others were only grazed periodically development (Schultz et al., 1955; Archer, or by small resident herds of animals. It is 1995; O’Connor, 1995; Cummings et al., diffi cult to determine the evolutionary 2007). Woody plant or invasive species infl uence of grazing on current rangelands dominance could be inhibited by harsher of North America, but it is likely that the micro-environmental conditions in a heavily introduction and ultimately the grazed community that may limit seedling confi nement of livestock 100–150 years ago establishment, or invasion could be inhibited has considerably changed the structure and by consumption of seedlings by livestock, function of rangeland ecosystems, the level of which depends on the specifi c infl uencing the invasion processes that we species of herbivore and invader. Many plant are observing today. seedlings are frequently more susceptible to Most of the settlers were inexperienced mortality from grazing than mature plants in managing the relatively dry rangelands of as immature plants can be more palatable North America. Consequently, initial stock- with less defense mechanisms, such as ing rates were often extremely high, secondary chemicals or thorns, and may be 26 S.D. Fuhlendorf et al.

less tolerant of defoliation (Taylor et al., therefore at low CO2/O2 ratios; and (iv) 1997; Cummings et al., 2007). invasion of many of these species has been Grazing also has a major infl uence on accompanied by a 30% increase in other disturbance processes. For example, atmospheric CO2 over the past 200 years grazing alters fi re regimes through the (Archer et al., 1995; Smith et al., 2000). removal of fuel required to ignite and carry a While the relationship between C3 plants fi re (Fuhlendorf and Smeins, 1997; and CO2 is theoretically sound and Fuhlendorf et al., 2008). For fi re-sensitive experimentally demonstrated, the ultimate species, reduction of fuel from grazing may role of these changes in the invasion be the greatest eff ect of grazing on the processes remains largely unstudied (Smith invasion process. Similarly, pre-settlement et al., 2000). We know that many aspects of grazing was largely interactive with fi re, plant physiology are altered by elevated creating a shifting mosaic where grazing levels of CO2 and many of these factors have followed a fi re, which would result in a the potential to alter ecosystems and continuum of grazing intensity and thus landscapes, but studying the eff ect on large variability in plant composition and and complex areas is challenging. Direct structure. Modern range management has comparisons of native and invasive species decoupled fi re and grazing and developed in controlled environments have demon- technologies to promote spatially uniform strated that invasive species benefi ted more utilization so that current grazing patterns than native species from elevated CO2 have little resemblance to evolutionary levels, which could promote invasion (Song patterns. Evolutionary patterns of dis- et al., 2009). Elevated CO2 has been shown turbance are understood to be critically to alter species and potentially communities important to biodiversity, but evolutionary and landscapes, but it remains diffi cult to patterns of disturbance may also be critically develop accurate predictions of future important to invasion ecology if the same change. However, this suggests that focus- processes that promote diversity also ing management on ‘pristine’ conditions promote or retard invasion. at the time of settlement may not return these plant communities to a desired com- position. Increase in atmospheric trace gases Climatic infl uences Increases in atmospheric trace gases such as CO2 can alter the relationship between Many long-standing ecological principles are woody and herbaceous plants to favor based on the premise that the dominant woody plant dominance (Polley et al., 1997). vegetation on a landscape is largely dependent Similar arguments can be made for some upon the climate (Clements, 1916). Th e non-woody invasive plants (Smith et al., importance of short-term and long-term 2000). Th ese arguments are based on weather patterns on the increase in woody observations that indicate: (i) many invasive plants over the past 100–200 years has been species possess the C3 photosynthetic debated in the ecological literature (Miller pathway, whereas some rangelands are and Wigand, 1994; Belsky, 1996). A series of dominated primarily by plants that possess studies in the deserts of the southwestern the C4 photosynthetic pathway; (ii) USA suggest a variety of climate-related increased atmospheric CO2 may confer a causal factors are associated with the increase signifi cant advantage to C3 species relative in woody plants, including short-term to C4 species with respect to physiological droughts, long-term downward trends in activity, growth, and competitive ability; precipitation, and long-term warming and/or (iii) C4 grasslands appear to have evolved at cooling trends (Neilson, 1986; Conley et CO2 concentrations below 200 ppm and al.,1992; Archer, 1994). However, grazing has Linking Disturbance Regimes, Vegetation Dynamics and Plant Strategies 27

been identifi ed as a confounding factor that by assuming that removal of human-caused may be at least as important as climatic disturbance will cause native plant com- patterns as a causal factor of woody plant munities to become more similar to increase (Conley et al., 1992). historical accounts, as humans were shaping Analyses of fossil pollen, pack-rat plant communities for thousands of years middens, and stable isotopes of organic prior to European settlement. matter suggest that populations of woody plants varied with fl uctuations in climate over the past 20,000 years (Van Devender Land use legacies: cultivation, and Spaulding, 1979; Nordt et al., 1994; Hall development, and fragmentation and Valastro, 1995). For instance, mean annual precipitation and proportion of Over 100 years ago, most of North America winter precipitation explained 62% of the was settled and society focused on developing variability in the relative abundance of and improving landscapes through the shrubs across temperate grasslands and introduction of agriculture. Government savannas of North America (Paruelo and policies required homesteaders to cultivate Lauenroth, 1996). From about 600 to 150 portions of their land to demonstrate years ago, the earth experienced a cooling improvements (see Morris, Chapter 3, this period known as the Little Ice Age when volume). In many cases these lands were conditions were cooler and moister than unsuited for cultivation and nestled in a large they were both before and after this period. landscape dominated by rangelands. Much of It has been suggested that this climatic this cropland has been abandoned or restored anomaly favored the establishment of leaving plant communities that are con- grasslands and the warming since has been a siderably diff erent from untilled prairies in signifi cant factor associated with the terms of soil properties, vegetation product- increase in woody plants (Neilson, 1986). ivity, and plant composition. Obviously, dynamics in climate patterns Studies of restoration or recovery of would result in ecosystem changes that may Great Plains grasslands in North America, include shifts in general physiognomy. that were cultivated in the early 1900s, Climate-driven vegetation change often have resulted in highly variable conclusions. lags behind the change in climate because Many of these ‘restored’ croplands have not vegetation established under one climate recovered after 30–70 years (Burke et al., regime can persist for tens to hundreds of 1995; Fuhlendorf et al., 2002) with higher years until disturbed (Prentice, 1986). Th e bare ground and lower production, soil pristine grassland described by many carbon, and soil nitrogen than uncultivated settlers may have been established under prairies. Some evidence suggests that these very diff erent climatic conditions than those marginal lands that have been abandoned that exist today. Change in climate over the or restored serve as a source of invasive past 1000 years, particularly when con- species for the larger landscape. At a sidered with other changes (i.e. altered minimum, these lands need to be con- grazing and fi re regimes, and carbon sidered independent of uncultivated lands dioxide), suggests that the increase in woody and in some regions (e.g. Southern Great plants could have potentially been initiated Plains of North America) previously culti- before settlement (Smeins, 1984). Change vated land may be a dominant landscape in climate to conditions more appropriate feature. for woody plants may have occurred before Habitat fragmentation and other human settlement, and human disturbances could alterations of ecosystems and landscapes have acted as the catalyst that facilitated the are deemed important factors as environ- directional increase in woody plants. Th is is mental degradation. Human facilitated another indication that management should plant invasions are one consequence of not focus on ‘pristine’ historical conditions these alterations (Lonsdale, 1999; With, 28 S.D. Fuhlendorf et al.

2002). Among the human activities and Land use legacies: improved forages alterations that promote plant invasions are gone bad roads (Trombulak and Frissell, 2000; Christen and Matlack, 2009), housing in For over a century, North American rural (Gavier-Pizarro et al., 2010), exurban agronomists have imported plants from (Lenth et al., 2006), and urban areas (Song many environments in an attempt to increase et al., 2005), and cropland abandonment forage and food production (Ball et al., 2002; (Vilà et al., 2003). Processes directly linking Barnes et al., 2003). For the most part, these plant invasions to human activities include introductions have contributed positively to intentional introductions of ornamental agriculture, and most species introductions plants for landscaping (Mack and Erneberg, do not create unforeseen problems. However, 2002), soil and vegetation alteration forage species are often selected for their associated with site development (Hobbs ability to establish and persist in diverse and Huenneke, 1992; Wania et al., 2006), environments, elevating the potential for and road construction and maintenance invasion into unintended ecosystems. Forage activities that create fertile safe sites for species are often planted and managed invasive plants to germinate and establish as monocultures because they respond (Greenberg et al., 1997; Trombulak and favorably and uniformly to fertilizer, are Frissell, 2000; Barbosa et al., 2010). persistent in monoculture, and are often less Landscape structure, land-use pattern, palatable than native plants. Primarily grown and land ownership pattern are key in monocultures for livestock grazing, exotic determinates of the extent and severity of forage species are selected for traits that exotic species invasions. Severity and extent promote persistence under intense grazing. of encroachment of woody plants into Rapid maturation, secondary chemicals, and grassland is greater in landscapes containing fungal associations are among the traits that large numbers of smaller, intermingled enhance persistence when these species are patch types by providing isolated patches grown in grazed monocultures, but these juxtaposed with pockets of seed sources same persistence traits can facilitate (Coppedge et al., 2001). When only a few dominance in plant communities with seed-producing woody plants become diverse plant species in which native and established with avian seed dispersers domestic herbivores preferentially select (Holthuijzen and Sharik, 1985) and under among multiple plant species. an altered fi re regime in which fi re spread is Forage plant invasions may in part be interrupted by fragmentation, a feedback explained by the plant’s evolutionary history cycle results in dense stands of woody as shaped by fi re and grazing. Native plants vegetation (Fuhlendorf et al., 1996). Th is from the Great Plains evolved under a fi re– same pattern holds true with invasive non- grazing interaction (Fuhlendorf et al., native plants with cattle as seed vectors 2009a) whereas many introduced forages (Brown and Carter, 1998). Fragmented evolved with heavy grazing pressure and ownerships in a landscape can result in may be less dependent on fi re. Th erefore, severe invasions because the incentive to when grazing and fi re are decoupled and/or control the invasion on a single ownership is fi re is removed, the new environment does relatively small (i.e. individual managers are not resemble the evolutionary environment responsible for small portions of total of native species, and exotic forages gain damage caused by the invasive species) competitive advantage because grazers whereas escalating cost of control is passed preferentially select the more palatable to neighbors by a manager who chooses not native species. Grazers shift their preference to control the invasion and whose land from individual plants to patches (i.e. serves as a propagule source (Epanchin-Niell burned patches over unburned patches) in et al., 2010). landscapes that include burned and Linking Disturbance Regimes, Vegetation Dynamics and Plant Strategies 29

unburned patches (Fuhlendorf and Engle, Conclusion 2004). Th is suggests that individual plant traits that promote persistence and Th e long-standing emphases placed on dominance under grazing alone (i.e. when vegetation dynamics, life history traits, the grazing animal is allowed full selectivity disturbance ecology, population ecology, and among plants) would become less important landscape ecology are strongly informative as patch-level selectivity increases among to the study and management of invasive burned and unburned patches. So, the fi re– species. Evaluation of literature on species grazing interaction that functioned in the invasion patterns indicates that generalized evolution of Great Plains grasslands resulted conclusions are diffi cult because they are in selection patterns at the patch level highly scale dependent like most other depending on time since fi re. Th is contrasts ecological processes (Fuhlendorf and Smeins, with the traditional agroecosystem model 1999; Fridley et al., 2007; Fuhlendorf et al., that minimizes preferential selectivity by 2009a). To more easily interpret results, using monocultures of plants that possess ecologists often simplify their approach in an traits that promote persistence. When experiment or observation by focusing on introduced forage species possess these one factor while purposefully excluding the traits and escape into diverse native eff ects of other factors or ecological inter- communities, grazing animals select native actions. Although the simplifying approach species and the introduced forage species is eases analysis and interpretation, it also competitively favored and increases in excludes inherently complex yet critically abundance. Some highly palatable species important processes or traits, often resulting (i.e. lucerne) do not become dominant in in conclusions that may not scale up. Th e grazed ecosystems because they lack many eff ect of grazing, for example, is commonly of the persistence traits of other introduced viewed as a simple binary response (i.e. often forages. either yes or no regarding defoliation) and in

Box 2.1. Case study: Lespedeza cuneata Lespedeza cuneata (Dum.-Cours.) G. Don. is a perennial legume, introduced from eastern and central Asia for forage production, erosion control, and land reclamation. Intentionally introduced, L. cuneata is now considered invasive throughout the eastern and midwestern USA. Invasion into disturbed habitat and rangeland is rapid, displacing native species and forming dense patches strongly dominated by L. cuneata (Eddy and Morre, 1998; Brandon et al., 2004; Cummings et al., 2007). Dominance by L. cuneata often reduces biodiversity and biomass production of other species, as well as important ecosystem processes, such as altering nitrogen cycling (Price and Weltzin, 2003; Garten et al., 2008). Several important life history and plant traits of L. cuneata contribute to its successful establishment and invasion in competitive and stressful environments. A primary mechanism of L. cuneata invasion may be its ability to suppress native vegetation by intercepting solar radiation and shading neighbor- ing plants (Brandon et al., 2004). Large amounts of leaf area contribute to dense, light-limited under- story herbaceous canopies. Once established, monocultures of L. cuneata will limit light energy available to other plant species. Greater leaf area than native species also increases acquisition of

CO2, compensating for reduced photosynthetic rates compared to other native species (Allred et al., 2010). The effi cient use of biomass allocated to foliage within L. cuneata, compared to native species, can also aid in stressed environments or adverse conditions. Higher specifi c leaf area (more leaf area for less biomass investment) in L. cuneata may sustain growth and resource acquisition if resources become limiting (Allred et al., 2010). Continued 30 S.D. Fuhlendorf et al.

Box 2.1. – Continued Introduced primarily as a forage species, L. cuneata is highly nutritious and palatable in early growth stages, but grazers shift to other forage plants when palatability and digestibility decline with rapid plant maturation (Donnelly, 1954; Schutzenhofer and Knight, 2007). Condensed tannins also decrease digestibility and can cause gastrointestinal discontent in ruminants (Mosjidis, 1990). These herbivory avoidance traits reduce herbivore pressure on L. cuneata thereby providing it competitive advantage in species mixes, which assists in successful invasion. Simplifi ed disturbance regimes that promote uniform grazing distribution have aided in the spread of L. cuneata by providing an environment where grazing avoidance traits are a benefi t. Cummings et al. (2007) found that the invasion rate of L. cuneata was four times greater on grasslands in which the fi re–grazing interaction was not present than those in which it was present (Fig. 2.4). Under the interaction animals selected for fresh growth in burned patches and were able to maintain plants, including L. cuneata, in pala- table regrowth stages. Herbivory avoidance and competition traits characteristics of L. cuneata are not as benefi cial under intense focal grazing created by the fi re–grazing interaction. cover (%) cover Sericea lespedeza

Fig. 2.4. Sericea lespedeza invasion over time in the traditional management and patch-burn treatments (Cummings et al., 2007). In 1999, the treatments had similar levels (p>0.05) of invasion. By 2000, the Sericea invasion in the traditional management treatment was higher than in the patch-burn treatment. In the spring of 2003, the traditional treatment pastures were burned in entirety, so at that point both treatments had the same amount of fi re over the pasture. After this point, invasion in the traditional management treatment followed the same linear increase of 2% per year, while the patch-burn treatment displayed fl uctuating invasions. Modifi ed from Cummings et al. (2007).

Lespedeza cuneata is now widespread throughout the eastern portion of the Southern Great Plains where uniform cattle grazing is an objective. While control with herbicides is common (Koger et al., 2002), the use of new management strategies based on life history traits and historical disturbances can be more effective (Cummings et al., 2007) and economically and ecologically advantageous. In this case, restoring the pattern of historical disturbances for a site benefi ted native species and limited an invasive species. Linking Disturbance Regimes, Vegetation Dynamics and Plant Strategies 31

isolation of other disturbance eff ects. How- resistance to disturbance. Understanding ever, the eff ect of grazing is more complex the invasion of species requires an and is infl uenced by a myriad of factors (e.g. understanding of the interplay between number, distribution, and species of these traits and the environment in which herbivore; climate; resources; disturbances) the species is invading. Second, the beyond the simple removal of biomass. environment is constantly changing and Assuming that grazing and fi re patterns on many of these changes can enhance or small plots are similar to large complex inhibit the invasion process. We often use landscapes is obviously erroneous. Incor- historical landscapes and disturbance porating other factors, as well as interactions regimes as our baseline and frequently among factors, leads to a more appropriate, restoration of disturbance processes can and often diff erent, under standing of the minimize invasions. However, sometimes, system (Fuhlendorf and Engle, 2001). invasions are facilitated by historical Evaluation of invasive species literature leads disturbance and critical feedbacks may occur to similar conclusions where small-scale that promote even more invasions. Finally, experimental studies may do little to explain studies of vegetation dynamics and assembly large-scale landscape pat terns. Scientifi c rules of plant communities are relevant to understanding and manage ment of invasive invasive species ecology and management. plants on landscapes requires some Th e concepts of thresholds, non-linear innovative and alternative experimental dynamics, and ecosystem resilience are approaches. central to the dynamics of invasive species. In general, the approach used to study Prediction of invasion and management of interactions, processes, and traits related to these processes requires innovative invasive species is frequently simplifi ed to approaches that focus on multiple spatial understand current invasions and to predict and temporal scales. future invasions. Th is simplifi cation is often employed by studying single species in small and controlled environments (e.g. small plots) over short time periods. Although References studying species traits and processes at fi ne spatial and temporal scales informs Allred, B., Fuhlendorf, S., Monaco, T. and Will, R. our understanding of plant invasions, (2010) Morphological and physiological traits in eliminating natural complexity and variation the success of the invasive plant Lespedeza limits the application to managing invasions. cuneata. Biological Invasions 12, 739–749. Simple studies of plot-level herbicide Archer, S. (1995) Herbivore mediation of grass- applications can lead to conclusions that woody plant interactions. Tropical Grasslands 29, 218–235. cannot be extrapolated to large scales Archer, S.R. (1994) Woody plant encroachment (Fuhlendorf et al., 2009b). Although into southwestern grasslands and savannas: studying vegetation dynamics, ecosystem rates, patterns and proximate causes. In: Vavra, processes, and invasions at broader spatial M.P., Laycock, W.A. and Pieper, R.D. (eds) and temporal scales will inherently increase Ecological Implications of Livestock Herbivory complexity and variation, only actively in the West. Society for Range Management, restoring complexity will complete our Denver, Colorado, pp. 13–68. understanding of system dynamics. Archer, S., Schimel, D.S. and Holland, E.A. (1995) While species invasions are complex and Mechanisms of shrubland expansion: land use, climate, or CO2? Climatic Change 29, 91–99. extrapolation across scales and regions Axelrod, D.I. (1985) Rise of the grassland biome, should be done with caution, some central North-America. Botanical Review 51, generalizations are possible. First of all, 163–201. species invasions are best understood by Ball, D.M., Hoveland, C.S. and Lacefi eld, G.D. focusing on the species traits, such as (2002) Southern Forages. Modern Concepts for dispersal, establishment, and response and Forage Crop Management. Potash & Phosphate 32 S.D. Fuhlendorf et al.

Institute and the Foundation for Agronomic three invasive plant species. Biological Research, Norcross, Georgia. Invasions 11, 453–465. Barbosa, N.P.U., Wilson Fernandes, G., Carneiro, Clements, F.E. (1916) Plant Succession: An M.A.A. and Júnior, L.A.C. (2010) Distribution of analysis of the development of vegetation. non-native invasive species and soil properties Carnegie Institution of Washington, Washington, in proximity to paved roads and unpaved roads DC. in a quartzitic mountainous grassland of Conley, W., Conley, M.R. and Karl, T.R. (1992). A southeastern Brazil (rupestrian fi elds). computational study of episodic events and Biological Invasions 12, 3745–3755. historical context in long-term ecological Barnes, R.F., Nelson, C.J., Collins, M. and Moore, processes: climate and grazing in the Northern K.J. (2003) Forages: an Introduction to Chihuahuan desert. Coenoses 7, 55–60. Grassland Agriculture. Iowa State Press, Ames, Coppedge, B.R., Engle, D.M., Fuhlendorf, S.D., Iowa. Masters, R.E. and Gregory, M.S. (2001) Belsky, A.J. (1996) Viewpoint: Western juniper Landscape cover type and pattern dynamics in expansion: is it a threat to arid northwestern fragmented southern Great Plains grasslands, ecosystems? Journal of Range Management USA. Landscape Ecology 16, 677–690. 49, 53–59. Cummings, D.C., Fuhlendorf, S.D. and Engle, D.M. Belsky, A.J. and Blumenthal, D.M. (1997) Effects of (2007) Is altering grazing selectivity of invasive livestock grazing on stand dynamics and soils in forage species with patch burning more effective upland forests of the interior west. Conservation than herbicide treatments? Rangeland Ecology Biology 11, 315–327. & Management 60, 253–260. Blumenthal, D., Mitchell, C.E., Pyšek, P. and Davies, K.F., Holyoak, M., Preston, K.A., Offeman, Jarošík, V. (2009) Synergy between pathogen V.A. and Lum, Q. (2009) Factors controlling release and resource availability in plant community structure in heterogeneous invasion. Proceedings of the National Academy metacommunities. Journal of Animal Ecology of Sciences 106, 7899–7904. 78, 937–944. Brandon, A.L., Gibson, D.J. and Middleton, B.A. DeSantis, R.D., Hallgren, S.W., Lynch, T.B., Burton, (2004) Mechanisms for dominance in an early J.A. and Palmer, M.W. (2010) Long-term successional old fi eld by the invasive non-native directional changes in upland Quercus forests Lespedeza cuneata (Dum. Cours.) G. Don. throughout Oklahoma, USA. Journal of Biological Invasions 6, 483–493. Vegetation Science 21, 606–618. Briske, D.D., Fuhlendorf, S.D. and Smeins, F.E. Donnelly, E.D. (1954) Some factors that affect (2006) A unifi ed framework for assessment and palatability in sericea lespedeza, L. cuneata. application of ecological thresholds. Rangeland Agronomy Journal 46, 96–97. Ecology & Management 59, 225–236. Eddy, T.A. and Moore, C.M. (1998) Effects of Brooks, M.L., D’Antonio, C.M., Richardson, D.M., sericea lespedeza (Lespedeza cuneata Grace, J.B., Keeley, J.E., DiTomaso, J.M., (Dumont) G. Don) invasion on oak savannas in Hobbs, R.J., Pellant, M. and Pyke, D. (2004) Kansas. Transactions of the Wisconsin Effects of invasive alien plants on fi re regimes. Academy of Sciences, Arts, and Letters 86, BioScience 54, 677. 57–62. Brown, J. and Carter, J. (1998) Spatial and temporal Elton, C.S. (1958) The Ecology of Invasions by patterns of exotic shrub invasion in an Australian Animals and Plants. University of Chicago tropical grassland. Landscape Ecology 13, Press, Chicago, Illinois. 93–102. Epanchin-Niell, R.S., Hufford, M.B., Aslan, C.E., Burke, I.C., Lauenroth, W.K. and Coffi n, D.P. (1995) Sexton, J.P., Port, J.D. and Waring, T.M. (2010) Soil organic matter recovery in semiarid Controlling invasive species in complex social grasslands: implications for the conservation landscapes. Frontiers in Ecology and the reserve program. Ecological Applications 5, Environment 8, 210–216. 793–801. Fridley, J.D., Stachowicz, J.J., Naeem, S., Sax, D.F., Burkhardt, J.W. (1996) Herbivory in the Seabloom, E.W., Smith, M.D., Stohlgren, T.J., Intermountain West: An overview of evolutionary Tilman, D. and Holle, B.V. (2007) The invasion history, historic cultural impacts and lessons paradox: reconciling pattern and process in from the past. Idaho Station Bulletin 58. Idaho species invasions. Ecology 88, 3–17. Forest, Wildlife and Range Experiment Station, Fuhlendorf, S.D. and Engle, D.M. (2001) Restoring Moscow, Idaho. heterogeneity on rangelands: ecosystem Christen, D.C. and Matlack, G.R. (2009) The habitat management based on evolutionary grazing and conduit functions of roads in the spread of patterns. Bioscience 51, 625–632. Linking Disturbance Regimes, Vegetation Dynamics and Plant Strategies 33

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ecological effects of roads on terrestrial and and changes in land-cover in the Mediter- aquatic communities. Conservation Biology 14, ranean region. Global Change Biology 9, 18–30. 1234–1239. Twidwell, D., Fuhlendorf, S.D., Engle, D.M. and Walker, B.H., Ludwig, D., Holling, C. and Peterman, Taylor, C.A. (2009) Surface fuel sampling R.M. (1981) Stability of semi-arid savanna strategies: linking fuel measurements and fi re grazing systems. Journal of Ecology 69, 473– effects. Rangeland Ecology and Management 498. 62, 223–229. Wania, A., Kuhn, I. and Klotz, S. (2006) Plant Van Devender, T.R. and Spaulding, W.G. (1979) richness patterns in agricultural and urban Development of vegetation and climate in the landscapes in Central Germany – spatial southwestern United States. Science 204, 701– gradients of species richness. Landscape and 710. Urban Planning 75, 97–110. Vilà, M., Burriel, J.A., Pino, J., Chamizo, J., Llach, With, K.A. (2002) The landscape ecology of E., Porterias, M. and Vives, M. (2003) invasive spread. Conservation Biology 16, Association between Opuntia species invasion 1192–1203. Land-use Legacy Effects of Cultivation on Ecological 3 Processes

Lesley R. Morris

US Department of Agriculture, Agricultural Research Service, Utah State University, USA

Introduction species that accompany crop seeds as contaminants, thereby creating a wide Land-use legacies are long-lasting changes species pool of potentially invasive plants in ecosystems following human utilization (Mack, 1989). Th e connection between of resources. Cultivation for crop production cultivation, land abandonment, and exotic is known worldwide for creating land-use species invasion is increasingly relevant legacies that can persist for decades, because the extent of old fi elds is increasing centuries, and even millennia (Foster et al., worldwide (McLauchlan, 2006; Hobbs and 2003; McLauchlan, 2006). Th ese legacies are Cramer, 2007). important because they represent funda- Certainly, not all exotic species in old mental changes in basic ecosystem processes fi elds are detrimental and some even assist such as plant species reproduction and in the re-establishment of native species colonization, soil nutrient cycling and following disturbance, called ‘secondary availability, and soil water movement and succession’ (Lugo, 2004). However, exotic availability. When these basic processes are species can also arrest secondary succession interrupted, old fi elds can become more on old fi elds and dominate these sites diffi cult to manage and potentially impos- for decades to over half a century (Stylinski sible to restore to the pre-disturbance native and Allen, 1999; Elmore et al., 2006; plant community. Cultivation (plowing, Standish et al., 2007). Exotic annual grasses, seeding, and harvesting a crop annually) in particular, are known to arrest secondary involves both soil disturbance and the succession of old fi elds around the world sowing and harvesting of the intended crop. (Cramer et al., 2008). For example, Sac- When cultivation ceases, the abandoned charum spontaneum prevents tropical forest fi elds (‘old fi elds’ hereafter) are commonly regeneration in Panama, Avena barbata dominated by exotic and invasive plant limits the reestablishment of eucalypt species. Among the temperate grasslands of woodlands in southwestern Australia, and the world, for instance, the introduction and Bromus tectorum restricts shrubland spread of exotic species is consistently regeneration in the western USA (Mack, linked to the cultivation of crops such as 1989; Hooper et al., 2005; Standish et al., cereal grains, legumes, and forage grasses 2006). Exotic shrubs and forbs can also from western Europe (Mack, 1989). Th e crop dominate old fi elds even after attempts to species can become invasive in their new establish more desirable species through habitats as Sorghum halepense did in seeding (Doren et al., 1991; Stylinski and Argentina and Elettaria cardamomum did in Allen, 1999; Banerjee et al., 2006). Because Sri Lanka and south India (Heywood, 1989; cultivation involves several anthropogenic Mack, 1989). Cultivation also brings the disturbances coupled with introductions unintended introduction of exotic plant of exotic plant species, cultivation is

© CAB International 2012. Invasive Plant Ecology and Management: 36 Linking Processes to Practice (eds T.A. Monaco and R.L. Sheley) Land-use Legacy Effects of Cultivation on Ecological Processes 37

inextricably linked to invasive species pogenic, varies in magnitude, size, and management. In this chapter, I will identify frequency (Pickett and White, 1985; Myster, how the anthropogenic disturbances 2008). However, unlike natural disturbance, associated with cultivation diff er from cultivation also varies locally and regionally natural ones, how these exotic disturbances with human history. In fact, the current create land-use legacies through alteration composition of plant species on old fi elds of biotic and abiotic processes, and what may have more connection to historical land managers might do to help repair these use and management than current climate processes. or resource conditions (Motzkin et al., 1999; Buisson and Dutoit, 2004). Th us, under- standing old fi elds requires a holistic Linkages Between Cultivation and approach that encompasses ecological, Exotic Plant Invasion historical, and spatial variables in a new way (Benjamin et al., 2005). Disturbance Anthropogenic disturbance versus magnitude includes both intensity (amount natural disturbance of biomass destroyed and degree of soil disruption) and severity (biological eff ect) Ecologists recognize that disturbance (e.g. (Pickett and White, 1985). Cultivation can fi re, fl ood, and herbivory) is a natural have diff erent disturbance magnitude on ecosystem function (Pickett and White, adjacent old fi elds in the same soil type due 1985; Hobbs and Huenneke, 1992). Natural to historical diff erences in equipment and disturbance regimes are characterized by practices between farmers (Coffi n et al., their distribution, frequency, rotation 1996; Buisson and Dutoit, 2004). Dis- period, predictability, area disturbed, and turbance magnitude can also vary over time magnitude (or severity) (Pickett and White, because cultivation methods vary over time. 1985). Th ere is growing recognition that For example, cultivation intensity was anthropogenic disturbance is diff erent than greater with early conventional tillage natural disturbance because it can involve methods (continuous plowing) until con- both modifi cations to these natural dis- servation tillage (reduced plowing) was turbance regimes and/or the introduction of developed in the western USA (Schillinger new disturbances (e.g. soil excavation and and Papendick, 2008). Conversely, dis- plowing) (Stylinski and Allen, 1999). In turbance severity can be greater in more comparison to the natural disturbance recent times due to technological advances regimes under which ecosystems evolve, in plowing, fertilization, irrigation, and anthropogenic disturbances are introduced subsoil drainage practices (Doren et al., to the ecosystem and accompany changing 1991; Benjamin et al., 2005; Standish et al., human activities (McIntyre and Hobbs, 2006). 1999). Th erefore, anthropogenic disturbance Likewise, the size and frequency of regimes are unpredictably linked to eco- cultivation disturbance is aff ected by human nomic and social factors, rather than any history (Benjamin et al., 2005). For example, environmental cycle (Lugo, 2004). Culti- innovations in fertilization led to clearing vation involves several anthropogenic larger areas for cultivation in Australia disturbances including plowing, fertiliza- (Standish et al., 2006). Th e size of an old tion, and irrigation. Th ese cultivation dis- fi eld is important for native species because turbances are outside of the magnitude and colonization tends to be higher at the edges frequency of natural disturbance regimes and decrease into the center of fi elds and can result in modifi cation of the (Standish et al., 2007). Disturbance fre- ecosystem and irreversible legacy eff ects, quency depends on the historical duration such as species loss, because the environment of cultivation (how long it was in use) and can become mismatched with the species’ the age since abandonment (time since traits (Byers, 2002). cultivation ceased). Longer duration in Disturbance, whether natural or anthro- cultivation can produce more depleted 38 L.R. Morris

native seed banks, more need for control of community development in old fi elds if exotic species, more soil compaction, more vegetation is composed of native species erosion, and more organic matter loss with life histories that overlap, and thus can (Burke et al., 1995; Jenkins and Parker, compete with them for space and resources 2000; Standish et al., 2007). Although after abandonment (Meiners et al., 2007). exceptions have been reported, species For example, in tropical forests where exotic richness and similarity to the pre- species are not shade tolerant, they can disturbance plant community generally dominate an old fi eld until native trees close increase with time since abandonment the canopy (Fine, 2002). However, ecological (Motzkin et al., 1999; Öster et al., 2009). processes like nutrient cycling or soil water Ecosystem properties (e.g. system movement may be altered by cultivation to structure, resource base, and landscape an extent that old fi elds become foreign to characteristics) and plant community the adaptations that native species formed properties (e.g. life history and competitive over their evolutionary history (Byers, ability of resident plant species) can make 2002). Native species, therefore, may no some ecosystems more susceptible than longer have a prior-resident advantage over others to land-use legacies (Pickett and exotic species (Byers, 2002). Th ese altered White, 1985; Kruger et al., 1989; Cramer et ecological processes may now favor exotic al., 2008). Th e severity of cultivation species that evolved with frequent disturbances can be greater in resource cultivation disturbance (Byers, 2002). Th us, limited systems versus resource rich ones. exotic species from habitats with a long Invasion of exotic species is enhanced by history of cultivation can become successful disturbance when it increases the availability dominants in disturbed land where of a limiting resource (Hobbs, 1989). For cultivation has a shorter history (Hobbs and example, in light-limited forest systems, Huenneke, 1992). For example, exotic clearing land for cultivation elevates the annuals from Mediterranean regions that limited resource (i.e. solar radiation at the were historically cultivated for centuries soil surface) that assists invasion of shade- have successfully dominated the newly intolerant exotic species (Meiners et al., disturbed grasslands of California, USA 2002). Finally, landscape position plays a (Hobbs and Huenneke, 1992). Th e plant role in the severity of cultivation disturbance community composition of an old fi eld will (Fig. 3.1). For example, species diversity is largely depend upon the suite of life histories often higher in old fi elds when they are and adaptations found in the resident native adjacent to a native plant community vegetation (Hobbs and Huenneke, 1992). (Lawson et al., 1999; Hooper et al., 2004). In Th erefore, in old fi elds, generalizations contrast, old fi elds that are adjacent to about the alteration of ecological processes currently cultivated fi elds are subject to a and the ecological principles that govern persistent species pool of potentially their management are always relative to invasive plants and additional disturbances human history and inherent ecosystem and (e.g. pesticide and herbicide drift or fi re) plant community properties in which (Cramer et al., 2008). cultivation takes place (Kruger et al., 1989; Plant community properties include life Cramer and Hobbs, 2007). history and competitive ability of the resident species. Life history refers to an organism’s rate of growth, allocation of How Land-use Legacies are Created assimilates and structure, and timing of life cycle events (Pickett and White, 1985). How To explain how land-use legacies are created, a plant community responds to disturbance Cramer et al. (2008) proposed a stepwise depends on life history traits and competitive model of degradation in ecological processes ability of the resident plant species (see, created by cultivation (Fig. 3.2). As the James, Chapter 8, this volume). Th e presence duration, intensity, and extent of cultivation of exotic species may not greatly alter plant disturbance increases, biotic, and then Land-use Legacy Effects of Cultivation on Ecological Processes 39

(a)

(b)

Fig. 3.1. Aerial photographs depicting different landscape positions of old fi elds. (a) Old fi elds surrounded by native vegetation (photo from United States Geological Survey, 1957). (b) Old fi elds surrounded by active agricultural fi elds (photo from Google Earth, 2010). 40 L.R. Morris

Biotic Abiotic threshold threshold

FunctioningFunctioning

Ecosystem processes

DegradedDegraded

Low Magnitude and intensity

Fig. 3.2. Model of stepwise degradation in old fi elds (adapted from Cramer et al., 2008 with permission of Elsevier). As the duration, magnitude and intensity of cultivation increases, ecosystem function will become more degraded. Crossing biotic thresholds will require less intervention to repair function. Crossing abiotic thresholds will require greater intervention and investment, and may not return to pre- disturbance function. abiotic thresholds are crossed. Th resholds munity, which remains in a degraded state are defi ned by the ability of the plant beyond the timeframe of a single human community to recover without human lifetime (Cramer et al., 2008). Any attempt intervention. When a threshold is crossed, at restoration of old fi elds, therefore, the ecological processes cannot repair requires identifi cation of the key biotic and themselves, and human intervention is abiotic processes that have been com- required (Whisenant, 1999; Cramer et al., promised and evaluation of the potential for 2008). Biotic processes involve reproduction, their repair (Hobbs and Cramer, 2007). seed dispersal, seed banking, and com- Likewise, the magnitude, frequency, and petition; while abiotic processes involve size of cultivation disturbances are all physical characteristics of soils, soil fertility, important for assessing the extent of and hydrology. When a biotic threshold is damage to ecological processes, and thereby, crossed, secondary succession of the plant defi ning expectations and goals for how community may be altered and novel much energy and resources are applied to assemblages of exotic species become repair them. established. With biotic thresholds, there is potential to return to the original plant community, but the rate of return is slowed. Biotic processes and thresholds When the duration, intensity, and severity of cultivation cross both the biotic and the Th resholds controlled by biotic processes abiotic thresholds, return to the historical require manipulation of vegetation for plant community is further stalled and the recovery (Whisenant, 1999). In old fi elds, changes may be irreversible. In these cases, biotic thresholds are governed by biological succession is arrested and old fi elds can processes such as reproduction, seed remain with a persistent exotic plant com- dispersal, seed banking, and resource com- Land-use Legacy Effects of Cultivation on Ecological Processes 41

petition (Cramer et al., 2008). When tain more species that reproduce vegetatively thresholds for these biotic processes are than larger old fi elds (Fernández et al., crossed, the old fi elds can become repro- 2004). ductively limited, become dispersal limited, and have altered competitive interactions Seed limitations can favor exotic plants for both resource exploitation and establish- ment space. Crossing thresholds associated Seed banking and dispersal are two processes with any or all of these biotic processes has that contribute to seed limitation on old the potential to favor exotic over native fi elds. In order to form a store in the soil, plant species. seeds must survive beyond the duration and intensity of cultivation (Standish et al., 2007). Native species of environments Reproductive limitations can favor exotic without frequent disturbance do not plants typically form long-lasting seed banks, or Both sexual and asexual reproduction of their seed banks are depleted by longer and native plant communities can be restricted higher cultivation duration and intensity, due to historical cultivation. For example, respectively (Cramer et al., 2008). Exotic pollinator populations may have declined crop and pasture weeds, on the other hand, under historical pesticide use, drift from form persistent soil seed banks that are nearby agricultural fi elds, or be inhibited by likely to dominate the seed pool long after the level of fragmentation in the landscape fi elds are abandoned (Ellery and Chapman, around old fi elds (Hobbs and Yates, 2003; 2000; Buisson and Dutoit, 2004; Cramer et Krug and Krug, 2007). If exotic plant species al., 2007). Th erefore, seed banks in old fi elds are wind pollinated, this may only become a have often changed from predominantly limiting factor to the native plant com- native to exotic species. Furthermore, some munity. Further, because cultivation crop weeds remain in the seed bank even removes entire plants, species that reproduce though they are not present in the above- primarily vegetatively are less likely to ground vegetation because the conditions occupy old fi elds (Fernández et al., 2004; are no longer conducive for their growth Standish et al., 2007; Dyer, 2010; Morris et (e.g. plowing), but their seed remains al., 2011). Old fi elds with longer duration available for germination and emergence and higher intensity of cultivation tend to once soil is disturbed again (Buisson and have few if any vegetatively reproducing Dutoit, 2004; Banerjee et al., 2006). species because soil cultivation destroys Ultimately, altered seed bank composition most of the root system (Jenkins and Parker, and the specifi c germination requirements 2000). Th is life history trait may make these of exotic species can arrest secondary species less able to compete for establishment succession on old fi elds to pre-disturbance space with exotic species that reproduce plant community composition (Hobbs and primarily by seed. Th erefore, plant com- Huenneke, 1992). position can be altered from a diverse com- When seed banks and vegetative repro- munity of sexual and asexually reproducing ductive parts have been destroyed, the species to a less-diverse one dominated by secondary succession of old fi elds to native species that rely primarily on seed for species is heavily dependent upon native establishment (Fernández et al., 2004; seed dispersal (Hooper et al., 2004). Morris et al., 2011). Exotic species typically However, seed dispersal of native species is rely upon and produce more seeds than often limited in old fi elds. Wind-dispersed native plant species (Mason et al., 2008). species are sometimes the few native plants However, size and landscape position can to consistently reestablish in old fi elds also interact with this pattern. For example, (Standish et al., 2006; Morris et al., 2011). smaller old fi elds in the Grazalema Wind-dispersed seeds are typically more mountains of southern Spain, that are likely to be carried into old fi elds than those surrounded by forests and shrublands, con- reliant on animals, whose populations may 42 L.R. Morris

have declined due to habitat fragmentation, resource exploitation and establishment land clearing, or historical pesticide use and space to exotic species over the native pest control (Hobbs and Yates, 2003; Dyer, species that have evolved life history traits 2010). Plants with generalized seed dispersal under conditions without frequent and mechanisms are more likely to provide intense disturbance (Hobbs and Huenneke, greater seed rain into an old fi eld (Fine, 1992). Th e competitive eff ects of exotic 2002). However, plant species abundance in grasses and forbs create an important the surrounding landscape can be equally as barrier to native plant reestablishment in important as dispersal mode in secondary old fi elds (Hartnett and Bazzaz, 1983; Holl succession (Lawson et al., 1999; Ruprecht, et al., 2000; Riege and del Moral, 2004; 2006). Native species seed dispersal into old Standish et al., 2007). Exotic grasses fi elds usually declines with distance from compete with tree seedlings for moisture intact vegetation despite the diversity of and nutrients and decrease the performance dispersal mechanisms (Hooper et al., 2004; of native woody species (Hooper et al., Cramer et al., 2008). Th e exceptions are 2002). Exotic grasses can also diff er from when animals assist in long distance seed the native community in root and shoot dispersal or when remnant trees are found morphology, phenology, and water uptake in the old fi elds (Hooper et al., 2004; Ganade, due to diff erent photosynthetic pathways 2007). Dispersal limitation is more likely to (e.g. C3 versus C4 grasses) and other traits aff ect late successional species, whose seeds traits associated with life form (e.g. annual tend to be larger and disperse via animals, versus perennial) (Mack, 1989; Davis et al., than early successional species, which have 2005). All of these factors enable exotic smaller seeds that spread easily (Bonet and grasses to successfully compete with native Pausas, 2004; Santana et al., 2010). Large- species and arrest secondary succession in seeded species can also become dispersal many ecosystems. For example, the exotic limited because the majority of animals in perennial grass Saccharum spontaneum ssp. old fi elds are smaller (e.g. birds or bats) and spontaneum inhibits forest regeneration on carry smaller seeds (Hooper et al., 2005). old fi elds in Panama (Hooper et al., 2005). Consequently, many herbaceous species Similarly, competition with the exotic with limited or unassisted animal dispersal rhizomatous grass Agrostis giganta was the can be under-represented in old fi elds (Dyer, primary barrier to native tree colonization 2010). Seed dispersal plays an important of old fi elds in rainforests of the Pacifi c role in determining the trajectory of the Northwest, USA (Riege and del Moral, plant community within old fi elds. Plant 2004). In the arid western USA, the communities are more likely to remain in a exotic annual grass Bromus tectorum has persistent degraded state where native seed replaced the native perennial bunchgrasses dispersal is limited and there are few on old fi elds where it can persist for over dispersal mechanisms (Cramer et al., 2008). half a century after cultivation ceased Conversely, exotic species have an increased (Daubenmire, 1975; Elmore et al., 2006). In likelihood of long distance dispersal over old fi elds in Australia, invasive annual native species (Mason et al., 2008). grasses like Avena barbata have remained dominant for up to 45 years after cultivation (Standish et al., 2007) (Fig. 3.3). Altered competitive interactions can favor Cultivation can also alter competitive exotic plants interaction for resource exploitation by As described above, many of the exotic altering the soil microfl ora and fauna. Soil species that invade old fi elds evolved under microbes are fundamental in decomposition, harsh conditions with frequent anthro- nutrient cycling, and nutrient availability pogenic disturbances, like cultivation, which for plants, which can all infl uence plant means they have adapted life history traits secondary succession by aiding plants to for disturbance tolerance (Mack, 1989). Th is acquire soil nutrients (Biondini et al., 1985; often confers a competitive advantage for Paul and Clark, 1989). Cultivation and Land-use Legacy Effects of Cultivation on Ecological Processes 43

(a)

(b)

Fig. 3.3. Photographs of old fi elds and exotic species. (a) Old fi eld in southern Florida (photo by Joy Brunk, Everglades National Park), USA, where Schinus terebinthifolius (on the left) has excluded the native sawgrasses (in foreground). (b) Old fi eld in Western Australia where Avena barbata (in foreground) has arrested succession of the eucalypt woodland (on the right, foreground) (reproduced from Cramer et al., 2008 with permission from Elsevier). fertilization can reduce abundance and years to recover in the Central Plains of the diversity of soil microbial communities, USA. However, recovery of microbial which also require decades to centuries to populations may be heavily infl uenced by recover (Lovell et al., 1995; Buckley and plant species composition and abundance as Schmidt, 2001; Steenwerth et al., 2002). For well as time since abandonment. For example, Burke et al. (1995) reported that example, while some have found decreasing soil microbial biomass between cultivated microbial abundance with time since and noncultivated soils required at least 53 abandonment, others found old fi elds with 44 L.R. Morris

dominance of exotic annual grasses to have original plant community is greater than similar soil microbial biomass regardless of when there is additional degradation of time since abandonment (Steenwerth et al., abiotic processes, but the rate of recovery is 2002; Kulmatiski and Beard, 2008). One slowed (Cramer et al., 2008). Whether or not study suggested that cultivation increased to intervene when a biotic threshold is non-native plant establishment because it crossed depends upon the amount of change decreased microbial abundance, not because from the pre-disturbance plant community it increased resource conditions (Kulmatiski that society will accept, and how much time and Beard, 2008). Although the mechanism is needed for its return (Cramer et al., 2008). was unclear, presumably the advantage from For example, even without dominance by plant–microbe feedbacks was benefi tting exotic species, unassisted secondary suc- the exotic plants over the native species cession to pre-disturbance calcareous grass- (Klironomos, 2002; Kulmatiski and Beard, land communities on old fi elds in Europe is 2008). estimated to take over 50 years (Fagan et al., Cultivation management practices (e.g. 2008; Öster et al., 2009). pesticides, herbicides, and fertilizers) have Although it may seem that seed limitation also been shown to reduce mycorrhizal can be overcome simply by adding more populations and community structure in seeds of the native species, there are many soils around the world (Douds and Millner, other environmental and biological factors 1999; Buckley and Schmidt, 2001; Oehl et that infl uence their establishment. Sown al., 2003). Because mycorrhizal associations species can establish in old fi elds, but can improve mineral nutrition, water recruitment from the seedings is generally absorption, and drought tolerance in plants, lower (Öster et al., 2009; L.R. Morris and their reduction can infl uence secondary T.A. Monaco, unpublished results). Arid succession in old fi elds (Janos, 1980; Allen areas, in particular, have had limited success and Allen, 1984; Auge, 2001). Plant– in establishing native plant communities on mycorrhizal relationships can confer com- old fi elds with reseeding (Roundy et al., petitive advantages to some plant species in 2001; Banerjee et al., 2006). Many of the low-nutrient environments where they are failures are blamed on low precipitation (see adapted to acquire nutrients through Hardegree et al., Chapter 6, this volume). mycorrhizal associations (Allen and Allen, Successful revegetation of old fi elds in 1984). However, invasive and native ruderal extremely arid areas of the Intermountain species that colonize old fi elds tend to be Region, USA, have been improved by using non-mycorrhizal (Reeves et al., 1979; Janos, irrigation along with seeding (Roundy et al., 1980; Kulmatiski and Beard, 2008). A legacy 2001). However, seeding is still considered of high nutrient concentrations from an unreliable method in arid regions, fertilization in old fi elds, particularly with especially when seed banks of exotic species phosphate, can reverse the competitive compete with seed banks of native species at advantage to favor non-mycorrhizal exotic all times of the year, regardless of irrigation species (Standish et al., 2006). Th ere is (Banerjee et al., 2006). Furthermore, increasing interest in understanding and granivore preference for native seeds over using mycorrhizal associations in restor- exotic ones as well as preferential herbivory ation; however, higher levels of infection of seedlings has direct and indirect eff ects with mycorrhizal fungi are not always linked on native species recruitment (Lawson et al., with higher levels of native vegetation 1999; Holl et al., 2000; Riege and del Moral, recovery on old fi elds (Richter et al., 2002). 2004; MacDougall and Wilson, 2007; Standish et al., 2008). In an eff ort to reestablish native Recovery of biotic processes vegetation on exotic-plant dominated land- scapes (including old fi elds), repair strategies When only a biotic threshold has been are often employed that include additional crossed, the potential to return to the exotic disturbances (e.g. plowing, disking, Land-use Legacy Effects of Cultivation on Ecological Processes 45

drill seeding) (Whisenant, 1999). Although Abiotic processes and thresholds considerably less research has examined the legacy of the exotic disturbances associated Th resholds controlled by abiotic processes with these repair strategies on ecological require physical manipulation of the soils processes, they may have similar con- for recovery (Whisenant, 1999). Cultivation sequences as cultivation (DePuit and practices often involve creating a uniform Redente, 1988; Kettle et al., 2000). For soil environment both physically (e.g. land example, because exotic plants can create leveling) and in nutrients (e.g. fertilization) persistent seed banks, the most eff ective to maximize crop production (Homburg and way to manage the exotic plant population Sandor, 2011). In old fi elds, these practices may be to manage seed banks (Marushia and impact abiotic processes associated with Allen, 2011). A typical method to manage an soil structure, surface microtopography, exotic species seed bank is through annual nutrient distribution, nutrient content, and disking (or plowing), which repeatedly mixes hydrology. As with biotic processes, crossing the soil and breaks down its stability and a threshold associated with any of these structure (Whisenant, 1999). Th is repair abiotic variables also has the potential to strategy may further degrade the abiotic favor exotic over native plant species. processes in the old fi eld past abiotic thresholds. Soil structure and microtopography can favor Th erefore, many researchers have begun exotic plants to call for more innovative methods that take advantage of the facilitative Where an abiotic threshold has been crossed, interactions among plants and emphasize the establishment niche of the native species using species that stabilize and modulate pool may no longer exist (Cramer et al., 2008). ecosystem function, called ‘foundation Cultivation can increase soil com paction (i.e. species’ (Buisson and Dutoit, 2004; Ellison bulk density) and soil surface penetration et al., 2005; Banerjee et al., 2006; Prevéy et resistance that create physical barriers to al., 2010). Planting patches of woody native seedling emergence and impede root growth species in old fi elds, for example, can (Unger and Kaspar, 1994). For example, facilitate seed dispersal and recruitment cultivation can create a plow pan (a compacted via animals as well as serve as catchments layer 20–25 cm beneath the soil surface) that for wind-dispersed seed (Holl et al., 2000; restricts root growth (Buschbacher et al., Buisson and Dutoit, 2004; Hooper et al., 1988; Uhl et al., 1988). A plow pan may be 2005). Th is strategy enables foundational less encumbering on exotic annuals with species to facilitate secondary succession shallow root systems. Soil compaction has by ameliorating harsh climatic conditions been shown to favor exotic species in old while concentrating nutrients under their fi elds in eastern US forests, possibly because canopies (Allen, 1988; Holl et al., 2000; the exotic species were horticultural intro- Banerjee et al., 2006; Prevéy et al., 2010). ductions for ground cover on highly disturbed Planting desirable species directly under sites (e.g. rock pathways) with traits allowing the canopy edge of trees and shrubs has them to thrive in compacted soils (Parker et also been shown to enhance their growth al., 2010). In contrast, Kyle et al. (2007) due to higher mycorrhizal infection, less reported soil compaction on old fi elds in competition from exotic grasses under western US steppe was negatively correlated shaded conditions, and increased moisture with exotic species while native species were and nutrients (Holl et al., 2000). Some unaff ected. Th is advantage could be related to suggest that only a few foundational enhanced root growth of the exotic species in species may need to be planted to initiate loosened soil, which allows more access to soil recovery of species diversity in old fi elds resources earlier in its life cycle (Kyle et al., because of their critical role in structuring 2007). the plant community (Buisson and Dutoit, Leveling of the soil surface and reducing 2004). small-scale heterogeneity also has been 46 L.R. Morris

implicated in reducing seedling establish- al., 1993; Robertson et al., 1993). Th is eff ect ment by decreasing the occurrence of is further enhanced with the application of depressions for water capture (Whisenant et fertilizers (Standish et al., 2006). Soil al., 1995). Cultivation also can reduce the homogenization on old fi elds due to number of safe sites for seed germination cultivation can contribute to the dominance and establishment. Th ere may be as many as of early successional species, like exotic 15-fold more wind-dispersed seeds than annual species (Robertson et al., 1993; vertebrate-dispersed seeds in an old fi eld, Standish et al., 2006). Homogeneous soil but the density of established wind- resources can favor exotic plant species that dispersed seed can be lower due to microsite tolerate regularly high competition for limitations (Ingle, 2003). Flinn (2007) resources across the landscape as opposed to showed that small scale heterogeneity of native species that may be more competitive post-cultivation forest fl oors had limited when resources are patchy (Hutchings et al., suitable sites for native fern establishment, 2003; Standish et al., 2006). and sowing the seed had little eff ect. Th e loss of microtopographic heterogeneity Soil nutrient availability can favor exotic may represent a larger limitation to seed- plants ling establishment or plant survival than soil quality in some systems (Flinn, 2007). Old fi elds often have altered soil organic However, there are fewer studies on carbon and soil nutrients in comparison to establish ment limitation connected to similar noncultivated sites (McLauchlan, abiotic properties of the soil, perhaps 2006). Even prehistoric old fi elds in because so many old fi elds are already known southwestern North America have been to be seed limited, or because individual found to contain lower soil organic carbon species interact diff erently depending upon 1000 years after cultivation ceased (Sandor their life history traits (Bakker and Berendse, et al., 1986). Th e loss of soil organic matter is 1999; Flinn and Vellend, 2005; Standish et linked to further degradation in soil al., 2007). structure; namely increased compaction that restricts soil water movement, nutrient cycling, and microbial activity (Whisenant, Soil nutrient distribution can favor exotic 1999; McLauchlan, 2006; Homburg and plants Sandor, 2011). Soil nutrients can be lost Cultivation practices not only homogenize through annual crop harvesting, reduction the soil physically, but also homogenize soil of soil organic material, and loss of top- resources, and patterns of nutrient cycling soil from wind and water erosion. In across old fi elds that last for decades contrast, soil organic carbon and nutrients (Robertson et al., 1993; Standish et al., can be increased in old fi elds through 2006). Homogenization is more profound the application of either organic (e.g. in nutrient-limited ecosystems where soil manure and ash) or inorganic fertilizers nutrients tend to be patchy in the (McLauchlan, 2006). Th e introduction and environment and are concentrated under use of synthetic fertilizers played a role in shrubs in zones known as ‘islands of expanding cultivation agriculture into areas fertility’ (Bochet et al., 1999; Whisenant, where it would not have been possible 1999). Th ese islands of fertility are otherwise, and where legacies are often mechanically homogenized and diff used greater (Li and Norland, 2001; Standish et across the whole soil surface during al., 2006; Cramer et al., 2008). Still, without cultivation (Charley and West, 1975). On a fertilization, long-term diff erences in soil landscape scale, processes like nitrogen nutrients can remain for decades to centuries mineralization may be similar between old (McLauchlan, 2006; Homburg and Sandor, fi elds and noncultivated land but their 2011). Even when ecosystems are gaining distribution is altered because the spatial soil organic matter, the rate of recovery does aggregation has been removed (Bolton et not always match the rate of loss (Ihori et al., Land-use Legacy Effects of Cultivation on Ecological Processes 47

1995). Remarkably, the recorded losses in Norland, 2001). Th e exotic and invasive organic matter, nitrogen, and phosphorus shrub, Schinus terebinthifolius, responds on a mass concentration basis can be similar more favorably to the additional P found in between soils of old fi elds abandoned for abandoned rock-plowed fi elds where it centuries and modern old fi elds (Homburg competitively excludes the native sawgrass and Sandor, 2011). In addition, the species, Cladium jamaicense, which is more agricultural crop itself can infl uence soil- adapted to low nutrient soils (Li and nitrogen levels because some nitrogen-fi xing Norland, 2001). Many exotic species plants (e.g. legumes) can add nitrogen to the competitively exclude native species when soil (McLauchlan, 2006). Likewise, the soil nutrients are elevated, but increased vegetation composition following abandon- fertility does not always confer a competitive ment, native or exotic, can infl uence both advantage to exotic species. For example, on the retention and loss of nutrients in the old fi elds in Australia with residually elevated soil (Lugo, 2004). P levels, reduction of P was not enough to Fertilization can elevate soil nutrients for manage the exotic Avena barbata because it decades to centuries (Dupouey et al., 2002; could still obtain more P than the native McLauchlan, 2006; Standish et al., 2008). species at all levels (Standish et al., 2008). Fertilization of old fi elds usually elevates the Th ese alterations in soil fertility can soil nutrients that are typically limiting in change nutrient cycling and nutrient uptake the native system, most typically nitrogen by plants, conditions which can be created (N), phosphorus (P), and potassium (K). and perpetuated through feedbacks by the However, the potential for P retention is exotic plants themselves (see Eviner and much higher than N because it has a greater Hawkes, Chapter 7, this volume). Th ere is sorption in many soils, and fewer loss debate over whether the plant–soil pathways (e.g. leaching) than N (McLauchlan, relationships found in exotic-dominated old 2006; Grossman and Mladenoff , 2008). fi elds are due to land-use legacies or plant- Indeed, elevated P legacies have been found growth infl uences (Kulmatiski et al., 2006). in Roman agricultural terraces from AD 50 In other words, are exotic species passengers to 250 (Dupouey et al., 2002). Elevated or drivers of the increases in soil fertility nutrient content can make it easier for (MacDougall and Turkington, 2005)? It can exotic species to colonize and compete with be diffi cult if not impossible to separate native species (Li and Norland, 2001; these feedbacks in scientifi c studies (Woods, Zimmerman et al., 2007; see Grant and 1997). However, there is some literature Paschke, Chapter 5, this volume). For that highlights the role of the cultivation example, some old fi elds in Puerto Rico have disturbance in altering soil fertility, and maintained exotic species dominance for 22 thereby community composition, because years due to altered soils from liming soil structural and nutrient modifi cations practices (Lugo and Brandeis, 2005; are also found in old fi elds that are not Zimmerman et al., 2007). Nutrient addition dominated by exotic species (Motzkin et al., in combination with soil disturbance can 1996; McLauchlan, 2006). Physical soil increase the establishment and growth of disturbance without nutrient addition may exotic species more than nutrient addition or may not increase invasion potential alone (Hobbs and Atkins, 1988; Li and (Hobbs, 1989). Hobbs (1989) found that soil Norland, 2001). For example, in the disturbance promoted the establishment of southern coastal state of Florida in the USA, exotic species, but their performance areas were opened for cultivation by the increased more when soil disturbance was introduction of synthetic fertilizers and combined with nutrient addition. As rock-plowing, which crushed the natural discussed above, ecosystem and plant oolitic limestone bedrock to create a growth community properties may determine which medium and increase soil depth for vegetable has greater infl uence (Pickett and White, production (Doren et al., 1991; Li and 1985; Hobbs, 1989). 48 L.R. Morris

Altered hydrology can favor exotic plants (Cramer and Hobbs, 2002; Standish et al., 2006). Th e replacement of deep-rooted Th e land-use legacies of cultivation also plants with shallow-rooted annual crops include changes in hydrological processes increased groundwater recharge and raised such as soil moisture avaliability, soil water- the water table, bringing salts to the surface holding capacity, runoff and infi ltration, and (Whisenant, 1999). Concentrations of salt groundwater fl ows. Cultivation legacies can on soil surfaces reduced infi ltration by have a greater eff ect on diff erences in soil dispersing clays, which further reduced water movement between plowed and porosity of the soil (Whisenant, 1999). Th is unplowed sites than the diff erences in soil legacy not only threatens the agricultural water movement between two diff erent soil output and restoration potential of the series (Schwartz et al., 2003). In fact, soil fi elds, but also the remnant vegetation in hydraulic conductivity can remain aff ected the surrounding landscape (Cramer and for well over 25 years after cultivation Hobbs, 2002). Similarly, the soil surfaces of ceases, and such alterations may be very abandoned sugarcane fi elds in Puerto Rico, diffi cult to restore to pre-disturbance that were created out of drained wetlands, function (Fuentes et al., 2004). Water are at a lower elevation than the surrounding availability can also be reduced by soil forest due to burning off of the peat for compaction in old fi elds (Standish et al., sugarcane harvests. Th is has led to longer 2006). Plowing has been shown to reduce periods of fl ooding and higher risk of infi ltration rates (Giff ord, 1972; Tromble, saltwater intrusion (Zimmerman et al., 1980). Recovery potential, for infi ltration 2007). As I mentioned previously, when rates on old fi elds with livestock grazing, is exotic disturbances are outside of the lower than is predicted for livestock grazing magnitude and frequency of natural alone (Giff ord, 1982). Finally, irrigation is disturbance regimes, they can result in implicated in both raising water tables and modifi cations with irreversible legacy lowering them during intensive pumping eff ects, such as native species loss, because (Whisenant, 1999; Elmore et al., 2006; the environment becomes mismatched with Cramer et al., 2007). the species’ traits (Byers, 2002). Sometimes Altered hydrology can impact secondary exotic species are the only plants in the plant community succession on old fi elds at community that are adapted to these new the local and the landscape scale. Ground- environments. water pumping can lower the water table under which the native plant community was dependent, inviting exotic plant species Recovery of abiotic processes to colonize and dominate old fi elds that are more dependent upon precipitation, and Once both the biotic and abiotic thresholds therefore modify the entire hydrologic cycle have been crossed, restoration involves in soils (Elmore et al., 2006). Compacted remediation of the physical environment, soils in old fi elds have been blamed for poor which requires intense eff ort and increased plant germination and establishment due to capital (Whisenant, 1999; Cramer et al., less pore space for water movement, root 2008). For example, where soil structure and growth, and water storage (Cramer et al., fertility had been altered in the rock-plowed 2007; Hillhouse, 2008). At the landscape areas of southern Florida, USA, the most scale, irrigation can also alter soil hydrology. eff ective way to manage the invasive shrub Irrigation can cause soil damage through Schinus was to completely remove topsoil (Li anthropogenic salinization (Whisenant, and Norland, 2001). Restoration that 1999; McLauchlan, 2006). A dramatic prevents reinvasion by exotic species is a example of this is the wheatbelt region of lengthy process with uncertain outcomes western Australia where expeditious land and risks. Even with intense eff orts, return clearing for cultivation raised the water to the pre-cultivation plant community and tables and increased surface soil salinity ecosystem state may not be feasible Land-use Legacy Effects of Cultivation on Ecological Processes 49

(Standish et al., 2008). For example, when be maintained by exotic plant–soil feedbacks wetlands or groundwater have been drained, (Kulmatiski, 2006; Whisenant, 1999). it may not be realistic, nor is it always Th erefore, in order to take advantage of suffi cient to re-water these areas because foundational species infl uence, it may be soil properties (e.g. pH, salinity, and nutrient more eff ective to plant patches containing levels) have also been altered (Bakker and foundational species and other native Berendse, 1999; Cramer et al., 2007). Clearly, seedlings (Holl et al., 2000; Hooper et al., abiotic and biotic processes are linked 2005; Banerjee et al., 2006; King and Hobbs, through feedbacks, but there must be a 2006). In addition to ameliorating the baseline of primary abiotic structure and environment for other desired plants, function in order for biotic process to be foundational species may improve the reestablished (Whisenant, 1999; King and function of compromised abiotic processes Hobbs, 2006). like water infi ltration, nutrient cycling, and Attempts to directly repair altered soil soil structure (King and Hobbs, 2006). nutrient availability have met varied success Few studies have examined how land-use (Blumenthal et al., 2003; Corbin and legacies in soil aff ect recovery in ecosystem D’Antonio, 2004). When soil nitrogen is processes during restoration (Anderson, elevated, strategies to reduce it have included 2008). More research, and explicit con- manipulations of the C:N ratio of soils via sideration of a priori legacies in soils, is organic carbon additions (e.g. sawdust) to needed and will help improve restoration promote microbial immobilization of N eff orts (Callaham et al., 2008). Under- (Blumenthal et al., 2003). Similarly, where standing the initial soil conditions at time of systems have elevated P, N addition has been abandonment (e.g. nutrient levels and used to lower the P:N ratio (Gough and mycorrhizal abundance) is an important Marrs, 1990; Fagan et al., 2008). Others fi rst step for increasing restoration success have found that simply seeding desired because these initial conditions can infl uence species into areas with elevated soil fertility the trajectory of secondary succession more is more eff ective than attempting to reduce than the restoration treatment (Anderson, soil fertility with straw mulch (Kardol et al., 2008). 2008). In some cases, plant species not included in the pre-disturbance community may be more suited to the new conditions at The Debate about When and How to the site, such as plants with functional traits Intervene and life histories more similar to the dominant exotic plants (Whisenant, 1999). It is important to acknowledge the global Another approach to repairing abiotic debate about whether old fi elds dominated processes involves using biotic intervention by exotic species should be restored back to with foundational species that exert control native plant communities, back to cultural over abiotic processes (Buisson and Dutoit, agricultural landscapes, or if there should be 2004; Ellison et al., 2005; Banerjee et al., intervention at all (Foster et al., 2003; Marty 2006; King and Hobbs, 2006; Prevéy et al., et al., 2007; Hobbs et al., 2009). Intervention 2010). Both carbon and nitrogen pools can depends upon the length of delay in recovery recover, although very slowly, in cultivated and amount of divergence of the old fi eld soils once perennial grasses replace the from the historical plant community that is annual species that dominate in initial deemed acceptable by society (Cramer et al., stages of abandonment (Burke et al., 1995). 2008). Th ere are also considerations about However, the traditional agronomic ‘if the cure is worse than the disease’ when it approaches to native plant restoration (e.g. comes to removing exotic species, especially landscape-wide herbicide application, when the management techniques involve plowing, and seeding) are unlikely to work additional exotic disturbance (Zavaleta et in exotic plant-dominated communities al., 2001; Hobbs et al., 2009). For example, where abiotic processes are altered and may removal of unwanted exotic species may not 50 L.R. Morris

necessarily bring the ecosystem back to its secondary succession on old fi elds when pre-disturbance condition, and may move it they are better adapted to cultivation toward a less desirable one (Pierson et al., disturbances than native species (Mack, 2007; Seastedt et al., 2008). In fact, there is 1989; Hobbs and Huenneke, 1992). a risk that compounded disturbances will Cultivation legacies can aff ect both biotic lead to longer-term alterations and new and abiotic processes of ecosystems and can, ecological surprises (Paine et al., 1998). therefore, infl uence every aspect of Unfortunately, the legacies of management management and restoration. Management and repair strategies have seldom been and restoration of old fi elds requires studied explicitly, even though some of indentifying the suite of ecological processes them involve further cultivation dis- that have been compromised – biotic, turbances that can also alter the same biotic abiotic, or both (Hobbs and Cramer, 2007). and abiotic processes discussed in this Biotic processes, such as reproduction and chapter (DePuit and Redente, 1988; Kettle seed dispersal, will be diffi cult to repair in et al., 2000). Still, even when there is a old fi elds, but less diffi cult than abiotic ones. reintroduction of historically natural Old fi elds that are relatively small and disturbances, the altered systems on old surrounded by native plant communities fi elds can remain relatively unchanged may have a better chance for unassisted (Motzkin et al., 1996). succession to take place if an abiotic However, not intervening in secondary threshold has not been crossed, although it succession is not strongly favored either will likely proceed slowly. Abiotic thresholds, (Standish et al., 2007). Furthermore, the such as changes in soils that infl uence reasons for and the goals of restoration nutrient uptake in plants, may not be eff orts are not always the same worldwide reversible with repair strategies, especially and will depend upon social values as well as at large scales (Cramer et al., 2008). It may economic needs (Hobbs et al., 2009). For be necessary to consider using plants with example, some may seek to restore old fi elds traits that are more adapted to the new to the pre-disturbance plant community for conditions at the site, even if they are not a historical conservation and biodiversity part of the pre-disturbance community while others may want more palatable forage (Whisenant, 1999). Th e level of required than the exotic species off er so the land can intervention increases when more than one be used for grazing (Marty et al., 2007; process is compromised and when several Hobbs et al., 2009). In some places, like the processes interact to create feedbacks tropics, reestablishing trees in old fi elds (Hobbs and Cramer, 2007). Because diff erent plays an important role in providing processes may be interacting simultaneously fast-growing carbon sinks for carbon at diff erent sites, eff ective strategies will sequestration while in other places it is more have to be site specifi c (Cramer et al., 2008). about human survival (Hobbs and Cramer, Old fi elds containing exotic species can 2007; Hobbs et al., 2009). Th erefore, be hybrid systems where some of the plant intervention in secondary succession on old species and ecological processes of the pre- fi elds can have a wide variety of goals, which cultivation system remain; or they can be also infl uences how and what types of repair completely novel systems with funda- strategies will be used. mentally new species assemblages and altered biotic and abiotic processes (Seastedt et al., 2008; Hobbs et al., 2009). Hybrid Conclusion systems, where both biotic and abiotic thresholds have not been crossed, may have Cultivation involves exotic disturbances more potential for restoration to the pre- that leave land-use legacies for decades, disturbance system, or at least its structure centuries, or millennia (Foster et al., 2003; and function (Hobbs et al., 2009). Novel McLauchlan, 2006). Exotic species can systems, where both biotic and abiotic become a long-term problem and arrest thresholds have been crossed, have less Land-use Legacy Effects of Cultivation on Ecological Processes 51

potential for return to pre-cultivation they relate to soil disturbance and soil biological function. Such novel systems may require activity. Vegetatio 60, 25–36. creative and innovative restoration alter- Blumenthal, D.M., Jordan, N.R. and Russelle, M.P. natives that do not include a return to the (2003) Soil carbon addition controls weeds and facilitates prairie restoration. Ecological historical ecosystem (Hobbs et al., 2009). Applications 13, 605–615. Bochet, E., Rubio, J.L. and Poessen, J. (1999) Modifi ed topsoil islands within patchy Acknowledgements Mediterranean vegetation in SE Spain. Catena 38, 23–44. I wish to extend my thanks to Tom Monaco, Bolton, H., Smith, J.L. and Link, S.O. (1993) Soil Roger Sheley, and Rachel Standish for their microbial biomass and activity of a disturbed helpful comments on early drafts of this and undisturbed shrub-steppe ecosystem. Soil chapter. I also want to thank Rachel Standish Biology and Biochemistry 25, 545–552. for her help with the photo in Fig. 3.3. And, Bonet, A. and Pausas, J.G. (2004) Species richness and cover along a 60-year chronosequence in thanks to Jonathan E. Taylor and Alice old-fi elds of southeastern Spain. Plant Ecology Clarke at the Everglades National Park, USA, 174, 257–270. for their assistance in locating an image of Buckley, D.H. and Schmidt, T.M. (2001) The the Hole-in-the-Donut used in Fig. 3.3. structure of microbial communities in soil and the lasting impact of cultivation. Microbial Ecology 42, 11–21. References Buisson, E. and Dutoit, T. (2004) Colonisation by native species of abandoned farmland adjacent Allen, E.B. and Allen, M.F. (1984) Competition to a remnant patch of Mediterranean steppe. between plants of different successional stages: Plant Ecology 174, 371–384. mycorrhizae as regulators. Canadian Journal of Burke, I.C., Lauenroth, W.K. and Coffi n, D.P. (1995) Botany 62, 2625–2629. Soil organic matter recovery in semiarid Allen, M.F. (1988) Belowground structure: a key to grasslands: implications for the Conservation reconstructing a productive arid ecosystem. In: Reserve Program. Ecological Applications 5, Allen, E.B. (ed.) Reconstruction of Disturbed 793–801. Arid Lands: An ecological approach. Westview Buschbacher, R., Uhl, C. and Serrao, E.A.S. (1988) Press, Boulder, Colorado, pp. 113–135. Abandoned pastures in eastern Amazonia. 2. Anderson, R.C. (2008) Growth and arbuscular Nutrient stocks in the soil and vegetation. mycorrizal fungal colonization of two prairie Journal of Ecology 76, 682–699. grasses grown in soil from restoration of three Byers, J.E. (2002) Impact of non-indigenous ages. Restoration Ecology 16, 650–656. species on natives enhanced by anthropogenic Auge, R.M. (2001) Water relations, drought and alteration of selection regimes. Oikos 97, 449– vesicular-arbuscular mycorrhizal symbiosis. 458. Mycorrhiza 11, 3–42. Callaham Jr, M.A., Rhoades, C.C. and Heneghan, Bakker, J.P. and Berendse, F. (1999) Constraints in L. (2008) A striking profi le: soil ecological the restoration of ecological diversity in knowledge in restoration management and grassland and heathland communities. Trends science. Restoration Ecology 16, 604–607. in Ecology and Evolution 14, 63–68. Charley, J.L. and West, N.E. (1975) Plant-induced Banerjee, M.J., Gerhart, V.J. and Glenn, E.P. (2006) soil chemical patterns in some shrub-dominated Native plant regeneration on abandoned desert semi-desert ecosystems of Utah. Journal of farmland: effects of irrigation, soil preparation, Ecology 63, 945–963. and amendments on seedling establishment. Coffi n, D.P., Lauenroth, W.K. and Burke, I.C. (1996) Restoration Ecology 14, 339–348. Recovery of vegetation in semiarid grassland Benjamin, K., Domon, G. and Bouchard, A. (2005) 53 years after disturbance. Ecological Vegetation composition and succession of Applications 6, 538–555. abandoned farmland: effects of ecological, Corbin, J.D. and D’Antonio, C.M. (2004) Can carbon historical and spatial factors. Landscape addition increase competitiveness of native Ecology 20, 627–647. grasses? A case study from California. Biondini, M.E., Bonham, C.D. and Redente, E.F. Restoration Ecology 12, 36–43. (1985) Secondary successional patterns in a Cramer, V.A. and Hobbs, R.J. (2002) Ecological sagebrush (Artemisia tridentata) community as consequences of altered hydrological regimes 52 L.R. Morris

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approach. Cambridge University Press, Zavaleta, E.S., Hobbs, R.J. and Mooney, H.A. Cambridge, UK. (2001) Viewing invasive species removal in a Whisenant, S.G., Thurow, T.L. and Maranz, S.J. whole-ecosystem context. Trends in Ecology (1995) Initiating autogenic restoration on and Evolution 16, 454–459. shallow semiarid sites. Restoration Ecology 3, Zimmerman, J.K., Aide, T.M. and Lugo, A.E. (2007) 61–67. Implication of land use history for natural forest Woods, K. (1997) Community response to plant regeneration and restoration strategies in invasion. In: Luken, J.O. and Thieret, J.W. (eds) Puerto Rico. In: Cramer, V.A. and Hobbs, R.J. Assessment and Management of Plant (eds) Old Fields: Dynamics and Restoration Invasions. Springer-Verlag, New York, pp. of Abandoned Farmland. Island Press, 56–68. Washington, DC, pp. 51–74. Resource Pool Dynamics: Conditions That Regulate Species 4 Interactions and Dominance

A. Joshua Leffl er1 and Ronald J. Ryel2

1 US Department of Agriculture, Agricultural Research Service, USA 2 Department of Wildland Resources and the Ecology Center, Utah State University, USA

Introduction plants has focused on gross availability of resources and does not eff ectively address A primary obstacle in the restoration of the variable timing and spatial distribution native plant communities dominated by of resources that are closely tied to plant invasive exotic species is the limited capacity requirements and performance (Th eoharides managers have to infl uence ecological and Dukes, 2007). In this chapter we detail processes that perpetuate the undesired resource pools as a common currency for con dition. Although much attention has species interactions and demonstrate that been given to understanding how dis- under standing how plants modify resource turbances lead to invasion (Sher and Hyatt, availability for other species can lead to an 1999; Davis et al., 2000; Davis and Pelsor, improved framework for making decisions 2001), the physiological and morphological to manage invasive species. diff erences between invasive and non- Understanding resource pool dynamics invasive plant species (Pyšek and Richard- is fundamental to understanding per- son, 2007), the existence of multiple stable formance of invasive species and the ecosystem states and the transitions among susceptibility of ecosystems to plant them (Sutherland, 1974; Westoby et al., invasion. Plants do not merely use required 1989; Briske et al., 2003), and feedbacks resources such as light energy, water, or that keep invasive plants entrenched on nutrients, but modify the availability of the landscape (Schlesinger et al., 1990; these resources to other plants. Plants are Klironomos, 2002; Sperry et al., 2006), ‘ecosystem engineers’ (sensu Jones et al., scientists have not linked ecological theory 1994) or organisms that modify the supply and management (Gurevitch et al., 2011). If of resources to other organisms. Plants scientists cannot mechanistically describe modify resource availability through their how basic ecological principles directly morphology or structure (i.e. shading other infl uence plant invasions, sound manage- plants or intercepting rainfall) and through ment strategies cannot be developed to their physiology (i.e. uptake of nitrogen); eff ectively prevent invasion and restore and because plant species are unique, the landscapes to a more desired state. One infl uence each has on resources is also underlying process known to be critically unique (Eviner, 2004). Th e infl uence of important to community structure, eco- plants on resource pool dynamics gained system function, and exotic plant invasion is recognition in the fi eld of environmental resource-use dynamics (Davis et al., 2000; biophysics (Campbell, 1977) and more Davis and Pelsor, 2001). However, much of recently in ecohydrology (Baird and Wilby, our current view of resource acquisition by 1999; Eamus et al., 2006).

© CAB International 2012. Invasive Plant Ecology and Management: Linking Processes to Practice (eds T.A. Monaco and R.L. Sheley) 57 58 A.J. Leffl er and R.J. Ryel

Resource availability is widely recognized focus on interactions between plants and as a primary factor controlling species resources, which necessarily addresses assembly in a plant community and con- three early phases of invasion, namely, sequently infl uences invasion. Th e niche colonization, establishment, and spread concept (Hutchinson, 1957) and the (Th eoharides and Dukes, 2007). While principle of competitive exclusion (Gause, numerous conceptual frameworks for 1934) suggest communities are composed invasion have been proposed (Gurevitch et of species that spatially and temporally al., 2011), the resource dynamic framework segregate resource use and that niche is especially applicable to management. requirement diff erences are suffi cient to maintain species diversity in communities (Adler et al., 2010). While stochastic events Resource Pools and Fluxes (i.e. disturbance and seed arrival) clearly play a role in community assembly (Gleason, Necessary resources for plants, ‘substances 1926; Hubbell, 2001), resource availability required for and consumed during growth’ determines which species persist (Tilman, (Field et al., 1992), have been recognized 1982) and increase when rare (Crawley et al., for over 200 years. Th e resource pools 1996; Th eoharides and Dukes, 2007; concept (Fig. 4.1a) comes from mid-20th- Gurevitch et al., 2011), which is required for century ecologists who described the invasion. Th e fl uctuating resource avail- transfer of energy and materials among ability hypothesis (Davis et al., 2000; Davis species in ecosystems (Ovington, 1962; and Pelsor, 2001) links disturbance, Bormann and Likens, 1967; Odum, 1969). resources, and plant community invisibility, Here, we emphasize that a resource pool is a and is broadly conceptualized as niche measurable, limited, and temporally and opportunity (Shea and Chesson, 2002). spatially variable quantity of a resource. Accordingly, invasive species exploit All pools are infl uenced by fl uxes, rates resources that are liberated following a of addition or subtraction of resources to disturbance. We need, however, to link or from the pool (Chapin et al., 2002). theoretical concepts of resources, com- Modulators are physical or chemical munities, and invasive species with practical properties (i.e. temperature, soil pH) that management. Th e infl uence of plants on aff ect a pool or fl ux but are neither consumed resources pools and the likelihood of nor depleted (Field et al., 1992). Modulators invasion suggest managers can manipulate can aff ect plants (e.g. low temperature resource pools to achieve a desired outcome. inhibits nitrogen uptake) but plants can also Understanding the roles that individual aff ect modulators (e.g. root exudates reduce species play in a community is the critical soil pH). link between theory and management. Resource pools are critical for plant In this chapter, we fi rst present the community assembly and invasion. Most concept of a resource pool and the fl uxes simply, invasion is a process by which a that contribute to or subtract from the pool; species increases in abundance when rare next we describe how plants infl uence pools (Crawley et al., 1996; Th eoharides and and fl uxes and the consequences these Dukes, 2007; Gurevitch et al., 2011). While modifi cations have for other species; then several processes may be operating to allow we examine how species interactions are an increase in population size such as enemy mediated by resource pools; and fi nally we release or propagule pressure (Gurevitch et describe links between resources pools, al., 2011), the basic resources required for invasion, and management. We conclude the growth and reproduction must be present. chapter by describing three scenarios that Consequently, given adequate seed supply illustrate the importance of considering and lack of enemies, invasion will be a resource pools when making decisions about resource-driven process (Crawley et al., the management of specifi c invasive species. 1996; Davis et al., 2000; Davis and Pelsor, Invasion is a multi-step process; here we 2001). Resource Pool Dynamics 59

(a) Evapotranspiration Precipitation (Peek and Forseth, 2003). Predictability is important to community assembly and plant invasion because invasive species often take advantage of high resource availability that is outside the normal range of variation experienced by the species in Run-on Soil H2O Runoff the undisturbed community (Sher and Hyatt, 1999). Pools and fl uxes of resources are not independent. Soil pools of water and nitrogen are inherently linked because Percolation below roots plants must take up nitrogen in solution (Nye and Tinker, 1977; Barber, 2001). Hence N supply to plants is strongly (b) Evapotranspiration Precipitation aff ected by diff usion rate and plant growth may become N-limited when soils remain wet (Ryel et al., 2008). Linkages among resources suggest that changes in pool size or fl ux of one resource can infl uence the Grass pool or fl ux of another resource. Con- Run-on Runoff sequently, even minor disturbances can infl uence a suite of resources and have a Shrub cascading eff ect on the ecosystem. Th ese types of changes likely cause abrupt transitions among community states on the landscape (Sutherland, 1974; Westoby Percolation below roots et al., 1989; Briske et al., 2003). Fig. 4.1. (a) A typical ecosystem ‘bucket’ model Critical to linking resources and com- describing a pool of soil water and the fl uxes that munity assembly to invasive species modify pool size. (b) Pools and fl uxes of soil water management is our assertion that resource as ‘perceived’ by and modifi ed by two species, a pools can be species specifi c (Fig. 4.1b). Two shrub (solid lines) and a grass (dotted lines). species, which diff er in their morphology Shrubs as larger individuals access a larger and physiological tolerance, essentially volume of soil water, transpire more water, and may ‘perceive’ diff erent sizes of the same resource be more tolerant of low soil water potential and pools. Consequently, modifi cation of a allow less water to percolate below the roots. Grasses have a smaller pool of soil water and are resource pool by each of these species will better at slowing water movement over the soil diff erentially infl uence resource availability surface. Precipitation and run-on are assumed for other species and ultimately coexistence. equal for each species in this model. Management actions can be taken to eff ect the same result; intentional removal or addition of a resource can infl uence com- Invasion is possible because resource munity structure (Perry et al., 2010; Dickson pools, fl uxes, and modulators vary con- and Foster, 2011). siderably in space and time. Some variation is predictable such as seasonal patterns of light, temperature, and precipitation, or the Plant Modifi cation of Resource Pools spatial arrangement of soil nitrogen with respect to plants (Burke, 1989; Hooker et al., Many organisms can be considered eco- 2008). Other variation is diffi cult to predict system engineers (sensu Jones et al., 1994). such as tree-fall gaps in a forest canopy Ecosystem engineers are organisms that (Hubbell et al., 1999) or discrete patches modify the supply of a resource for other of nitrogen derived from animal waste plants in the community. While the classic 60 A.J. Leffl er and R.J. Ryel

example of an animal engineer is the beaver paths along the surface of roots and through (Castor canadensis), plant species that soil macropores created by roots (Beven profoundly modify the infl ux and effl ux of and Germann, 1982; Li and Ghodrati, 1994), resources, as well as microclimate and or through the process of hydraulic disturbance regimes, can also be considered redistribution among soils of diff erent water ecosystem engineers (Fig. 4.2). potential (Richard and Caldwell, 1987; Plants considerably modify the infl ux of Caldwell et al., 1998; Leffl er et al., 2005). resources. Th e canopy of a plant can intercept Plants infl uence nutrient infl ux through the precipitation preventing it from entering deposition of leaves or roots and organic the pool of soil water (Bosch and Hewlett, molecules from root exudates (Haynes, 1982; Brown et al., 2005; LaMalfa and Ryel, 1986; Porazinska et al., 2003; Zhang et al., 2008), or promote condensation of fog 2008). Plants reduce the fl ux and modify the (Azevedo and Morgan, 1974; Ewing et al., spectrum of light between above- and 2009), essentially creating precipitation below-canopy environments (Fahnestock when it would not have occurred. Plants and Knapp, 1994; Hubbell et al., 1999; Kobe, slow the fl ow of water over soil surfaces 1999). preventing erosion but also increasing Plants modify the effl ux of resources. local infi ltration (Schlesinger et al., 1990; Even in ecosystems with considerable Bhark and Small, 2003; Ludwig et al., 2005). precipitation, water use by plants will Roots can encourage deep recharge of water substantially dry soils (Duff et al., 1997). At through the creation of preferential fl ow the scale of a watershed, plant transpiration

Precipitation

Evapotranspiration

Run-on

Litter-fall

Infiltration Through-fall

Stem flow Nutrient uptake

Root Runoff exudates

Hydraulic redistribution Subsurface flow Stream

Fig. 4.2. Plants as ecosystem engineers. Arrows indicate fl uxes associated with pools of soil water and nutrients. Plants modify precipitation by intercepting it, channeling it along stems, or slowing overland fl ow of water. Plants modify soil water through transpiration and hydraulic redistribution. Plants modify soil nutrient pools through deposition of litter, nutrient uptake, and root exudates. Resource Pool Dynamics 61

can be observed in diminished stream fl ow 2005; LaMalfa and Ryel, 2008). Shrub- during daylight hours (Bond et al., 2002). reduction treatments in semi-arid lands Plants cast shadows, which reduce (Mueggler and Blaisdell, 1958; Wambolt evaporation (LeMaitre et al., 1999; Zhang and Payne, 1986) slow water use (Sturges, and Schilling, 2006) and cool the soil surface 1973, 1993; Inouye, 2006; Prevéy et al., (D’Odorico et al., 2010; He et al., 2010). 2010a, b), but result in carbon and nitrogen Harvest of trees in semi-arid Australia causes loss (Tiedemann and Klemmedson, 1986; upward movement of groundwater and soil McClaran et al., 2008). salinization because plants are no longer using the water resource (Peck, 1978; McFarlane and Williamson, 2002). Plants Modifi cation of a Resource Pool by use considerable mineral nutrient resources; One Species has Consequences for when plants senesce in late summer, their Others lack of use results in an accumulation of inorganic nitrogen in the soil (Booth et al., Species are linked by common resource pools 2003b; Walvoord et al., 2003; Sperry et al., in an ecosystem (Bormann and Likens, 2006; Hooker et al., 2008; Adair and Burke, 1967; Odum, 1969). Consequently, when a 2010). Fire is an important disturbance in resource pool is modifi ed, other species in many ecosystems; the density and size of the community will have access to a larger or plants infl uences frequency and intensity of smaller pool (Jones et al., 1994). For fi res and consequently the loss of carbon and example, it is often diffi cult for seedlings to nitrogen during fi res (Cleary et al., 2010; establish when other plants have prior Wan et al., 2001). access to soil water (Gross and Werner, Plant modifi cation of resource infl ux and 1982; Prevéy et al., 2010b); these seedlings effl ux alters fl uxes among resource pools. essentially experience microsite drought, Rate and depth of water infi ltration is a even though soil moisture conditions at the function of soil water content (Brooks et al., landscape scale are favorable for germination 1991), which depends on plant transpiration. and establishment. Conversely, hydraulic Th e slowing of surface water movement by redistribution allows deep-soil water, which plants (Schlesinger et al., 1990; Bhark and is accessed by shrubs and trees, to be used Small, 2003; Ludwig et al., 2005) reduces by shallow-rooted grasses (Richard and export of carbon and nitrogen from the Caldwell, 1987; Dawson, 1993; Caldwell et point of precipitation to downstream eco- al., 1998; Ludwig et al., 2004). systems. Plants infl uence nutrient cycling Th e failure of a species to use a resource by altering the C:N ratio of soils (Hooper pool creates a pool that can be used by other and Vitousek, 1998; Eviner et al., 2006) and species. Shrub removal results in an increase consequently the production of nitrate, in soil water availability (Sturges, 1973, which is highly mobile and easily leached 1993) but also leads to the establishment of (Miller and Cramer, 2004). invasive annual species such as cheatgrass Finally, land use and management can (Bromus tectorum) and western salsify infl uence resource pools. Livestock grazing (Tragopogon dubius), which require increased and timber harvest remove nutrients stored soil water for successful establishment – in plant tissue (Th ompson Tew et al., 1986; (Prevéy et al., 2010a, b). Soil NO3 Milchunas and Lauenroth, 1993). Grazing concentration increases following plant can either increase or decrease soil pools of senescence in autumn or removal of plants water, carbon, and nitrogen (Milchunas and (Booth et al., 2003b; Walvoord et al., 2003; Lauenroth, 1993; Naeth and Chanasyk, Sperry et al., 2006; Hooker et al., 2008; Adair 1995; Schuman et al., 1999; Donkor et al., and Burke, 2010), which favors species or 2006; Bagchi and Ritchie, 2010), and genotypes most capable of using the increases light availability (Fahnestock and excessive resource (Rosenzweig, 1971; Knapp, 1994); harvest of forests reduces Maron and Connors, 1996). For example, interception of precipitation (Brown et al., nitrogen infl ux is enhancing invasion by 62 A.J. Leffl er and R.J. Ryel

non-native Phragmites in the northeastern unpublished data). Because these nitrogen- USA (Holdredge et al., 2010). enhanced soils are riparian, this excess N Modifi cation of resource pools by plants can be exported to downstream ecosystems at small spatial and temporal scales can (Binkley et al., 1992; Stein et al., 2010), supersede large-scale infl uences. For altering N pools in other communities. example, soil water content can vary with plant communities (Caldwell et al., 2008). Replacement of a conifer forest with a All Species in a Community do not deciduous forest type may increase water Rely on Identical Resource Pools yield by 30%, locally equivalent to 300 mm of additional precipitation (LaMalfa and Species coexist when they are not limited by Ryel, 2008). In southwestern Australia, the same niche dimension (Gause, 1934; waterlogging of soils following conversion of Hutchinson, 1957; Hardin, 1960; Chesson, savanna to cropland is so profound that crop 2000). Th is basic tenet of ecology theory yields are negatively correlated with indicates that species diff er in how they use precipitation (McFarlane and Williamson, resources or perceive resource pools. Th ese 2002). Tree-fall gap studies indicate the diff erent resource requirements led Elton profound community change that results (1958) to propose that diverse communities following sudden increases in solar fl ux are more resistant to invasion because there (Hubbell et al., 1999). are no niche opportunities (Shea and Plants can also infl uence physical and Chesson, 2002); all resources are exhausted chemical properties to the benefi t of other by the species in the community. individuals of the same species. So called Th e soil water pool is commonly viewed positive feedbacks can promote invasion by as divided among plant functional types. plants with apparently diff erent life Walter (1971) proposed a two-layer model of histories. In the Chihuahuan desert of North water availability, near-surface water and America, encroachment by creosote bush water at depth. Ehleringer et al. (1991) (Larrea tridentata) into grasslands increases expanded this concept, demonstrating that minimum soil temperature and facilitates diff erent life forms (i.e. grasses and trees) further shrub establishment (D’Odorico et draw water from diff erent depths. Some al., 2010; He et al., 2010). In much of the functional groups, such as shrubs, relied on Great Basin, frequent fi res in cheatgrass- water from a range of soil depths while other dominated sites not only kill young shrubs, species were closely tied to near-surface – they create a near-surface supply of NO3 water (annual species), or only used surface that promotes rapid growth of this invasive water if it was reliably available (Williams grass (Booth et al., 2003b; Sperry et al., and Ehleringer, 2000). 2006). Th e recent concept of growth and Management for one species can maintenance pool water (Ryel et al., 2008, enhance resource pool size for other species 2010) is especially relevant to management. in a community. Intentionally planting the Rather than viewing water among diff erent nitrogen-fi xing shrub bush lupine (Lupinus soil depths as divided among functional + arboreus) doubled soil NH4 and altered the groups, soil water is separated into pools for species composition of a coastal plant diff erent functions and explicitly links water community (Maron and Connors, 1996). and nutrient availability. Th e growth pool of Th e exotic nitrogen-fi xing tree Russian water exists when soil water potential is olive (Elaeagnus angustifolia), initially high enough to allow diff usion of nutrients planted for stream-bank stabilization, to the root surface. Th is water is used by all increases soil N in riparian areas in the species and growth will cease when nutrients southwestern USA (DeCant, 2008) and can no longer be extracted. Th e maintenance fi xed nitrogen has been observed in pool is the water that remains in the soil cottonwood (Populus angustifolia) trees near after growth has stopped. Species will diff er Russian olive in southern Idaho (Leffl er, in the timing of their requirements of these Resource Pool Dynamics 63

pools (Fig. 4.3). Annual plants do not require unaff ected. Consequently, a land manager a maintenance pool; they senesce when the can examine the resource pool requirements growth pool is exhausted. Perennial plants of desired and undesired species and take need a maintenance pool to persist through action to infl uence appropriate resource the growing season and the amount of water pools. in this pool is a function of soil water content Th e soil nitrogen pool may also be and root depth (Fig. 4.3). Th is concept can be perceived diff erently by each species in a extended further. A winter annual plant community. While soil nitrogen is typically requires a ‘germination’ pool of water during concentrated in the uppermost soil layers the autumn. If this resource is limited or (Barber, 2001; Hooker et al., 2008), it exists absent, population growth of winter annual in multiple organic and inorganic forms. species will suff er, yet perennial and summer Species diff er in their ability to acquire annual species in the community would be organic nitrogen (mostly as amino acids),

Annual Perennial Forb Shrub grass grass

– Soil Soil NO3 H2O (v/v) (ppm) 0 0.25 0 12 0

Perennial May

1 Annual

Depth (m) July

2

crit (MPa) –0.5 –2.5 –2.0 –5.0

May H2O (mm) 39 126 250 600

July H2O (mm) – 62 10 36

– Fig. 4.3. Soil water and NO3 pools available to different life forms in a semi-arid ecosystem. The growth pool, water at high potential available in the near-surface soil, is similar for each species but maintenance pools differ considerably based on root depth and critical water potential (Ψcrit, Ψ which causes uncontrolled xylem cavitation). For soil water content in annual grass, perennial grass, and shrub communities: values represent the difference between soil water content in the rooting zone during May, the soil water content derived from Ψcrit, and depth of water extraction from Peek et al. (2005), Ryel et al. (2010), and Leffl er et al. (2005), respectively. For May, we assumed water content of 30% to 40% throughout the soil profi le in calculations. Calculations of Ψcrit and soil water content in July based on moisture release curve in Leffl er et al. (2002) and data in Ryel et al. (2010). Forb data are based on – Prevéy et al. (2010a) and an assumed root depth of 1 m. For soil NO3 : profi les in annual and perennial – communities to 1 m derived from Hooker et al. (2008); below 1 m values are hypothetical but NO3 reservoirs have been observed previously (Walvoord et al., 2003; Sperry et al., 2006; Graham et al., – 2008). Soil NO3 values are appropriate for spring and early summer. Drawings not to scale. 64 A.J. Leffl er and R.J. Ryel

+ – NH4 , and NO3 (Weigelt et al., 2005), 1991; Brooker and Kikvidze, 2008) because potentially allowing specialization on these they infl uence resource pools. Drought diff erent forms (McKane et al., 2002; Ashton directly limits water availability and et al., 2010). McKane et al. (2002) observed indirectly limits nutrient availability by species to diff er in timing, depth, and form reducing diff usion (Nye and Tinker, 1977; of uptake, and Aanderud and Bledsoe (2009) Barber, 2001); herbivory or disease can slow observed partial separation of N-form plant water use or increase light for sub- uptake between native and invasive grasses canopy species (Fahnestock and Knapp, in an oak woodland in California, USA. 1994; Naeth and Chanasyk, 1995; Heil et al., Finally, solar fl ux is a critical resource 2000); predation can have the opposite divided among species based on intensity. eff ect by removing herbivores that defoliate Early successional species, grasses, forest plants (McClaren et al., 2008). Th us, even canopy emergent trees, and others require apparently top-down processes (Schmitz, high intensity light while other species 2008) can ultimately have bottom-up eff ects cannot tolerate direct sun (Hubbell et al., through their alteration of resource pools or 1999; Kobe, 1999). Consequently, high-light fl uxes. Th is resource-based framework species create an environment only for low- eliminates the distinction between abiotic light species (Messier et al., 1998), essentially stress such as drought caused by physical preempting this resource from other species processes (i.e. lack of rain) or by other requiring high light. Grass cover reduces individuals (i.e. available water was used) available light and limits forb biomass in and focuses on the mechanistic interaction tallgrass prairie (Turner and Knapp, 1996), among individuals mediated by resource and MacDougall and Turkington (2005) dynamics (Grime, 1977; Welden and observed subordinate species to respond to Slauson, 1986). increased light rather than increased N in an We present this perspective because oak savanna. direct competition for a resource by two individuals is rare, especially for below- ground resources. First, many apparently Interactions Among Species competitive interactions are separated in are Mediated by Resource time. Individuals may directly compete for Pools soil water if they germinate together and grow at similar rates, but only after reaching Th e competition paradigm has dominated sizes where resource acquisition is limited ecological theory since inception of ecology by other individuals (i.e. density dependence, as a discipline (Gause, 1934; Clements, Goldberg et al., 2001). Commonly, however, 1936; Tilman, 1982; Goldberg and Barton, germination is separated by several days in 1992). Accordingly, species compete directly these individuals. Consequently, one or indirectly for a set of resources (Pianka, individual germinates and begins growing 1987) and the theory implies that the into an environment previously modifi ed likelihood of competitive exclusion increases (i.e. drier or shaded) by the species that with increasing niche overlap (Gause, 1934; germinated earlier. Annual and perennial Hardin, 1960; Johansson and Keddy, 1991). species are often described as in competition While competition exists (Goldberg and (Corbin and D’Antonio, 2004; Blank, 2010), Barton, 1992), many authors question the but perennial species are biologically active importance of competition in structuring long after annual species have senesced and plant communities (Hutchinson, 1957; therefore they no longer modify water, Welden and Slauson, 1986; Rees et al., 1996; nitrogen, or light availability for perennial Damgaard and Fayolle, 2010). Th e crux of species. Perennial plant species, however, the argument against the importance of may experience an environment previously competition is that other factors such as modifi ed by an annual species (e.g. low soil abiotic stress, predators, disease, or dispersal N or soil moisture) to the detriment of can play a relatively stronger role (Grace, perennial species performance and sub- Resource Pool Dynamics 65

sequent fi tness. Second, other competitive species are in the regional pool, capable of interactions are only temporary. For dispersing to the site, and establishing. example, trees or shrubs and grasses are Alternatively, if established plant species are often described as in competition (Facelli using that resource, other species will often and Tembly, 2002), but can only compete for be inhibited. water until the woody plants have access to Because resources are dynamic, com- deep soil water (Ehleringer et al., 1991), and munities are constantly changing. Modern grasses cannot eff ectively limit light to ecology recognizes that communities are mature trees. Taller statured trees, however, neither completely deterministic (Clements, create a shade environment for the grasses; 1936) nor stochastic (Gleason, 1926) col- thus the distribution of grass species across lections of species (Sutherland, 1974; May, the landscape may be aff ected by the 1977). Rather, a community is constantly shade patterns of trees. Finally, interactions changing within environmental constraints. among species can shift from generally Th ese changes are generally small such as a negative (i.e. competition) through neutral series of wet summers increasing grass to generally positive (facilitative) depending abundance and/or biomass in a shrub/ on abiotic conditions (Callaway and Walker, grass community. Occasionally, changes in 1997; Maestre et al., 2003). Individual resources are large, shifting communities to aridland plants growing close together is diff erent states (Sutherland, 1974; Westoby often interpreted as competition for water et al., 1989; Briske et al., 2003) such as fi re (Welden and Slauson, 1986; Keddy, 1989), removing the shrub component of an but when water was scarce, shading of ecosystem and making resources available grasses by shrubs reduced leaf temperature for the establishment of grasses and forbs. and transpiration rate, and promoted grass Whenever resource pools change through survival (Maestre et al., 2003). natural processes or management, changes Th e processes of community assembly in the community will likely follow (Davis and succession are often thought to depend and Pelsor, 2001). Predictable and repeat- on availability of species, availability of sites able patterns of resource dynamics result in for establishment, and performance of stability of plant assemblages, while species at those sites (Pickett et al., 1987; unpredict able or directional trends in Sheley et al., 2006). For a species to enter a resources will alter species assemblages. A community it must be present in the region key factor in community change is resource and be capable of dispersing to the site. Th is fl ux outside the typical range of variation requirement may depend on availability (Sher and Hyatt, 1999). of plants to disperse propagules across the landscape, or the presence of vectors (Th eoharides and Dukes, 2007). Sites must Resource Pool Dynamics and contain the resources necessary for ger- Invasive Species min ation and establishment, and resource availability must coincide with plant Changing resource availability, whether phenology for germination and establish- caused by natural processes or management, ment. A site previously occupied by another will ultimately result in changes in plant seed that germinated likely lacks the neces- community composition. Th ere is a growing sary water, nitrogen, or light required for body of evidence that altered resource the newly germinating individual. Finally, availability is a key factor infl uencing persistence requires diff erent resource pools colonization of a site by an invasive species than establishment. A shrub may establish (Th eoharides and Dukes, 2007). In many in a favorable site, but will not persist in the cases, the trigger for invasion is a discrete community if deep water is not available. An disturbance that alters resource availability ecosystem that has a largely untapped either directly, or disrupts plant communities resource such as deep-soil water will support to such an extent that supply and use of a a species that requires that resource if those single or multiple resources is altered (White 66 A.J. Leffl er and R.J. Ryel

and Pickett, 1985; Hobbs and Huenneke, over 100 years (Elmore et al., 2006; see 1992; Sher and Hyatt, 1999; Adair et al., Morris, Chapter 3, this volume). 2008). Plant invasion following disturbance and Disturbance is often a prerequisite for a change in resource availability is a natural invasion (Hobbs and Huenneke, 1992; process that would occur even if exotic Davis et al., 2000), and it can occur through species were absent. Succession following natural or anthropogenic processes. Studies disturbance can proceed linearly from short- often attribute plant invasion to land uses lived, rapidly growing species that can such as livestock grazing (Young et al., 1972; quickly colonize bare soil to larger, slow- Chambers et al., 2007), cultivation for crops growing species, tolerant of high plant (Williamson and Fitter, 1996; see Morris, density (Tansley, 1935; Clements, 1936; Chapter 3, this volume), and recreation Connell and Slayter, 1977); or non-linearly (Vilá and Pujadas, 2001). While these land through multiple stable states (Sutherland, uses have profoundly changed plant com- 1974; Westoby et al., 1989; Briske et al., munities, from the perspective of altered 2003). Th ese previous invasions, however, resource availability, natural processes such were by local or regional species; con- as fi re (Zedler and Scheid, 1988; Chambers temporary invasion is often by exotic et al., 2007) or drought (Allen and Breshears, species, creating novel communities with 1998) can also promote long-term vege- resource pool dynamics that do not have tation change. Th e nature of the distur bance historical or evolutionary precedent at the is less important to plants than the con- site of invasion. Th e same basic ecological sequences that the disturbance has for processes operate in native and exotic plant resource pools. communities (Gurevitch et al., 2011), but All disturbances are not equivalent; some there has been insuffi cient time for an have large infl uence on resource pools, while evolutionary response by native species to others have only minor consequences. exotics (Crawley et al., 1996). Drought can infl uence landscapes, and if Invasion involves a process of matching persistent, can alter the nature of a plant changes in resource pools with species that community for decades to centuries. Th e can take advantage of those resources. Only 1950s drought in the southwestern USA a small fraction of exotic species are invasive shifted the woodland-forest ecotone by over (Williamson and Fitter, 1996; Levine et al., 2 km in distance and 200 m in elevation 2003). Many of these invasive exotic species (Allen and Breshears, 1998). In addition, the can be broadly described as having an infl uence of livestock grazing can range from r-selected life history (MacArthur and minimal to intense. Light grazing likely Wilson, 1967) or as ruderals (Grime, 1977). removes only a small portion of plant bio- Many invasive exotic species can be mass and has minimal eff ect on soil water, categorized as ‘acquisitive’ on the leaf while heavy grazing reduces transpiration economics spectrum (Wright et al., 2004) and water infi ltration (Naeth and Chanasyk, and traits such as rapid growth, early 1995; Chanasyk et al., 2004) and increases repro duction, and high fecundity are light availability (Fahnestock and Knapp, almost universally found in invasive species 1994). Even light grazing, however, can (Pyšek and Richardson, 2007), but other create bare patches or depres sions in soils traits such as germination timing, leaf area, where water may accumulate, providing and biomass diff er little between native opportunities for exotic plants to establish if and invasive exotic species (Pyšek and seed sources are available (Hobbs and Richardson, 2007), and a universal com- Huenneke, 1992). Old-fi eld studies illustrate plement of invasive traits has yet to be found the infl uence of former land cultivation (Tecco et al., 2010). Th e key to understanding practices on soil structure, water-holding if an exotic species will become invasive is capacity, and nutrient content. In many not the traits it possesses alone, but if those cases, the infl uence of agriculture on sub- traits allow the exotic species to exploit an sequent plant communities is evident for available resource pool (Heger and Trepl, Resource Pool Dynamics 67

2003; Roscher et al., 2009). Hence recent fundamental ecological principles (Sheley studies link invasion to increased water et al., 2006; Gurevitch et al., 2011) that are availability following shrub removal (Prevéy responsible for community assembly and et al., 2010a, b) and increased N availability plant invasion. following herbivory by insects (Brown, Resource pools or fl uxes can be directly 1994). managed. Manipulation of nitrogen Th e resource pool perspective for plant availability through the addition of labile invasion has several consequences for man- carbon such as sucrose, sawdust, or wood age ment. In addition to taking advantage of chips can favor the slow growth of native favorable resource pools during establish- species, a trait more important when ment, an invasive species will further modify nitrogen is not abundant (Perry et al., 2010). these pools, promoting subsequent domin- Supplemental water can be applied when ance, impact, and spread. Consequently, desirable species are active and invasive invasion by few individuals of one species annual plants have already senesced, or may promote spread (D’Odorico et al., 2010; supplemental water can be injected at soil He et al., 2010) or secondary invasion by depths that native species can access but other species. Removal of shrubs creates invasive species cannot. Th ese eff orts, a deep-soil pool of water that can be tapped however, would be costly and indirect by exotic annual and perennial forbs, management of resource pools may be a including Tragapogon dubius or Centaurea better option. maculosa (Hill et al., 2006; Kulmatiski et al., Resource pools and fl uxes can be managed 2006), which are invasive in western North indirectly through the manipulation of America. Also, because transitions among plants and animals. Grazing, mowing, or stable community states can be infl uenced burning can be used to remove nitrogen by resource pools (i.e. abiotic mechanisms, from soils (Marrs, 1993; Augustine, 2003; Briske et al., 2006), opportunities for Härdtle et al., 2006), especially if plants high restoration of native species may be more in nitrogen can be targeted for removal. likely when favorable resource conditions Plants that contain a high C:N ratio can be exist for the establishment of the desirable established, which will ultimately increase species. For example, abundance of cheat- the C:N ratio of the soil and further promote grass in the Intermountain West, USA can immobilization of N by soil microbes (Zink vary considerably among years at the same and Allen, 1998; Perry et al., 2010). Reducing site (Griffi th and Loik, 2010), and cheat- the biomass of shrubs, grasses, and forbs grass population die-off s can occur, creating can increase near-surface soil water, while establishment opportunities for other reducing shrub and tree biomass can species, including the native shrub big increase availability of deep-soil water sagebrush. (Sturges, 1993; Leffl er et al., 2005; Prevéy et al., 2010a, b). Alternatively, managers can increase near-surface soil water availability Managing Resource Pools in the early summer by reducing the biomass to Infl uence Invasive Plant of annual species that rapidly use water in Abundance the spring (Harris, 1967; Booth et al., 2003a). Manipulation of resource pools can be a Successful management of invasive powerful tool to manage invasive plant species through resource manipulation will species. Resource availability can be man- require careful identifi cation of the critical aged by either directly manipulating the resource pools. In most instances, the resource pools or their fl ux, or indirectly by resources required for plant establishment managing plants and animals to aff ect a are diff erent from the resources required pool or fl ux. A resource pool approach to for long-term persistence (Holt, 2009). management of invasive species is However, it is long-term resource modifi - especially powerful because it is based on cation by desired species that will reduce 68 A.J. Leffl er and R.J. Ryel

the probability of invasion. Hence, adopting Soil water pools in the sagebrush steppe a systems approach (Odum, 1994) to ecosystem resource pool management will require separate identifi cation of the resource Native sagebrush steppe ecosystems in pools that are necessary for establishment western North America contain a mixture and persistence of desired species. For of perennial tussock grasses and big example, native grasses in Oregon, USA sagebrush (Artemisia tridentata), with a have high germination rate, low survival variety of less dominant perennial forbs. immediately following germination, but Native annual species are relatively rare in high survival once established (James et al., this ecosystem (West and Young, 2000). 2011). Consequently, the critical resource Much of this ecosystem type is now pool for these native grasses may be dominated by nearly monotypic stands moisture at the site of germination in the of big sagebrush or cheatgrass (Bromus spring. Moisture status can be improved by tectorum), an invasive winter annual removal of competing vegetation or species. Perennial grasses have largely been ensuring seed contact with the soil. lost from this ecosystem in many areas; a Management of a community for long- change partially attributed to limited term resistance to invasion will also require grazing tolerance and altered fi re regimes a systems approach. A resistant community (Chambers et al., 2007). will have few available resources that an Soil water recharge in these ecosystems is invasive species can exploit (Elton, 1958; dominated by autumn, winter, and spring Davis et al., 2000; Shea and Chesson, 2002). precipitation (Caldwell, 1985) that results in Th is low available-resource state arises not accumulation of water during non-growth from a diversity of species, but from a periods, top-down recharge of a shallow diversity of resource-use patterns that growth pool (surface to 30 cm), and a deeper coincide with spatial or temporal aspects of maintenance pool (to 1–1.5 m). When not the establishment and growth of potential intensively disturbed, the intact perennial invasive plant species. Rather than grass/forb/shrub community rapidly uses approach ing an invasive species problem the growth pool in the spring, draws the with the goal of establishing plant species maintenance pool to low water potentials that are matched for a specifi c use, managers during summer, and little soil water remains need to consider the resource pools present in the autumn (Seyfried et al., 2005). on the landscape and establish desired Regeneration of sagebrush or grasses occurs species that will fully use the resource pools. in small gaps where resources are available While this approach may not maximize in the spring, and may be enhanced by larger short-term gain, it will enhance probability gaps formed by recently deceased tussock of long-term sustainability. grasses or shrubs. While all species use water from the growth pool, the deeper water is used Management Scenarios primarily by sagebrush and can be drawn to water potentials well below the wilting point We present three scenarios to illustrate the of grasses and forbs. When the water use of resource pools in ecosystem manage- potential of the growth pool is insuffi cient ment. Scenarios are based on conceptual for nutrient diff usion, the grasses and forbs models, developed from relevant studies, produce seed and become dormant (Ryel et which are the core tools for understanding al., 2010). Sagebrush, however, relies on the resource pools dynamics as a function of maintenance pool for limited physiological vegetation composition, climate, and land activity throughout the summer and into management practices. Th e scenarios the autumn. Consequently, the maintenance describe specifi c situations, but the reason- pool is largely depleted as well. Invasion and ing and approach can be applied to other dominance by exotic species is diffi cult situations. because few resources remain for them to Resource Pool Dynamics 69

exploit (Chambers et al., 2007; Prevéy et al., managed to limit impact on growth and seed 2010a, b). production, and sites should be allowed When perennial grasses and forbs are enough time to recover from use. If perennial removed, growth of the remaining sagebrush grasses are removed, seeding should be is temporarily enhanced because it has undertaken very quickly, especially in shrub exclusive access to the growth pool (Ryel interspaces where growth pool water may be et al., 2010). Higher sagebrush biomass more available, or seed could be added increases demand on the maintenance pool, regularly as a precaution. Large-scale treat- which is unaff ected by the loss of the grasses. ments to reduce shrub density should not be Consequently, the maintenance pool cannot undertaken if the growth pool is or can be supply adequate water to allow sagebrush to dominated by annual grasses. If treatments persist through the summer; xylem are necessary, perennial grasses should be cavitation (Sperry and Hacke, 2002) and established fi rst. death of branches occurs. Plants in this condition are experiencing severe late- summer drought, but not from lack of Soil N in annual and perennial grass precipitation; rather water stress arises from communities excessive use of the maintenance pool due to greater shrub biomass. In subsequent years, Grasslands of the Great Plains are dense use of the growth pool begins to diminish communities composed of cool and warm due to declining shrub condition. season species. Cool-season grasses can At this point, the stands are very remain green under snow, are physiologically susceptible to invasion from annual species active in the spring, and senesce during late such as cheatgrass that can easily exploit spring and early summer; warm-season the underutilized growth pool. Once grasses are physiologically active in response cheatgrass biomass is suffi cient to carry fi re to summer rains. In these systems, there are the remaining mature sagebrush will be active species throughout the growing killed; a dense monoculture of cheatgrass season and soil-N fl uctuations are minimal often follows. Th ese dense, contiguous (McCulley et al., 2009). – stands of cheatgrass prevent establishment When plant growth ceases, soil NO3 of sage brush seedlings, as resource-rich increases because plants are no longer using gaps are no longer available. Cheatgrass this resource (Booth et al., 2003b; Sperry et very eff ectively depletes the growth water al., 2006; Adair and Burke, 2010), but pool, but uses little of the maintenance nitrifi cation can continue at low soil water pool. As such, this water can accumulate in potential (Low et al., 1997). Although deeper soil layers and provide resources to grazing can have long-term positive or subsequent invaders such as deeper-rooted negative eff ects on soil N (Milchunas and biennial species (Prevéy et al., 2010a, b), or Lauenroth, 1993), an immediate eff ect of may eventually connect to deep ground- excessive grazing is reduced uptake of N by water pools where water quality may be perennial grasses, which can result in high – low (e.g. saline) and the resulting mixed soil NO3 availability (Booth et al., 2003b; water inhibits plant growth (Ryel et al., Sperry et al., 2006; Adair and Burke, 2010). 2010). Th ese systems become invasible by a species Sagebrush steppe as a model system that can respond to high N availability with indicates the importance of managing the vigorous growth. growth soil water pool. When this pool Short-lived annual species typically have becomes available on a large scale, invasion high relative growth rates (Wright et al., by annual grasses is likely if propagules are 2004; James, 2008) and can grow rapidly in available in the regional species pool. Th e high-N environments (Maron and Connors, simplest management option is to ensure 1996). In many cases, annual grasses will that perennial grasses and forbs are not respond to increased N availability to a removed; grazing should be carefully greater extent than perennial grasses. If an 70 A.J. Leffl er and R.J. Ryel

invasive annual species is in the regional Establishment and maintenance water species pool, it can more eff ectively take pool in riparian forests advantage of the increased N availability following perennial species loss. In Alteration of stream fl ow has led to – subsequent years, high soil NO3 in the late dramatic declines in riparian forests of the summer and autumn will become common western USA (Webb and Leake, 2006). and serve to perpetuate annual grass Th ese groundwater-dependent ecosystems, dominance. historically composed of a cottonwood Th is model of annual grass invasion (Populus sp.) overstory and an understory of illustrates the importance of careful willow (Salix sp.), have been replaced by the management of soil N pools. In an intact invasive shrub salt cedar (Tamarix chinensis) system, N fl uctuations are minimal, but and the invasive tree Russian olive the N resource fl uctuates in space and time (Elaeagnus angustifolia). Gallery forests, in a disturbed system, creating windows of which once extended for hundreds of meters opportunity for invasive species (Davis et from the main channel of large rivers such al., 2000; Davis and Pelsor, 2001). Of as the Rio Grande in North America, are course, the best management option is to now restricted to narrow bands less than 50 prevent loss of perennial species in the m wide (Howe and Knopf, 1991). Th ese fi rst place, ensuring that soil N remains forests have been lost because of upstream low, a diffi cult proposition in regions with dams, channelization, and water diversion high N deposition (Matson et al., 2002). If for culinary and agriculture use (Webb and a system is already in a poor state, however, Leake, 2006). other options need to be considered to Water in riparian forests is supplied by reduce soil N. both local and upstream sources (Leffl er and Several methods have been used, each Evans, 1999). Th e growth pool can be intending to increase the ratio of carbon to recharged by local precipitation or fl ooding nitrogen in soils degraded by annual of the forest during spring snowmelt several grasses. Most commonly, addition of hundred kilometers away. Th e growth pool carbon-rich compounds such as sugar or also may be maintained well into the sawdust stimulates immobilization of summer where monsoon precipitation is mineral nitrogen by bacteria (Zink and substantial. Th e maintenance pool, however, Allen, 1998; Klironomos, 2002; Mazzola et is supplied almost exclusively by ground- al., 2011). Carbon addition can be eff ective water derived from off -site sources. Large at limiting N availability to annual grasses, trees such as older cottonwoods require this but it must be combined with seeding of maintenance pool and are highly vulnerable perennial grasses to ensure annual species to xylem cavitation when groundwater is simply do not recolonize and dominate the limited (Tyree et al., 1994; Leffl er et al., limited soil N pool. More importantly, 2000). Th us, older individuals may die from however, would be the timing of carbon drought stress. New recruits, however, face addition. Th is treatment should be applied additional challenges because they require to specifi cally deny N to annual species the disturbance of a scouring fl ood, when they require it for establishment in essentially an establishment pool of water, the autumn. Favoring perennial grasses to coincide with a short window of seed over annual species will naturally begin to viability (Bradley and Smith, 1986). In the – stabilize soil NO3 as well. It is also pos- absence of these fl oods, largely eliminated sible to remove nutrients by removing by dams, cottonwoods fail to establish. If biomass with fi re, mowing, or grazing they do emerge, root extension must keep (Perry et al., 2010). Th ese eff orts, however, pace with accelerated groundwater decline will likely remove only a small amount of (Scott et al., 1997; Stella et al., 2010) or the the excess N. seedlings will die. Resource Pool Dynamics 71

Salt cedar and Russian olive modify the resource pools allows for assessing growth pool to the detriment of cottonwoods functionally important elements of and willows. Salt cedar is tolerant of necessary and limiting resources for native increasingly saline soils caused in part by species and exotic species that either are, or runoff from agriculture (Shafroth et al., may become invasive. From this view, it is 1995). Th is species deposits salt on the soil apparent that the invasion process requires surface when it sheds leaves in the autumn resources to become available and the (Shafroth et al., 1995). Because salt presence of an exotic species capable of eff ectively lowers the soil water potential, using the resource. Th e concept of resource cottonwood and willow seedlings experience pool dynamics allows managers to better drought even when water is abundant, i.e. understand the implications of manage- the growth pool shrinks for the native ment decisions on invasion by exotic species. Not only are water pools altered, species, assess the role of factors outside Russian olive is capable of nitrogen fi xation management infl uence such as climate and (Katz and Shafroth, 2003) and its growth is weather on invasion, and comprehend the not limited by soil nitrogen pools. challenges in restoration of a degraded Consequently, diff usion of N in low water community. Resource availability will potential soils does not limit its access to the always vary in space and time, but invasion growth pool and it can promote rates of is likely to occur when these dynamics are water table decline in excess of that which abruptly disrupted or changed directionally native, non-nitrogen fi xing species can (i.e. nitrogen becomes more available over survive. time). Moreover, management to transition Th e key to managing riparian forests invaded communities to a more desirable for native cottonwood and willow species state should be taken when resource is periodic fl ooding to allow establishment availability favors that transition. and occasional maintenance of a high water Relevant management decisions can be table to allow persistence. Th is requires made when resource pools are considered manage ment primarily of off -site water to rather than merely applying land manage- enhance the maintenance pool. Conse- ment treatments to reduce the abundance of quently, region-wide water management certain species or functional types, or plans are essential because so many modifying land uses. For each decision, a stakeholders compete for access to this manager needs to ask, ‘How do I need to limited resource. Active removal of invasive change water and nutrient pools to achieve species is also necessary for successful my desired objective and how will other restoration because salt cedar and Russian management activities aff ect pools of water olive have a substantial infl uence on the and nutrients?’ Management tools can growth pool of water. therefore be more varied and targeted, potentially giving managers additional options and more eff ective treatments. Summary

While the assertion that resource Acknowledgements availability infl uences invasion (Davis et al., 2000; Davis and Pelsor, 2001) is not new, We thank T.A. Monaco and R.L. Sheley for we articulate a view of resource dynamics conceiving of this project and their valuable that is critical for management of invasive- editorial input. Previous versions of this dominated communities and for initial manuscript were improved with the help of prevention of invasion. Th e construct of J.J. James and E. Espeland. 72 A.J. Leffl er and R.J. Ryel

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Thomas A. Grant III and Mark W. Paschke

Department of Forest, Rangeland, and Watershed Stewardship, Colorado State University, USA

Introduction to apply ecological principles to the manage- ment of ecosystem processes. Prior to the 19th century biological invasions An ecosystem is a representation of the were probably infrequent, natural events biotic elements and abiotic variables that that contributed to a diverse and ever interact at a given place and time. Any evolving patchwork of species and eco- dramatic change in the species composition, systems. Th e dramatic increase in the human either above- or belowground, will likely population and our infl uence over the aff ect the composition and function of the earth’s biotic and abiotic systems has created whole. An understanding of the direct and a world that is dominated by the eff ects of indirect impacts of invasive species on a humans (Vitousek et al., 1997). In general, system is critical to delineate if we are to humans have rapidly increased the manage natural and utilitarian ecosystems movement of fl ora and fauna across what eff ectively and with the goal of conserving were once natural barriers to migration or ecosystem diversity and function. Invasion survival and the consequences are now by exotic plants represents a fundamental termed ‘biological invasions’ (Elton, 1958). change in a system and a unique management All organisms infl uence and modify their challenge for human society, both ethically environment, whether through competitive and economically. From a Clementsian interactions that alter species composition, framework of successional dynamics, fol- the addition or removal of resources, or the lowing invasion, a late-seral community of transformation of habitat and large-scale native species may no longer be a possible ecosystem processes. To successfully manage outcome of long-term ecosystem develop- invasive species, it is critical that we ment, because the whole system has changed understand how these species interact and and its trajectory is diff erent (Clements, modify their novel environment and use this 1916). Within a Gleasonian individualistic knowledge to improve the effi ciency and perspective of plant dynamics, the key eff ectiveness of management practices. Th is players have changed and they will infl uence chapter will specifi cally address the impacts the system in a unique and diff erent manner of invasive plants on the soil, including the (Gleason, 1926). History is no longer a physical and chemical composition of soil, guide to the interaction of species and litter decomposition and biogeochemical systems, therefore how they will interact and cycling of nutrients, and soil microbial the resulting community composition is communities. Th e primary emphasis will unknown. As the plant community com- focus on the impacts of invasive species on position changes due to the invaders, so will these systems and utilizing this knowledge many of the ecological processes that link the

© CAB International 2012. Invasive Plant Ecology and Management: Linking Processes to Practice (eds T.A. Monaco and R.L. Sheley) 79 80 T.A. Grant III and M.W. Paschke

above- and belowground elements of the composition, succession, and invasion. ecosystem (Wardle et al., 2004), including In-depth research has provided quantitative decompo sition, nutrient cycling, rhizosphere information about how plants, animals, soil, exudation, microbial composition and and the atmosphere interact and ecosystems function. It is likely that the change in plant are modifi ed, but the focus on competition community will infl uence the amount and and resource effi ciency has not completely type of animal herbivory, with cascading addressed many questions about vegetation eff ects on soil properties, root growth dynamics (Klironomos, 2002), especially responses from herbivory, and nutrient concerning the paradox of invasion. inputs from animal defecation. Lastly, a Competitive ability has been used to describe change in vegetation can alter the disturbance why one organism, species, or population regime of a system, particularly fi re cycles displaces another or is more successful and intensity. Overall, any striking change in in the capture of a resource and the sub- the plant community will have diverse eff ects sequent eff ect on its survival and fecundity. throughout the whole realm of plant–soil– Obviously some species are better at microbe interactions. acquiring a resource than another, but can Invasive species tend to be highly the simplistic focus on one nutrient, growth successful through their rapid growth and habit or reproductive strategy explain the reproduction, high fecundity, great dispersal diversity of ecosystems or form a rationale abilities, and apparent predisposition to for successful invasion by an exotic species? establish in disturbed areas. Many invasive By focusing on a species’ multi-scale direct species have been removed from their and indirect interactions within an eco- evolutionary constraints or have been system and the subsequent impacts of released from enemies and are often found a change in biodiversity, we may see proliferating in disturbed areas that have the outcomes of competition are rooted available resources. Th ese aggressive and in the interactions and feedbacks of a opportunistic traits do not facilitate their complex system. It may be possible to view dominance in the species’ native range and these interactions as the mechanism that this conundrum has been called the ‘invasive structures a community and facilitates plant paradox’ (Rout and Callaway, 2009). It invasion. Th is concept removes the idea that may be possible that invasive species one species independently is a superior dramatically modify the novel habitat to competitor and focuses on an organism or increase their fi tness or negatively aff ect species’ interactions with its environment as their neighbors and competitors. Most forming the mechanisms for the outcome of organisms attempt to do this in their competitive scenarios. Th eoretically, com- struggle to reproduce and survive, but are petition is removed from the attributes of usually constrained by resources or other an individual and placed in the realm of a species, especially in the long evolutionary complex and dynamic web of direct and timeframe of an ecosystem’s development. indirect interactions of varying strengths Th ese interactions, both direct and indirect, that include both resource and non-resource may be part of a plant–soil feedback that connections (Hierro and Callaway, 2003). facilitates invasion through either positive Given this systems-based approach or or negative interactions. Feedbacks may supposition, the focus of invasive species facilitate fecundity of the invader, negatively management should be on the system’s infl uence competitors, or modify soil to the interactions and the species’ impacts and extent that the system is fundamentally not just the inherent capabilities of the diff erent and less hospitable to certain species, because the exotic entity cannot organisms. avoid being a constituent of a dynamic and Much ecological research has focused on inherently interconnected system. factors such as competition, dispersal, Invasive species have dramatically resource use effi ciency, herbivory, and impacted most ecosystems throughout the predation to understand plant community world and caused a rapid change to their Invasive Plant Impacts on Soil Properties 81

diversity and function. Although plant community has changed from a diverse species invasions are a natural part of polyculture, represented by many diff erent ecosystem development and succession, the functional traits, to a simplifi ed system that rate and scale of exotic invasions have is dominated by functionally similar increased due to anthropogenic modifi cation organisms. Broadly speaking, the energy to the environment and movement of fl ow of the system increases as the biomass propagules in an increasingly globalized and net primary productivity of the invaded world (Mack et al., 2000). Restoration and system are modifi ed. Th e consequences control eff orts on the most problematic are directly and indirectly proliferated invasive species are marginally successful at throughout the ecosystem. Aboveground small scales, but not at landscape levels. Th e the changes in the plant community are impacts of invasive species to agricultural obvious, but it is at the plant–soil interface production, recreation, and natural and belowground that scientifi c inquiry has ecosystems are vast and costly, yet these begun to focus on questions concerning are only the obvious outcomes of invasion the eff ects of invasive plant litter on and effi cient management requires an decomposition and nutrient cycling, and the understanding of how an exotic species subsequent changes in microbial community successfully invades a system. To improve structure and function (Hawkes et al., 2005). management or control of exotic species it is Many of these activities are strongly critical to understand the direct and indirect regulated by microorganisms, which will impacts of species on basic ecosystem also change as their primary habitat, the processes and how these changes infl uence rhizosphere, is modifi ed by changes in our ability to manage systems. Exotic species species-specifi c root types, architecture, fundamentally modify nutrient cycling, biomass, and physical and chemical char- litter decomposition, soil microbial com- acteristics of the soil. Dramatic changes in munities, and physical properties of soil; biochemistry of plant root exudates occur as simple removal of undesirable species or functionally and biochemically diverse revegetating areas by seeding with native species are replaced by one dominant or species may be ineff ective in restoring functionally similar species. Th is cascade of systems to pre-invasion states or former interactions begins with the invasion by ecosystem functioning. By addressing the an exotic species and in all likelihood novel conditions and processes that have most plant–soil–microbe interactions are developed due to dominance by an exotic impacted, although the degree and direction species we can develop a clearer under- is variable. Wolfe and Klironomos (2005) standing of how to eff ectively manage propose three linkages that are directly systems, determine realistic restoration impacted by invasive species: (i) plant goals, and possibly prevent additional community composition and ecosystem invasions. Th e purpose of this chapter is to processes; (ii) plant community composition review the processes and interactions by and soil community composition; and (iii) which exotic plant species alter the physical, soil community composition and ecosystem chemical, and microbial characteristics of processes (Fig. 5.1). Exotic plant invasion soil and provide a basis for applying these directly aff ects these three primary linkages concepts to innovative soil management and potentially modifi es the plant com- strategies that can reduce dominance of munity, soil community, and ecosystem invasive plants and conserve diverse, processes and function. In addition, functioning natural ecosystems. numerous indirect interactions occur and dramatically increase the complexity of managing plant invasions based on eco- Impacts of Invasive Species on Soil logical principles and successional theory (Krueger-Mangold et al., 2006; Sheley et al., Th e impacts of invasive species are pervasive. 1996). Current research is attempting to In many invasions, the whole plant understand the impacts of invasion on the 82 T.A. Grant III and M.W. Paschke

Plant community composition

2

1 Soil community composition

3

Ecosystem processes/properties

Exotic species

Fig. 5.1. Impacts of an invasive plant on the direct interactions and linkages between the plant community, soil community, and ecosystem processes. Adapted/redrawn from Wolfe and Klironomos (2005). Anagallis arvensis L. (scarlet pimpernel) illustration by Robert H. Mohlenbrock (Robert H. Mohlenbrock @ USDA-NRCS PLANTS Database / USDA SCS. 1991. Southern wetland fl ora: Field offi ce guide to plant species. South National Technical Center, Fort Worth, Texas, USA). soil and apply this information to the biochemically diff erent material is added to feedbacks between the complex systems the system, often in larger quantities than that infl uence the invasion of exotic species pre-invasion (Ehrenfeld, 2003). Within the (see Eviner and Hawkes, Chapter 7, this litter:soil interface, the microbial community volume). will adapt to the new conditions and energy sources, producing diff erent decomposition rates and fl uxes of nutrients. Indirect eff ects Direct and indirect interactions between of increased litter layers and reduced solar plants and soil radiation could include modifi cations to soil moisture, temperature, and microsites for Th e impacts of invasive species will infl uence plant establishment. Disturbance regimes, both the potential vegetation and the ability such as the infl uence of fi re on litter, soil to restore ecosystems to any semblance of nutrients, or nascent plants, have been their pre-invasion vegetation state and shown to be indirectly aff ected by invasion function. To comprehensively understand and could have long-term feedbacks that the impacts of an invasive species, it is alter the successional trajectory and critical to provide an overview of the direct potential vegetation of the community. and indirect interactions of an exotic species Change in the plant-community composition in its novel environment. Figure 5.2 may also indirectly aff ect the amount or type attempts to illustrate a simplifi ed scenario of herbivory. Grazing can stimulate root of the potential interactions between a plant growth and infl uence soil bulk density and and soil. Aboveground, the type and quantity nutrient inputs from fecal matter. of plant litter will likely change with Additionally, the feeding preferences of invasion. Decomposition will be modifi ed as animals will likely change with a shift in Invasive Plant Impacts on Soil Properties 83

Aboveground Biomass/NPPmass/NPP LitterLitter QQualityuality aandnd QQuantityuantity

Soil PhysicalPhysical PropertiesProperties Interface (pH, temp., moisture, and DDecompositionecompositii ion Nutrients/OMNutrients/O aggregation)regation)

IInvertebratesnvertebrates Belowground Biomass/NPPmass/NPP (comminution,(comminution, herbivory,herbivory, anda predation)predation)

RootR ExudatesE d MMicro-Organismsicro-Organisms (nutritional and ((pathogens,pathogens, mutualisms, allelopathic ) aandnd ddecomposers)ecomposers)

Fig. 5.2. Potential direct and indirect interactions in soils impacted by exotic plant invasion. Solid and dashed lines represent direct and indirect interactions, respectively. Illustration by Janet Wingate and reprinted with artist’s permission. vegetation. Belowground, nutrient inputs ecosystem’s above- and belowground inter- from the litter and microbial activity are connectedness increases so does our ability indirectly modifi ed by invasion, as are the to develop management programs based on habitats for microorganisms due to the ecological principles. Historically, much change in the physical and biochemical of vegetation management consisted of composition of roots. Th e rhizosphere will simplistic, one-dimensional approaches, change as the root architecture becomes such as removing an unwanted species or more homogenized by physically and augmenting desirable species. Th is approach functionally similar roots of the invading ignored each species’ infl uence on other species. Biochemically, root exudates and aspects of the system (microsites, litter decomposing root masses will modify the decomposition, soil nutrients, or microbial rhizosphere and energy sources for communities) and ultimately the system as microorganisms, therefore infl uencing the a whole. Ecosystems are collections of diversity and function of soil microbes. numerous interacting organisms and abiotic Understanding direct and indirect conditions that cannot easily be com- interactions develops a framework for the partmentalized or isolated. Progressive, myriad of connections that drive a system process-based management can utilize and provides a conceptual basis to apply ecological principles to manipulate eco- ecological principles to the management of system processes and direct a system to a natural systems. Th e complexity of ecological desired state, possibly with signifi cantly systems has long made it diffi cult to less disturbance than historical eradication understand how an invasive plant or anthro- and revegetation programs. Th e following pomorphic management action modifi es an sections of this chapter will highlight specifi c ecosystem, but as our knowledge of an examples related to direct and indirect 84 T.A. Grant III and M.W. Paschke

impacts from invasive plants on litter Decomposition occurs through three decomposition, nutrient cycling and biogeo- primary pathways: leaching of soluble chemistry, microbial community diversity materials, the comminution (the physical or and function, invertebrates, physical mechanical breakdown) of biomass, and the properties of soil, and allelopathy. Th rough- conversion of fi xed carbon to CO2, H2O, or out this discussion on the impacts and energy via oxidation or catabolism (Seastedt, interactions of invasive plants, the concept 1984). Th ese pathways reduce complex, of utilizing ecological principles to develop organic structures into simpler compounds, innovative management practices will be which are used as an energy source for soil highlighted. Management of ecological pro- fauna and fl ora, aff ect the physical properties cesses can include the manipulation of of the soil (i.e. soil organic matter), and nutrient inputs (litter) or cycling rates, ultimately provide nutrients for plants. pathogens or plant-growth promoting Nitrogen is an essential nutrient for plants microbes, invertebrate herbivory, and and is one of the most commonly limiting chemical interference (allelopathy). Th e con- nutrients in the decomposition of plant sequences of process-based management of matter, because it regulates both the growth systems will probably not be immediate or and turnover of microbial communities obvious, but through the subtle manipu- (Heal et al., 1997). Nitrogen is often the lation of ecological processes it may be focus of invasive plant studies, because it is possible to restore diversity and resilience to a dominant driver of plant growth and is ecosystems. essential to both plants and microbes. Th e infl uence of soil carbon:nitrogen ratios on decomposition and microbial diversity Invasive plant litter and began to be acknowledged in the 1920s decomposition (Heal et al., 1997). Currently, lignin:nitrogen and polymer:nitrogen ratios are also used A logical starting place to study the eff ects of to analyze decomposition pathways. In an invasive species on the soil system is the general, leaf litter with higher N content plant litter being added to the environment. decomposes more rapidly due to pre- Every species has a unique biochemical ferential colonization by bacteria and fungi composition (Ehrenfeld, 2006) and varying populations (Melillo et al., 1982), although amounts of potential litterfall that will C:N ratios are dynamic as decomposition infl uence the system diff erently. Th e progresses and microorganisms turnover. quantity and quality of plant litter will Th e microbial mineralization of N and directly impact the potential nutrients, subsequent death and decay of microbes is biogeochemical cycling, and microbial the primary source of N for most plants diversity of a site. Litter decomposes at (Knops et al., 2002; Schmidt et al., 2007), diff erent rates based on the quality or type although this traditional view of microbial of litter and the environment, which will control of N cycling has recently been directly aff ect the amount and type of challenged by the concept of plants being nutrients fl owing into the soil. Invasive able to directly compete with microbes for plants have been shown to generally have organic N that has been depolymerized by greater biomass or net primary productivity the extracellular enzymes of microbes than adjacent native species (Ehrenfeld, (Schimel and Bennett, 2004; Chapman et al., 2003), greater rates of decomposition 2006). An understanding of the decom- (Ehrenfeld, 2003; Ashton et al., 2005), and position of plant matter is necessary for the higher amounts of nitrogen (N) in litter, development of ecosystem management especially if the species is capable of practices that acknowledge the complex symbiotic N fi xation with bacteria. A review interactions between plants, soil, and of numerous studies by Ehrenfeld (2003) microbes; and ultimately how these pro- found that 9 of 13 papers reported increased cesses infl uence plant communities and litter decomposition with invasion. successional dynamics. Invasive Plant Impacts on Soil Properties 85

Plants directly aff ect nutrient cycling, can alter N cycling and create a self- edaphic characteristics, soil fauna, and perpetuating positive feedback system due microorganism communities of an eco- to its litter quality and quantity has been system by their litter (quantity and quality) thoroughly evaluated, although more recent (Wardle et al., 2004; Georgieva et al., 2005; research integrates the mechanisms Chapman et al., 2006) and root exudates facilitating the feedback processes by released to the rhizosphere (Bais et al., incorporating the analysis of soil micro- 2006). Wardle et al. (2004) proposed the organisms (Hawkes et al., 2005, 2006) and concept of plants as the integrator of above- fauna (De Deyn et al., 2003), edaphic and belowground feedbacks, but emphasized characteristics (Goslee et al., 2003; Grant et the diffi culties in understanding the al., 2003), and the diff erences between mechanisms due to the complexity of organic and inorganic N use by plants organisms and environments involved. Due (Schimel and Bennett, 2004; Chapman et al., to the increased net primary productivity 2006) into the complex systems. (NPP) of many invasive species (Ehrenfeld, Process-based management of natural 2003), litter quality and amounts represent systems requires understanding ecological a starting point for studies of plant invasions processes and developing innovative and their impacts. Diff erent species have techniques that utilize scientifi c principles markedly divergent litter qualities and these to achieve land management objectives. Th e physical characteristics modify their manipulation of litter quantities or qualities nutrient composition, decomposition rates, may be a possible management technique to and potentially the biogeochemistry and reduce the fl ow of energy through an microbial diversity of a system. Additionally, invaded system (Fig. 5.2). Th e goal of leaf litter and root exudates of invasive manipulating litter inputs would be to stress plants may contribute to indirect chemical the invasive species through limiting interference or allelopathy (Bonner, 1950; resource inputs and possibly increasing Muller, 1966; Bais et al., 2003). In the competition with species that are adapted to context of invasive plants, changes in the lower resource levels. Th is method is quality and amount of litter within a system predicated on the assumption that reduced may represent a critical tipping point for an resources would add suffi cient stress to the invasive species to modify many aspects of invasive species to infl uence vegetation the system, including the rate of dynamics. Management of litter inputs decomposition (sensu Ehrenfeld, 2003), the would require knowledge of the system fl ux of nutrients (Evans et al., 2001), and the and how invasive species have modifi ed microorganisms involved (Wardle et al., decomposition, microbial communities, and 2004). nutrient cycling; otherwise it may be Th e decomposition environment can be ineff ective or have unintended consequences. dramatically diff erent between invaded and Two potential strategies for litter manage- non-invaded regions of the same ecosystem. ment should be evaluated further: (i) Studies have shown that in the eastern removal of litter; or (ii) addition of low hardwood forests of the USA, the litter of quality litter (i.e. high C:N ratio). Th e invasive species decomposes more rapidly removal of litter in an invaded system could than native plant litter and invaded eco- reduce nutrient inputs and potentially stress systems decompose litter faster regardless the invasive species because they have high of the litter’s origin (Ashton et al., 2005). A productivity and subsequent nutrient separate study did not determine diff erences demands. Experiments incorporating car- in decomposition between sites dominated bon amendments into soils of invaded by either native or invasive species, but systems or to reduce invasion in areas found that the loss rates of phosphorus, undergoing revegetation or restoration have lignin, and trace elements from litterbags had limited success in reducing N availability was reduced in invaded sites (Pritekel et al., and invasive plants (Perry et al., 2010). 2006). Th e concept that an individual species Augmentation with low quality litter may 86 T.A. Grant III and M.W. Paschke

achieve similar goals as soil carbon amend- plant or microbes drive these changes in ments, although the technique requires nutrients has long been debated, but will experimentation and outcomes could take only briefl y be discussed here as an years to be measurable. Another possible introduction to the topic. At the center of ecological consequence of litter manipulation this question are two competing yet likely would be the changes to microsites for seed co-occurring processes: (i) microbes control germination or establishment. Th e addition N cycling and plants can only access of litter could modify the microsites inorganic nutrients that remain after (temperature, sun light, moisture) and make microbial turnover (Knops et al.’s (2002) them less hospitable for germination by microbial N loop); and (ii) plants are actively invasive species, although this requires competing with microbes for organic N that in-depth knowledge of an invasive species’ is made available due to extracellular autecology. Th e inherent connection depolymerization by microbes (Schimel and between plant litter and plant-available Bennett, 2004; Chapman et al., 2006). nutrients emphasizes the importance of Knops et al.’s (2002) theory infers that the considering ecological principles related to type of litter and its quantities do not aff ect decom position as a potential method to nutrient cycling, at least not as much as site manage ecosystem processes and direct a or species-specifi c impacts on N inputs and system towards a desirable vegetation state. losses that are based on factors such as fi re, leaching, atmospheric deposition, and symbiotic N fi xation. Additionally, because Nutrient cycling and biogeochemistry the microbial-N loop controls the fl ux of N, Knops et al.’s (2002) theory places less In the mid-20th century a mechanistic importance on issues of litter quality and understanding of decomposition led to the quantity, therefore invasion by an exotic formation of many modern ecological would not greatly aff ect N levels, assuming theories and subsequently stressed the that there is no loss or gain of species capable importance of nutrient transformations and of symbiotic N fi xation or modifying major cycling in the maintenance of ecosystem disturbance cycles. An alternate theory function (Heal et al., 1997). Plant prod- posits that plants compete with microbes uctivity and diversity is inexplicably linked for organic N and that the consequences of a to nutrients, although understanding this change in species due to invasion will have interaction has proven diffi cult. Historically, great ramifi cations on the system’s bio- a majority of research on invasive species geochemistry, both as the litter type and and biogeochemical cycling of nutrients has quantity change, but also as the net primary focused on N, since it is a primary limitation productivity (NPP) and nutrient require- of productivity in terrestrial ecosystems ments of the invasive species are modifi ed. (LeBauer and Treseder, 2008). Many Given the latter scenario, the subsequent invasive species form symbiotic relationships impacts of an invasive plant species’ with N-fi xing bacteria and this interaction is dominance on an ecosystem will include all capable of dramatically increasing the aspects of nutrient cycling, including eff ects amount of N in a system (Ehrenfeld, 2003). on the richness, diversity, and functioning In contrast, other invasive species can of the microbial community. reduce the amount of N symbiotically fi xed A review of the impacts of exotic plants by native species (Wardle et al., 1994) and on nutrient cycling by Ehrenfeld (2003) thus decrease the amount of plant-available provides an excellent summary of our N in a soil. Changes in plant species current knowledge. An overarching theme of composition due to invasion will likely this review is that general trends are diffi cult modify a system’s biogeochemistry as to identify and both positive and negative changes occur in the quantity and quality of impacts from invasion are found for C, N, litter inputs, root architecture and exudates, and water (Ehrenfeld, 2003). Th is lack of and microbial communities. Whether the uniformity should not be a surprise given Invasive Plant Impacts on Soil Properties 87

the incredible variability of natural systems of plant-available, inorganic N and the and the context dependency of ecological rates of mineralization, nitrifi cation, and theory. Regardless of the variability atmospheric fi xation (sensu Ehrenfeld, highlighted in Ehrenfeld’s (2003) review, the 2003), although exceptions exist. Th e following list includes many important increased NPP of invaded areas may also generalizations concerning the impacts of require higher amounts of N to support exotic species on soils: larger standing crops, but also deposit greater amounts of litter. Th e cyclical nature of plant productivity, decomposition, 1. Exotic plants often have greater nutrient availability, and plant nutrient aboveground biomass, net primary pro- requirements are important to consider ductivity, higher shoot to root ratios, and when assessing how invasive species aff ect a faster growth rates than co-occurring native system. species. A study of wiener-leaf or saltlover 2. Exotic plant litter usually decomposes (Halogeton glomeratus) invasion in the cold faster than co-occurring native species. deserts of Utah, USA documents distinct 3. Exotic plant soils generally have more increases in nitrate and phosphorus con- extractable inorganic N than soil from centrations in the invaded area compared to co-occurring native species. the adjacent ecotone and uninvaded areas 4. Soils of exotic species frequently have (Duda et al., 2003). Interestingly, the increased rates of N mineralization and ecotone had signifi cantly less ammonium nitrifi cation. than the adjacent invaded or uninvaded 5. Exotic plants capable of symbiotic N areas, which could signify higher mineral- fi xation can dramatically aff ect N cycling. ization and nitrifi cation rates. Duda et al. 6. Exotic species can aff ect symbiotic and (2003) also detected much higher salts (Na, non-symbiotic N-fi xing microorganisms Ca, K), organic matter, and bacterial that are associated with native plants. functional diversity (BIOLOG substrate 7. Exotic species can infl uence the spatial analysis) in the H. glomeratus infestations. distribution and temporal fl ux of nutrients, Due to a limited experimental design, the even if overall quantities of nutrients are authors could not rule out the possibility not aff ected. that wiener-leaf preferentially invaded soils 8. Both positive and negative changes in with specifi c nutrient and microbial soil carbon, N, and water are associated with characteristics, although they state that ‘the exotic species. possibility of pre-existing gradients fails to When considering the eff ect of an invasive explain the patterns in our data.’ Cheatgrass plant on nutrient dynamics, it is important (Bromus tectorum) is one of the most to assess all aspects of nutrient cycling. Th e problematic and widespread noxious weeds following examples will focus on N due to in the western USA. A study in Canyonlands the breadth of research on this essential National Park of Utah, USA found that nutrient. Th e plant–soil N cycle receives cheatgrass invasion was associated with a additions from atmospheric fi xation by decrease in the amount of plant-available N bacteria or lightning, transformations by and N mineralization, and increased N mineralization, immobilization in living immobilization (Evans et al., 2001). Th e tissues, and losses by denitrifi cation and researchers found that the changes to N leaching. Each of these components can be dynamics were linked to increased litter impacted by a change in the dominant quantity, changes in the quality of litter and vegetation and diff erent species will have in the amount of soil organic matter directly unique impacts on N cycling, including modifying soil carbon to nitrogen ratios. indirect eff ects related to a species Surprisingly, the impacts of cheatgrass on changing disturbance regimes that can the N cycle occurred within 2 years of dramatically aff ect N (i.e. fi re). In general, invasion. Additional research has docu- many invasive species increase the amount mented higher nutrient levels in invaded 88 T.A. Grant III and M.W. Paschke

habitats of Europe for several invasive Vitousek and Walker, 1989). N2 fi xation by species and prescribed the eff ect to increased M. faya increased the amount of N in the NPP of the invasive species compared to system and dramatically altered ecosystem native vegetation (Vanderhoeven et al., development by increasing the amount of 2005; Dassonville et al., 2007). In a exotic species following the decline of M. comparison of fi ve exotic species (Fallopia faya (Vitousek and Walker, 1989; Adler et al., japonica, Heracleum mantegazzianum, Prunus 1998) and increasing the potential for fi re serotina, Rosa rugosa, and Solidago gigantea) (Adler et al., 1998). Only by a thorough in Belgium, Vanderhoeven et al. (2005) study of the impacts of invasion and the found signifi cant increases in potassium and discovery of a novel source for N input to the manganese in the invaded sites compared to system was it possible to understand the the adjacent uninvaded areas. Another study long-term impacts of M. faya on vegetation in Belgium by Dassonville et al. (2007) succession, ecosystem function, and large- concluded that F. japonica increased nutrient scale disturbances (fi re). cycling and topsoil fertility. Th ese few Russian olive (Elaeagnus angustifolia) is examples illustrate the variability of invasive an intentionally introduced tree species that species’ impacts to the many diff erent is currently getting a great amount of aspects of nutrient cycling and the capacity attention in the western USA. It invades for invasive species to modify ecosystem riparian systems and modifi es large scale function. biogeochemical cycling. Th e species also Novel nutrient acquisition strategies may forms symbiotic associations with Frankia aid invasion by an exotic species, especially spp. and has leaf and litter N levels that are if the system being invaded is limited in a nearly double that of the native cottonwoods specifi c nutrient or lacks a species capable of (Populus spp.) (Katz and Shafroth, 2003). a unique strategy (i.e. actinorrhizal nitrogen Control or restoration of Russian olive- fi xation). Due to the organism’s inherent invaded areas has proven diffi cult and has ability to obtain nutrients, an empty niche primarily focused on chemical and mech- may be available for the exotic species to fi ll. anical methods, although managing rivers Invasive species that are capable of to simulate historic fl ood regimes and dinitrogen-fi xing symbioses with bacteria promote recruitment of native cottonwoods may have an advantage in the novel is being tested (Lesica and Miles, 2001). Th e environment and the diazotrophic relation- signifi cant amount of N added to invaded ship may have large-scale and long-term areas may have long-term eff ects on impacts on nutrient availability, vegetation revegetation success or promote invasion by composition, and disturbance dynamics of exotic forbs or grasses. Similar issues have the system. Th e invasion of the Hawaiian been noted with invasive N2-fi xing black Islands by Myrica faya (fi retree) represents a locust (Robinia pseudoacacia) in Europe and pivotal point in our scientifi c understanding parts of Asia (Weber, 2003). of how an invasive species can impact an Strategic management of nutrients to ecosystem, modify nutrient dynamics, and reduce invasive plant populations is context alter successional processes. Myrica faya is a specifi c and requires knowledge of how a small tree from the Canary and Azores target species or plant community will react islands that invaded relatively young to changes in nutrient cycling (Perry et al., volcanic substrates in Hawaii beginning in 2010). Managing weeds by manipulating the late 1800s. Th e invasive tree is capable of ecological processes attempts to utilize the forming symbiotic relationships with inherent transformations and interactions N2-fi xing actinomycetes (Frankia spp.). No of an ecosystem to achieve a management other species in this ecosystem develops goal. Nutrient cycling is inexplicitly linked actinorrhizal symbioses and this relationship to litter inputs, decomposition, and dramatically changed the inputs and amount microbial communities. Some species will of biologically available N in these nitrogen probably not be aff ected by anthropogenic limited systems (Vitousek et al., 1987; attempts to reduce nutrients and stress Invasive Plant Impacts on Soil Properties 89

invasive species, while others may become theory, much less management of plant less competitive and make it easier to invasions. Considering the role of establish desirable species. Overall, microorganisms in the mineralization of nutrients are only one component of the nutrients, decomposition, N fi xation, soil complex interactions that determine aggregation and aeration, and their positive vegetation dynamics. Generally, when or negative growth eff ects on plants, it is managing early seral, r-selected weedy essential to assess the impacts of invasive species, reducing nutrient availability and species on microbial communities. Field and increasing competition from aggressive greenhouse experiments have documented native species may be a successful approach. diff erent soil microbial communities in the Conversely, some invasive species do not fi t soils of diff erent plant species (Kourtev et into the early-seral concept and may not be al., 2002) and the infl uence of unique plant aff ected by nutrient reductions (i.e. long- species on the diff erentiation of the soil lived rhizomatous species such as Acroptilon microbial community (Westover et al., repens). Hypothetically, adding nutrients 1997). Recent research has documented and revegetating with aggressive early-seral, how the soil microbial community diff ers native species could assist in establishing between native and invasive plant species species that can compete with the more (Hawkes et al., 2005, 2006; Klein et al., 2006) K-selected invasive species or at least create and that these changes in composition and an opening for establishment of competi- function can aff ect nutrient cycling and tive species. Applying Davis et al.’s (2000) availability. It is broadly acknowledged that theory of fl uctuating resources and the soil microorganisms mediate or regulate subsequent invasibility of an ecosystem to nutrient cycling in the soil and our recent the manipulation of nutrients and revege- understanding of the indirect eff ects of tation (pseudo-invasion) with native species plant invasion on microbial composition may provide eff ective methods to restore highlights the importance of understanding dominance of native plants in certain the cascade of eff ects invasion will cause on situations. Manipulation of a system’s bio- soil microbe composition, nutrient cycling, geochemistry to direct plant community and potentially pathogen accumulation dynamics must be based on a strong (Fig. 5.2). Th ese eff ects can infl uence the understanding of ecological principles that dominance of invasive species through drive processes, otherwise unintended con- feedback cycles, but will also aff ect our sequences may occur. ability to target specifi c ecosystem processes to achieve management goals in an effi cient and timely manner. If the impacts of Microbial communities invasion fundamentally alter the com- position of microorganisms and ultimately Th e soil is often represented as a black box in the biogeochemical functioning of eco- ecological experiments and studies of plant systems, it becomes critical to recognize invasion (Kardol et al., 2006; Kulmatiski and these changes and adapt management Beard, 2011). Th e diversity of prokaryotic practices to the novel conditions created by species (bacteria and archaea) is potentially the invasive species. Th e following examples in the millions, while only approximately highlight how plant invasion can change the 4500 species have been identifi ed (Torsvik et soil microbial community and indirectly al., 2002). Approximately 170,000 soil aff ect ecosystem processes. organisms have been identifi ed with the Th e impacts of invasive species on soil largest group being fungi (Wall and Virginia, microorganisms are often determined by 1997; Wall and Moore, 1999). Our limited microbial mediated changes to the system’s understanding of soil organisms and their biogeochemistry and research has frequently interactions with plants and the environ- focused on N due to plants’ heavy reliance ment has made it diffi cult to incorporate on this essential macronutrient. A recent these complex systems into ecological study by Hawkes et al. (2005) documented 90 T.A. Grant III and M.W. Paschke

increased amounts of nitrifying bacteria and 2004) depending on the role of fungi in unique DNA signatures (restriction length plant nutrient uptake, nutrient patterns based on polymerase chain reaction immobilization and turnover of fungal (PCR) methods) in experimentally grown hyphae, and plant root responses to fungal monocultures of exotic grasses compared to infection (carbon exudation). Busby et al. (in monocultures of a native grass, forb, or press) have suggested that the recovery of polycultural mixtures of exotic and native AMF communities after invasion by exotic species. Th e authors related these changes in cheatgrass may be dependent on the species the microbial community to diff erent plant identity of native plant used for restoration. compositions and linked the impacts of the Wolfe and Klironomos (2005) provide an exotic grass(es) and modifi ed soil microbial excellent overview of specifi c invasive community to functional changes in the species and their documented eff ects on the system’s nutrient cycling. A fi eld-based structure and function of native soil com- study in the forests of the northeastern USA munities. Based on the examples provided, documented diff erent microbial community invasive species generally decreased AMF, composition and function in the soils of two fungi abundance, or diversity, although the invasive and one native understory species results varied between species (Wolfe and (Kourtev et al., 2002). Th e modifi cation of Klironomos, 2005). soil communities was strongest in the An interesting feedback system was rhizosphere soils, but surprisingly the recently documented in India in which an impact was also documented in nearby bulk exotic, invasive plant (Jack in the bush, soil. Using canonical correlation analysis the Chromolaena odorata) promotes the growth study found that changes in the function of of a native, generalist soil pathogen the soils were correlated to changes in the (Fusarium sp.) and subsequently creates a microbial composition and structure. A negative eff ect on native plant species greenhouse experiment by the same (Mangla et al., 2008). Th e root exudates of C. researchers and with the same plant species odorata were shown to promote Fusarium replicated the fi eld results and also identifi ed growth in non-invaded soils and activated increased nitrifi cation rates and pH in the carbon (AC) reduced the promotion of the soil of one exotic species (Kourtev et al., fungi by the root exudates. Th is unique 2003). feedback pathway illustrates the complexity Th e belowground diversity of arbuscular and variability in the reaction between mycorrhizal fungi (AMF) has been directly plants and microorganisms, regardless of related to the functioning and stability of their origin (i.e. home versus foreign). plant communities. At low AMF diversity, Although not a direct impact of an invasive plant community composition has been species on the soil, several experiments have shown to fl uctuate greatly (lack of stability), documented exotic species experiencing less while high AMF diversity promoted greater negative impact from microbial pathogens nutrient capture and productivity (van der (Mitchell and Power, 2003) or accumulating Heijden et al., 1998). We expect exotic, pathogens at a slower rate than native invasive species to have diff erent AMF species (Klironomos, 2002; Eppinga et al., communities than neighboring native 2006) and have hypothesized that this species due to inherent physiological and release from enemies contributes to the phenological diff erences between plant invaders’ success. species, but invasive species have also been Microorganisms represent an incredible shown to cause changes in the fungal breadth of diversity, although our under- diversity of co-occurring native species standing of species and functional groups following invasion (Hawkes et al., 2006). Th e is limited. Th e regulation of decomposition consequences of change to the AMF and nutrient cycling by microbes requires community of an invaded area are unknown, in-depth research and elucidation if but are potentially an important mechanism we intend to manage ecosystems through in successful plant invasion (Callaway et al., the manipulation of ecological processes. Invasive Plant Impacts on Soil Properties 91

Molecular techniques (PCR and DNA com- with diff ering mesh sizes to exclude certain munity profi ling techniques and sequencing), soil biota (Seastedt, 1984; Huhta, 2006), fatty acid analysis (phospholipid fatty acids, while others have applied chemicals (i.e. fatty acid methyl esters), and carbon naphthalene) or X-rays to eliminate substrate utilization (i.e. Biolog plates and arthropods (Newell et al., 1987). Both substrate induced respiratory responses biocide methods have been shown to have (SIR)) methods are beginning to classify and non-target eff ects on soil fungi (Newell et describe the functional traits of micro- al., 1987), although the results of litterbag organisms. As this knowledge base expands studies have overwhelmingly shown that and the methods become more consistent soil fauna increase the decomposition of and less expensive, the classifi cation of a plant litter (Huhta, 2006) and can have plant community’s microbially regulated variable eff ects on nutrient cycling. A well- nutrient cycling will make it possible to replicated fi eld study in Ohio used incorporate microbial communities in land electroshocking of soil to reduce earthworm management practices. When we understand populations without non-target eff ects on how an invasive species modifi es the microarthropods, nematodes, or micro- microbial community, and subsequently organisms (Bohlen et al., 1995). In the decomposition and biogeochemistry, it will context of invasive plant research and be realistic to attempt to modify micro- management, the role of soil invertebrates organisms in ways that enhance desirable is relevant due to their infl uence on plant communities and suppress unwanted decomposition rates and nutrient availability. species. Th is will probably focus on managing Comminution aff ects the size and surface N fl uxes and availability or utilizing species’ area of litter, which consequently modifi es specifi c pathogens. Ecological principles many factors that regulate invertebrates guided by system and species-specifi c and microorganisms, including: microsites, knowledge can lead innovative management predator–prey relationships due to size practices that manipulate processes to limitations, and access to water or nutrients. achieve land management goals, although, Although bacteria and fungi are the primary because microbes are the smallest and most decomposers of organic matter, soil fauna numerous, their interactions may be the are intricately involved in the physical most complex to grasp for management processing and movement of plant biomass purposes. Th is novel type of management (fecal excretion) (Davidson and Grieve, will require a massive increase in our 2006), and the direct or indirect regulation understanding of plant–microbe inter- of soil microorganism communities by their actions and the potential feedbacks that feeding habitats (Seastedt, 1984). Due to could ensue when we begin to tinker with diffi culties identifying and describing soil complex systems. fauna, the organisms are frequently grouped into the following functional groups based on Brussaard et al. (1997). Macrofauna Soil invertebrates consist of root herbivore insects, termites, ants, and earthworms. Mesofauna include Soil fauna play an important role in many mites, collembola, and enchrytraeids. Micro- ecological processes including decomposition fauna are protozoas, ciliates, and nematodes. of biomass via comminution, root herbivory, Th e macro- and mesofauna generally movement of nutrients, and modifi cation of decrease particle size of plant litter and bacterial and fungal communities through indirectly increase surface area (habitat for feeding activities. Th ese activities can microorganisms) and mobilization of infl uence plant succession directly and nutrients. Microfauna and some mesofauna indirectly, although the impacts of soil fauna (mites and collembola) graze on fungal and how to incorporate these eff ects into spores and bacteria. Soil fauna can directly ecological management remain controversial impact plants by root herbivory, spreading or unknown. Early studies utilized litterbags of pathogens, and modifi cation of soil 92 T.A. Grant III and M.W. Paschke

microorganism populations through decom position and possibly nutrients), and predation. Invasive species will directly predation of microorganisms (pathogens, N change the inputs to the soil (litter, roots, fi xers, decomposers) (Fig. 5.2). Th e and exudates) and therefore the habitat and utilization of soil fauna to infl uence any of food sources for soil fauna are likely to be these ecological processes will require modifi ed with invasion. In order to thorough knowledge of the specifi c plant– adequately understand the impacts of soil system to successfully achieve manage- invasive plants, the interactions of soil ment goals. invertebrates with microorganisms, the rhizosphere, and the physical properties of soil must be incorporated into ecological Soil physical properties studies. Previous studies and several review papers (Seastedt, 1984; Huhta, Th e physical properties of soil are the results 2006) provide a general framework for of long-term interaction between a region’s incorporating the eff ects of soil fauna into geology, climate, and biota. Hans Jenny plant–soil interactions and plant suc- (1941) fi rst described the formation of soil cessional dynamics. Many studies are based with a simple function: upon litter quality (i.e. labile or resistant) and initial soil N levels as important factors soil = f (climate, organisms, topography, in the interactions of microfl ora and -fauna. parent material, and time) Huhta (2006) summarized the role of soil fauna in N-limited systems as ‘generally Each of these factors will aff ect the enhance(ing) decomposition and mineral- composition of soil and subsequent plant– isation, whereas in the presence of excess soil feedbacks. Th e incredible variability in nitrogen they have little eff ect.’ soil development and interactions with Soil fauna may alter succession due to plants will make it diffi cult to identify selective predatory eff ects upon the consistent impacts on the soil due to plant dominant species, therefore releasing the invasion, but an understanding of the soils, sub-dominant plant species from com- plants, and microbial species involved petition and facilitating the succession and can identify novel ecological interactions development of a more diverse and hetero- and improve our management practices geneous vegetation community (i.e. less minimizing invasive species or restoring dominance by any species based upon native fl ora. As the medium for plants, Simpson’s evenness index) (De Deyn et al., microbes, and nutrients to interact, the soil 2003). In the context of invasive species, is often represented as a black box due to its Mayer et al. (2005) documented increases complexity and poorly understood systems. in decomposition when macro-detritivores Yet it is also the arena in which many of the were allowed access to litterbags and the impacts from a change in biodiversity or amount of decomposition correlated posi- dominance by an invasive species will be tively with increasing cover of an invasive manifested. Impacts of invasion on the grass (Festuca arundinacea). In general, our soil’s physical properties are highly variable understanding of the interactions between and probably species and site specifi c, but soil invertebrates and invasive species is too can include the following characteristics: limited to base management practices upon. soil moisture content, salinity, pH, organic Th e addition or removal of soil fauna are matter content, soil aggregation, and micro- possible, although the practice must be climate eff ects. based on ecological principles and a sound Invasive species frequently have greater under standing of how the management will biomass than surrounding native vegetation aff ect the system and target species. Th e and therefore probably require more primary manners in which soil fauna can nutrients and water, but also increase litter infl uence vegetation are through root inputs to the system. Consequently, these herbivory, comminution of litter (increased dramatic changes to the productivity, water Invasive Plant Impacts on Soil Properties 93

usage, nutrient cycling, and inputs to the saltcedar on water resources remain unclear, system will aff ect many aspects of the soil, but the species has signifi cant eff ects on soil including pH. In New Zealand, studies have salinity (Nagler et al., 2008) and riparian shown that mouse-eared hawkweed forest structure. (Hieracium pilosella) decreased soil pH by Soil aggregation can be used as a segway approximately 0.5 units (McIntosh et al., to generically describe soil quality, because 1995). Conversely, Kourtev et al. (2003) more stable aggregates are less prone to documented increased soil pH in greenhouse erosion and hold greater amounts of water, incubations of the exotic Japanese stiltgrass nutrients, and carbon (Batten et al., 2005). (Microstegium vimineum) compared to a Soil structure or aggregation is frequently native blueberry. Ehrenfeld’s (2003) review studied through the quantifi cation of highlights the variability of soil pH following glomalin, a glycoprotein that is produced by occupation by an exotic species. A arbuscular mycorrhizal fungi (AMF) and is continental-scale study of soil microbial positively correlated with the stability of soil composition in North and South America aggregates (Lutgen and Rillig, 2004). found that diff erences in plant species Because many invasive plants aff ect soil richness and diversity were largely explained microbial communities, it is of interest to by soil pH and plant community (Fierer and determine if the impacts cascade to soil Jackson, 2006), although the paper did not aggregate stability or soil quality. A study of directly address invasive species. chemically and mechanically controlled Water use by invasive species may alter spotted knapweed (Centaurea stoebe) evapotranspiration and overall water usage infestations found that total glomalin levels rates, which can cause changes in soil and AMF hyphal lengths were negatively moisture content, water table levels, and correlated with percent cover of the invasive salinity. Th e impacts of saltcedar (Tamarix plant, but did not detect a reduction in spp.) in the southwestern USA are hotly aggregate stability (Lutgen and Rillig, 2004). debated, because the major river systems it Th e authors stated that soil aggregate water infests supply water to millions of people stability was initially high at the study sites and have numerous contractual and inter- and that ‘spotted knapweed may exert a national obligations. Th e prolifi c invasive deleterious eff ect on soil structure’ in areas species was thought to use more water than with lower initial stability. Preliminary native species and was targeted for evidence of the invasive Jack in the bush eradication by states and municipalities in (Chromolaena ordorata) improving soil order to salvage water for anthropocentric structure through the promotion of earth- uses. Early measurements of saltcedar water worm activity was documented in eucalyptus usage (>200 gallons per tree per day) may plantations in the Congo (Mboukou- have been inaccurate and overestimated the Kimbatsa et al., 2007). Th e variability of economic benefi ts of control methods responses in soil aggregate stability (Owens and Moore, 2007). Water salvage following invasion exemplifi es the species experiments that focused on saltcedar specifi city of impacts and the importance of control have not been as successful as considering initial soil conditions and the expected (Shafroth et al., 2005). Th e impacts wide variety of ecosystems when assessing of saltcedar on riverine and groundwater are an exotic species impact. diffi cult to measure consistently due to issues of scale. Plant water use measurements are conducted at the leaf, stem, plant, or Allelopathy and invasive plants ecosystem scale and comparisons across scales can be inconsistent, although a Although the concept of allelopathy or general trend for greater water use by chemical interference between plants has invasive species in drier, hotter climates and long been postulated, many diffi culties have at larger scales has been documented been encountered in the detection and (Cavaleri and Sack, 2010). Th e impacts of quantifi cation of this elusive interaction. 94 T.A. Grant III and M.W. Paschke

Th e success of many invasive plant species plants into the scientifi c realm in the early has been attributed to allelopathy, primarily 1800s. In 1832, a system of crop rotation through the soil matrix, and therefore it is was developed by the botanist A.P. important to consider the potential impacts DeCandolle based upon his research into the of allelopathic invasive species on soil and interspecifi c inhibitory eff ects of certain potential management or restoration. agricultural species upon others (Bonner, Allelopathy specifi cally describes the release 1950). Th e early theories on chemical of a chemical into the environment by a interference or allelopathy were dismissed plant or microorganism, via exudation, by many researchers as information con- volatization, or transformation of biomass, cerning the depletion of soil nutrients and which has a direct or indirect positive or competition for these minerals and water negative eff ect on another species (Rice, formed the prevailing theory in plant 1984). Currently, the discussion of allelo- interactions (Bonner, 1950). Additional pathy focuses on negative or inhibitory research in the early to mid-1900s detected eff ects on a plant due to the release of a toxic substances in or around many plant chemical by another species and usually species, including: the leaves of a desert ignores potential stimulatory eff ects. If shrub (Encelia farinosa), the leaves and roots chemical interference can be determined to of black walnut (Juglans nigra), the roots of aff ect plant growth and the availability of smooth brome (Bromus inermis), and the water or nutrients within natural plant soils of peach and rubber tree plantations communities, the consequences of allelo- (Bonner, 1950). Th e major problem in pathic induced changes in vegetation com- determining that chemical inhibition was position or succession within ecosystems infl uencing the vegetation composition and should be considered in addition to the plant community succession related to the traditional view of ecological theory that is lack of evidence connecting the known based primarily upon the competition for phytotoxins to their release from the plant, resources (Bonner, 1950; Muller, 1966). A accumulation in the soil, and the mechanism major point of contention concerning that negatively aff ects the surrounding allelopathy has assumed that the eff ects are vegetation. Many of the early experiments direct and therefore measurable. We could detect phytotoxic chemicals, but they emphasize the importance of understanding could only be correlated with the inhibition that the impacts of potentially allelopathic of neighboring plants. Direct evidence root exudates are most likely weak, indirect regarding the release of the chemical and and will occur over long time frames and how it interacts with the soil and neighboring involve multiple scales, including inter- plants is still proving diffi cult to document. actions with microbes and subsequent Th e fact that most allelopathic eff ects are changes to soil chemistry. For these reasons, weak relative to other factors suggests that we promote the concept of soil chemical any impacts on neighboring plants would ecology over allelopathy, because it play out on longer time scales than are highlights the complexity of interactions in typically considered in ecological studies. a plant–soil system and removes much of Slow and chronic antagonistic eff ects would the historical controversy surrounding be diffi cult to document against a backdrop allelopathy (Inderjit and Weiner, 2001). of other competitive processes. Th e fi eld of Potential allelopathic interactions in allelopathy was thus heavily criticized in the chick pea (Cicer arietinum) was fi rst noted by mid-1900s due to the correlative nature of Th eophrastus around 300 BC (Rice, 1984). the studies and the vast amount of research Although many years have passed, there is supporting competition for resources as the still much confusion about both the primary driver in plant interactions and defi nition of allelopathy and its detection in successional dynamics. natural ecological systems. Agricultural Th e subject of allelopathy came into the problems related to ‘soil sickness’ brought forefront of science again in the 1960s, the issue of chemical interference between primarily due to the work of C.H. Muller on Invasive Plant Impacts on Soil Properties 95

the bare zones surrounding the aromatic mechanical, or biological control methods shrubs of coastal California. Muller’s (1966) and occasional follow up with revegetation well-known paper indicated that certain with native plants or restoration of a specifi c shrubs produce phytotoxic substances that ecosystem process (i.e. hydrology). Although could inhibit the establishment of seedlings these control techniques have proven (intra- and interspecifi c) and therefore are eff ective in some situations, they are also important factors in the diversity and cost prohibitive and frequently have non- successional changes in a plant community. target or unintended eff ects. Novel eco- Although Muller had considered the eff ects system management practices attempt to of herbivory and granivory on the bare utilize our understanding of ecosystem zones, his work and allelopathy as a whole interactions and processes to manipulate came under intense scrutiny with the plant communities towards a desired publication of Bartholomew’s (1970) paper outcome. Th e manipulation of soil nutrients, concerning the role of animals in the bare microorganisms, invertebrates, and soil zones between the shrub and grassland chemistry may provide low impact methods communities. to achieve resource management objectives, To this day, the separation of chemical although many contingencies exist when interference and resource competition is working with specifi c species and eco- still the largest methodological hurdle in the systems. Th e immediate results of these drive to detect and understand allelopathy novel practices may not be as dramatic as (Muller, 1966; Weidenhamer, 1996; Wardle eradication methods, but by attempting to et al., 1998; Romeo, 2000; Ridenour and work within an existing system we reduce Callaway, 2001; Inderjit and Callaway, 2003). the amount of disturbance and may promote Th e importance of conducting appropriately more resilient and stable ecosystems in the controlled laboratory experiments (Inderjit long term. Here we off er a few illustrative and Dakshini, 1995) (i.e. realistic toxin examples. concentrations) and similar fi eld-based experiments (Inderjit et al., 2001) is essential to isolating chemical interference Carbon addition and fertilization from resource competition. Experiments on allelopathy should include density- High N availability has been shown to dependent factors (competitors and/or facilitate invasion by exotics (sensu Perry et chemicals) and methodologies that attempt al., 2010), primarily because N is a major to separate resource competition from limiting resource in most communities chemical interference, such as activated (LeBauer and Treseder, 2008) and promotes carbon or resource addition and removal the rapid growth of early-seral species, often (Romeo, 2000; Inderjit and Callaway, 2003). to the detriment of late-seral native species. Methodologically, allelopathy research Nitrogen levels have been dramatically needs dramatic improvement in fi eld-based increased worldwide due to anthropogenic studies and more in-depth knowledge of the fi xation and increased agricultural growth of chemical agents and their potential multi- N-fi xing legumes (soybeans and lucerne) functionality over the ecologically meaning- (Vitousek et al., 1997). Long-term research ful time spans that weak interactions are in the shortgrass steppe of Colorado, USA likely to play out. has shown that fertilization with N increased the abundance of annual grasses and forbs compared to perennial plant growth in Select Methods to Manage Soil control and carbon-addition treatments Processes for Invasive Species (Paschke et al., 2000). Control methods that Control extend the focus beyond eradication and directly address the causes of invasion (i.e. Traditional management of invasive species high N availability) are more likely to prevent has focused on eradication with chemical, reinvasion and achieve long-term manage- 96 T.A. Grant III and M.W. Paschke

ment objectives (Perry et al., 2010). Many Activated carbon as a tool for methods exist to reduce N in an ecosystem management including carbon addition, burning, grazing, biomass and top soil removal. Carbon In recent years, activated carbon (AC) has addition causes N to be temporarily immo- had resurgence as an experimental treat- bilized in microorganisms, while the other ment to study allelopathic interactions methods remove N from the system. A between plants and potentially as a tool to reasonably large body of evidence has shown minimize the impacts of some invasive that carbon addition reduces or prevents plants. Activated carbon is known to sorb plant invasion, primarily by reducing the large, organic molecules indiscriminately growth of invasive species that have high and therefore is a ‘blunt’ tool for studying nutrient demands and therefore increasing the highly complex chemical interactions in the competitiveness of native species, or the plant–soil interface. It is produced from disrupting feedbacks between exotic plants charcoal, wood, or nutshells and has an and nutrient cycling (Perry et al., 2010). incredibly high surface to volume ratio. Carbon is usually added in the form of sugar, Th e large surface area and pore volume gives sawdust, or wood chips and sequesters N by the compound the ability to sequester increasing microbial growth and activity, organic molecules, including phytotoxic thus causing more N to be immobilized in root exudates and organic nutrients. Many microbes. Alpert (2010) provides an suspected allelopathic compounds are excellent overview of the results of carbon secondary metabolites (Muller, 1966), such addition in diff erent ecosystems, amounts as sesquiterpene lactone or polyacetylenes, of carbon required, and a timeline of and presumably would be bound by AC. impacts. Th e monetary cost and application Experimentally, AC has been shown to be of carbon to a system can be expensive and eff ective in modifying competitive out- disruptive, especially on a large scale, comes between native and invasive plants although the method may be practical for (Mahall and Callaway, 1992; Callaway and the management of important, high value Aschehoug, 2000; Abhilasha et al., 2008), sites. although many methodological hurdles exist Conceptually similar to carbon addition, in our understanding and utilization of AC. removal of litter or biomass may be a In addition to AC’s eff ects on chemical potential tool to reduce invasive plant interference, recent research on the impacts populations, especially if litter and nutrient of AC to microbial communities support the cycling in the invaded habitat are enhanced compound’s role in aiding the restoration of or accelerated by the exotic species’ presence. native plant communities beyond a treat- Th e removal of plant material will reduce ment consisting of only seeding with native nutrient inputs to the system and reduce the species (Kulmatiski, 2011). competitiveness or growth of species that Due to the physical and chemical have high nutrient demands. Another properties of AC, additional ‘non-target’ potential technique may be the addition of eff ects are known to occur and can confound low quality litter to a system. Th e decom- experimental results (Lau et al., 2008). A position of low nutrient biomass will aff ect study by Ridenour and Callaway (2001) nutrient cycling, microbial communities, found that AC decreased the rate of water and could have benefi cial manage ment loss in soil. It has been hypothesized that AC outcomes. If applied in conjunction, these could decrease microbial activity due to the two methods may be a low impact, cost- sequestering of organic compounds and a eff ective method to slowly reduce popu- subsequent reduction in bacterial transform- lations of invasive plants. Th e challenge ations of N (Kulmatiski and Beard, 2006). however is to concomitantly establish Kulmatiski and Beard (2006) found that desired native species in the site with organic N and C were decreased in the lowered nutrient availability. presence of AC, but inorganic nitrate Invasive Plant Impacts on Soil Properties 97

increased, possibly due to a decrease in or invertebrate communities. Although microbial activity. A separate study found an research in this area is limited and many alteration of the microbial community contingencies exist, the manipulation of soil composition with the addition of AC and an microbial or invertebrate communities is a increase in carbon sequestration compared potential method to minimize or to other organic adsorbing compounds modify the negative consequences of included in the experiment (Pietikainen et invasion or assist in the restoration of al., 2000). invaded communities (Boyetchko, 1996). A recent review of AC by Lau et al. (2008) Inoculation with N-fi xing bacteria has long documented experimental artifacts in the been used to promote plant growth or higher use of AC and variable species-specifi c N levels in the soil and it may be possible to responses to the amendment. Th e study add soil microfauna or fl ora that will found an increase of plant biomass with AC promote native species or have a detrimental addition in most of the species studied, eff ect on invasive plants, potentially via increased potting soil pH, and positive or pathogen accumulation. Similar to biological negative changes in the amounts of specifi c control with insects, the addition of nutrients. Plant-available N was signifi - pathogens from an invasive species’ native cantly increased in the presence of AC even habitat may help to suppress the species. though the amount of N in the AC was Another method is to reduce or remove relatively small (0.549% to 0.637% depend- microbes or invertebrates with X-rays or ing upon the source of AC). Methodo- chemicals, possibly to reduce pathogen logically, the confounding eff ects of AC may accumulation on native or rare species, be minimized or ‘controlled’ by an under- which have been shown to accumulate standing of the treatment’s eff ect on pathogens more rapidly than exotic plants individual species and in interspecifi c (Klironomos, 2002; Mitchell and Power, competitive scenarios with and without the 2003). Th is action must be specifi cally addition of AC and fertilizer (Lau et al., tailored to a known ecological interaction 2008). Due to the indiscriminate nature of with the invasive species. De Deyn et al. AC, its value as an experimental or more (2003) showed how invertebrates enhanced importantly a management tool depends secondary succession of European grassland heavily on the experimental design utilized vegetation by suppressing dominant early and an evaluation of the interactions with seral species. Increasing an invaded com- the species being studied. An important munity’s succession towards later-seral factor limiting the use of AC in management vegetation may promote diversity and slowly settings is the need to incorporate it into diminish the dominance of invasive species soils to be eff ective. Regardless of the without physically disturbing the system. It problems and contingencies associated is possible to manage earthworm popu- with AC use, it is one of few tools available lations using electroshocking without non- to potentially minimize the impacts of target eff ects (Bohlen et al., 1995) and invasive plants in the soil or alter the indirectly aff ect microbial activity (Binet et microbial community (Kulmatiski, 2011) al., 1998) and potentially nutrient cycling. and additional research may yield practical Th e impacts of earthworms, native and applications for the compound. invasive, have been overlooked in many ecosystems and may provide novel manage- ment strategies. Naturally any application of Manipulation of soil microbial and these practices must have a well-defi ned goal invertebrate populations and understanding of the plant–soil– microbial interactions. Th e impacts of invasive species may be A novel concept promoted by Harris managed by progressive methods that (2009) utilizes fungal:bacterial ratios across directly or indirectly modify soil microbial dominant vegetation groups as a measure of 98 T.A. Grant III and M.W. Paschke

a system’s successional development and Conclusion suggests that the manipulation of microbial communities may be used to enhance the Invasive plant species impact all aspects of restoration of degraded systems towards a soil, including litter decomposition, nutrient specifi c late-seral plant community. Th e use cycling, soil fauna and fl ora, and the physical of microbes as a ‘restoration shortcut’ characteristics of soil. Th e complex inter- probably depends upon whether micro- actions between these organisms and organisms facilitate vegetation dynamics or entities require in-depth research to clarify are followers of the change (Harris, 2009). our understanding of plant–soil feedbacks, In all likelihood, this will be species and yet recent research is beginning to provide system dependent, but the concept of information that can be used to manage microbial manipulation is worthy of basic ecological processes for the control of additional research and management appli- invasive species. Historically, invasive plant cations especially if known pathogens or species were managed by a top-down, benefi cial symbioses can be appropriately command and control approach based on applied. Few studies have researched the removing the species or propagules, often impacts of microbial inoculation on invasive repeatedly. Although this approach can be plants or restoration projects and the eff ective in the short term, it creates outcomes have been mixed (Stacy et al., disturbance and frequently promotes 2005; Rowe et al., 2007, 2009; Abhilasha et re invasion. Additionally, chemical and al., 2008). Another approach may be to mechanical control methods are expensive, manage the species composition of plant energy intensive, and may have undesirable communities either through plant species eff ects on the environment and perpetuate additions or removals in order to exact disturbance cycles. desired changes in soil microbial com- As our understanding of dynamic munities (Boyetchko, 1996; Busby et al., in biological systems increases, we have the press). Most of these novel management knowledge and skills to manage invasive practices are based on the Natural Enemies species through the manipulation of basic theory (Elton, 1958), Enemy Release ecosystem function or processes. Th e Hypothesis (Mitchell and Power, 2003; modifi cation of nutrient availability due to Levine et al., 2006; Mitchell et al., 2006), microbial function or inputs from decom- or an understanding of how a species position can infl uence plant community accumulates pathogens (Eppinga et al., dynamics, although it will not be immediate 2006), and will require extensive knowledge and will rarely cause the complete removal of the system and proper experimentation of an unwanted species. Similarly, a change prior to use as a management tool. in microbial pathogens or levels of root An important premise of this type of herbivory will be unlikely to create a resource management is a relatively clear dramatic diff erence in vegetation, at least in understanding of the organisms and the short term. Soil amendments or interactions involved, otherwise unintended fertilization can alter microorganisms and outcomes will occur. Clear examples of non- plant dynamics leading to diff erent suc- target eff ects by management have occurred cessional trajectories and plant com- in biological control with insects and munities. Vegetation, soils, and microbial belowground manipulations must strive to systems are interconnected and their avoid these mistakes (Louda et al., 1997; linkages are beginning to be understood. Pemberton, 2000; Pemberton and Cordo, Potential manage ment methods are based 2001). It is critical to avoid introducing on tinkering in the incredibly complex potentially harmful or invasive micro- feedbacks between plants, soil, and microbes organisms into systems in which our in order to fi nd a balance or stability wherein comprehension is limited (van der Putten et higher species diversity is reached, while al., 2007). preventing a non-native species from Invasive Plant Impacts on Soil Properties 99

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Stuart P. Hardegree, Jaepil Cho, and Jeanne M. Schneider

US Department of Agriculture, Agricultural Research Service, USA

Introduction rangeland site in southern Idaho, USA that receives 295 mm of average annual Precipitation, solar radiation, wind speed, air precipitation. Weather is even more variable temperature, and humidity are principal at the level of seasonal distribution (Fig. 6.1; drivers controlling energy and water fl ux Table 6.1), which is the relevant scale for in plant communities. Climate is defi ned as critical phases of germination, emergence, the long-term average representation of and seedling establishment. these variables, and their seasonal pattern. Individual plants respond to their local Rangelands are generally characterized by an micro-environment in which weather inputs arid or semi-arid climate with plant com- are moderated by interactions with other munities dominated by grassland, shrub- plants, the soil surface, and lower soil layers steppe, and savanna vegetation. Gross (Campbell and Norman, 1998). Soil develop- climatic variability generally determines the ment is itself aff ected by weather and climate, suitability of both native and introduced but also by the resident vegetation, under- plant materials for rangeland restoration and lying parent material, topography, soil age, rehabilitation (Shown et al., 1969; Shifl et, and other factors, all of which are highly 1994; Barbour and Billings, 2000; Vogel variable over space (Jenny, 1980). Th e soil et al., 2005; USDA, 2006). Unfortunately, micro-environment is most variable in near- the micro-environmental requirements for surface layers that respond quickly to changes germination, emergence, and seedling in atmospheric variables (Flerchinger and establishment are much more restrictive than Hardegree, 2004). Deeper soil layers can store the longer term climatic requirements for precipitation and buff er establishing plants maintenance of mature plant communities from unfavorable weather conditions, but the (Call and Roundy, 1991; Peters, 2000; majority of mortality events among seeded Hardegree et al., 2003). species occur relatively near the soil surface Weather is also the combined expression at, or before, the relatively vulnerable stage of of climatic variables, but over a relatively seedling emergence (James et al., 2011). short time scale. Rangelands are relatively Existing plant cover shades the soil surface dry in a climatological sense, but exhibit and provides protection from wind-driven high spatial and temporal variability in evaporation, but competition for moisture weather (Rajagopalan and Lall, 1998). from invasive weeds can be a dominant Figure 6.1 demonstrates annual and limiting factor for water availability in the seasonal variability in precipitation for a seedbed. Seedbed preparation and planting

© CAB International 2012. Invasive Plant Ecology and Management: Linking Processes to Practice (eds T.A. Monaco and R.L. Sheley) 107 108 S.P. Hardegree et al.

Hydrologic year

Fig. 6.1. Annual (black bars) and March–May precipitation (white bars) for Boise, Idaho, USA. Mean annual precipitation and standard error of the mean (SE) for this location is 295 ± 11 mm but the standard deviation of the mean (SD) is 68 mm with a coeffi cient of variation (CV) of 24%. March–May precipitation is relatively more variable with a mean and SE of 93 ± 6 mm, an SD of 38 mm, and CV of 41%.

Table 6.1. Mean monthly precipitation (mm) for rangelands with similar annual precipitation. Numbers in parentheses represent the standard deviation of the mean. Windhoek Neuquen Ivanhoe Boise (USA) Urumqi (China) (Namibia) (Argentina) (Australia) January 35 (20) 8 (6) 64 (60) 13 (20) 33 (43) February 27 (17) 9 (7) 77 (61) 17 (25) 30 (42) March 31 (18) 19 (13) 70 (65) 26 (36) 29 (36) April 29 (18) 31 (20) 24 (22) 22 (29) 20 (28) May 33 (26) 33 (23) 4 (7) 19 (18) 29 (25) June 22 (18) 33 (23) 2 (7) 26 (26) 23 (19) July 7 (8) 28 (21) 1 (2) 18 (19) 24 (20) August 7 (12) 21 (21) 0 (1) 15 (17) 24 (20) September 14 (16) 24 (17) 2 (3) 19 (23) 23 (19) October 20 (15) 21 (14) 10 (18) 28 (25) 31 (31) November 34 (16) 18 (10) 18 (21) 19 (22) 24 (23) December 36 (22) 13 (8) 27 (27) 19 (29) 24 (27) Annual 295 (68) 258 (74) 299 (141) 241 (144) 314 (131)

methods are designed to optimize micro- ation seldom addresses the issue of weather environmental conditions for planted species, variability per se. Hardegree et al. (2011) to increase the number of favorable microsites surveyed the rangeland planting literature for germination and establishment, and to for the western USA and observed that less mitigate or control competition for water and than 60% of studies reported weather other resources from undesirable species conditions, and less than half were replicated (Roundy and Call, 1988; Call and Roundy, for year eff ects that would have included 1991; Sheley et al., 1996, 2006; Krueger- weather as a variable factor. In studies that Mangold et al., 2006). did report weather conditions, successful Published research on rangeland restor- establishment was almost always associated Soil Micro-environment for Rangeland Restoration 109

with what would be considered average or Weather, Climate, and Successional above average precipitation during the study Processes period. Th is implies that climatic thresholds exist, below which, management actions Site availability and optimization of have little eff ect on establishment success. seedbed micro-environment In practice, most rangeland restoration activities make limited use of weather and Disturbance is the principal ecological climate infor mation. Climate and gross soil process aff ecting site availability for diff erences are generally considered only in rangeland plant establishment (Sheley et al., the selection of appropriate plant materials 2010). Soil disturbance can favor invasive based on the potential distribution of over desirable species, but some type of mature-plant com munities. Seedbed prepar- mechanical disturbance is generally neces- ation and planting techniques are designed sary to create safe sites for establishment of to optimize seedbed micro-environment but desirable, late-successional species (James these methodologies are used prescriptively et al., 2010). Seedbed preparation and regardless of historical or potential weather planting methods are designed to reduce conditions. Weather information is mostly water loss and mitigate adverse thermal used retrospectively to explain seeding conditions in the seed zone. Th is is generally failure. accomplished through mechanical disturb- Rangeland restoration practices are ance, soil fi rming and surface modifi cation, generally viewed in the context of a suc- control of seeding depth, and less frequently, cessional model that recognizes the causes the application of soil surface amendments of succession and ecological processes that (Roundy and Call, 1988; Sheley et al., 1996). are amenable to management (Pickett et al., Soil surface modifi cation is often used to 1987; Westoby et al., 1989; Whisenant, increase water availability to the seed. 1999; Bestelmeyer et al., 2003; Roundy, Alternative seedbed preparation strategies 2005; Krueger-Mangold et al., 2006; Sheley can alter seedbed micro-environment by et al., 2006). Sheley et al. (2010) and James improving seed–soil contact, reducing et al. (2010) outline these processes, and the the amount of surface area subject to ecological principles that defi ne management evaporation, improving infi ltration and alternatives for invasive plant management water holding capacity, and creating and rangeland restoration. All of the microsites that capture and retain water ecological processes underlying succession (McGinnies, 1959; Roundy et al., 1992). are directly or indirectly aff ected by weather Animal trampling, land imprinting, pitting, and climate. Sheley et al. (2010) also describe furrowing, and rolling treatments have all a generalized planning cycle that includes been used to create microsites that capture initial site assessment, monitoring, adaptive or preserve moisture, and to improve management, and reassessment of manage- hydraulic conductivity by pressing surface- ment eff ects. Site assessment and adaptive applied seeds into the soil (Hyder and Sneva, management are two steps in the planning 1956; McGinnies, 1962; Haferkamp et al., cycle that can be signifi cantly improved by 1987; Winkel and Roundy, 1991; Roundy et incorporation of weather and climate al., 1992). Post-disturbance soil fi rming can information. In the following sections, we improve hydraulic conductivity to drilled use the ecologically based model described seeds by reducing soil surface area and soil by Sheley et al. (2010) to structure a macroporosity (Hyder and Sneva, 1956; discussion of weather and climate impacts McGinnies, 1962). Various studies have also on some key processes aff ecting succession, reported diff erential establishment success the micro-environmental underpinnings of relative to the position of soil surface some important restoration management features that aff ect local micro-environment tools, and strategies to incorporate weather (Bragg and Stephens, 1979; Eckert et information into the restoration-planning al., 1986; Roundy et al., 1992). Surface cycle. modifi cation, however, can also cause seed 110 S.P. Hardegree et al.

burial beyond establishment depth (Kincaid Monsen and Stevens, 2004; Lambert, 2005; and Williams, 1966; Slayback and Renney, Ogle et al., 2008a, b). Th e decision to 1972; Winkel et al., 1991a). broadcast or plant seeds, however, is often Mulch application can reduce water loss mandated by topographic complexity and and moderate soil surface temperatures but economic considerations. is generally considered too expensive for extensive rangeland use (Lavin et al., 1981; McGinnies, 1987; Ethridge et al., 1997). Seeding rate and optimization of species Hardegree et al. (2011) found that mulch availability application improved seedling establishment in only 62% of 21 rangeland seeding studies Rangeland restoration is often conducted in surveyed. Mulch application has a more areas that have lost their source of native consistent record, however, for erosion plant materials, and availability of desirable control and soil stabilization, which can species can only be addressed by seeding. have positive secondary eff ects on soil Seeding rate recommendations are linked to microsite availability (Bautista et al., 1996; microclimatic considerations as increased Brockway et al., 2002; Benik et al., 2003). seed numbers increase the probability of Depth of planting is a critical factor in seeds reaching safe microsites, irrespective successful plant establishment as microsite of active depth management (Harper et al., favorability is depth dependent (Young et 1965; Roundy et al., 1992; Chambers, 1995). al., 1990; Winkel and Roundy, 1991; Most individual studies reporting eff ects of Chambers and MacMahon, 1994; Ott et al., seeding rate on establishment success are 2003). Th e physical rationale for depth not replicated suffi ciently to evaluate recommendations is based on a tradeoff interactions with annual and seasonal between increased water availability and variability in weather conditions. Th e increased energy requirements for combined literature, however, supports the emergence as a function of depth (Roundy concept that higher seeding rates may and Call, 1988; Call and Roundy, 1991). enhance the likelihood of successful initial Evidence for depth eff ects is generally establishment (Vogel, 1987; Sheley et al., limited to relatively small but detailed 1999; Wiedemann and Cross, 2000; Williams studies conducted in the laboratory, and less et al., 2002; Eisworth and Shonkwiler, 2006; frequently, in the fi eld (Kinsinger, 1962; Hardegree et al., 2011). Broadcast seeding Vogel, 1963; Hull, 1964). Th ere are many rates are generally recommended at 2–3 rangeland seeding studies that compare times the rates for planted seeds to increase establishment success of broadcast and the likelihood that suffi cient seeds fi nd safe planted seeds (e.g. Nelson et al., 1970; Wood sites for establishment (e.g. Nelson et al., et al., 1982; Haferkamp et al., 1987; Ott et 1970; Wood et al., 1982; Haferkamp et al., al., 2003), but very few have actually 1987; Ott et al., 2003). characterized post-planting seed depth (Winkel and Roundy, 1991; Winkel et al., 1991a, b). Th ese studies generally support Species performance and competition for the hypothesis that very small seeds resources establish more frequently from near-surface seed placement, larger seeds require soil Climate is the primary criterion for selection cover for maximal performance, and seed of acceptable plant materials for rangeland performance drops dramatically below some restoration. Th is is generally acknowledged threshold of depth (Hull, 1948; Stewart, in most seeding guides in the form of tables 1950; Douglas et al., 1960). Seeding depth that list species and cultivar suitability as a recommendations can be fairly specifi c, but function of mean annual precipitation (e.g. are based on rules of thumb regarding Jensen et al., 2001; Lambert, 2005; Ogle et seeding depth as a function of seed size al., 2008b). Seeding guides may also cite (Plummer et al., 1968; Jensen et al., 2001; climatic thresholds below which active Soil Micro-environment for Rangeland Restoration 111

seeding practices are not recommended due establish ment (Plummer et al., 1968; to the low probability of success (Anderson Roundy and Call, 1988; Monsen and Stevens, et al., 1957; Jordan, 1981). Vegetation 2004). Th e following serve as examples from distribution as a function of climate is the rangelands in the western USA. Spring is basis of most recommendations for native generally the most favorable establishment plant materials (e.g. Barbour and Billings, period in Mediterranean-coastal and semi- 2000; USDA, 2006), but cultivar-specifi c arid interior rangelands that are subject to recommendations can also be based on signifi cant summer drought (Douglas et al., plant materials evaluation and development 1960; Nord et al., 1971; Harris and by local and regional government, and Dobrowolski, 1986). Th e summer monsoon academic organizations (Schwendiman, is a critical establishment period in many 1956; Harlan, 1960; Alderson and Sharp, arid desert environments (Jordan, 1981; 1994; Asay et al., 2003). Plant materials are Abbot and Roundy, 2003; Hereford et al., often selected or bred for superior 2006). Plant establishment generally occurs establishment, growth, and production in late-spring through early summer in within a targeted region or climatic regime relatively more moderate precipitation (Schwendiman, 1958; Johnson and Asay, zones in the Great Plains (Robertson and 1995; Asay et al., 2003). Current plant Box, 1969; Hart and Dean, 1986; Ries and materials development and evaluation Hofmann, 1996) and late-spring through programs are increasingly focused on para- early fall in some higher elevation mountain meters related to species performance and sites (Hull, 1966; Lavin et al., 1973). Post- establishment under alternative conditions planting microclimate must be favorable for of weather and climate (Aguirre and initial germination and emergence, but also Johnson, 1991; Johnson and Asay, 1995; needs to remain favorable during the Arredondo et al., 1998; Jensen et al., 2005). vulnerable period of seedling establishment Asay et al. (2001) have argued that the (Hyder et al., 1971; McGinnies, 1973; Frasier relatively harsh climatic conditions on et al., 1987; Abbot and Roundy, 2003; James rangelands may preclude the eff ective use of et al., 2011). many native plant materials, and that it may Dormant-fall seeding is a commonly be prudent to plant more easily established recommended practice in the Intermountain non-native species. Indeed, biodiversity and West, USA. Th is practice places seeds in the restoration objectives may require multiple- ground well in advance of the optimal growing year strategies for replacement of non-native season to take advantage of all opportunities species only after initial site stabilization and for germination, emergence, and growth suppression of annual weed competition during favorable periods in the winter and (Bakker et al., 2003; Cox and Anderson, spring (Plummer et al., 1968; Nelson et al., 2004). Biodiversity and restoration object- 1970; Hart and Dean, 1986; Monsen and ives may need to be addressed only in years Stevens, 2004). Dormant-fall seeding is also when climatic conditions are amenable recommended when wet spring weather (Holmgren and Scheff er, 2001; Hardegree et precludes the use of mechanical seeding al., 2003; Cox and Anderson, 2004; Hardegree equipment and to mitigate eff ects of and Van Vactor, 2004). unpredictable spring weather (McGinnies, Species performance can be optimized by 1973; Hart and Dean, 1986). Th e timing of selecting the most appropriate season for seeding may also be dependent on seasonal planting. Optimal planting season is patterns of weed establishment and timing determined by the historical pattern of requirements of essential weed control precipitation and temperature, and the measures (Robocker et al., 1965; Klomp and anticipated phenological development of Hull, 1972). Eiswerth and Shonkwiler (2006) planted species, existing desirable conducted meta-analysis of a large number of vegetation, and resident weeds. Th e general rangeland seedings and confi rmed the objective is to get seeds in the ground before relative benefi ts of fall/winter-dormant the most favorable season for plant seeding on interior rangeland locations in 112 S.P. Hardegree et al.

Nevada, USA. Very few experimental studies resource utilization under conditions of of seeding-season eff ects, however, are water stress, and when soil temperatures are replicated in more than 1 or 2 years (Hull, low in the fall, winter, and early spring 1948, 1974; Douglas et al., 1960; Robocker et (Harris and Wilson, 1970; Harris, 1977; al., 1965; Ries and Hofmann, 1996). Fall- Melgoza et al., 1990; Roundy et al., 2007; dormant planting was found to be superior to Hardegree et al., 2010). Chemical or spring planting in 73% of the individual mechanical weed control is frequently studies reviewed by Hardegree et al. (2011) necessary for successful establishment of for rangeland seeding in the Great Basin, desirable plant species in areas that are USA, although the majority of these studies aff ected by invasive weeds (e.g. Evans and were conducted in years of relatively favorable Young, 1978; Humphrey and Schupp, 2002; precipitation. Early emerging seedlings can Mangold et al., 2007). Hardegree et al. (2011) take advantage of favorable conditions, but observed that out of 52 rangeland seeding are also more vulnerable to periods of studies surveyed that included an evaluation drought, extreme cold, and other mortality of weed control treatments, all but two factors (James et al., 2011). concluded that weed control was either Seedbed preparation and planting necessary, or at least benefi cial to successful methods often include strategies to reduce establishment. competition for water by undesirable plants (Gonzalez and Dodd, 1979; Ott et al., 2003; Mangold et al., 2007). Cheatgrass (Bromus Assessment, Monitoring, and tectorum L.) is a dominant annual weed that Adaptive Management Tools has invaded millions of hectares of rangeland in the Intermountain West, USA (Knapp, State-and-transition probabilities and 1996). Water availability in the seedbed is assessment of weather variability greatly aff ected by competition from cheatgrass (Fig. 6.2), which is relatively more Th e need for rangeland restoration begins effi cient than native perennial grasses for with a perception that an existing vegetation both initial establishment and subsequent state is undesirable relative to some

0.25

0.20

0.15

0.10

0.05 Volumetric water content (0–60 cm) water Volumetric

0.00 1 Jan 1997 1 Jan 1998 1 Jan 1999 1 Jan 2000 1 Jan 2001

Fig. 6.2. Volume-averaged percent water content over the depth range of 0–60 cm on a loamy-sand (sandy, mixed mesic Xeric Haplargid) soil type in southwestern Idaho, USA. Soil water content was measured with time-domain-refl ectometry sensors under replicated (n=3) and interspersed bare soil (open circles) and cheatgrass (closed circles) cover plots. Soil Micro-environment for Rangeland Restoration 113

alternative state. Ecological site descriptions for the example of sagebrush-steppe utilize state-and-transition models to defi ne vegetation in the Great Basin region, USA. the range of vegetation states that currently Th is vegetation type historically maintained exist, that may develop after additional site a disturbance cycle with a period of 60–100 disturbance, or that may be achievable years that included wildfi re, and a post-fi re through management (see Brown and successional sequence that moved through a Bestelmeyer, Chapter 1, this volume). Current native bunchgrass phase, and increasing state-and-transition models acknow ledge shrub component (Wyoming big sagebrush; that the potential trajectories between Artemisia tridentata Nutt. ssp. wyomingensis undesirable and desirable states are limited Beetle and Young). Introduced annual by weather variability and that some grasses, such as cheatgrass, have disrupted transition pathways may require a specifi c this system over millions of hectares and perhaps infrequent series of climatic resulting in a 5–10 year cycle of wildfi re events before a successful change in state can followed by immediate and persistent occur (Westoby et al., 1989; Batabyal and annual weed cover. Cheatgrass can be Godfrey, 2002; Bestelmeyer et al., 2003; expected to establish adequately on most Briske et al., 2008). sites and most years on sagebrush-steppe Figure 6.3 shows a simplifi ed schematic rangeland (Roundy et al., 2007), therefore, of alternative states and transition pathways the fi re/cheatgrass cycle represented in Fig. 6.3 is relatively robust under almost any weather scenario. Sheley et al. (2011) discuss the initial evaluation of undesirable site conditions, NativeDiversity and the use of rangeland health assessment to identify the ecological processes in need 5 of repair (Pyke et al., 2002). We recommend 5 1 2 that an additional tool for initial assessment be the evaluation of historical weather PostFire 4 IntroducedPerennials variability. Th is variability may provide 5 valuable perspective on the probability that proposed management actions will have the 1 3 desired eff ect on a given ecological process 4 in any given year. A minimal evaluation of weather should AnnualWeeds include an assessment of mean annual and monthly patterns of precipitation and temperature. Information on climatological Fig. 6.3. State-and-transition model for means will be necessary for selection of sagebrush-bunchgrass rangeland in southern appropriate plant materials, and will defi ne Idaho, USA that has been disturbed by introduced the seasonality of conditions favorable for annual weeds. Transition 1 represents fi re events, plant establishment and growth. Climat- which can happen in any given year. Transition 2 ological means also will provide baseline represents natural recovery of native plant conditions for interpreting deviation from diversity in the absence of introduced annual average conditions in any given establishment weeds. Transition 3 represents type conversion if year. Annual variability can be assessed by introduced weed seeds are present. Transition 4 organizing and ranking historical data of the represents a transition to an introduced perennial grass community, which may be possible only in type displayed in Fig. 6.1 for the Boise moderate to favorable precipitation years. precipitation record. Ranking can be used to Transition 5 represents a transition to a diverse assess where a given year falls in the spectrum native community that may be possible only in of potential weather conditions and is favorable precipitation years with annual weed especially useful in interpreting the relative control (if necessary). success of previous management actions. 114 S.P. Hardegree et al.

Monitoring and adaptive management not seem to work in that particular year. Lessons learned from successful manage- Monitoring is a critical part of ecologically ment actions should also be weighed in the based restoration planning as most manage- context of relative favorability in weather in ment actions involve a relatively high degree that year. Multi-year evaluation should be of uncertainty (Sheley et al., 2010). Th is considered when comparing alternative uncertainty is exacerbated by weather management treatments, and multi-year variability, and a general lack of information treatments may be necessary to achieve on how weather impacts the relative success acceptable levels of establishment success at of alternative management strategies. Most a given site. individual studies in the range planting literature are insuffi ciently replicated to extract valid inferences about weather Long-term weather forecasting eff ects (Hardegree et al., 2011). Th e majority opportunities of range planting studies, until fairly recently, have not measured critical environ- Th e most useful potential technology for mental factors aff ecting success, such as soil enhancing establishment success lies in temperature and water relations, but only development and utilization of relatively report relative treatment eff ects (Call and long-range weather forecast technology Roundy, 1991; Vargas et al., 2001). Range specifi c to rangeland planting applications planting studies also tend to extrapolate (e.g. Barnston et al., 2000, 2005; Garbrecht results obtained from atypical sites and and Schneider, 2007; Lim et al., 2011). weather conditions over larger areas (Cox Similar technology is in relatively common and Martin, 1984), and are seldom replicated use for more traditional agricultural in multiple seeding years to account for applications (Doblas-Reyes et al., 2006; inter-annual weather variability (Casler, Baigorria et al., 2008; O’Lenic et al., 2008). 1999). It is logistically diffi cult to obtain Long-term weather forecasts in many fi eld data that replicates inter-annual rangeland areas are often merely synoptic variability. Unfortunately, previous studies descriptions of historical weather patterns do not include enough commonality in and not based on physical or empirical experimental design features to be subject prediction of future weather conditions. to any detailed meta-analysis of general Even low-resolution weather forecasts, weather eff ects (Durlak and Lipsay, 1991; however, would increase the probability of Michener, 1997; Osenberg et al., 1999; successful native plant establishment if Gurevitch et al., 2001). It may be possible to management decisions at the time of develop guidelines, however, for establishing seeding could be based on the anticipation some common experimental design features of favorable conditions for seed germination, for future studies that may be amenable to emergence, and seedling establishment more sophisticated meta-analysis. (Hardegree et al., 2003; Hardegree and Van Th e stochastic nature of weather Vactor, 2004). Weather forecasts could be variability may require adoption of new used to initiate contingency plans in areas paradigms for monitoring and evaluating that have been previously identifi ed for alternative management strategies. Spe- restoration, and for which pre-management cifi cally, both restoration failure and success logistics of equipment, personnel, and plant must be evaluated in the context of weather materials are in place (Westoby et al., 1989; conditions during the period of establish- Bakker et al., 2003). ment. Adaptive management alternatives Currently available long-term forecast should be viewed in the context of weather information is probabilistic in format. ranking during the establishment season Probability of Exceedance (PoE) distributions being evaluated. If the seasonal conditions defi ne the potential deviation of future were signifi cantly below average, it may not weather conditions (typically over seasons) be necessary to abandon strategies that did from the long-term mean (Barnston et al., Soil Micro-environment for Rangeland Restoration 115

2000; Schneider and Garbrecht, 2006; repository for US weather information, but Garbrecht and Schneider, 2007; Lim et al., also has links to global weather resources, 2011). In simpler terms, the forecasts including the World Meteoro logical predict the odds for whether the weather Organization that maintains a list of will shift toward wetter or drier, warmer or meteorological data sources by country cooler, when compared to the reference (www.wmo.int/pages/members/members_ period of record. Th ese predictions apply to en.html). NOAA has also been developing relatively large spatial domains and are Regional Climate Centers that consolidate currently subject to relatively high predictive state weather and climate data (www. errors for many parts of the globe. Long- ncdc.noaa.gov/oa/climate/regionalclimate term forecast predictions work better in centers.html), and can frequently provide some areas than others and are generally advice on data access and quality assurance more accurate for temperature than for questions. precipitation (Schneider and Garbrecht, Many fi eld sites will not have suffi cient 2006; Livezey and Timoveyeva, 2008; Lim et local information for detailed character- al., 2011). In the USA, predictive accuracy is ization of site variability. Some areas will relatively higher where extreme weather have additional regional resources for events and sustained deviations from interpolating weather records in areas average are associated with the El Niño that do not have local monitoring. Inter- Southern Oscillation (ENSO) phenomenon. polated data will not be as accurate as actual Long-term weather predictions in the USA observations but may be suffi cient for are relatively more accurate in the south- many practical applications. Tools and east, southern Texas, desert southwest, databases for interpolation of historical California, Pacifi c Northwest, and northern weather data are available in much of the Rocky Mountains, but are relatively less USA through sites such as the DAYMET US accurate in the Intermountain Great Basin. Data Center, which is provided by the Other areas of the world, particularly Montana State University, Numerical Ter- tropical regions in or adjacent to the Pacifi c radynamic Simulation Group (www.daymet. Ocean, enjoy a stronger and more reliable org/default.jsp) and the USDA Natural predictive signal. Th e Australian Govern- Resources Conservation Service National ment Bureau of Meteorology has developed Water and Climate Center (www.wcc.nrcs. a successful forecast system for ENSO- usda.gov/climate/prism.html). related impacts on agriculture and water Long-term (covering weeks to multiple resources in eastern and southern Australia months) weather forecast information is (Lim et al., 2011). increasingly available in many countries. Th ese forecasts are created through ensemble analysis of multiple global climate Weather and climate data resources models, and analysis of statistical relation- ships between weather and oceanic state It can be challenging to locate accurate, (e.g. ENSO) in diff erent parts of the globe continuous, long term, site-specifi c weather (Barnston et al., 2000, 2005; Lim et al., data. Much of the data available over the 2011). Many nations and nation groups internet has not been quality assured, or have developed, or are developing, regional may be a derivative product representing predictive capabilities such as the European data averaged or interpolated over large Centre for Medium-Range Weather Fore- areas. casts (www.ecmwf.int). Long-term forecasts Th e global database for historical weather for precipitation and temperature for data can be accessed through the US Australia are available through the Department of Commerce, National Oceanic Australian Bureau of Meteorology (www. and Atmospheric Administration (NOAA), bom.gov.au/climate/ahead), and for much of National Climate Data Center website (www. the globe through the International ncdc.noaa.gov/oa/ncdc.html). Th is site is the Research Institute for Climate and Society 116 S.P. Hardegree et al.

(http://portal.iri.columbia.edu/portal/ References server.pt/). Th e principal repository of long- term forecast data and information in the Abbott, L.B. and Roundy, B.A. (2003) Available USA is the NOAA National Weather Service’s water infl uences fi eld germination and Climate Prediction Center (www.cpc.ncep. recruitment of seeded grasses. Journal of noaa.gov). Range Management 56, 56–64. Aguirre, L. and Johnson, D.A. (1991) Root morphological development in relation to shoot growth in seedlings of four range grasses. Incorporating weather variability and Journal of Range Management 44, 341–346. long-term forecasts in rangeland Alderson, J. and Sharp, W.C. (1994) Grass Varieties restoration efforts in the United States. Agricultural Handbook No 170. US Department of Agriculture, Washington, As a fi rst step, assessment of historical DC. weather variability should be made for the Anderson, D., Hamilton, L.P., Reynolds, H.G. and rangeland site of interest. Monthly total Humphrey, R.R. (1957) Reseeding Desert precipitation and average daily air tem- Grassland Ranges in Southern Arizona. Arizona perature are relatively easy to obtain and are Agricultural Experiment Station Bulletin 249, Tucson, Arizona. suffi cient to provide a baseline for assess- Arredondo, J.T., Jones, T.A. and Johnson, D.A. ing rangeland restoration options, and (1998) Seedling growth of Intermountain for considering the possible use of long- perennial and weedy annual grasses. Journal of term forecasts. Th e World Meteorological Range Management 51, 584–589. Organization produces a product repre- Asay, K.H., Horton, W.H., Jensen, K.B. and Palazzo, senting the most recent 30-year period, A.J. (2001) Merits of native and introduced revised every decade (currently 1981–2010). Triticeae grasses on semiarid rangelands. It is highly desirable to fi nd the most local Canadian Journal of Plant Science 81, 45–52. monthly data possible, or barring that, a Asay, K.H., Chatterton, N.J., Jensen, K.B., Jones, good interpolation for the area of interest, T.A., Waldron, B.L. and Horton, W.H. (2003) Breeding improved grasses for semiarid for these 30 years. Once obtained, monthly, rangelands. Arid Land Research and seasonal, and annual data can be totaled Management 17, 469–478. (precipitation) and averaged (air tem- Baigorria, G.A., Jones, J.W. and O’Brien, J.J. perature) to establish baseline data for (2008) Potential predictability of crop yield using ranking of current and historical time an ensemble climate forecast by a regional periods relative to average conditions for a circulation model. Agricultural and Forest given time period. Meteorology 148, 1353–1361. Forecasting applications are generally Bakker, J.D., Wilson, S.D., Christian, J.M., Li, X.D., applied to 3-month intervals rather than Ambrose, L.G. and Waddington, J. (2003) individual monthly predictions (e.g. January– Contingency of grassland restoration on year, site, and competition from introduced grasses. February–March, February–March–April, Ecological Applications 13, 137–153. etc.). It is critical to understand that forecasts Barbour, M.G. and Billings, W.D. (2000) North are probabilistic predictions and subject to American Terrestrial Vegetation. Cambridge relatively large errors. Th e authors recom- University Press, New York. mend considering only those precipitation or Barnston, A.G., He, Y. and Unger, D.A. (2000) A air temperature forecasts predicting a shift in forecast product that maximizes utility for state- odds of at least 8% (Schneider and Garbrecht, of-the-art seasonal climate prediction. Bulletin 2006), with more serious consideration for of the American Meteorological Society 81, larger forecast shifts, especially during 1271–1279. moderate to strong ENSO conditions. Such Barnston, A.G., Kumar, A., Goddard, L. and Hoerling, M.P. (2005) Improving seasonal strong forecasts are rarely off ered for prediction practices through attribution of precipitation but may represent a reasonable climate variability. Bulletin of the American guide for the effi cient expenditure of range- Meteorological Society 86, 59–72. land restoration resources when they do Batabyal, A.A. and Godfrey, E.B. (2002) Rangeland occur. management under uncertainty: a conceptual Soil Micro-environment for Rangeland Restoration 117

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Valerie T. Eviner1 and Christine V. Hawkes2

1 Department of Plant Sciences, University of California, USA 2 Section of Integrative Biology, The University of Texas, USA

Introduction al., 2008; Harris, 2009; Eviner et al., 2010). In this chapter, we explore the mechanisms Th ere are countless examples of management driving plant–soil feedbacks in invaded projects that have attempted to decrease or systems, and potential management tools to eradicate invasive species at a site, only to alter these feedbacks to be more benefi cial have them rapidly recolonize within a few to natives over invasive species. years. While this is often attributed to reinvasion through propagules remaining at the site, or high propagule pressure from the The role of plant–soil feedbacks in surrounding landscape (Leung et al., 2004; shaping plant communities and plant Lockwood et al., 2005), this also may be due invasions to invasive species changing site conditions to favor conspecifi cs over native species. Plant–soil feedbacks occur when a shift in Many studies have documented that invasive plant community composition changes soil plants can impact numerous soil properties conditions, and these altered soil conditions and processes (Leffl er and Ryel, Chapter 4, further alter the plant community. Positive this volume; Ehrenfeld, 2010), and that feedbacks occur when a given plant species invader impacts on soil can infl uence com- alters the soil in a way that promotes its own petitive dynamics between plant species, persistence and growth (either directly by often favoring the invaders (Callaway and enhancing its own growth and that of Aschehoug, 2000; Reinhardt and Callaway, conspecifi cs, or indirectly via greater 2006; Batten et al., 2008; Kulmatiski et al., inhibition of the growth of other species 2008; reviewed in Eviner et al., 2010). Some compared to conspecifi cs). Conversely, a of the eff ects of invasive species on soils can plant species can alter a soil to its own persist after the invader has been removed, detriment, or in a way that promotes other making the system more susceptible to species more than itself, resulting in a reinvasion (reviewed in Eviner and Hawkes, negative feedback. Recent research has 2008; Kulmatiski and Beard, 2011). In these shown that plant–soil feedbacks can play an cases, restoration eff orts must be focused important role in shaping succession, not only on removing invasive species, but species coexistence, species dominance, also counteracting their eff ects on soil range expansion, and the success of invasive characteristics and processes (Heneghan et species (reviewed in Bardgett et al., 2005;

© CAB International 2012. Invasive Plant Ecology and Management: 122 Linking Processes to Practice (eds T.A. Monaco and R.L. Sheley) The Effects of Plant–Soil Feedbacks on Invasive Plants 123

Kardol et al., 2006; Manning et al., 2008; van Invader plant–soil feedbacks enhance der Putten et al., 2009). resilience of invaded state Plant–soil feedbacks are of particular relevance in understanding and managing Invasive species that generate positive species invasions, because positive feedbacks feedbacks are of particular concern for are more common in invaded communities, conservation and restoration, because they while negative feedbacks are more prevalent often create a barrier to the reintroduction in native communities (Klironomos, 2002; of native species. Regardless of what factors Kulmatiski and Kardol, 2008; Kulmatiski et precipitated the initial success of an invader, al., 2008; van der Putten et al., 2009). Of established invasive species can alter the soil particular concern are cases of ‘invasional and create a ‘novel ecosystem,’ an alternative meltdown’ (Simberloff and Von Holle, 1999), stable state that is diffi cult, if not impossible when one invasive species changes the soil to revert back to the native state (Suding et to enhance not only itself, but also the al., 2004; Seastedt et al., 2008; Farrer and invasion of other non-native species. For Goldberg, 2009; Hobbs et al., 2009; example the invasion of Bromus tectorum Hardegree et al., Chapter 6, this volume). enhances invasion of Taeniatherum caput- Th e degree of persistence versus medusae, and invasion of Taeniatherum reversibility of invader impacts on soils and increases invasion of exotic forbs (reviewed associated ecosystem processes is a critical in Eviner et al., 2010). Similarly, the invasive component of restoration potential. Some Bromus inermis alters the soil microbial of the changes caused by invasive species community to enhance the growth of the may be rapidly reversible upon removal of invader Euphoria esula (Jordan et al., 2008). the invader and do not require additional While, on average, invasive species are management. For example, decreased soil more likely than native species to create water availability caused by high plant positive (or less negative) feedbacks, there transpiration rates should reverse quickly are many exceptions to this general trend. once the invasive plant species is removed. Many plant–soil feedbacks are highly In contrast, alterations to soil properties species-specifi c, so that a given invasive such as soil structure, water infi ltration, species may negatively impact a subset of water holding capacity, carbon storage, and native species, but not all of them, and nitrogen cycling rates may persist for diff erent invaders are likely to impact months to decades, even with active diff erent native species (Casper and Castelli, management (van der Putten et al., 2009). 2007; Manning et al., 2008). In contrast to In these cases, reinvasion is likely to take the general trends, some invasive species place before soil conditions can be restored, create soil conditions that generate negative particularly if the altered state favors the feedbacks to conspecifi cs, while some native invasive plant species relative to native species create positive soil feedbacks to species. For example, extensive erosion as a conspecifi cs and negative feedbacks to result of invasion of Centaurea maculosa invaders (Kulmatiski et al., 2004; van der (Lacey et al., 1989) can take decades to Putten et al., 2007). For example, in a shrub- centuries to reverse via soil formation steppe ecosystem in Washington State, USA, processes and the gradual buildup of organic the native perennial grass, Pseudoroegneria matter by the restored plant community. spicata, alters soil in a way that decreases its Such cases highlight the importance of own growth, but has even stronger negative disrupting invader–soil feedbacks early in eff ects on the invasive species Centaurea the invasion process. diff usa, reducing invader cover from 18% to While invader-induced feedbacks may 5% (Kulmatiski et al., 2004). Promoting the create a stable invaded state in an invader’s specifi c native species that decrease the new range, these feedbacks do not always abundance of invasive plants can be a operate in the invader’s home range. For promising fi rst step in restoration of native example, the negative eff ects of C. diff usa on plant communities. its neighbors are much stronger in its 124 V.T. Eviner and C.V. Hawkes

invaded range than its home range (Callaway these mechanisms can be important in any and Aschehoug, 2000). In their home ranges, given invasion, and little is known about the invasive species are usually subject to their relative importance or the extent to the same negative plant–soil feedbacks which they strengthen or counteract one common to native plants in general another to create overall positive versus (Reinhart et al., 2003; reviewed in Reinhart negative feedbacks. While some mechanisms and Callaway, 2006). Th e existence of have similar management approaches for controls over invaders in their home ranges counteracting their associated feedbacks (e.g. through soil feedbacks, natural enemies, (Table 7.1), eff ective management will competitive eff ects, or of the evolution of require knowledge of how feedbacks are neighbor resistance to allelochemicals), generated by a given invader. Identifying the suggests that there may be long-term mechanisms driving feedbacks for an potential to control invaders in their new invasive species is often not straightforward, ranges through approaches such as bio- and even when they can be identifi ed, control agents or selection for native plant management of these feedbacks is still species that are resistant to the invader largely in the experimental stage. Th is eff ects. Both of these will likely happen over chapter highlights promising approaches to the long term, even without active manage- managing plant–soil feedbacks, but we ment. With increasing time since invasion, recognize that continued research on invaders tend to lose their initial advantage management strategies is required, both due to escape from negative interactions in across invasive species and sites, to improve the new range (Hawkes, 2007), and the these management tools and our ability to invaders’ impacts on the native community predict which approaches will be most decrease (Strayer et al., 2006; reviewed in eff ective for a given invader. Diez et al., 2010). Alternatively, the invasive species may evolve to have increased competitive ability, which can strengthen Litter both its negative impacts on native species and positive feedbacks to conspecifi cs. For Plant litter dynamics are an important example, an invader that benefi ts from its driver of plant community structure and own litter buildup may evolve to have more ecosystem processes (reviewed in Ehrenfeld recalcitrant litter, strengthening the positive et al., 2005). In general, increased litter feedback (Eppinga et al., 2011). Because few accumulation often decreases plant diversity studies have documented long-term impacts in herbaceous communities (Grime, 1979; of invasive species on communities and Foster and Gross, 1998). Litter alters surface ecosystems (reviewed in Strayer et al., 2006), and soil microclimate, directly inhibiting the we are still unable to predict whether long- establishment of select species (Facelli and term presence of a specifi c invader will Pickett, 1991) or enhancing key plant control versus enhance invasion through herbivores or pathogens (Lenz et al., 2003; changes in the strength and direction of reviewed in Flory and Clay, 2010). Th ese feedbacks. physical eff ects of litter are often initially more important than associated nutrient feedbacks, which can take longer to develop Mechanisms of Feedbacks, and (Amatangelo et al., 2008). While invasive Potential Management Tools plant species can aggressively compete for resource uptake, in some cases, the litter, Plant–soil feedbacks can be mediated rather than the live plant, is directly through many mechanisms, including plant- responsible for the invader’s impacts on the induced changes to soil structure, chemistry, plant community and soil conditions (Farrer and biota (Ehrenfeld et al., 2005; reviewed in and Goldberg, 2009; Holdredge and Casper et al., 2008), as well as the litter layer Bertness, 2011). Th ere are many examples (Farrer and Goldberg, 2009). A number of where litter accumulation drives both The Effects of Plant–Soil Feedbacks on Invasive Plants 125

Table 7.1. Mechanisms that drive plant–soil feedbacks and potential management tools.

Mechanism driving feedbacks Potential management tools Potential limitations Litter Litter removal through mowing, Timing is critical – may enhance or control invasion burning, grazing/trampling Not always possible at all sites or across broad scales Allelochemicals Activated charcoal Also can impact nutrient availability, or inhibit the activity of other compounds

Removal of invaders + time Reinvasion can reintroduce the allelochemicals quickly

Plant non-sensitive ‘transition’ Limited information available on which species may not species be sensitive, and which can transition to the ultimate desired community Soil microbial Activated charcoal Also can impact nutrient availability, or inhibit the activity community of microbes

Plant ‘transition’ species which can Interactions of plants and microbes are highly species- tolerate invader soil or promote specifi c, so there is limited information on which pathogens of invader species are impacted by, or resistant to given changes in the microbial community, and limited information on which species can transition to the ultimate desired community

Soil inoculation Local sources that specifi cally enhance native growth are usually not readily available, not always effective in establishing desired microbes Nitrogen Removal of N – burning, mowing, Disturbance may also enhance invaders and decrease grazing natives in the short term

Carbon additions to sequester Mixed effectiveness in sequestering soil N; have been soil N shown to enhance, as well as inhibit invasion by some species

Topsoil removal Can disrupt native microbial community and seed bank

Plant native species that take up Often invaders are more aggressive than natives in high quantities of N taking up N, but has been effective in conjunction with burning and carbon additions

Salinity Promote leaching of salts out of Can take a long time to reverse salinization, irrigation the upper layers of soil; method water may also add salts depends on soil, but can include increased water additions or promotion of soil drainage through increasing soil pores (through roots or soil organisms)

Plant salt-tolerant natives Will not necessarily reverse salinity, or aid in reestablishment of natives that were on the site before it became saline

Remove topsoil Also removes nutrients, soil microbes, and seeds 126 V.T. Eviner and C.V. Hawkes

invasive plant species’ impacts and feed- strong impacts on which species benefi t, backs, including: Typha × glauca invasion because timing of these disturbances can into wetlands with associated increases in also greatly impact seed production (Pollak nitrogen availability and decreases in light and Kan, 1996; DiTomaso et al., 2006; and native species diversity and abundance Holdredge and Bertness, 2011). (Farrer and Goldberg, 2009); Taeniatherum caput-medusae invasion into western US rangelands where its recalcitrant litter Allelochemicals inhibits the germination of other species, leading to monotypic stands (Young et al., A number of studies have suggested that 1971); and Microstegium vimineum invasion some invasive species decrease the per- of northeastern US forests where the formance of native plant species through physical litter barrier inhibits native tree the release of allelochemicals: organic seedling establishment and reduces seedling compounds that are either directly phyto- survival through enhanced vole activity toxic, or inhibit the activity of microbes that (Flory and Clay, 2010). Litter buildup can are symbiotic with plants (Wardle et al., also promote fi res, further leading to eco- 1998; Ridenour and Callaway, 2001; system alterations that may benefi t invasive reviewed in Bais et al., 2006). For example, over native species (reviewed in Davies and Alliaria petiolata can decrease the growth of Svejcar, 2008). Plant litter inputs are also native plant species by releasing compounds one of the main mechanisms driving species’ that decrease arbuscular mycorrhizal fungi impacts on soil chemistry, structure, and (Stinson et al., 2006). Similarly, Carduus biology (reviewed in Eviner and Chapin, nutans releases compounds that inhibit 2003a). nodulation and nitrogen fi xation in legumes, Litter accumulation does not always which is likely the cause for this invasive benefi t invasive plants. In some cases, the plant species decreasing the growth of a accumulation of litter from invasive plants neighboring legume species (Wardle et al., may also benefi t native species. In California, 1993, 1994). Other invaders that negatively USA, coastal sage scrub, accumulation of impact native communities by releasing invasive grass litter benefi ts a suite of allelochemicals include: C. maculosa and C. invasive grasses, but also enhances growth diff usa (Callaway and Aschehoug, 2000; of native shrubs by enhancing soil moisture Callaway and Vivanco, 2007; Th orpe et al., availability (Wolkovich et al., 2009). In other 2009), Fallopia × bohemia (Murrell et al., cases, native species may negatively aff ect 2011), and Acroptilon repens (Stermitz et al., invasive plant species through native litter 2003). accumulation. Th e invasive M. vimineum, for example, which benefi ts from its own litter, Management has lower seedling survivorship in patches where the litter of native species builds up Th e most direct way to manage allelo- (Schramm and Ehrenfeld, 2010). Litter can chemicals is to add compounds that can play a key role in shaping the community, sequester these allelochemicals, thus but the relative feedbacks to invasive and inhibiting their impact on soil microbes and native species may need to be considered in native plants. Activated carbon, also known litter management strategies. as activated charcoal, is highly absorptive due to its high density of micropores and sequestration of compounds through ionic Management bonding or adsorption (reviewed in Grazing, mowing, and burning are eff ective Kulmatiski, 2011), which has resulted in its for litter removal and often increase native common use for chemical purifi cation and species in invaded stands (Sheley et al., pollutant removal from water and air. 2007; reviewed in Holdredge and Bertness, Additions of activated carbon have been 2011). Th e seasonality of litter removal has eff ective in decreasing the negative impact of The Effects of Plant–Soil Feedbacks on Invasive Plants 127

invasive species on native species in a removing all invasive plant material from a number of systems (Callaway and Aschehoug, site. 2000; Ridenour and Callaway, 2001; Because the eff ects of allelochemicals Kulmatiski and Beard, 2006; Callaway and are species-specifi c, another potential Vivanco, 2007; Lau et al., 2008; Th orpe et al., restor ation approach is to plant native 2009; Kulmatiski, 2011). For example, the species that are not susceptible to these native grass Festuca idahoensis, when grown compounds (Perry et al., 2005; Alford et al., with the invasive C. maculosa, grew 85% 2009). Plant species are being tested for larger with activated carbon than without innate resistance to the allelochemicals of (Ridenour and Callaway, 2001). It is the invasive C. maculosa. Th e establishment important to note that activated carbon of these resistant species can prevent additions on their own are often not Centaurea from reinvading and may suffi cient to decrease the abundance of eventually facilitate the establishment of invasive plants – clearing of invasive plants native species that are susceptible to these along with native seed planting is frequently allelochemicals (Callaway and Aschehoug, required. In ex-arable fi elds in Washington 2000; Callaway and Vivanco, 2007; Th orpe State, USA, that were dominated by invasive et al., 2009). plants for decades, the combination of Th e allelopathic eff ects of invasive species clearing of invasive vegetation, a single on native species may decrease with time, as application of activated carbon, and native native species adapt to these inputs (Callaway seed additions shifted dominance from et al., 2005). Allelochemicals can have invasive to native plants, and this was stronger impacts on heterospecifi c neighbors maintained even after 6 years (Kulmatiski, in their invaded ranges, compared to their 2011). home ranges (Bais et al., 2003; Callaway et Allelochemicals can be highly species- al., 2008; Th orpe et al., 2009), suggesting specifi c in their impacts, which likely that there has been ongoing selection for accounts for the fact that additions of resistance in the home range. Over time in activated carbon vary in their eff ectiveness the new range, the inhibitory eff ects of in controlling invasive species, and may allelochemicals may decrease as native promote some, but not all native species species similarly evolve resistance to invasive (Lau et al., 2008; reviewed in Kulmatiski, species (Callaway et al., 2005; reviewed in 2011). Activated carbon additions also can Strayer et al., 2006). Breeding of resistant increase the prevalence of some invasive native plant genotypes may be a potential species (Kulmatiski and Beard, 2006; Lau et management approach. With time since al., 2008). Beyond the species-specifi c invasion, the impacts of allelochemicals on nature of activated carbon impacts, its use is the soil microbial community can also vary. far from straightforward because it not only Comparisons of sites that had been invaded sequesters allelochemicals, but also alters by A. petiolata for 20–50 years, demonstrated nutrient availability, rates of nutrient that resistance of the microbial community cycling, and the soil microbial community to allelochemicals increa sed over time. In the (Lau et al., 2008; Kulmatiski, 2011). longest invaded sites, Alliaria populations Allelochemicals generally are short-lived decreased allelo chemical inputs, further in the soil (hours to days) (Blair et al., 2005; decreasing overall impacts of the invasion on Reigosa et al., 2006), suggesting that the microbial community (Lankau, 2011). activated carbon may be most useful to Th ese cases suggest that invasions that are minimize the eff ects of invaders currently at facilitated by allelochemical inputs may be a site, or early in restoration, when it can controlled over the course of 4 to 5 decades sequester allelochemicals from newly through the strong selection imposed by invading individuals. To ameliorate potential allelochemicals on the native plant and longer term legacies of allelochemicals microbial communities. However, this deposited through plant litter (Reigosa et al., selection may be at the cost of decreased 2006), best practices should include diversity (e.g. Lankau, 2011). 128 V.T. Eviner and C.V. Hawkes

Soil microbial community Callaway et al., 2008; Wolfe et al., 2008). As with the general soil microbial community, Th e soil microbial community frequently the strength and direction of feedbacks from mediates soil feedbacks associated with mycorrhizal fungi are context-dependent, invasive plant species, but their specifi c based on factors such as the identities of the eff ects can be diffi cult to predict. For invader and fungi, the ecosystem, and the example, the soil microbial community is identities and life stages of neighboring altered by invasion of Aegilops triuncialis into native plant species (reviewed in van der an herbaceous serpentine community, Heijden and Horton, 2009). For example, in leading to decreased growth and fl owering California grasslands, USA, the invasive forb time of one native forb, Lasthenia californica, Carduus pycnocephalus grows best in soils but not other native species (Batten et al., without arbuscular mycorrhizal (AM) fungi 2008). Similarly, in the Great Plains, the soil and its growth decreases AM fungal dens- microbial community is altered by the ities in soil, resulting in reduced colonization invasion of Agropyron cristatum, B. inermis, of native roots and decreased growth of the and Eu. esula; each invader benefi ts from the native forb Gnaphalium californicum (Vogel- changes it induces, but only a subset of sang et al., 2004; Vogelsang and Bever, native species are aff ected by the altered soil 2009). Other invaders, such as C. maculosa, community of each invasive plant species appear to tap into existing native mycorrhizal (Jordan et al., 2008). Th e lack of apparent networks, essentially parasitizing resources, generality, and thus unpredictability, of which results in substantial growth benefi ts invasive species eff ects on soil microbial (Marler et al., 1999; Callway et al., 2003; communities may be partly due to our poor Carey et al., 2004). Invasive plants can also understanding of the specifi c microbial alter the composition of AM fungi taxa and mechanisms responsible for the mycorrhizae that colonize native plant observed feedbacks. roots. Invasion of annual grasses in Our understanding of the mechanisms California, USA (Hawkes et al., 2006; underlying soil microbial community feed- Hausmann and Hawkes, 2009), as well as C. backs is most developed for pathogens and maculosa in Montana, USA (Mummey et al., symbionts. Some invasive plant species are 2005), can shift the AM fungal community successful because they have escaped soil infecting native plant roots to substantially pathogens common in their native range, overlap with that of the exotic plants. While and this release from pathogens makes them the mechanisms driving invasive plant more competitive against native species, eff ects on mycorrhizal communities are which commonly experience negative feed- often unknown, in some cases, invasive backs with the soil pathogen community plants that are less reliant on mycorrhizal (reviewed in Reinhardt and Callaway, 2006). fungi may release inhibitory compounds However, invasive plants can exacerbate this that can broadly reduce the abundance of feedback, because their leachates enhance mycorrhizal fungi in soil (Stinson et al., the pathogens of native species (Mangla et 2006; Callaway et al., 2008; Wolfe et al., al., 2008). In other cases, invasive plants 2008). In other cases, invasive plants can may be successful because the benefi t associate with a subset of the mycorrhizal obtained from local mycorrhizal mutualists community, such as fungal generalists is greater than the negative eff ects of (Moora et al., 2011) or those fungi most pathogens in the new range (Klironomos, benefi cial to the invader (Zhang et al., 2010), 2002). which may promote the selected fungal taxa Invasive plant species are usually over others. In these cases, the network of colonized by local mycorrhizal fungi and can mycorrhizal fungi supported by invasive have direct eff ects on the composition and plants may create a priority eff ect (reviewed abundance of the mycorrhizal community in Hausmann and Hawkes, 2010). that can feed back to the plant community Co-invasion by plants and their mycor- (Hawkes et al., 2006; Stinson et al., 2006; rhizal fungi may also facilitate plant invasion The Effects of Plant–Soil Feedbacks on Invasive Plants 129

success through positive feedbacks, such as the microbial community. Litter removal, or with Pinus species and ectomycorrhizal inputs of activated carbon to deactivate key fungi in New Zealand (Dickie et al., 2010). plant metabolites have been eff ective in Where ectomycorrhizal associates are managing invasive species, but may also spatially limited, the spread of exotic Pinus promote some invaders (reviewed in species can also be limited (Nuñez et al., Kulmatiski, 2011). As described above, 2009). More than 200 species of ecto- activated carbon can inhibit the impacts of mycorrhizal fungi have been introduced to allelochemicals on the microbial community. new ranges worldwide; these fungi are For example, in a case where a tropical largely associated with plantation forestry invasive plant increases generalist soil (Vellinga et al., 2009) and thus the spread of pathogens, addition of activated carbon the ectomycorrhizal fungi and their plant decreases pathogen spore numbers and hosts may be linked in many cases. increases native plant growth (Mangla et al., 2008). Few studies have assessed the impacts of Management disturbance regimes on invader–soil feed- As described above, when invader-induced backs. In Portuguese coastal dunes, fi re feedbacks are strong enough to prevent the decreased AM fungal colonization in all original native species from persisting long species, and rhizobial colonization in native, enough to alter soil conditions, a multi-stage but not invasive legumes. Overall, fi re successional approach can be employed by enhanced invasive species performance by initially planting species that are more changing invader–soil biota feedbacks from tolerant of the invaded soil conditions. Th is neutral to positive, and native species is feasible because most plant–microbial feedbacks from negative to neutral (Carvalho interactions are species-specifi c. Once the et al., 2010). Th e impacts of disturbance initial plantings ameliorate the invaded soil regimes on plant–soil feedbacks may be legacies, the target native community can be important to consider, because it may result reestablished as seeds or transplants (Jordan in disturbance events that were meant to et al., 2008). Th is plant-induced change to control invaders having the unintended the microbial community may take time. consequence of strengthening of invader– While changes in plant species can impact soil feedbacks. While this example demon- some components of the microbial com- strates that disturbance further strengthens munity within weeks to months, the invader feedbacks, disturbance may be microbes mediating plant–soil feedbacks eff ective in disrupting invader–soil feed- can persist unchanged for at least a growing backs in other cases. season (Kulmatiski and Beard, 2011). A number of studies have investigated A more aggressive approach would be to the potential to inoculate invaded soils with plant native species that culture soil desirable microbial communities (reviewed pathogens that decrease the growth of in Vessey, 2003; Schwartz et al., 2006). On invasive species (Knevel et al., 2004). Finding highly degraded soils with a depauperate soil and promoting such native species could be community, inoculation with microbes, a key tool for disrupting invasive species’ particularly mycorrhizal fungi, can enhance positive feedbacks within the soil plant establishment unless soil conditions community. Even without intervention, the are too stressful (Kardol et al., 2009). Where strength of negative feedbacks on invasive sites dominated by invasive plants have an species increases with time since intact soil community, inoculation can be establishment, suggesting that over the long more complicated because the inoculated term, the soil microbial community may microbial community may be inhibited by decrease the dominance of invasive species the microbes already present (Kardol et al., (Diez et al., 2010). 2009; Mummey et al., 2009). However, in Another management option is to other cases, inoculation into intact com- interfere with the plant inputs that shape munities can be eff ective. In the tallgrass 130 V.T. Eviner and C.V. Hawkes

prairie of the Central Plains, USA, for invasive species decrease N availability, such example, inoculation with AM fungi as the invasion of Ae. triuncialis into increased the cover of native grasses over grasslands of California, USA (Drenovsky weedy plants (Smith, M.R. et al., 1998). and Batten, 2007), Bromus tectorum into Th e eff ectiveness of microbial inoculation western US shrublands (Bradley et al., 2006), in controlling invasive plants is also and A. cristatum into the northern Central complicated by the ecological specifi city of Plains of the USA (Christian and Wilson, interactions between plants and their 1999). Whereas shifts in the soil microbial microbial communities. Inoculation can community tend to have species-specifi c increase or decrease plant growth, depending impacts on plant growth, enhanced N on the identity of inoculated microbes, the availability often will increase the per- plant species, and the environmental formance of most plants when grown alone conditions (reviewed in Harris, 2009; in an invaded soil (Casper et al., 2008). Mummey et al., 2009), which supports the However, in mixed communities, increased use of local microbes for inoculation eff orts. soil N can shift plant community com- Th e source of inocula can have a strong position through selection for species that impact on restoration success. Most are more competitive (Clark et al., 2007; commercial inocula contain generalist AM Suding et al., 2008). Of particular concern is fungi that may not support the native plant that invader-induced increases in soil N community and may decrease soil mycor- availability will feed back to enhance rhizal diversity (reviewed in Harris, 2009). invasion, because soils with high N While generating native inoculum can be availability are more susceptible to plant challenging, it can be critical for eff ective invasion (reviewed in Heneghan et al., 2008; results. For example, on degraded shrub- Suding et al., 2008). lands in Spain, the biomass of plants was While the amount of soil N can be a key twice as high when inoculated with a mixture regulator of plant species composition, of indigenous AM fungi compared to invasive species may also change the timing inoculation with an exotic AM fungus and location of N availability. For example, (Requena et al., 2001). Pre-inoculation of leaching from Bromus tectorum litter native seedlings with desirable AM fungi redistributes soil nitrate deep in the soil may further help to minimize the AM fungal profi le, where native grasses cannot access taxa associated with exotic species it, thus increasing N availability to Bromus. (Mummey et al., 2009). Th is enhances Bromus growth at the expense of the native grasses (Sperry et al., 2006). Bromus also alters the timing of soil-N Nitrogen availability, with high soil-N availability occurring after the senescence of Bromus While many nutrients are critical in (Adair and Burke, 2010). Similarly, invasion regulating plant growth, interactions of exotic grasses into Hawaiian woodlands between invasive plants and nitrogen (N) greatly alters the seasonality of soil-N are particularly important because N is the availability. Grass invasion shifts most net-N most commonly limiting nutrient to plant mineralization from the dry season to the growth in temperate terrestrial ecosystems, wet season due to grass impacts on soil and as such, has strong impacts on plant organic matter enhancing wet-season N species composition and diversity (Eviner cycling, and grass impacts on microclimate and Chapin, 2003b; reviewed in Clark et al., decreasing dry-season N cycling rates (Mack 2007; Suding et al., 2008). On average, and D’Antonio, 2003). Invasive plants also invasive compared to native plant species, can alter the form of N available. For enhance N availability through increases in example, in California grasslands, USA, decomposition and N mineralization rates invasive grasses increase the soil nitrifi er (Ehrenfeld, 2003; Corbin and D’antonio, population, and thus nitrifi cation rates 2004; Liao et al., 2008), although some (Hawkes et al., 2005). Conversely, the The Effects of Plant–Soil Feedbacks on Invasive Plants 131

invasion of Andropogon garanus into removes the native seed bank and microbial Australian grasslands inhibits nitrifi cation community, which will need to be restored. (Rossiter-Rachor et al., 2009). While there While eff ective, topsoil removal can only are not clear examples of native versus be used in smaller restoration projects, invasive species performance being impacted and is limited to sites accessible to heavy by the form of N, the relative amount of N machinery. available as ammonium versus nitrate has Additions of biologically available carbon, been shown to alter competition between such as sawdust or sugar (as opposed to the species (reviewed in Marschner, 1986; more inert activated carbon), can fuel Crabtree and Bazzaz, 1993). growth of soil microbes, thus sequestering N in microbial biomass. Th is approach has been eff ective in reducing a number of Management invasions, and seems to be particularly Soil nitrogen can be removed by repeated eff ective in inhibiting grasses (rather than disturbances, including burning, grazing, or forbs or shrubs), and in shifting dominance mowing and removal of vegetation, and this from invasive annual to native perennial decrease in N can cause a shift from species (reviewed in Perry et al., 2010). dominance by competitive, weedy species, However, its eff ectiveness in reducing soil-N to a more diverse plant community (reviewed availability and controlling invasive species in Marrs, 1993; Walker et al., 2004; Perry et is variable, and often short-lived. In some al., 2010). In many cases, these techniques cases, adding carbon can actually enhance N are also used to directly decrease the availability and/or invasive species prevalence of invaders (e.g. by removing (Blumenthal et al., 2003; Krueger-Mangold invasive plants before they set seed), and et al., 2006; Corbin et al., 2007; Eviner and while the timing of disturbance may have Hawkes, 2008; reviewed in Alpert, 2010; minimal impacts on N removal, it will be Eviner et al., 2010; Perry et al., 2010; critical in infl uencing which plant species Kulmatiski et al., 2011). Th e amount and reestablish (Pollak and Kan, 1996; DiTomaso type of carbon needed to sequester N can et al., 2006; Holdredge and Bertness, 2011). vary by species and site (Blumenthal et al., In the long term, repeated disturbances can 2003; Prober et al., 2005). In some cases, the reduce N availability, but in the short term, amount of carbon needed may be prohibitive N availability can be enhanced immediately due to expense and logistics, suggesting that after disturbance, and in some cases, this it may be a tool appropriate to small, high- increase in N can be sustained during the intensity restoration sites, but may not be next few disturbance cycles (reviewed in feasible across large areas (Perry et al., Perry et al., 2010), making the system 2010). Even when it is eff ective in vulnerable to reinvasion if invasive species sequestering N, much of this N is re-released propagules are present. within a few months to a few years, so this In extremely high fertility sites, such as technique is most often eff ective in those that have been fertilized for years, it conjunction with quickly restoring native can take decades to adequately restore target plant species (reviewed in Perry et al., 2010). N cycles and the plant community through Planting native species that decrease N grazing, burning, or mowing (reviewed in availability through high N uptake has Walker et al., 2004). In these extreme cases, decreased the prevalence of some invaders. topsoil removal (also known as sod-cutting) In rangelands of northwestern USA, can rapidly remove accumulated nutrients planting Secale cereale or the native perennial and organic matter, as well as soil microbes grass Elymus elymoides decreased available and many invasive plant propagules in the soil N, shifting competitive dominance from seed bank (reviewed in Marrs, 1993; Walker the invasive C. maculosa to the native late- et al., 2004). Topsoil removal is the most seral species Pseudoroegneria spicata (Herron eff ective method of quickly and reliably et al., 2001). Eff ective management often removing N (Perry et al., 2010), but it also requires a combination of approaches; using 132 V.T. Eviner and C.V. Hawkes

disturbance and/or carbon additions to conductivity levels. However, in many cases, temporarily decrease available N, in longer term decreases in salinity will require combination with fostering plants that can restoration of historic fl ood regimes and/or maintain low soil-N availability. For ground water-table levels (reviewed in example, in Australian grasslands, the Ladenburger et al., 2006). Where invasive combination of burning, carbon additions, plants have redistributed salts to be and seed additions of a native grass with concentrated at the soil surface, such as high-N uptake (Th emeda triandra) was Mesembryanthemum crystallinum (Vivrette required to eff ectively decrease weed cover and Muller, 1977), topsoil removal may be and reduce soil nitrate to levels found on required. native-dominated sites (Prober and Lunt, 2009). Alternatively, if soil N can be adequately reduced by carbon additions or Disturbance as an important feedback disturbance, low-N adapted plants can be pathway introduced, and their low litter quality can feed back to maintain or further decrease While this chapter focuses on plant–soil low-N availability (reviewed in Perry et al., feedbacks, invasive species can also greatly 2010). alter disturbance regimes to benefi t In instances where invasive plant species themselves. For example, B. tectorum in the may inhibit nitrifi cation through Great Basin, USA (Knick and Rotenberry, allelochemicals, activated carbon may be 1997), T. caput-medusae in the western USA eff ective in binding these allelochemicals (Davies and Svejcar, 2008), and invasive and increasing nitrifi cation rates (reviewed grasses in Hawaii, USA (D’Antonio and in Lau et al., 2008). In contrast, when Vitousek, 1992), can increase fi re frequency, invasive plant species enhance nitrifi cation thus enhancing their own growth at the rates, commercial nitrifi cation inhibitors expense of native species. Brassica nigra in can be used. Th ese are commonly added to California grasslands, USA, enhances fertilized agricultural sites (Prasad and herbivory of the native bunchgrass, Nassella Power, 1995), and have been eff ective in pulchra, by small mammals, and this eff ect decreasing some invasions, while enhancing extends 30 m away from invaded patches native species (Young et al., 1997, 1998). (Orrock et al., 2008).

Challenges in Understanding and Soil salinity Managing Feedbacks A few invaders have been shown to increase It is clear that invasive plant species can soil salinity, thus decreasing the performance alter the soil in a way that benefi ts their own of native competitors. Examples include performance, and in these cases, their Tamarix species (Smith, S.D. et al., 1998; eff ective eradication may require inter- Ladenburger et al., 2006), Carpobrotus edulis ference with invader–soil feedbacks. How- (Kloot, 1983), and Halogeton glomeratus ever, the study of feedbacks is still a relatively (Harper et al., 1996; Duda et al., 2003). new fi eld, and eff ective management Conversely, invasion of brackish marshes by requires a better predictive ability of Phragmites australis decreases salinity (cited feedback mechanisms and their relative in Ehrenfeld et al., 2005). importance, specifi city, context-dependence, and spatial and temporal patterns (reviewed Management in Ehrenfeld et al., 2005). Key challenges to understanding and managing feedbacks are Natural fl ooding and/or high rainfall can discussed below. leach salts from soils in the short term, and can be used in conjunction with promoting 1. Relative importance of diff erent feedback native species tolerant of higher electrical mechanisms. Clearly, many mechanisms can The Effects of Plant–Soil Feedbacks on Invasive Plants 133

drive plant–soil feedbacks, and these grasslands (reviewed in Ehrenfeld et al., mechanisms vary in both how specifi c their 2005; Ehrenfeld, 2010). Similarly, the impacts are for various native species and strength and magnitude of feedbacks are which management approaches will likely be likely to vary across space and time, and eff ective. Although some broad management depending on which species are interacting techniques are similar in method (e.g. (reviewed in Bardgett et al., 2005; Eviner et planting transitional species that can al., 2010). While some species consistently tolerate invasive species’ soil legacies), the generate negative soil feedbacks to con- selection of native species will depend on specifi c species across sites, the direction the underlying mechanisms. Individual and magnitude of feedbacks can diff er by invasive species may have multiple mech- site for other species (Casper et al., 2008). In anisms driving soil feedbacks that must be a particularly interesting example from an considered when attempting to ameliorate annual-herb-dominated community in the invasive soil legacies. For example, B. UK, eight plant species signifi cantly diff ered tectorum alters the amount and distribution in their eff ects on soil properties, which of available soil N (Sperry et al., 2006), as then fed back to impact the relative growth well as disturbance regimes (D’Antonio and of these species. N enrichment did not Vitousek, 1992), and micro bial communities impact the eff ects of these species on soil (Belnap and Phillips, 2001; Hawkes et al., properties, but the interaction of N 2006). Th e relative importance of diff erent enrichment with plant eff ects on soils feedback mechanisms likely varies with the greatly altered plant growth responses to specifi c invasive species, native species, and species-specifi c changes to soils (Manning et site conditions. Where multiple mechanisms al., 2008). are at play, selection of restoration approaches will require knowing if one key Time is a particularly important driver of feedback mech anism can be targeted, or if context-dependence of plant–soil feedbacks. management of each feedback pathway is Vulnerability to pathogens diff ers with life required. stage for a given plant species, and in the 2. Specifi city of feedback mechanisms. As extent to which mycorrhizal fungi can be reviewed in this chapter, many feedbacks negative or benefi cial (reviewed in Bardgett depend on the identities of the invasive and et al., 2005; Casper and Castelli, 2007; van native species. Soil feedbacks from one der Heijden and Horton, 2009). Th e length invader can impact a number of native of time an invader has been at a site has species, while a second invader in the same large impacts on the extent to which it ecosystem can have feedbacks that aff ect an changes the soil (reviewed in Strayer et al., entirely diff erent set of native species. As we 2006), and can thus alter feedbacks increase the number of well-developed case (reviewed in Bardgett et al., 2005). For studies of invader feedbacks, we will example, B. tectorum has its strongest improve our understanding of the types of positive feedback in its third generation on a native plants that are more sensitive to given soil (Blank, 2010). Changes over time specifi c changes in the soil physical, may be due to the accumulation of impacts chemical, and biotic environment. (e.g. accumulation of soil organic matter) or 3. Context-dependence of feedbacks. Many shifts in the relative strengths of positive studies have shown that the impacts of versus negative feedback pathways (e.g. soil plant species on soils vary with environ- symbionts versus pathogens), as occurs mental conditions and the amount of time during succession (Kardol et al., 2007). an invader has been present (reviewed in Another key time-related concern is the Ehrenfeld et al., 2005; Strayer et al., 2006; persistence of invader eff ects on soils, even Eviner and Hawkes, 2008). For example, in after invasive plant species have been North America, B. tectorum can increase removed from a site. When an invader has rates of N cycling in cool deserts, and been at a site for decades to centuries, its decrease N-cycling rates in warmer arid impacts on soil microbes, organic matter, 134 V.T. Eviner and C.V. Hawkes

and nutrients can persist long after the thunbergii on the soil microbial community invader has been removed (reviewed in were not proportional to its relative Eviner and Hawkes, 2008; Eviner et al., abundance in the mixture (Elgersma and 2010). Even when an invader has been at a Ehrenfeld, 2011). site for a short duration, its impacts on soil may persist long enough to interfere with native plant restoration (Grman and Suding, 2010). Summary 4. Relative importance of feedbacks versus other drivers of invasion. Th ere are many While there is still much to learn about the potential mechanisms driving invasions role of plant–soil feedbacks in exotic plant (Th eoharides and Dukes, 2007), and a species invasions, they clearly do play an number of these may be operating integral role in some systems, and must be simultaneously. It is critical to compare the addressed to restore resilient native com- relative importance of soil feedbacks to munities. Despite the considerable variation other mechanisms such as competition in the eff ects of invasive species across space (Casper and Castelli, 2007), propagule and time within a specifi c area, current tools pressure (Eppstein and Molofsky, 2007), for altering plant–soil feedbacks show and release from aboveground natural considerable promise and will be improved enemies such as herbivores and pathogens with more case studies and collaborations (Mitchell and Power, 2003; Agrawal et al., between land managers and researchers. 2005). For some invasive species, factors While few ‘rules of thumb’ for management such as competition and climate are more are available from this emerging fi eld, some important than soil-feedback eff ects (e.g. general principles do apply: Yelenik and Levine, 2011). In other cases, • As with other mechanisms of invasion invasive plant species dominance may be (e.g. high propagule availability), the maintained by the combination of asym- most effi cient management approach will metric competition generated through early be to quickly eradicate new infestations germination and the negative feedbacks to of invasive species, before they are able native plants generated by soil legacies to alter soil conditions to benefi t them- (Grman and Suding, 2010). More work will selves. be required to understand the role of soil • Testing potential approaches to manage feedbacks relative to other mechanisms in plant–soil feedbacks (Table 7.1) without invasive species success. knowing the mechanism driving the 5. How prevalent does an invader need to feedback can be risky. Th e species-specifi c be to induce feedbacks? It is often assumed nature of most of these feedback that the impacts of a plant species on the mechanisms indicates that many of these soil are proportional to its biomass in the management techniques have a chance to community (Grime, 1998; Parker et al., promote, rather than control invasive 1999), but recent work has shown that some species. In cases where the mechanisms invasive species can have signifi cant impacts are not known, trials should be small- on soils even when they are relatively rare. scale and well monitored, before they are For example, in a river fl oodplain in New applied to broader areas of invasion. Zealand, non-native plants made up less • Th e mechanisms driving plant–soil than 3% of plant community biomass, but feedbacks, and the strength and direction had signifi cant impacts on soil carbon, of these feedbacks can change greatly microbial biomass, and microbial com- over time since invasion, as well as site- munity structure (Peltzer et al., 2009). to-site. Th us, even when management Similarly, varying proportions of native and has successfully disrupted invader–soil invasive plant litter demonstrated that the feedbacks at one site, preliminary trials eff ects of litter of the invasive Berberis under diff erent conditions (or at sites The Effects of Plant–Soil Feedbacks on Invasive Plants 135

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Wolkovich, E.M., Bolger, D.T. and Cottingham, K.L. D.E. (1998) Nitrogen interactions with (2009) Invasive grass litter facilitates native medusahead (Taeniatherum caput-medusae shrubs through abiotic effects. Journal of ssp. asperum) seedbanks. Weed Science 46, Vegetation Science 20, 1121–1132. 191–195. Yelenik, S.G. and Levine, J.M. (2011) The role of Young, T.A., Evans, R.A. and Kay, B.L. (1971) plant-soil feedbacks in driving native-species Germination of caryopses of annual grasses in recovery. Ecology 92, 66–74. simulated litter. Agronomy Journal 63, 551–555. Young, J.A., Clements, C.D. and Blank, R.R. (1997) Zhang, Q., Yang, R., Tang, J., Yang, H., Hu, S. and Infl uence of nitrogen on antelope bitterbrush Chen, X. (2010) Positive feedback between seedling establishment. Journal of Range mycorrhizal fungi and plants infl uences plant Management 50, 536–540. invasion success and resistance to invasion. Young, J.A., Trent, J.D., Blank, R.R. and Palmquist, PLoS ONE 5, e12380. Species Performance: the Relationship Between Nutrient 8 Availability, Life History Traits, and Stress

Jeremy J. James

US Department of Agriculture, Eastern Oregon Agricultural Research Center, USA

Introduction component to restoring invasive plant- infested systems. However, the role soil Diff erences in species performance (i.e. how nutrient availability plays in our ability to a species captures and utilizes resources to restore invasive plant dominated systems is maintain and increase population size) not as straightforward as it initially appears. infl uence the rate and direction of plant Th e objectives of this chapter are to: (i) community change (Sheley et al., 2006). examine current paradigms and assumptions Species performance is determined by a about how nutrient availability infl uences number of interacting factors. Th is includes the relative performance of invasive and resource supply rates, physiological traits native plants; (ii) develop the argument that that determine how a species aff ects and the infl uence of nutrient availability on responds to the environment, life history species performance is mediated by life traits that determine patterns of birth, history traits and stressors such as drought mortality, and growth of individuals in a and herbivory; and (iii) describe how these population, as well as abiotic and biotic principles that determine species per- stressors such as herbivory or drought. formance in diff erent nutrient environ- Researchers and land managers long have ments can be used to develop more eff ective recognized that our ability to predict and invasive plant management strategies. manage the spread of invasive species directly depends on our understanding of the processes that diff erentially impact the Current Paradigms of Nutrient performance of invasive and native plants. Availability and Invasion In nutrient-poor systems across the globe, increases in nutrient availability increase Background the susceptibility of ecosystems to invasion (Mack et al., 2000). Th is general and Th eoretical and empirical work has widespread response indicates that increases established a positive relationship between in nutrient availability are likely favoring resource availability and habitat invasibility. the performance of invasive species over For example, in nutrient-poor systems native species. domin ated by slow-growing species, inva- Having recognized this, restoration sions have commonly been attributed to practitioners have long considered soil increased soil nitrogen (N) availability nutrient management to be a fundamental following disturbance (McLendon and

© CAB International 2012. Invasive Plant Ecology and Management: 142 Linking Processes to Practice (eds T.A. Monaco and R.L. Sheley) The Relationship Between Nutrient Availability, Life History Traits, and Stress 143

Redente, 1991; Kolb et al., 2002; Brooks, in invasive plant-dominated coastal and 2003). Due to their faster growth rates and interior grasslands, as well as sagebrush ability to rapidly take up N, fast-growing steppe communities in North America, did invasive species are thought to be more not facilitate reestablishment of perennial competitive than slow-growing native grasses (Morghan and Seastedt, 1999; species in high N soils (Norton et al., 2007; Corbin and D’Antonio, 2004; Huddleston Mazzola et al., 2008; Mackown et al., 2009; and Young, 2005; Mazzola et al., 2008). Th e Perry et al., 2010). In contrast, slow-growing discrepancies between these studies and the native species, with their greater investment current paradigm of soil N eff ects on invasive in belowground structures and ability to and native plant performance have been recycle and store N (Chapin, 1980; Fargione explained a number of ways including lack of and Tilman, 2002), are thought to be favored long-term eff ects of treatments on soil N, under low N conditions. Th ese observations the type of treatments used to reduce soil N, have led to the logical assumption that and environmental conditions such as managing soils for low nutrient availability drought (Blumenthal, 2009; Brunson et al., will facilitate restoration of invasive plant- 2010; Perry et al., 2010). While some dominated systems. variation among studies could be due to Based on the positive relationship these factors, the large discrepancies in between nutrient availability and habitat overall conclusions reached in these research invasibility, land managers and researchers eff orts suggests the current conceptual have been applying a range of techniques to framework for understanding the role lower soil N prior to native plant restoration, nutrient management plays in our ability to including the use of cover crops, carbon restore invasive plant-dominated systems is amendments, and topsoil removal (Perry et incomplete. al., 2010). Following decades of manipu- lations, however, the eff ect of soil N management on the relative performance of Reconsideration of the nutrient native and invasive species has been mixed. availability–invasibility paradigm For example, work in some grassland systems has shown that reductions in soil N Th e development of the current nutrient availability may reduce the growth of management paradigm is largely centered invasive species and facilitate the establish- on the assumptions that traits such as ment of native species (Morgan, 1994; fast growth make invasive species better Blumenthal et al., 2003). Likewise, reduction ‘adapted’ to high nutrient environments in soil N availability to invasive annual than native species while slow growth grass-dominated coastal prairie and old- makes native species better ‘adapted’ to low fi elds in North America and pasture in nutrient environments than invasive Australia lowered soil N availability and species. However, over the last decade or so, facilitated reestablishment of perennial there have been major advances in under- grasses (Alpert et al., 2000; Paschke et al., standing the plant physiological and 2000; Prober et al., 2005). While these morphological traits underpinning the studies did not identify how decreasing soil diff erent ecological strategies displayed by N facilitated establishment of native species, native and invasive species as well as the these results have been largely attributed to ecological trade-off s associated with these an overall lower N requirement of native traits (Weiher and Keddy, 1999; Diaz et al., plants or to changes in competitive 2004; Wright et al., 2004). Th ese insights relationships among species as N availability are instrumental for determining how our declined. ability to predict nutrient availability eff ects A number of other studies, however, have on species performance can be improved. not observed the expected eff ect of lowering For example, current plant trait frame- soil N on invasive and native species works suggest plant ecological strategies performance. For example, lowering soil N may be more clearly diff erentiated based on 144 J.J. James

how a species balances the trade-off between when environmental stress or herbivory resource acquisition and resource con- pressure is high. Th erefore, in a restoration servation, rather than growth potential scenario where both native and invasive (Weiher and Keddy, 1999; Diaz et al., 2004; species are recruiting from the seed bank, Wright et al., 2004). Under this framework, invasive species may maintain an initial some species maximize resource con- growth advantage even in low nutrient soils. servation by making mechanical and If this initial growth advantage allows chemical investments in tissue that increase invasive species to interfere with the growth tissue lifespan and decrease tissue loss due and survival of native species, then native to herbivory or environmental stress species will have little opportunity to (Lambers and Poorter, 1992; Westoby et al., capitalize on traits that ultimately will 2002; Wright et al., 2004). While these provide them an advantage in nutrient-poor investments decrease relative growth rate soils. Given this potential and the variable (RGR), these traits increase mean nutrient results seen in soil N manipulation studies, residence time, allowing a greater duration it may be helpful to examine a quantitative of return on nutrients captured and greater synthesis of the literature and determine if nutrient conservation (Berendse and Aerts, the balance of evidence supports the current 1987). Over the long term, these traits are paradigm or suggests a need to further expected to be advantageous in low nutrient refi ne a predictive framework for the role soils (Berendse, 1994; Aerts, 1999). On the soil nutrient management plays in restoring other hand, some species make thin, short- invaded systems. lived tissue. While this decreases tissue life span and increases tissue susceptibility to herbivory or environmental stress, it allows Evaluation of the nutrient availability– these species to construct more absorptive invasibility paradigm root and leaf surface area per unit of biomass allocated to tissue, resulting in more rapid A recent study used meta-analysis to resource capture (Lambers and Poorter, evaluate the degree to which soil N 1992). Th ese traits are expected to provide management diff erentially impacts growth an advantage in nutrient-rich soils where and competitive ability of invasive annual resource capture is favored over resource and native perennial grasses (James et al., conservation. 2011). Meta-analysis is a useful tool to Th is research suggesting tissue con- quantitatively synthesize results of inde- struction strategy (thin, short-lived versus pendent studies (Gurevitch and Hedges, thick, long-lived) infl uences the balance 1993). In the experiments analyzed in this between resource acquisition and resource meta-analysis, all species were started from conservation and, consequently, dif- seed and thus provide useful insight into ferentiates the ecological strategies of how soil conditions may impact initial plants, has major implications for predicting growth and competitive interactions and set how we expect soil N management to impact the stage for long-term plant community restoration eff orts. Specifi cally, because changes in a restoration setting where both certain invasive species invest little in tissue native and invasive species are starting as protection and defense, over the short term seed. Here three hypotheses were tested: (i) they may achieve greater growth rates than decreasing soil N availability has a greater native species even in low nutrient soils negative eff ect on biomass and tiller (Burns, 2004; Garcia-Serrano et al., 2005; production of annual compared to perennial James, 2008). In addition, the advantages grass seedlings; (ii) however, seedlings of native species have in low nutrient soil, fast-growing species, including invasive including greater nutrient mean residence annual grasses, maintain a higher RGR than time and greater protection from seedlings of slow-growing species in low N environmental stress or herbivory, are only environments, allowing invasive annual going to be manifested through time and grasses to construct more biomass and The Relationship Between Nutrient Availability, Life History Traits, and Stress 145

tillers than perennials in the short-term; Consistent with the greater growth rate, and (iii) as a result, lowering soil N biomass and tiller production by invasive availability will not alter the competitive annuals compared to native perennials in advantage that annual grass seedlings have low N soil, and in support of the third over perennial grass seedlings. hypothesis, lowering soil N availability did To test these hypotheses, 70 comparisons not alter the competitive advantage invasive were extracted from 25 studies and eff ect annual grass seedlings have over native size was calculated as the natural log perennial grass seedlings. While all point response ratio (ln RR) where: estimates of competition parameters were negative, indicating plants competed under T ln RR lnz lnTT ln (8.1) low and high N, confi dence intervals for low T zb b and high N overlapped substantially (Fig. 8.2), suggesting N availability had little and Tz and Tb are means of the response variable for two diff erent treatment groups eff ect on competition intensity. Moreover, (Hedges et al., 1999). Th e parameter ln RR the competition studies showed that annual estimates the proportional diff erence between treatment groups. An ln RR of zero (a) Biomass indicates the response variables do not vary between groups, and a positive or negative ln RR indicates the response variable is larger for the Tz or Tb group, respectively. Figures 8.1–8.3 show the results of this quantitative synthesis. Figure 8.1 shows the eff ect of soil-N availability on the diff erence in (a) biomass, (b) number of tillers, and (c) relative growth rate (RGR) of invasive (b) Tiller annual grasses and native perennial grasses growing in low and high N environments. Confi dence intervals on biomass diff erences (annual – perennial) for high and low N do not overlap (Fig. 8.1a). Th is supports the fi rst hypothesis that reducing soil N decreased annual grass biomass by a greater proportion than perennial grass biomass. Th e confi dence interval on the RGR (c) RGR diff erence (fast growing species – slow growing species) is greater than 0 for low N (Fig. 8.1c). Th erefore, fast growing invasive annual grasses maintained higher RGR under low and high N compared to slow- growing perennial grasses. As a result of this higher RGR, invasive annual grasses produced more biomass and tillers than perennial grasses when N availability was low (Fig. 8.1a, b). Th is supports the second hypothesis that seedlings of fast-growing Fig. 8.1. Most likely parameter values (dots) and species, including invasive annual grasses, 95% confi dence intervals (bars) on log response ratios (ln RR) where: (a) ln RR = ln(perennial maintain a higher RGR than seedlings of grams per plant) – ln(annual grams per plant); (b) slow-growing species in low N environments, lnRR = ln(perennial tillers per plant) – ln(annual allowing invasive annual grasses to construct tillers per plant); and (c) ln RR = ln(RGR of fast more biomass and tillers than perennials in growing species) – ln(RGR of slow growing the short term. species). 146 J.J. James

Fig. 8.3. Most likely parameter values (dots) and 95% confi dence intervals (bars) on log response ratios (ln RR). The ln RR describes effects of nitrogen availability on biomass production of perennial grasses growing with invasive annual grass neighbors. Specifi cally, ln RR = ln(biomass of perennial plants grown with annual grasses Fig. 8.2. Most likely parameter values (dots) and under high N) – ln(biomass of perennial plants 95% confi dence intervals (bars) on log response grown with annual grasses under low N). Separate ratios (ln RR) from addition series studies estimates are provided for the three types of included in the meta-analysis. More negative competition designs analyzed in the meta-analysis values indicate more competitive effects of (addition series, replacement series, and simple neighbors on target plant biomass. addition). grass neighbors had a stronger competitive increasing N availability facilitates invasion, eff ect on both annual and perennial targets decreasing N availability should facilitate compared to perennial neighbors and that restoration of systems dominated by the competitive eff ects of annual neighbors invasive plants (McLendon and Redente, on perennials did not decrease with lower N 1992; Alpert et al., 2000; Blumenthal et al., availability (Fig. 8.2). In addition, perennial 2003). Th e meta-analysis described here targets competing with annual grass rejects this assumption for invasive annual neighbors did not incur a net cost in biomass grasses. In addition, because many invasive production in high-N environments (Fig. plant species display traits that favor 8.3). Point estimates and confi dence resource acquisition over resource con- intervals describing the diff erence in servation compared to their native perennial plant biomass when perennial counterparts (Grotkopp et al., 2002; James plants were grown with annual neighbors and Drenovsky, 2007; Leishman et al., under high N compared to when perennial 2007), these fi ndings likely apply to other plants were grown with annual grass invasion scenarios. Th is means it is very neighbors under low N were positive or likely that when invasive and native species overlapped zero in the diff erent types of are establishing from seed, soil nutrient competition designs used. Th is indicates management alone will not directly perennials produced either more biomass facilitate establishment of native species. when grown with annual grass competitors While this conclusion is supported by an under high N compared to when grown with array of studies showing soil N management annual grass competitors under low N, or failed to facilitate restoration of invasive that there was no detectable diff erence in plant-dominated systems (e.g. Corbin biomass produced under high and low N. and D’Antonio, 2004; Huddleston and Th e current soil N management frame- Young, 2005; Mazzola et al., 2008), this work rests on the assumption that because conclusion is at odds with studies that have The Relationship Between Nutrient Availability, Life History Traits, and Stress 147

demonstrated that soil N management capture rates, it increases the ability to facilitates restoration of systems infested by conserve resources because low SLA species a range of invasive plants (e.g. Alpert et al., tend to have longer leaf lifespan than high 2000; Paschke et al., 2000; Prober et al., SLA species (Westoby et al., 2002). A high 2005). An important question then is how SLA, on the other hand, allows quicker can we use these results and current return on nutrients and dry matter invested understanding of plant life history trade- in leaves, resulting in rapid resource capture off s to reconcile these discrepancies and but poor resource conservation due to improve our ability to predict how nutrient shorter leaf lifespan. Variation in SLA availability impacts species performance? among species, therefore, represents a key axis describing diff erent plant ecological strategies that range between rapid resource Infl uence of Nutrient Management on capture and eff ective resource conservation Species Performance: the Role of Life (Diaz et al., 2004; Wright et al., 2004; History, Drought, and Herbivory Leishman et al., 2007). While we would expect resource con- Advances in understanding traits underlying servation strategies to become increasingly variation in plant ecological strategies have important as soil nutrient availability been made. Th ese advancements can declines, favoring low SLA species, the meta- improve our ability to understand and analysis demonstrated that fast-growing, predict how soil nutrient management may high SLA species, including invasive annual infl uence restoration outcomes. Potential grasses, maintained greater growth rates in growth rate long has been a key trait used to low-N soils than slow-growing species. characterize diff erent plant ecological Th ese diff erences in initial RGR were strategies (e.g. Grime, 1977; Goldberg and refl ected in the greater biomass and tiller Landa, 1991; Loehle, 2000). When grown production by invasive species in low-N soils under optimum conditions, species that compared to native species. High-SLA dominate nutrient-rich sites typically have a species also may produce thinner and less higher RGR than species that dominate dense root tissue compared to low-SLA nutrient-poor sites (Lambers and Poorter, species, resulting in a lower specifi c root 1992). While a high RGR allows rapid growth length (SRL) (Elberse and Berendse, 1993). and resource capture, the ecological Construction costs per unit tissue dry advantage a low RGR provides is less clear. weight are similar among species (Poorter Early interpretations focused on advantages and Bergkotte, 1992). As a result the thinner directly due to low RGR, such as the potential or less dense root and leaf tissue generally for low RGR species to function closer to produced by invasive compared to native their physiological optimum in infertile soils species allows invasive species to maintain a compared to high RGR species (Grime and greater rate of return on biomass allocated Hunt, 1975; Chapin, 1980). Th e current to leaves and roots and ultimately grow consensus, however, argues that traits faster than native species. At the seedling associated with low RGR, not RGR itself, stage, this diff erence in tissue production have been the target of selection in nutrient- cost allows invasive species to preempt more poor systems (Lambers et al., 1998; Aerts, belowground resources than native species 1999). Specifi c leaf area (SLA) is the principle in nutrient-poor soils (James, 2008). trait infl uencing RGR variation among Th erefore, because of these diff erences in species (Lambers and Poorter, 1992). tissue construction strategies, at the Morphological and chemical adjustments seedling stage, slow-growing native species that increase leaf tissue toughness and have no direct advantage in terms of initial protection from abiotic stress or herbivores growth in nutrient-poor soils compared to decrease SLA and consequently, RGR invasive species. If restoration outcomes are (Poorter et al., 2009). While a low SLA largely determined during the seedling indirectly reduces growth and resource establishment phase, then soil nutrient 148 J.J. James

management alone likely will not infl uence et al., 2007). Similar to nutrient recycling restoration outcomes. traits, because these traits defl ect some Restoration outcomes, however, also are biomass away from growth functions (e.g. infl uenced by longer-term processes, not creation of leaf or root surface areas for just processes associated with the seedling photosynthesis and nutrient uptake) toward establishment phase. Internal plant nutrient nutrient conservation functions, these recycling, for example, contributes sig- traits place native seedlings at an initial nifi cantly to the long-term performance of disadvantage in nutrient-poor soils. natives in low nutrient soils (Killingbeck and Modeling and empirical work, however, Whitford, 1996). Plants have two primary demonstrate that over time, these traits sources of nutrients: soil nutrient pools and allow slow-growing species to accumulate nutrients retained within the plant. Th e more biomass and have faster population reliance on these two diff erent nutrient growth rates than fast-growing species in sources depends on a variety of factors, nutrient-poor soils (Berendse et al., 1992; including internal plant nutrient recycling, Berendse, 1994; Ryser, 1996). soil nutrient availability, and life history. Th e outcome of competition at the During tissue senescence, biomolecules are seedling stage, however, is not solely based broken down and the mineral nutrients are on resource capture diff erences, and other translocated to storage tissues, such as abiotic and biotic factors likely will be stems and roots. Internal nutrient recycling important in infl uencing the outcome of buff ers plants against variation in soil competition between invasive and native nutrient availability (Aerts and Chapin, species in nutrient-poor soils. In addition to 2000), impacts plant fi tness (May and increasing tissue life span and nutrient Killingbeck, 1992), and may be less costly residence time, the thicker, longer-lived leaf than acquiring and assimilating soil and root tissue constructed by native nutrients (Wright and Westoby, 2003). compared to invasive species can decrease Although the amount of nutrients resorbed tissue loss from drought. Structural (i.e. realized resorption) may vary annually, investments in leaf and root cells and tissue the maximum amount of nutrients that can xylem increase tissue density and are an be resorbed physiologically (i.e. potential opportunity cost in terms of investment of resorption) is considered an evolved ‘target biomass in surface area for resource capture. value’ (Killingbeck, 2004). As such, potential Th ese investments, however, allow plants to resorption is greater in plants from low- maintain physiological function as soils dry nutrient habitats compared to plants that compared to plants that invest less in these dominate nutrient-rich systems (Killing- structures. Th ese diff erences in drought beck, 2004). If native species resorb more tolerance may heavily infl uence competition nutrients from dying tissue in nutrient-poor outcomes in environments with low and soils than invasive species, then over time variable precipitation inputs (Goldberg and natives will accumulate larger nutrient Novoplansky, 1997; Chesson et al., 2004). pools, which should place them at a com- Likewise, mechanical investments in tissue, petitive advantage. that decrease susceptibility to drought, also In addition to nutrient recycling, native can decrease tissue palatability, making plants also conserve nutrients by investing these plants less susceptible to attack by heavily in structural support for leaf and generalist invertebrate herbivores. For root tissue. While these investments example, Buckland and Grime (2000) showed decrease SLA and SRL and lower rates of that in fi eld microcosms, fast-growing growth and resource capture, they increase species dominated low, moderate, and high tissue life span. As a consequence, these fertility soils in the absence of a generalist traits result in longer mean nutrient invertebrate herbivore. When a generalist residence time and a lower relative nutrient herbivore was added, the abundance of slow- requirement for native compared to invasive growing species increased in low-fertility species (Berendse and Aerts, 1987; Berendse soils but not in high-fertility soils. Th us, The Relationship Between Nutrient Availability, Life History Traits, and Stress 149

population dynamics of invertebrate 1. Managing environments for low nutrient herbivores that target fast-growing species availability will favor resource conservation with less structural investment in tissue, over resource capture by plants, favoring may play an important role in determining desired species over invasive species. outcomes of competition and assembly in 2. Initial establishment of desired species nutrient-poor soils (Fraser and Grime, 1999; needs to be successfully managed to realize Olofsson, 2001; Burt-Smith et al., 2003). any benefi t of nutrient management. 3. When desired species are establishing from seed, a suffi cient amount of stress that Designing Ecologically Based inhibits the performance of invasive species Invasive Plant Management will need to be applied to realize any benefi t Strategies Based on Principles of of nutrient management. Species Performance in Low and To understand how these principles can be High Nutrient Environments incorporated into a practical invasive-plant management program it may be useful Th eory, empirical work, and the results of to consider two contrasting restoration the meta-analysis described here all support scenarios. In both scenarios there is a the notion that increases in nutrient signifi cant invasive plant seed bank, but in availability favor invasive species’ per- scenario one, most native plants must be formance over native species (Huenneke et seeded, while in scenario two, a signifi cant al., 1990; Stohlgren et al., 1999; Davis et al., number of native plants are still present at 2000). Th ere also is evidence that invasive the site. species, once established may accelerate If native species must be seeded, they nutrient cycles and maintain high levels of cannot capitalize on key traits that provide nutrient availability, ensuring resource them an advantage in low nutrient soils, conditions continue to favor dominance of such as the ability to recycle stored nutrients invasive species (Liao et al., 2008). From or their greater mean nutrient residence these results, it is clear that invasive-plant time. Managing soils for low nutrients will managers should carefully examine how not provide a growth advantage to native management decisions and natural pro- compared to invasive species at the seedling cesses impact nutrient availability. Manage- stage. If herbivore pressure or abiotic stress ment decisions and natural processes is pronounced, the lower SLA and SRL of that increase nutrient availability will favor native seedlings may be advantageous due the performance of invasive over native to greater investment in lignin or other species. Th erefore, maintaining low compounds that increase tissue toughness. nutrient environments is a central goal of On the other hand, if stresses like drought ecologically based invasive plant manage- or herbivory do not occur, or cannot be ment programs. applied in a manner that preferentially However, the meta-analysis and our impacts invasives (e.g. livestock grazing at current understanding of traits underlying inappropriate times), the performance of variation in plant ecological strategies invasive species will have to be controlled via suggest nutrient management alone will not other means (e.g. herbicide) for a long be suffi cient for facilitating establishment of enough period to give native species an native species in invasive plant-dominated opportunity to establish and realize the systems. By considering the meta-analysis benefi ts of their resource conservation and the trait-based frameworks of plant traits. If this is achieved, managers should ecological strategies we can formulate some expect a long-term benefi t to soil nutrient principles and predictions about how and management. If, however, the initial when soil N management will positively performance of invasive species is not infl uence restoration outcomes. Th ree managed, managers should not expect to principles emerge: receive a benefi t from nutrient management. 150 J.J. James

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Joseph M. DiTomaso1 and Jacob N. Barney2

1 Department of Plant Sciences, University of California, USA 2 Department of Plant Pathology, Virginia Tech, USA

Introduction plasticity, than native species. Plants with greater adaptive qualities are expected to Most non-native plants in natural areas do maintain higher fi tness and perform better not out-compete native species or cause in a wider range of habitats and environments signifi cant impacts to ecosystem function (Rejmánek and Richardson, 1996; Alpert et (Rejmánek, 2000, 2011; Smith and Knapp, al., 2000; Rejmánek, 2011), particularly in 2001). It has been estimated that <10% of disturbed environments where conditions invasive species that have established and are in frequent fl ux (Claridge and Franklin, persist in natural areas actually transform 2002; Daehler, 2003). Th ese same traits that the ecosystem by changing the character, allow a species to be a successful invader condition, form, or nature of an area outside its native range also account for (Richardson et al., 2000). Th ere are many widespread distribution in its native range theories and reviews on why species become (Richardson and Pyšek, 2006). However, it invasive, including release from natural is important to recognize that increased enemies and herbivores in their native range adaptive qualities in non-native invaders do (Keane and Crawley, 2002; Daehler, 2003), not indicate a performance advantage over improved competitive ability through a shift native species within any single, defi ned in allocation from defense to growth environment (Daehler, 2003). (Blossey and Nötzold, 1995), and the Phenotypic adaptation not only enhances development of novel growth or functional the invasive potential of non-native species, forms in invasive species that have com- but also increases the challenge of manipu- petitive advantages over native species lating their performance in management and (Alpert et al., 2000). While these general restoration programs. Long-term sustainable theories may contribute to invasiveness in management programs will require per- non-native species, specifi c traits related to sistent competition from desirable species or reproduction, establishment, growth, com- increased environmental stress to reduce the petitive ability, susceptibility to herbivory abundance and performance of invasive and pathogens, and dispersal can all con- species. Th is chapter will discuss the various tribute to the success of invasive species in ecological processes operating at the plant specifi c habitats (Smith and Knapp, 2001; and ecosystem level associated with the Van Clef and Stiles, 2001; Ehrenfeld, 2004). invasibility of natural ecosystems, and the In general, most successful invasive management tools that can manipulate these plants tend to have greater adaptive processes to achieve a more desirable plant qualities, often referred to as phenotypic community.

© CAB International 2012. Invasive Plant Ecology and Management: 154 Linking Processes to Practice (eds T.A. Monaco and R.L. Sheley) Reducing Invasive Plant Performance: a Precursor to Restoration 155

Key Ecological Processes vegetation is multilayered, such as forests and riparian areas. An increase in light Resource acquisition and plant–soil availability can facilitate rapid invasion of resource interactions many woody species. For example, in an eastern US forest community, three invasive Resources that are most limiting in many woody species, Chinese tallow tree (Triadica environments include light, water, and soil sebifera (L.) Small), Amur honeysuckle nutrients, particularly nitrogen and phos- (Lonicera maackii (Rupr.) Herder), and phorus. Plant communities that are con- oriental bittersweet (Celastrus orbiculatus sidered to be most susceptible to invasion are Th unb.), often form a seed and seedling those with high availability of unused bank under shaded conditions of an resources (Davis et al., 2000; Smith and established forest. Once light becomes Knapp, 2001). Th is generally occurs in available through natural disturbance, the communities where the invasive species does seedlings of invasive species are poised to not encounter intense competition for rapidly respond and quickly occupy the resources from resident native or desirable canopy of the site (Jones and McLeod, 1989; species (Richardson and Pyšek, 2006). For Luken and Goessling, 1995; Marks and example, species-poor communities, often Gardescu, 1998). In contrast, the native caused by natural or human-related dis- species Celastrus scandens L. does not have a turbance, can have more free resources com- persistent seed or seedling bank, and is out- pared to undisturbed species-rich com- competed for light by its congeneric invasive munities. Although not always the case relative or other invasive species (Van Clef (Alpert et al., 2000; Maron and Marler, 2008), and Stiles, 2001). invasive species grow more quickly and are Even in grassland communities, light can more successful than native species when play a signifi cant role in altering the resources are high and stress is low, but native regeneration and growth of native grass and species often compete better with invasive forb species (Gerlach et al., 1998). In species when resource availability is low and California grasslands, seedlings of invasive stress is high (Alpert et al., 2000). Th us, the forbs, including yellow starthistle (Centaurea level of environmental stress can greatly solstitialis L.) and artichoke thistle (Cynara impact the relative performance of native or cardunculus L.), are sensitive to low light and desirable species compared to invasive plants. do not eff ectively compete for light in well When increased resources are available, established dense grasslands (DiTomaso et frequently through disturbance, invasive al., 2003; White and Holt, 2005). However, species are generally able to use them more grazing or mowing can increase light quickly and effi ciently than native species penetration resulting in rapid establishment (Funk and Vitousek, 2007; Sheley et al., and dominance of invasive forbs in these 2010; Rejmánek, 2011). Disturbance, ecosystems. whether natural or human-related, can vary Once established, invasive species can seasonally, causing resource pulses, defi ned dominate plant canopies, thus preventing by high resource fl uctuation and availability. light penetration to native species below. Th ese pulses in resource availability can be a Th is is most notable for many climbing vines key factor controlling invasibility of eco- that grow over adjacent vegetation, such as systems (Davis et al., 2000; Chambers et al., kudzu (Pueraria montana (Lour.) Merr. var. 2007), particularly when invasive species lobata (Willd.) Maesen & S. Almeida), propagule pressure to the system is high bittervine (Mikania micrantha Kunth), (Richardson and Pyšek, 2006). Japanese climbing fern (Lygodium japonicum L.), and old world climbing fern (Lygodium myriophyllum L.) (Bryson and Carter, 2004; Light Brooks et al., 2008). In a comparative study Light can be the principal limiting factor in with native woody species in riparian and many environments, particularly when the foothill habitats of California, USA, invasive 156 J.M. DiTomaso and J.N. Barney

species nearly always reduced canopy While high nitrogen conditions rarely openness to a higher level compared to the favor native species, reductions in available native woody vegetation (Sedna and nitrogen often increase the relative Rejmánek, unpublished data). Th e level of performance and competitive ability of openness under the invasive shrub Scotch native species, particularly perennial grasses broom (Cytisus scoparius (L.) Link) was (Alpert et al., 2000; Paschke et al., 2000; signifi cantly lower compared to the native Daehler, 2003). Th is is not always the case, manzanita shrub (Arctostaphylos viscida however. Under low nitrogen conditions, Parry) during the growing season. In invasive annual rangeland grasses acquired addition, light levels under fi g trees (Ficus more nitrate than the associated native carica L.) were an order of magnitude lower perennial grasses, resulting in equal or than the native woodlands, which were greater invasiveness compared to higher expressed in the virtual absence of other nitrogen conditions (Monaco et al., 2003). species below its canopy. Similar results In some cases, perennial invasive species were also found for other invasive species, with high plant uptake rates, such as musk including giant reed (Arundo donax L.), tree- thistle (Carduus nutans L.), can cause long- of-heaven (Ailanthus altissima (Mill.) term declines in soil nitrogen availability. Swingle), and Himalaya blackberry (Rubus Because its seedlings can tolerate low armeniacus Focke), which all dramatically nitrogen levels it creates an environment reduce native species richness and diversity that favors its own establishment (Wardle et in their understory compared to native al., 1994). Th is and other cases illustrate woody communities. the potential for some invasive plants to impact nutrient cycling to their advantage in a positive plant–soil feedback cycle Soil nutrients (Drenovsky and Batten, 2007). Although phosphorus is important in plant Young et al. (1998) showed that fertilizing − growth and performance, nitrogen rangelands with either nitrate (NO3 ) or + availability plays a greater role in deter- ammonium (NH4 ), enhanced seedling mining species dominance in most emergence and establishment of the invasive communities, including mesic heathland annual grass medusahead (Taeniatherum (Aerts and Berendse, 1988), abandoned caput-medusae (L.) Nevski). However, cropland (Wilson and Gerry, 1995), shrub- Monaco et al. (2003) did not fi nd a signifi cant steppe rangeland in North America (Paschke diff erence in either seedling growth or et al., 2000) and Australia (Snyman, 2002), nitrogen allocation between invasive annual arid shrublands (Milberg et al., 1999; grasses and native perennial grasses to the − + Monaco et al., 2003), and wetlands (Daehler, two forms of nitrogen (NO3 and NH4 ). 2003). In the western USA, studies have Th us, the form of nitrogen may give a shown that the addition of mineral forms of competitive advantage to either native or nitrogen to disturbed rangelands increases invasive species, depending on the specifi c the relative abundance of the invasive response of the species within the com- annual grass downy brome (Bromus tectorum munity. L.) compared with native perennial species Invasive nitrogen-fi xing species, mainly (Paschke et al., 2000). Th e negative eff ect of members of the Fabaceae, can alter soil nitrogen addition to grassland and rangeland chemistry, soil enzyme activity, and possibly communities is exacerbated by altered change the coupling of nutrient cycles disturbance regimes caused by human within the ecosystem (Caldwell, 2006). It activities (Daehler, 2003). Typically, nutrient has been proposed that natural nutrient enrichment shifts species composition to enrichment by nitrogen-fi xing shrubs can fewer relatively fast growing species, promote establishment of other invasive primarily annual grasses and robust species, particularly annual species, whose perennial species (Alpert et al., 2000). propagules are present (Alpert et al., 2000). Reducing Invasive Plant Performance: a Precursor to Restoration 157

Water fl ow diversions) have altered the stream fl ow characteristics of many river systems, In arid and semi-arid ecosystems, water is reducing water availability, and promoting the most important resource driving eco- the spread of the more drought-tolerant system processes and community com- invasive shrub saltcedar (Tamarix ramosis- position (Chambers et al., 2007). In general, sima Ledeb.) (Cleverly et al., 1997). increased water addition to mesic and xeric Th e timing of water use can also diff er plant communities, primarily through between invasive and native species. In arid precipitation, will increase susceptibility to environments that have few native annual invasion (Maron and Marler, 2008). In these grasses, non-native winter annual species same communities, drought stress generally such as downy brome, medusahead, and red limits invasibility, in part, due to the lower brome (Bromus rubens L.) can complete their drought tolerance of certain invasive species life cycle during short periods when water compared to many native species (Alpert et availability is high, prior to the dry season al., 2000; Daehler, 2003), which have evolved (Alpert et al., 2000). In more mesic environ- in that ecosystem. For example, diff use ments, non-native winter annual grasses knapweed (Centaurea diff usa Lam.) and can extract surface water early in the yellow starthistle, like many other invasive growing season, before many native forbs or thistles, produce roots that penetrate deep perennial grasses are active (Young et al., into the soil profi le, allowing them to utilize 2009). In comparison, late-season, deep- late-season, deep-soil moisture (Kiemnec et rooted, invasive forbs utilize late-season al., 2003; Enloe et al., 2004). In cool and wet moisture (Young et al., 2010a, b). Th is conditions, root penetration into the deep combination of early season, shallow, water soil profi le was greater for diff use knap- use by non-native winter annual grasses and weed compared to the native perennial late season, water use by deep-taprooted grass bluebunch wheatgrass (Pseudoroegneria perennials can explain why grassland spicata (Pursh) A. Löve), but under warm communities in California, USA, resist and dry conditions, there was no diff erence establishment of native species, which tend between the species (Kiemnec et al., 2003). to largely extract soil water in the middle of Under these drier conditions, the perennial the growing season (Kulmatiski et al., 2006). grass was more competitive for both Invasive species, such as yellow star- nutrients and water, as diff use knapweed thistle, can create locally dry conditions and was unable to utilize the late season maintain a signifi cantly drier soil profi le resources. than either an annual or perennial grass Some invasive species are more drought- community (DiTomaso et al., 2003; Enloe et tolerant than associated native species al., 2004; Gerlach, 2004). When infestations (Alpert et al., 2000). For example, the are dense, soil moisture in the following invasion success of buff elgrass (Pennisetum season does not recharge after subnormal or ciliare (L.) Link) is partially due to its ability even normal winter and spring precipitation to emerge following relatively low pre- (DiTomaso et al., 2003). cipitation levels under desert conditions (Ward et al., 2006). In Montana, spotted knapweed (Centaurea stoebe L. ssp. Reproduction and dispersal micranthos (Gugler) Hayek) maintained greater water potentials despite greater Pollination, seed production, and propagule transpiration compared to established dispersal play a fundamental role in the perennial grasses (Hill et al., 2006). Th is maintenance of natural communities and in adaptation to drought-prone areas was the potential invasion of non-native plant attributed to its deep root system, which species (Traveset and Richardson, 2006). provided access to more consistent deep-soil Short generation times, from germination moisture. In riparian areas in southwestern to reproduction, and high seed production USA, human activities (i.e. dam building, are characteristics of many invasive species 158 J.M. DiTomaso and J.N. Barney

that accomplish rapid dispersal and the impacted the San Francisco Bay estuary, buildup of large seed banks (Bryson and USA, as the hybrid is more vigorous than Carter, 2004; Rejmánek, 2011). Ideally, the either parent (Anttila et al., 1998; Ayres et mating system of a successful invader would al., 2004). include the ability to self-pollinate when infestations are small, but later, when Seeds and other propagules populations are large and pollinators are not limited, to increase genetic diversity and Th e seed banks of most perennial herbaceous polymorphism by outcrossing (Rejmánek, species, especially grasses, are typically small 2011). due to variable seed production and short- lived seeds (Chambers et al., 2007). By comparison, large long-lived seed banks can Pollination account for the persistence of many invasive Unlike many native species that have species (Alpert et al., 2000; Bryson and co-evolved with a suite of both specialist and Carter, 2004; Krueger-Mangold et al., 2006). generalist pollinators, invasive species Even species that reproduce by vegetative generally attract only pollination generalists means, such as giant reed, can develop in their introduced range (Bartomeus et al., rhizomes that remain viable for at least 16 2008), particularly those favored and weeks after isolation from the parent plant introduced by humans (e.g. honeybees). Such (Boose and Holt, 1999; Decruyenaere and insects can displace specialist pollinators Holt, 2001). However, long-lived seed banks and can often have negative eff ects on are not always the case with invasive species native species recruitment (Traveset and and often depend on the specifi c habitat. Richardson, 2006). In addition, invasive Many invasive woody or perennial grass species could competitively aff ect pollination species of riparian areas (e.g. Tamarix spp. of desirable plants, including wind-pol- and Cortaderia spp.) have small seeds with linated (through reduction in pollen quality) short-lived seed banks (DiTomaso, 1998; and animal-pollinated species (through Drewitz and DiTomaso, 2004). Th is is an reduction in pollen quantity and quality) adaptation to broad dispersal and rapid (Brown and Mitchell, 2001). Th is could occur germination on very wet mineral substrates through pollen swamping, which can reduce (Rejmánek, 2011). seed set through stigma or stylar clogging. Examples of pollinator competition include Propagule pressure the invasive purple loosestrife (Lythrum salicaria L.), which was shown to sub- Propagule pressure from either seeds or stantially reduce the seed set of the native reproductive vegetative fragments is congener Lythrum alatum Pursh (Brown and infl uenced by many characteristics, Mitchell, 2001), and the invasive police- including dispersal agents, the degree of man’s helmet (Impatiens glandulifera Royle), habitat fragmentation, and human activities which halved the seed set of the native (Alpert et al., 2000). Th e amount of Stachys palustris L. (Chittka and Schürkens, propagule pressure in an area is more 2001). important than biological characteristics When invasive species populate areas among species that account for invasive with congeneric native species, there is potential, such as adaptation to disturbance the possibility for hybridization. These and effi cient use of resources (Rejmánek, hybrids can dilute the native species gene 2011). For example, species that are highly pool and potentially lead to more vigorous competitive and widely invasive through a invasive hybrids (Brown and Mitchell, variety of characteristics may not become 2001). For example, hybrid vigor between established due to limitations in dispersal of the native cordgrass (Spartina foliosa propagules, but less invasive species may Trin.) and the invasive smooth cordgrass dominate an area if their propagule pressure (Spartina alternifl ora Loisel.) has dramatically is extremely high (Lonsdale, 1999). Once Reducing Invasive Plant Performance: a Precursor to Restoration 159

invasive species become well established and distance. Many native species have tight dominant, the propagule pool of desirable relationships with specifi c frugivore species may be much lower proportionally dispersers and are not generally known to be compared to the invasive, thus compromising dispersed long-distance by a single animal site rehabilitation eff orts (Krueger-Mangold vector (Meisenburg and Fox, 2002). In et al., 2006). contrast, invasive species typically attract more generalist vertebrate dispersers. Among the animal seed vectors, birds are Propagule dispersal the most important vertebrate responsible Th e seeds of plants, including invasive for long-distance dispersal of invasive species, are primarily dispersed long species (Rejmánek, 2011). distances by humans, animals, water, and Th ere are two mechanisms by which wind. Plants have many diff erent adaptations animals disperse fruits or seeds (Meisenburg or preadaptations for dispersal by these and Fox, 2002). Th e fi rst is through vectors (Rejmánek, 2011). Overall, however, ectozoochoric fruit or seed, which attach to invasive species have a seed dispersal animals with awns, hooks, barbs, or sticky advantage over many native species secretions. Th e second is by endozoochoric (Daehler, 2003). One reason for this fruit, which provide a food source for advantage is because human dispersal, animals. For invasive species that rely on either intentional or accidental, is more endozoochory for their dispersal, it is nearly common in invasive plants than in native always assured that a suitable vertebrate species, and is often the most signifi cant disperser will be available (Meisenburg and driver of invasions (Bryson and Carter, Fox, 2002). Fruits can possess an edible 2004; Richardson and Pyšek, 2006). For mesocarp or endosperm, or may contain an example, in California, USA, 63% of the edible large elaiosome or a fl eshy aril that most invasive 200 species were introduced surrounds the seed (Meisenburg and Fox, and spread intentionally as ornamentals, or 2002; Rejmánek, 2011). However, not all for other purposes such as erosion control, animals that eat fruit disperse the seed. livestock forage, or agricultural products Many animals can actually be seed predators (Bossard et al., 2006). Additionally, in a that destroy seed viability. In Florida, animal global evaluation, non-native species ingestion of endozoochoric fruits was the increased proportionally in nature reserves most common method of dispersal, repre- with higher number of visitors (Lonsdale, senting half of the invasive species evaluated 1999). Dispersal of invasive plant seeds and (Meisenburg and Fox, 2002). propagules can occur via many routes, Th e dominance of an animal-dispersed, including direct movement through planting invasive species can also impact the animal for ornamental, soil erosion, shading, forage, community within an area. For example, or crop use, or indirect movement as invasive gorse (Ulex europaeus L.) replaced contaminants of soil, vehicles, equipment, New Zealand plant communities formerly or clothing. dominated by the native kanuka (Kunzea In addition to human dispersal, seed ericoides (A. Rich.) Joy Th omps.). Th is shift dispersal by vertebrate animals is responsible in dominance has also changed the for the success of many invaders into both proportion of seed-dispersing mammals and disturbed and undisturbed habitats birds in the invaded region (Williams and (Rejmánek, 2011). Mammals, including Karl, 2002). both livestock and wildlife, can transport Water and wind dispersal can greatly seeds or fruit on their fur, by hoarding seeds impact the speed and opportunity for in caches, or following ingestion (Davies and invasion of non-native species. Buoyant Sheley, 2007). Fish and even insects, seeds, particularly those with winged particularly ants, can also disperse some appendages (e.g. Rumex spp., Sagittaria spp., species with edible seeds or fruit, but these Sesbania punicea (Cab.) Benth.), have the vectors typically only move seeds a short potential to be moved long distances by 160 J.M. DiTomaso and J.N. Barney

water (Rejmánek, 2011). Seed or fruit Disturbance movement in the atmosphere by wind is Disturbance is often cited as an important facilitated by pappi, or plumed or winged precursor for invasion of an ecosystem appendages. For example, cogongrass (Symstad, 2000). While invasions can take (Imperata cylindrica (L.) Beauv.) seeds can be place without disturbance, events that kill or dispersed by wind at distances up to 25 km damage extant organisms, or reduce their (Bryson and Carter, 2004). Wind can also biomass or performance, can strongly aff ect disperse plants through a tumbleweed habitat invasibility (Alpert et al., 2000). Th ere action, where the entire aboveground shoot are several types of natural and human- (e.g. Salsola tragus L., Bassia scoparius (L.) A.J. related disturbances that can aff ect Scott) or infl orescence (e.g. Panicum capillare vegetation. Th ese include grazing, fi re, fl ood, L.) breaks off from the stem and is dispersed and drought, which can occur either regularly by tumbling (Davies and Sheley, 2007). Although many invasive plants have well- or episodically at high or low intensity and at developed, long-distance dispersal mech- a landscape level or more locally (James et al., anisms, a number of species shed their seeds 2010). Such events reduce the level of in close proximity to their parent plant. competition among species and often lower Th ese species either lack a facilitated competitive stress by increasing resource dispersal feature or disperse seeds through a availability. An increase in resource mechanism that only moves seeds a short availability can occur when plant uptake rates distance (Cousens and Mortimer, 1995). For decline due to suppression of resident example, some fruit can disperse seeds a few vegetation, or when resource supply increases meters through explosive dehiscence of the at a rate faster than the resident vegetation fruit. Ballistic dispersal is often important can sequester it, which often follows burning in local expansion, but does not assist in (Davis et al., 2000). Th is can open niches long-distance propagule movement at the within eco systems and increase safe site landscape or regional scale. availability for the successful dispersal and establishment of invasive species. In many cases, disturbance is a Invasibility of ecosystems prerequisite for invasion (Smith and Knapp, 2001). Disturbance associated with human Invasion success not only depends on the activities generally gives invaders the largest processes directly infl uencing invasive performance advantage, particularly when species, but also the processes operating the disturbance intensity and intervals are within the plant community (Barney and very diff erent from the natural disturbance Whitlow, 2008). Invasibility is defi ned as the regime (Daehler, 2003). In general, an intrinsic susceptibility of an area to invasion increase in disturbance frequency and (Richardson, 2001), and depends on a intensity tends to favor invasive species number of factors, including identity and (James et al., 2010). As disturbance intensity characteristics of the invading species, increases, nutrient-cycling rates increase, evolutionary history, species diversity, com- and the ability of resident vegetation to munity structure, the strength of inter- sequester nutrients decreases (James et al., actions between species, propagule pressure, 2010). Th is results in enhanced nutrient assemblages of predators or pathogens, availability. Some habitats with a long disturbance and stress, resource availability, history of human disturbance may be less and growing conditions (Lonsdale, 1999; invasible because the resident vegetation Alpert et al., 2000; Daehler, 2003; Barney would have already been selected to perform and Whitlow, 2008). In particular, well under a frequent disturbance regime disturbance, habitat and species diversity, (Alpert et al., 2000). resource availability and open niches, and While invasions of non-native species are propagule pressure are critical aspects that most common in agricultural or urban sites, contribute to the invasibility of an ecosystem. where human disturbance is regular, they Reducing Invasive Plant Performance: a Precursor to Restoration 161

are least common in natural areas with low conditions of the specifi c habitat, and the rates of disturbance, such as deserts and propagule pressure is adequate to allow savannas (Lonsdale, 1999; Daehler, 2003). establishment (Tilman, 1997; Alpert et al., When disturbance frequency is maintained 2000; Davis et al., 2000; Rejmánek et al., at a low rate, species, particularly native 2005). Th us, even in areas considered ones, which tolerate stressful conditions (i.e. relatively invasion-resistant, introduction of limiting resources), will continue to an appropriately matched species at the dominate (Krueger-Mangold et al., 2006). right time can lead to invasion. For example, Non-native species that are effi cient com- the Mojave and Sonoran deserts are two petitors for limiting resources in undisturbed highly stressful, low resource environments plant communities will often be the most in North America. Although these eco- successful invaders and the worst weeds of systems have few non-native plants, they natural ecosystems (Rejmánek, 2011). have recently been invaded by African mustard (Brassica tournefortii Gouan) and buff elgrass (Rejmánek et al., 2005). Habitat and species diversity Successful invasion occurs when the fi tness A number of studies have shown that more diff erence between the resident vegetation diverse plant communities resist invasions and the non-native species favors dominance and use resources more completely (Case, of the latter or when niche diff erences (i.e. 1990; Tilman, 1997; Maron and Marler, resource acquisition characteristics) are 2008). However, other studies demonstrate large enough to allow the latter to establish that communities richer in native species despite lower population fi tness (Rejmánek, have more invasive species (Stohlgren et al., 2011). Conversely, invasion resistance 1998, 2003; Lonsdale, 1999). In these cases, occurs when the fi tness of the resident non-native species can more easily invade vegetation is greater than the non-native areas with high resource availability, which species or when niche overlap is large are often riparian areas or islands, and more between the resident vegetation and the available resources also promotes increased latter. In many cases, there is little diff erence native plant diversity (Stohlgren et al., 1998; in fi tness between the invasive and the Alpert et al., 2000; Richardson and Pyšek, resident vegetation and niche overlap is 2006). From these two concepts, it can be small. In these situations, the invader concluded that either unexploited available coexists with the resident vegetation and resources or near complete exploitation of does not cause detectable displacement of resources can occur in both low and high the desired species (Tilman, 1997; Rejmánek, diversity ecosystems, such that invasibility 2011). is not necessarily directly related to species Niche occupation is generally related to richness (Bulleri et al., 2008). More functional diversity within a site. Higher importantly, habitat diversity or hetero- functional diversity, often associated with geneity within a large scale ecosystem may species richness, can reduce resource provide more opportunity for suitable availability and provide greater resistance to habitat for both native and invasive species invasion compared to areas with low (Stohlgren et al., 2003; Richardson and functional diversity (Symstad, 2000; Dukes, Pyšek, 2006; Bulleri et al., 2008). Although 2001). However, even in areas with high loss of species diversity alone may not aff ect functional diversity, fl uctuations in resource community invasibility, communities with availability can occur from year to year, or fewer species are often more susceptible to within a year. Th ese episodic fl uctuations or invasion (Dukes, 2001). events can dramatically increase resources and greatly enhance invasion success (Davis et al., 2000; Rejmánek et al., 2005; Chambers Resource availability and open niches et al., 2007). Invasion occurs when the growth form and Although relatively uncommon, the traits of an invader are favored by the local existence of one species in an ecosystem, 162 J.M. DiTomaso and J.N. Barney

native or invasive, can facilitate the success its seed bank increases in proportion to that of other invasive species by creating an of the native or desirable vegetation. Th is unused niche. For example, the estab- leads to further decline in the resident lishment of invasive ruderal annual species species habitat as a result of seed swamping in California coastal prairies, USA, is by the invasive species (Daehler, 2003). enhanced by the native nitrogen-fi xing shrub Lupinus arboreus Sims (Maron and Connors, 1996). As another more specifi c Managing plant performance and interaction, the establishment of invasive ecosystem processes dandelion (Taraxacum offi cinale F.H. Wigg.) is facilitated by the presence of cushions A tremendous body of work exists formed by the native plant Azorella monantha determining both species characteristics that Clos in the alpine zone of the Chilean Andes contribute to successful invasion, and habitat (Cavieres et al., 2008). Th ese cushions characteristics that facilitate susceptibility to increased nutrient availability, which was invasion. However, few studies have bridged speculated to benefi t dandelion the chasm with empirical studies on the establishment. interactions between species and habitats (Barney and Whitlow, 2008). Heger and Trepl (2003) proposed a conceptual model of Propagule pressure the ‘key-lock approach’ that considers the While the invasibility of an environment by features of both the invasive species and the a non-native species depends on its habitat into a relational model. Th ey contend characteristics and the susceptibility of the that species with specifi c characteristics environment to invasion, the invasion (keys) fi t into specifi c habitats (locks) that process itself depends on conditions of end in successful invasion. Hence, not all resource enrichment or release and the habitats are equitably invasible, and there timely availability of invading propagules are no ‘invasive characters.’ Rather, a specifi c (Lonsdale, 1999; Davis et al., 2000). combination is required to facilitate suc- Propagule pressure is often a strong cessful invasion. determinant of habitat invasibility (Von Th e future of identifying mechanisms Holle and Simberloff , 2005). In plant along the invasion pathway will be founded communities that are considered less in integrated studies of species, habitats, resistant to invasion, fewer propagules are and the role of propagule pressure needed for a newly introduced species to (Richardson and Pyšek, 2006; Barney and establish and invasion rates are relatively Whitlow, 2008). Understanding the complex fast. In comparison, habitats with high interactions of species traits contributing to invasion resistance require much higher invasion, and the ‘weakness’ of the receiving rates of propagule supply to overwhelm the habitat can directly infl uence our ability to ecological resistance of the site (Rejmánek et successfully manage the invasion. According al., 2005; Von Holle and Simberloff , 2005; to the principles of ecologically-based Richardson and Pyšek, 2006). For example, invasive plant management (EBIPM), there the resident species richness of experimental are three general causes of succession: (i) riparian plots in California, USA, did not site availability; (ii) species availability; and have as great an eff ect on invader (iii) relative species performance (James et establishment as did the number of invader al., 2010; Sheley et al., 2010). Th e underlying propagules added to the plots (Levine, cause of invasion must fi rst be identifi ed to 2000). In another study in South Africa, the allow land managers to appropriately intensity of propagule pressure, rather than manipulate one or more of these factors to any environmental factor, best explained guide ecosystem change and modify or the invasion of three woody invasive species repair the ecological processes that favor a (Rouget and Richardson, 2003). Once the more desired community or vegetation invader population establishes and matures, trajectory. Reducing Invasive Plant Performance: a Precursor to Restoration 163

As previously discussed, invasibility of a In general, small scale, less frequent, particular environment and invasiveness of and low intensity disturbances tend to individual non-native species can depend favor establishment and growth of desirable on specifi c disturbance regimes, propagule species (James et al., 2010), while large characteristics, resource availabilities, and scale, more frequent and higher intensity community composition that favor the disturbances favor invasion by ruderal performance of the invader. However, these species (Alpert et al., 2000). Because nearly same processes can be manipulated and all native communities have evolved under controlled to infl uence the performance of a particular historic disturbance regime, both invasive and desired species. Th rough restoring a more natural disturbance these activities it can be possible to direct regime will generally favor native plants successional transitions that promote a (Daehler, 2003). Th is is not always the case, more desired plant community (James et however, as there are other disturbance al., 2010; Sheley et al., 2010). A critical regimes that can also favor native over element in this process is to initially identify specifi c invasive species. In contrast, the abiotic (i.e. disturbance factors, climate, suppressing natural disturbance regimes resource availability) or biotic (i.e. can also increase invasion (Alpert et al., competing vegetation) fi lters or stressors 2000). It is important to recognize that that contribute to invasion. With this even when natural disturbance regimes are understanding it is then possible to established, a large number of invasive manipulate these through modifi cation in species within or near the community will disturbance intervals and intensities, increase the probability that some of these resource availability, hydrological regimes, species will tolerate the historic disturbance management strategies, and propagule regime (Daehler, 2003). pressure, which can also be tied to Disturbances that can alter the com- revegetation eff orts (Krueger-Mangold et petitive relationships between native and al., 2006). invasive species include fi re, grazing, mechanical damage (e.g. mowing or tillage), and alterations in hydrological regimes. Disturbance regimes Diff erent types of disturbance can have Many ecosystems can be managed to favor a diff erent eff ects in the same habitat (Alpert particular set of species or community et al., 2000). For example, maintaining composition by manipulating the type, historical fi re regimes often favors native interval, and intensity of disturbance. While over invasive species (Daehler, 2003). In certain disturbance regimes often lead to the Great Basin region of western USA, plant invasions, they can conversely be burning increased the availability of soil intentionally infl uenced to direct a com- water and nitrate. While this often leads to munity toward a more preferred vegetation an increase in downy brome populations state by providing appropriate germination when native perennial species are in low conditions and sites, maintaining recruit- abundance, the opposite occurred when ment timeframes for resident species, there was a high population of native reducing invasive species performance, and perennial species. In this case, the native maximizing growth and reproductive perennial species responded positively to performance of native or desired species the increased availability of water and (Daehler, 2003; Krueger-Mangold et al., nitrate and limited the increase in downy 2006; James et al., 2010; Sheley et al., 2010). brome (Chambers et al., 2007). As another In addition, a particular disturbance regime example, altering the historic fl ood regimes can be used to suppress or manage an along rivers in southwestern USA is invasive species, and these types of strategies thought to have led to the establishment have been commonly employed by land and spread of invasive saltcedar (Sher et al., managers (DiTomaso, 2000; DiTomaso et 2000; Bay and Sher, 2008). By restoring the al., 2010). natural fl ood regime, it may be possible to 164 J.M. DiTomaso and J.N. Barney

reestablish the native woody vegetation Minimizing unnecessary disturbance, using while suppress further saltcedar encroach- weed-free hay and uncontaminated soil in ment. construction activities, and revegetating with high purity seed can also help prevent unintended introductions. Physical barriers Propagule pressure or properly timed grazing can reduce the risk Recruitment limitations are a critical aspect of moving invasive species propagules from of long-term ecosystem management one area to another. Th is is particularly true (Tilman, 1997). A particular species, for invasive species with seeds possessing whether invasive or desirable, will have a awns, hooks, or barbs that facilitate their greater chance of successfully establishing dispersal (Davies and Sheley, 2007). Although and colonizing a site when its propagule manipulation in grazing timing and intensity numbers are greater than competing species is more practical and easily accomplished occupying a similar niche (Von Holle and with livestock than wildlife, establishing Simberloff , 2005). Th us, shifting the relative barriers, such as fences or enclosures, can abundance towards desirable species will also infl uence wildlife grazing. For invasive generally require manipulating and plants that are wind dispersed, it may be managing the availability, frequency, and possible to promote taller vegetation that not abundance of desirable species propagules only reduces wind velocities but also relative to that of invaders over time (Davies physically intercepts seeds within these sites and Sheley, 2007; James et al., 2010; Sheley to form a barrier to long-distance dispersal to et al., 2010). Manipulating propagules can uninfested area (Davies and Sheley, 2007). occur by preventing seed recruitment and Th e dispersal of seeds through tumbling dispersal of invasive plants from adjacent action in wind can be limited with fencing areas or by enhancing desirable plant along the perimeter of the restored area recruitment, dispersal, or abundance of (Baker et al., 2010). desirable species via intentional seed intro- It may not be completely necessary to duction. prevent dispersal and new recruitment of Managing recruitment of invasive plant invasive plants to an area, provided that propagules, primarily seeds, will diff er desirable species disperse and successfully depending on the dispersal vector responsible germinate fi rst (James et al., 2010). Early for introduction. In many areas, introduction germination timing and establishment of of new propagules is through specifi c desirable species can greatly favor their corridors, such as roads, trails, streams, and success or, conversely, delayed dispersal of rivers, and winter-feeding areas for livestock invasive species can limit their success and and wildlife (Krueger-Mangold et al., 2006). promote plant community change to a more Th ese source sites should be identifi ed and desired vegetation state. prioritized for strategic management to reduce the risk of reintroduction or intro- Manipulating soil resources ducing new invasive species (Davies and Sheley, 2007). For example, water corridors Th e competitive relationships and relative can serve as very eff ective dispersal vectors. performance of desirable versus invasive As such, it may be critical to prevent species can shift dramatically under a introduction of invasive species seeds or particular set of resource conditions. vegetative propagules into water by Typically, invasive species dominate over prioritizing the management of their popu- native and other desirable non-native lations near water’s edge. species when resource availability, including Preventing or limiting movement of light, nutrients, and water, are high (Daehler, invasive species can be achieved by managing 2003). Presumably, a return to the natural propagules along the primary vectors or resource levels would favor native species corridors by several control options, but most that had evolved under historic conditions often by mechanical or chemical methods. compared to non-native invasive species. Reducing Invasive Plant Performance: a Precursor to Restoration 165

Th is assumes there is an adequate propagule Soil nutrient availability can also be pressure of desirable species and that manipulated to favor certain desirable disturbance levels are not too dissimilar species. Strategies to reduce nitrogen from natural conditions. availability can give an advantage to native In many natural environments, historical species and reduce the competitive ability of resource levels were low, and native species invasive species, as well as prevent invasion often evolved under relatively high stress and establishment of new invasive conditions. Increased disturbance and propagules (Alpert et al., 2000). Th ere are resource availability has favored invasive several ways of manipulating nitrogen over native species (James et al., 2010). availability in invaded ecosystems. In some Environmental manipulations that lower cases nitrogen can be added to the system to resource availability, particularly nitrogen, enhance native or desirable plants. For and increase stress can shift the community example, increasing the proportion of dynamics of an area in favor of native species nitrogen as ammonium or urea in fertilizer (Alpert et al., 2000; Krueger-Mangold et al., mixtures can restrict germination or growth 2006). In addition to manipulating total of ammonium- or urea-sensitive weeds and resources available to the ecosystem, enhance native species (DiTomaso, 1995; conditions that favor native or other desirable Monaco et al., 2003). However, this requires plants may also be achieved by altering the a thorough understanding of the response of intensity of the stress (i.e. heat of fi re and both desirable and invasive species to various level of grazing pressure) or the timing of forms of nitrogen. Alternatively, fertilizer resource availability (Daehler, 2003). For applications during the time when nutrient example, moderate prolonged resource stress uptake is maximal in desirable plant roots favors desired species over invasive species can increase their competitive ability against compared with short-duration and more invasive species that utilize nutrients later in intense stress (James et al., 2010). the year (DiTomaso, 1995). Another method Although possible, manipulating resource to selectively manipulate the timing of availability is often diffi cult to accomplish nutrient use in desired species is through and generally impractical at the landscape seed priming. Seed priming is a technique level. Reducing light availability can be where seeds are sown after they are partially accomplished by revegetating with taller hydrated and germination has begun, but herbaceous perennial or woody desirable emergence of the radicle has not yet occurred species. Invasive species, such as yellow (Baskin and Baskin, 1998). Th is strategy can starthistle and jubatagrass (Cortaderia jubata ensure resource preemption and improve the (Lem.) Stapf.), are suppressed by reduced coupling of nutrient cycling between plants light and occur only infrequently under the and soils (Krueger-Mangold et al., 2006). canopy of taller vegetation (DiTomaso et al., Other nutrient manipulations are 2003; Stanton and DiTomaso, 2004). In designed to reduce soil nitrogen availability. contrast, the performance of the invasive For example, fi res promoted by invasive perennial vine Cape ivy (Delairea odorata annual grasses can temporarily increase Lem.) is very poor under high light available nitrogen but will decrease total conditions and can be managed by increasing nitrogen in the system through volatilization light intensity (Robison et al., 2011). (Daehler, 2003). Over time and through Similarly, the availability and fl ow of water several fi re cycles in areas without a can be manipulated in some riparian areas preponderance of nitrogen-fi xers, a gradual where dams or the fl ow of rivers and streams reduction in nitrogen could promote a shift are controlled, and it has been proposed that away from annual grass dominance back manipulating hydrological regimes may be toward a late seral native community. Other eff ective for enhancing the establishment of more immediate methods of reducing native woody vegetation while managing available soil nitrogen can include carbon invasive saltcedar (Sher et al., 2000; Bay and amendments to the soil in the form of Sher, 2008). sawdust, mulch, or sucrose (Corbin and 166 J.M. DiTomaso and J.N. Barney

D’Antonio, 2004; White and Holt, 2005). be used to damage annual invasive species Th ese carbon sources reduce mineral nitro- and promote the rapid establishment of gen availability by promoting immobilization native plants in revegetation projects (Sheley by soil microbial populations, which can et al., 2010). Other mechanical techniques, eff ectively decrease invasive plant per- such as mowing, can be used in a timely formance and favor native perennial species manner to facilitate the growth of native (Perry et al., 2010). For example, by species and control non-native species. For experimentally reducing nitrogen with example, annual mowing of the invasive sucrose addition over a 4-year period in perennial tall oatgrass (Arrhenatherum elatius Colorado, USA, plots dominated by downy (L.) P. Beauv ex J. Presl & C. Presl) during its brome were converted to a community that fl owering phase, in late spring or early closely resembled the natural late-seral, summer, reduced its cover and biomass while shortgrass steppe vegetation (Paschke et al., increasing fl owering and growth of native 2000). perennial grasses in an Oregon, USA, prairie (Wilson and Clark, 2001). Under some situations, restrictions in Management strategies for reducing grazing may be necessary to allow desirable species performance of invasive plants vegetation to recover following management of invasive species, although this is not Management strategies should, ideally, always possible in areas occupied by target the most vulnerable or susceptible livestock. In contrast, restoration of historic stages in the life cycle of the invasive plant grazing regimes can often promote native while minimizing the negative eff ects on species diversity (Daehler, 2003). In native or desirable species. Th us, manage- rangelands, it is possible to adjust the ment strategies that directly infl uence key intensity, duration, and timing of livestock ecological processes, including altering grazing to favor particular species (Larson disturbance regimes, manipulating resource and Kiemnec, 2005). In addition, certain availability, limiting invasive seed dispersal classes of livestock preferentially graze on and propagule pressure, and enhancing specifi c plant life-forms or species. Sheep, desirable species germination, seedling for example, usually prefer forbs over grasses establishment, survivorship, growth, and and shrubs, cattle generally graze on grasses, reproduction (Table 9.1). Th e timing and and goats often favor shrubs (Popay and techniques used to achieve this objective are Field, 1996; Krueger-Mangold et al., 2006). critical to successful invasive plant manage- In sensitive riparian areas or grasslands, for ment. For example, rapid germination and example, goats have been used to graze on establishment of the invasive perennial spiny brush species, including wild species buff elgrass and crimson fountain- blackberries (Rubus spp.), sweet roses (Rosa grass (Pennisetum setaceum (Forsk.) Chiov.) spp.), and matagouri (Discaria toumatou occur soon after precipitation events in Raoul) without damaging the desired native many arid environments of the world. vegetation or forage species (Dellow et al., Monitoring areas of seedling development 1987; Holgate and Weir, 1987; Cossens et and initiating management programs within al., 1989). By manipulating timing and a short period following rainfall events can intensity of grazing, defoliation will be an eff ective strategy for preventing primarily target invasive species to give a subsequent establishment of large infest- competitive advantage to other plants in the ations (Ward et al., 2006; Rahlao et al., community (Krueger-Mangold et al., 2006). 2010). As an example, when medusahead was While soil cultivation is a disturbance that grazed by sheep at the ‘boot’ stage, just prior generally increases the dominance of many to exposure of the infl orescences, its cover invasive plants in grasslands and other was reduced by over 86%, while native forb natural areas (Stromberg and Griffi n, 1996), cover signifi cantly increased (DiTomaso et timely light disking and soil imprinting can al., 2008). Reducing Invasive Plant Performance: a Precursor to Restoration 167 re or hydrology regime (i.e. occurrence, occurrence, regime (i.e. re or hydrology fi owering of invasive species of invasive owering fl c mechanical control techniques to favor desirable desirable c mechanical control techniques to favor fi mechanical or hydrological disturbance) mechanical or hydrological timing, and intensity) frequency, timing, and intensity duration, or tillage) mowing species (e.g. species to increase stress on invasive recruitment of reseeding) seed production prevent early germination priming and establishment to achieve site or to restoration methods in areas close proximity along transport corridors Eliminate cause of disturbance (i.e. grazing, burning, grazing, Eliminate cause of disturbance (i.e. Restore desirable frequency, type, through livestock practices Modify grazing Speci timing, or technology herbicides or application rate, Selective Prevent species (e.g. desirable for Increase propagule numbers species seed production or recruitment of invasive Prevent herbicides to or selective timely grazing Use of mowing, or seed transplanting species, Timely seeding of desirable species with long-distance dispersal Management of invasive

• • • • • Management strategy • • • • desirable species desirable desirable species desirable desired plants species desirable species invasive invasive species propagules invasive dispersal of desirable facilitate species Reduced competitive ability of Reduced competitive ability of Reduced competitive Changes in nutrient cycling to favor Timely susceptibility of ecosystem to that Changes in animal communities Physical damage to desirable species damage to desirable Physical • Factors leading to invasiveness Factors ood control) fl seed production in invasive species seed production in invasive production of invasive species production of invasive congeners plant species or diluted native invasive gene pool propagules wind, animals) and equipment, soil, clothing, vehicles erosion or crop, ornamental, forage use, or nutrient, water) or provides pulse in or provides nutrient, water) invasive use by for resource availability species Pollen swamping of desirable species or of desirable swamping Pollen Hybrid vigor creates more aggressive Hybrid vigor creates more aggressive Long-lived seed banks or vegetative seed banks or vegetative Long-lived water, long-distance dispersal (i.e. Natural livestock, dispersal (i.e. Human facilitated Plant Ecosystem High fecundity of invasive species of invasive High fecundity • Propagule build-up of invasive species of invasive Propagule build-up Increases available resource (i.e. light, resource (i.e. Increases available production occurrence Intensity damage to of physical Higher level Reproduction timing Short times from germination generation to Dispersal Seed or vegetative Seed or vegetative Frequency Frequency Pollination pollinators increase seed Generalist Reduction in propagule production No historical Linkages between ecosystem processes and management strategies. ecosystem processes and management strategies. Linkages between natural regime) natural Disturbance (shift from Table 9.1. Table Process Characteristic Propagule pressure 168 J.M. DiTomaso and J.N. Barney xers or altering xers fi re) or management (i.e. herbicide, mowing, or mowing, herbicide, re) or management (i.e. fi re occurrence tillage) amendments or by removing nitrogen- removing amendments or by fi vegetation penetration compete with invader effectively species mixes conditions ideal establishment potential for maximize species seedling establishment shift successional community accelerate or germination of seeded species and establishment Addition of nitrogen forms that favor desirable species desirable that favor Addition of nitrogen forms through carbon Reduce nutrient availability Timely addition of nutrients to enhance desirable light to reduce canopy Planting desired vegetation with functionally similar species to Revegetation appropriate revegetation niche occupation by Diverse to seasons or years multiple Seeding in phases over control options to manage invasive Timely selective species to Early planting of desirable to late seral grazing to stress (i.e. exposure Timely seeding to follow to facilitate Ensure proper seedbed preparation

• • • • • • • • • • • xers or other xers fi resource year-round or resource year-round seasonally full resource acquisition year- round or seasonally can promote resource provider further invasions High availability of unused of unused High availability Species present do not provide Species present do not provide Invasive nitrogen- Invasive ciency, photosynthetic ciency, fi cient resource capture though physiological mechanisms (i.e. high mechanisms (i.e. physiological use ef water or or nutrient uptake) rates morphological characteristic (i.e. deep roots or climbing habit) fi Ef water Species diversity composition Resource availability Nutrients, light, Community Community Reducing Invasive Plant Performance: a Precursor to Restoration 169

Much like grazing, properly timed Herbicides are widely used tools for prescribed burning can also manage invasive control of invasive plants in many eco- species and select for more desired native or systems. Many herbicides selectively control desirable species. In California, USA, early specifi c groups or even taxa of species, while summer prescribed burning signifi cantly having little eff ect on others (Fig. 9.2). reduced the cover of yellow starthistle, Choosing the proper herbicide, applying it at medusahead (Fig. 9.1), and barb goatgrass the correct rate and at the most appropriate (Aegilops triuncialis L.), and also dramatically timing, and using application technology increased the cover of native perennial that maximizes its eff ectiveness and grasses by at least ten-fold (DiTomaso et al., selectivity can give successful invasive plant 1999, 2001, 2006; Kyser et al., 2008). control and minimize damage to non-target Periodic burning may be an important tool species (Krueger-Mangold et al., 2006). to maintain healthy perennial grass or fi re- tolerant shrub communities even after other control tools have been used to reduce Revegetation manipulations invasive plant populations. Early-seral species are favored by higher Manipulating propagule pressure, disturb- available soil nitrogen and can prevent the ance, and soil resources with management establishment of desirable late-seral native tools is critical to the restoration of invaded species through intense competition during natural ecosystems. However, it is just as the establishment phase. As a means to important to know if and when revegetation manipulate soil nutrient levels, Herron et eff orts are necessary and to understand the al. (2001) showed that soil nitrogen could revegetation methods and techniques that be reduced by planting an ephemeral cover favor successful management of invasive crop (Secale cereale L.) or the mid-seral species and restoration of a desired plant native bottlebrush squirreltail (Elymus community. elymoides (Raf.) Swezey). Th ese species Th e important question of whether or accelerated the successional change from a not it is necessary to actively revegetate weedy plant community dominated by invaded ecosystems takes on greater spotted knapweed toward a more desired importance when monetary resources are plant community. limited. In some situations, not using active

Fig. 9.1. Prescribed burn for control of medusahead. 170 J.M. DiTomaso and J.N. Barney

Fig. 9.2. Aerial application of the herbicide clopyralid for the selective control of yellow starthistle in a California, USA, grassland. management practices and allowing the rangelands (Chambers et al., 2007). Choos- ecosystem to passively recover may be ing the most appropriate species for enough (Holl and Aide, 2010). However, revegetation is among the most critical steps passive restoration often takes several years, to a successful restoration program. Species- and many agencies or land managers feel the rich seed mixes can increase the likelihood pressure to accelerate the process through of survival of some species under varying expensive active revegetation eff orts. For stressful environmental conditions and also example, passive restoration of forest provide more opportunities for occupying systems often requires more than 40 years complementary functional guilds with to recover without revegetation eff orts invaders (Krueger-Mangold et al., 2006). (Jones and Schmitz, 2009). In some cases, Introducing or maintaining high functional removing grazing and reducing the diversity can more fully utilize available frequency of fi re allows recovery of extensive resources, both spatially and temporally. areas of tropical dry forest, but even these Th is will increase the competitive infl uence practices require decades (Janzen, 2002). In of the desired species on the invasive species ecosystems that are heavily invaded, where and provide a greater level of resistance to the level of degradation becomes more reinvasion or dominance of an invasive intense and the native plant seed bank is species (Davies et al., 2007; James et al., reduced, active restoration that includes 2010). revegetation eff orts and stress manipu- In addition to choosing appropriate lations through herbicide use, tillage, species mixes, it is critical to consider the periodic fl ooding, prescribed burning, or relative propagule pressure between the timely strategic grazing are often necessary desired and invasive species within a site. to recover certain ecosystem functions Th is requires both management of the (Alpert et al., 2000; Holl and Aide, 2010). invasive species seed bank (control In most situations, revegetation eff orts strategies) and seeding at high enough rates include establishing propagules of perennial to off set invasive species seed numbers. species, either herbaceous or woody. Th is is When this is not possible, it may be particularly true for riparian, forests, and necessary to transplant seedlings or young woodland areas, but also for grasslands and plants of desired species into restoration Reducing Invasive Plant Performance: a Precursor to Restoration 171

sites (Kulmatiski et al., 2006). In this case, susceptibility to pests and pathogens or transplants must be able to establish root alleviate abiotic stress (Madsen et al., 2009). systems quickly to capture resource before the invasive species. Timing of revegetation eff orts can also Conclusions determine the success of a restoration program. In some cases, seeding should be Reducing invasive plant performance conducted soon after management depends on an understanding of the practices have been applied. Th is ensures characteristics and processes that make a rapid establishment of the desired plant plant invasive, as well as the aspects of an canopy and subsequent suppression of ecosystem that facilitate or enhance the invasive species. In other cases, seeding invasion process. With a foundational should be conducted when predicted knowledge of these concepts, it is possible to weather con ditions are most conducive to employ a number of tools to enhance the successful germination and establishment competitive ability of native and other of the desired species (Kulmatiski et al., desirable species, while minimizing the 2006). Although more expensive, seeding performance of invasive species. Th e various can also be conducted in phases throughout methods of manipulating propagule pres- multiple seasons or years to increase the sure of invasive plants, altering disturbance chances of encountering environmental regimes to favor desirable species, modifying conditions that maximize seedling estab- resource availability to support a more lish ment and survival (Krueger-Mangold et desired plant community, and employing al., 2006). As another alternative, seeding management tools that selectively damage diff erent species over a period of time, or suppress invasive species are critical termed ‘assisted succession,’ can increase aspects to a successful long-term passive or the long-term success of a restoration active restoration program. When active program through assisted succession (Cox revegetation is necessary in a restoration and Anderson, 2004). In this situation, eff ort, there are additional techniques that species known to eff ectively compete with can further enhance the probability of the invasive species are initially seeded. successfully achieving a desired outcome. Th is is followed by subsequent seeding of Th e practices should be incorporated into an other native species that provide added EBIPM program to maximize the probability ecosystem values. For example, estab- of successfully converting degraded land- lishment of crested wheatgrass (Agropyron scapes into healthy functional eco systems. cristatum (L.) Gaertn.) in the Inter- mountain Region of western USA initially decreased downy brome cover, and through References ‘assisted succession’ eventually led to the successful restoration of native sagebrush- Aerts, R. and Berendse, F. (1988) The effect of grassland steppe (Cox and Anderson, increased nutrient availability on vegetation 2004). dynamics in wet heathlands. Vegetation 76, Proper seedbed preparation can also 63–69. facilitate better germination and establish- Alpert, P., Bone, E. and Holzapfel, C. (2000) ment of seeded desirable species. For Invasiveness, invasibility and the role of example, large depressions trapped and environmental stress in the spread of non-native retained more seeds and moisture in arid plants. Perspectives in Plant Ecology, Evolution, environments, resulting in high seedling and Systematics 3, 52–66. Anttila, C.K., Daehler, C.C., Rank, N.E. and Strong, emergence compared with smaller D.R. (1998) Greater male fi tness of a rare depressions (Chambers, 2000). To further invader (Spartina alternifl ora, Poaceae) increase successful germination of desired threatens a common native (Spartina foliosa) species, it is possible to treat the seed (seed with hybridization. American Journal of Botany coating) with products that decrease 85, 1597–1601. 172 J.M. DiTomaso and J.N. Barney

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Jane M. Mangold

Department of Land Resources and Environmental Sciences, Montana State University, USA

Introduction Plant community – Undesired state For sites severely degraded by exotic invasive Site plants, simply controlling the weed to release availability desirable plants from competition may not be adequate. Introducing propagules of desired species through revegetation may be required. Revegetation is a resource-intensive Species Species performance availability venture that often results in less than optimum outcomes. Successional manage- ment (Pickett et al., 1987; Sheley et al., 1996; Krueger-Mangold et al., 2006), in which site availability, species availability, and species performance are manipulated to direct plant Plant community – Desired state communities from an undesir able state to a desirable state, may serve as a useful Fig. 10.1. Successional management can serve framework for assessing site conditions, as an ecological framework for planning and choosing exotic plant control methods and implementing revegetation of weed-infested plant revegetation strategies, and planning communities. In successional management, site follow-up management (Fig. 10.1). Designing availability, species availability, and species revegetation programs that are based on our performance are manipulated to direct plant best understanding of the primarily ecological communities from an undesirable state to a processes responsible for plant community desirable state (adapted from Pickett et al., 1987). dynamics at a given site, may initiate outcomes that more fully meet management objectives. Th is chapter will briefl y address the ecological theory for revegetation. Th en, Ecological Theory Behind the current revegetation strategies, limitations Necessity for Revegetation of those strategies, and new approaches for revegetation will be described in the context Over the past several decades, increased of site availability, species availability, and education and awareness concerning the species performance. impacts of invasive plants has resulted in

© CAB International 2012. Invasive Plant Ecology and Management: 176 Linking Processes to Practice (eds T.A. Monaco and R.L. Sheley) Using Current Technologies and Ecological Knowledge 177

increased eff orts to control them. When (D’Antonio and Meyerson, 2002). One major invasive plants are controlled via herbicides, challenge of revegetating invasive plant- grazing, mowing, biocontrol, or other infested areas is overcoming the abundance methods, open niches are created in the of invasive plant propagules and lack of plant community. Desirable species, released abundance of desired species propagules. from the competitive eff ects of the invasive Even when invasive plants are controlled, plant, often respond to the increase in site they often re-occupy the site due to a sizeable availability, and re-occupy the site (Sheley et seed bank that responds quickly to increased al., 2000). However, in plant communities site availability created by the disturbance of that have been dominated by invasive plants control itself (Hobbs and Huenneke, 1992; for some time, desirable species may be Reinhardt Adams and Galatowitsch, 2008). exceedingly rare or even completely absent Revegetation to address site and species from existing vegetation and the seed bank. availability must be integrated with control If invasive plants are controlled, but eff orts to address species performance with propagules of desirable species are not the goal of hindering performance of weedy present to occupy open niches, invasive species while establishing desirable species plants are likely to reestablish (Laufenberg and promoting their performance in subs- et al., 2005; Mangold et al., 2007a; Reinhardt equent years. Adams and Galatowitsch, 2008). In some cases the same invasive species reestablishes, but in other cases a diff erent, but no less Current Strategies for Revegetating troublesome invasive plant becomes Invasive-Plant Dominated Plant dominant. For example, at two sites in Communities western Montana, USA, the root weevil Cyphocleonus achates (Fahraeus) drastically Site availability decreased spotted knapweed (Centaurea stoebe L.) populations, but the invasive Disturbance is the primary ecological annual grass cheatgrass (Bromus tectorum L.) process that regulates site availability and is comprised 50–90% of the replacement commonly defi ned by the size, severity, and vegetation (Story et al., 2006). patchiness of the disturbance, time intervals Ecological processes such as nutrient between disturbance, and pre-disturbance cycling (Ehrenfeld, 2003; Rodgers et al., history (Pickett et al., 1987). Disturbance 2008), carbon storage (Ehrenfeld, 2003), typically results in an increase in resource and fi re regimes (Brooks, 2008) can be availability due to a decline in resource use altered in an invasive plant-dominated by plants that were killed by the disturbance system compared to the original system. In and an increase in resource supply rates such cases, multiple and incremental actions through plant decomposition, thereby must be undertaken to amend ecological increasing susceptibility to invasion (Davis processes so that the system is capable of et al., 2000). Disturbance is commonly supporting native vegetation and/or viewed negatively by land managers because desirable introduced vegetation. Active it contributes to weed invasion (Lozon and revegetation will likely be a necessary action MacIsaac, 1997). However, land managers to suppress reestablishment of invasive can design disturbances to manipulate site plants and initiate the reestablishment of availability to meet revegetation needs. systems dominated by desirable vegetation Designed disturbance has primarily been that meets management objectives (Borman accomplished via mechanical means (i.e. et al., 1991; Lym and Tober, 1997; DiTomaso, tilling, harrowing, chaining, disking, 2000). plowing), fi re, and herbicides, and is often While revegetation of invasive plant- part of controlling weedy plants at the site. infested rangeland sounds ideal in theory, it Th e primary goal of designed disturbance is is challenging in practice and often results in to prepare a seedbed that is conducive to invasive plant species remaining dominant seedling establishment, but a secondary 178 J.M. Mangold

objective is to control weedy species so that disturbance can be manipulated to promote desired species can establish and grow with the abundance of desired vegetation over minimal competition from weeds for soil invasive plants. For example, fi re has been resources and light. Prescribed fi re (Kyser used in fi re-tolerant native plant com- and DiTomaso, 2002; Emery and Gross, munities to reduce the seed bank of the 2005; Mangold et al., 2007a) and herbicides invasive forb yellow starthistle (Centaurea (Sheley et al., 2001, 2007; Monaco et al., solstitialis L.) (Hastings and DiTomaso, 2005; Davison and Smith, 2007) are 1996). However, properly manipulating more commonly used for controlling weedy disturbance will require further research species, while mechanical means are and understanding of what constitutes a typically used for seedbed preparation safe site for a suite of desired species. As site (Sheley et al., 1999; Th ompson et al., 2006; availability is better addressed in the future, Fansler and Mangold, 2011). it will be important to consider both the Site availability, as modifi ed through natural and human-directed disturbance designed disturbance, should strive to create regime (size, severity, frequency, patchiness, safe sites for desired species. While the and history) at a site, how they might be precise conditions that might defi ne a safe contributing to invasion, and how they site for a given species are not known (but might be manipulated during revegetation see Mangold et al., 2007b), it is generally to promote desired species (D’Antonio and assumed that a safe site meets the light, Meyerson, 2002; Isselin-Nondedeu et al., water, and nutrient needs of a species so 2006; Watts, 2010). that a seed or vegetative propagule can Seeding method focuses on placing seeds germinate, emerge, and establish (Harper et in a safe site, although it can be used to al., 1965). As mentioned above, mechanical modify species availability as well. Th e most means are often employed to create a common seeding methods are broadcast seedbed that combines a certain degree of seeding and drill seeding. Choice of seeding tilth and fi rmness to increase soil–seed method depends on factors like land contact (Sheley et al., 2008). In some cases a accessibility, topography, seedbed character- soil imprinter is employed to create surface istics, and economic constraints. Broadcast depressions that accumulate moisture, seeding is used where terrain is inaccessible organic matter, and seeds (Dixon, 1988). to machinery due to steepness, rockiness, Mechanically preparing a seedbed, however, and/or remoteness. Large areas may be can result in proliferation of annual weedy quickly and eff ectively seeded by plane or species, some of which can be invasive (J.M. helicopter, while smaller areas can be seeded Mangold, unpublished results). It is well with a hand spreader (Sheley et al., 2008). A established that disturbance results in an few of the major drawbacks of broadcast increase in nutrient availability, which seeding are the diffi culty in preparing invasive plants often capitalize on to a seedbeds and the inability to directly place greater degree than native, desired species seeds of desired species into probable safe (Huenneke et al., 1990). Th erefore, care sites. In general, broadcast seeding is should be taken to minimize disturbance as defi cient in providing suffi cient seed-to-soil much as possible while still placing desired contact that many species require for species propagules in a microenvironment adequate germination, emergence, and conducive to their growth. In regard to establishment. Because of the lack of disturbance frequency, severity, and size, adequate seed-to-soil contact with broadcast desired species will be favored when seeding, careful attention to site availability disturbances are less frequent, less severe, through seedbed preparation is necessary and smaller scale. (see discussion above). When direct seedbed Because natural and direct human preparation is not possible, fi re is believed disturbances are known to promote invasive to provide an adequate seedbed, and sites plants (Davis et al., 2000; D’Antonio and are often broadcast seeded following Meyerson, 2002), aspects of designed wildfi res or prescribed burns. In the absence Using Current Technologies and Ecological Knowledge 179

of fi re and when seedbeds cannot be Species availability prepared, seeds are typically broadcasted at rates two to three times higher than the Selecting species to address species recommended drill rate (Sheley et al., 2008). availability is often defi ned by management If the land is accessible to machinery and objectives for the site. For example, if is not too rocky, drill seeding using a no-till livestock grazing is the primary management drill is the preferred method of seeding goal at a site, revegetating with competitive, (Monsen and Stevens, 2004). A no-till drill grazing adapted, high production forage is a tractor-pulled machine that opens a grasses may be appropriate. Alternatively, furrow in the soil, drops seeds in the furrow the primary management goal may be at a specifi ed rate and depth, and rolls the habitat for specifi c wildlife species, which furrow closed (Fig. 10.2). Th e benefi t of this will require the establishment of native method of seeding is that the seed is placed forbs, grasses, and shrubs (USDI USFWS, into a predefi ned safe site (depth), thus 2001). In addition to management objectives, increasing desired species establishment. species are typically selected by matching Both site availability (i.e. safe sites created environmental conditions at the site such as with drill) and species availability (i.e. soil type, precipitation, temperature, and seeding rate) are controlled. Th ere are some elevation, with species preferences and drawbacks to drill seeding, which include: requirements. Ease of establishment is also (i) plants develop in rows resembling a crop; important as seedlings of planted, desired (ii) seeds can lodge in the drill and/or species will be competing with weed seedlings separate in the seed box based on weight emerging from the seed bank. Species that and size; and (iii) drill furrows can cause soil germinate quickly and have high growth erosion from water fl ow, but there are fairly rates have been good candidates for simple techniques for overcoming these, revegetating invasive plant-infested sites and most land managers are willing to (Sheley et al., 2008). negotiate the drawbacks in return for the Historically, a narrow suite of species improvement in species establishment (relative to total number of naturally occur- (Sheley et al., 2008). ring species) have been commercially

Fig. 10.2. Rangeland no-till drill seeder. Photo courtesy of Jane Mangold, November 2009. 180 J.M. Mangold

available for large-scale restoration of accommodate them will result in lower landscapes degraded by invasive plants, than expected establishment. Ideally, overgrazing, wildfi re, etc.; therefore, many there should be adequate propagules of degraded plant communities were seeded desired species to fi ll all available safe sites with introduced, competitive forage grasses. that are naturally occurring and/or created For example, to support managed grazed through designed disturbance (Satterth- systems, drought- and grazing-tolerant waite, 2007). introduced species such as crested wheat- grass (Agropyron cristatum L.), smooth brome (Bromus inermis Leyss.), and timothy Species performance (Phleum pretense L.) were widely established in North America. Th e use of native species Th ere are many well-established methods is being increasingly recommended to for controlling the performance of invasive restore functional and diverse plant com- plants in an eff ort to reduce their dominance munities while maintaining ecological in plant communities. Because several of stability and reducing the risk of seeding an these methods are discussed extensively in aggressive or invasive species (USDI and other chapters of this book (see DiTomaso USDA, 2002). Fortunately, plant material and Barney, Chapter 9, this volume), they availability continues to increase, likely due will not be discussed in detail here. However, to the increasing recognition that many it is commonly believed that performance of drastically disturbed landscapes will require invasive plants must be controlled, some- revegetation for successful restoration; times for multiple years, in order for availability of native species is especially successful establishment of more desirable improving due to extensive research eff orts vegetation through revegetation (Berger, (Shaw and Pellant, 2010) and increasing 1993; Wilson and Ingersoll, 2004; Fansler demand (USDI and USDA, 2002; Shaw et al., and Mangold, 2011). Invasive plant control 2005; O’Driscoll, 2007). can be achieved with a variety of mechanical, Properly distributing seeds for cultural, biological, and chemical methods, revegetation is heavily dependent on seeding but control will be highest when methods rates to facilitate colonization. Generally, address key ecological processes that are seeding rates should attempt to increase the regulating plant community dynamics at a frequency of desired propagules relative specifi c site. For example, if continual soil to weedy propagules. Typically a seeding disturbance is present at a site, then rate of 215 to 550 viable seeds m−2 is desired mechanical control of invasive plants would (Sheley et al., 2008). However, when probably not be the best option, and eff orts revegetating invasive plant-infested sites, should focus on the least disruptive method seeding rates two to three times higher are while working to minimize disturbance recommended (Sheley et al., 2008). One (Watts, 2010). In general, the performance study suggested seedling establishment of invasive plants should be addressed in an could be further improved by increasing eff ort to reduce the production of propagules seeding rates to 5 and 25 times the and decrease interference between invasive recommended rate when revegetating and desired plants. grasslands infested by invasive forbs like After the extensive eff orts implemented spotted knapweed (Centaurea stoebe L.) to prepare a site through designed dis- (Sheley et al., 1999). Ultimately, seeding rate turbance, select species, and seed species at should be determined by simultaneously the appropriate rate and with the appropriate reviewing processes that infl uence species method, follow-up management must be and site availability. For example, if an over- employed. Follow-up management should abundance of safe sites are present, then be designed to bolster the performance of seeding rates should be high. Conversely, seeded species and hinder the performance inundating the site with propagules where of invasive plants, and often takes the form there are not adequate safe sites to of herbicide applications to control reinvad- Using Current Technologies and Ecological Knowledge 181

ing weeds, deferring grazing to protect productive, there are fairly uniform soil and recently seeded species, and additional topographic conditions across a fi eld, and seeding to supplement areas of poor initial plant community dynamics are driven establishment. Unfortunately, follow-up primarily by manmade disturbances. In con- management is often overlooked (D’Antonio trast to cropping systems, natural systems and Meyerson, 2002). Because seeded vary considerably in soil characteristics, species mortality can be high during the fi rst elevation, aspect, plant community types, year or two following seeding (Jessop and etc., across any given area (Pickett and Anderson, 2007; Fansler and Mangold, Cadenasso, 1995), and a variety of dis- 2011) and reinvading weeds can be prevalent turbances (both manmade and natural) (Mangold et al., 2007a; Pokorny et al., 2010), direct vegetation change. Furthermore, monitoring to assess desired species plants native to such systems typically establishment compared to weedy species are slow growing and not highly productive establishment is critical. Additional seeding in comparison to cropping systems (Mac- and/or invasive plant control may be Arthur, 1962; Tilman and Wedin, 1991). necessary to further move the plant com- Until such diff erences between natural munity along a trajectory to a desired state. systems and cropping systems are recognized, Th e site must be managed into the future to and methods are developed to address prevent reinvasion and dominance by variation in biotic and abiotic traits and invasive plants. In many cases, manage ment processes across natural systems, revege- regimes that led to dominance by invasive tation will probably continue to result in plants prior to restoration must be changed more failures than successes. so that history does not repeat itself In most cases, invasion and subsequent (D’Antonio and Meyerson, 2002). loss of native vegetation is a slow process, taking years to decades to occur (Kowarik, 1995). At the same time, when revegetation Limitations of Current Strategies and is implemented, it usually occurs over a time Application of Ecological period of 1 to 2 years. Adequate control of Understanding invasive plant propagules prior to intro- ducing propagules of desired species may In spite of several decades of practice and take multiple years, from 2 to 5, depending research associated with revegetating weed- on the target species and its abundance in infested plant communities, revegetation the seed bank. Furthermore, legacy eff ects more often results in failure than success of invasive plant dominance on other (James and Svejcar, 2010). Th is may be due ecosystem components like soil microbes to the predominant use of traditional (Stinson et al., 2006) and nutrient cycling agronomic practices that may not be eff ective (Mack and D’Antonio, 2003) may take on landscapes that vary considerably across several years to decades to diminish before time and space. For example, a standard native species reestablishment and domin- revegetation operation on weed-infested ance are encouraged. rangeland in the western USA would likely Not only should pre-revegetation eff orts include a herbicide application, seedbed be extended beyond the typical 1 to 2 year preparation with a tiller, disk, harrow, or fi re, timeframes, but the evaluation of reveget- and seeding either with a broadcast seeder or ation outcomes should as well. Again, drill (Sheley et al., 2006; Mangold et al., revegetation is typically monitored for 1 to 2 2007a; Davies, 2010). Th is may be eff ective years, but experience suggests that shifts in in cropping systems, but less so on extensive, plant community composition, including natural lands, where the ecological processes that imposed through management, take directing vegetation change and plant species much longer and initial trends may or may traits are diff erent. For example, in cropping not be indicative of longer-term outcomes systems fertilizer and other soil amendments (Roche et al., 2008; Rinella et al., 2011). are used, crops are fast growing and highly Along with follow-up management to boost 182 J.M. Mangold

the performance of seeded species, three sites in northwestern Montana, USA. monitoring and evaluation should extend Th eir research sites occurred on ephemeral beyond a few years. wetlands that were dominated by invasive plants and had varying levels of disturbance by meadow voles (Microtus pennsylvanicus New Approaches for Revegetation Ord.) (site availability), remnant native that Integrate Our Best plants (species availability), and water Understanding of Ecological availability (species performance). At two Processes of the three sites, using augmentative restoration improved revegetation out- Th e lack of success in revegetation of weed- comes by combining treatments that would infested plant communities and the ever- maximize seedling establishment while increasing need for the restoration of lands minimizing costs. degraded by weeds suggests new approaches When designing disturbance to create are necessary. Combining our existing safe sites, abiotic and biotic soil char- knowledge and past experiences with novel acteristics must be considered. Soil approaches that consider the ecological ecological knowledge (SEK), which processes associated with site availability, acknowledges the interactions among species availability, and species performance principal components of the soil as well as will move us towards restoring plant com- feedback between aboveground and below- munities that meet management objectives ground ecosystem processes, is gaining and maintain ecosystem services in the near recognition as an important factor to and distant future. consider during restoration (Heneghan et al., 2008) (see Morris, Chapter 3, and Eviner and Hawkes, Chapter 7, this volume). Site availability Invasive plants can alter physical (e.g. water infi ltration capacity), chemical (e.g. Th e nature of natural systems is such that nitrogen mineralization, accumulation of we must address a mosaic of variable allelopathic compounds), and biological conditions across a large landscape. It will (i.e. mycorrhizal fungi associations) soil not suffi ce to treat the landscape as a characteristics. Recognizing feedbacks uniform fi eld across which large pieces of between invading plant species and soil equipment are driven to implement a may be crucial to restoring plant com- standardized disturbance treatment. munities to a more desired condition Instead, a site should be carefully assessed (Heneghan et al., 2008). For example, Vinton prior to any revegetation eff orts to ascertain and Goergen (2006) documented positive ecological processes in need of repair, and feedback between litter quality and nitrogen how site availability, species availability, and mineralization when restoring grasslands species performance may be modifi ed to invaded by smooth brome, an introduced remove invasive plants and create safe sites grass that demands more nitrogen than for desired plant communities. Th en native grasses. Others have found that soil disturbance treatments should parallel the legacies from plants that previously mosaic of biotic and abiotic conditions, occupied the soil can contribute to strong capitalizing on site and species availability priority eff ects of invasive plants on native and species performance where they are plants. Grman and Suding (2010) found intact and amending them with treatments that soil legacies created by invasive plants when necessary. For example, Sheley et al. decreased native colonizer biomass by 74%. (2009) tested ‘augmentative restoration’ in Such legacies could be due to changes in which they identifi ed and selectively resource availability (Mack and D’Antonio, repaired or replaced damaged processes (in 2003), buildup of allelochemicals (Orr et the context of site availability, species al., 2005), or alterations in soil microbial availability, and species performance) at communities including pathogens and Using Current Technologies and Ecological Knowledge 183

mycorrhizal fungi (Kourtev et al., 2002; microsite for colonization (Isselin-Nondedeu Hawkes et al., 2006). et al., 2006). Domestic sheep and goats may If soil legacies are present, site preparation also impact species performance and limit must include actions to remediate them to invasive plant species availability by grazing favor site availability free from such legacies. invasive forbs (Launchbaugh, 2006). At the When conditions of high soil nitrogen same time domestic livestock can also concentrations exist, it may be necessary to aggravate erosion processes and increase implement approaches for lowering nitrogen soil compaction, so care would need to be availability into site preparation, such as soil taken if they were to be used in site carbon addition, burning, grazing, topsoil preparation prior to aerial seeding on steep removal, and/or biomass removal (Perry et and rocky slopes. al., 2010). If allelopathic compounds are Extensive energy and resources have been present and inhibiting the establishment of devoted to developing and improving seeding seeded species, activated carbon could be technology so that propagules are directly incorporated into the soil prior to seeding to placed into a safe site. Equipment and sequester such compounds (Kulmatiski and methods for seeding natural systems in the Beard, 2006). Solarization, which is the western USA were being developed as early capturing of radiant heat from the sun to as the 1940s as ranchers and land managers cause biological, chemical, and physical attempted to increase forage production and changes to the soil, has been shown to kill rehabilitate overgrazed landscapes. Th e weed seeds or pathogens and improve standard rangeland drill seeder was designed seedling establishment (Moyes et al., 2005). in the early 1950s and remained the industry While it may be too expensive to implement standard for over 50 years. Some newer such site preparations over the entire rangeland drills, such as the Truax Rough landscape, critical areas could be identifi ed Rider™, combine technologies from the and more intensively prepared for seeding original rangeland drill with those of more than others. Most importantly, asses sing modern no-till agricultural drills (Kees, variation in biotic and abiotic character- 2006). While much progress has been made, istics such as soil texture and surface more is necessary. For example, controlling structure, nutrient and water cycling, and seeding depth, ensuring good seed-to-soil decomposition rates across the site as part of contact, and proper calibration for mixed site preparation will allow the implementation seed varieties, have all been identifi ed as of vital and appropriate treatments. problem areas (Kees, 2006). James and Unfortunately, many invasive plant- Svejcar (2010) compared seedling establish- infested lands occur on steep and rocky ment across fi eld plots where seeds were terrain that is not amenable to designed planted using a rangeland drill or seeded by disturbance via standard equipment. Just as hand. In the hand-seeding treatments, a using traditional agronomic practices often furrow similar in width and depth to the fails where sites are accessible, developing drill was made; seeds were placed, covered site preparation and seeding methods for with 1 cm soil, and gently compacted. revegetating rough, inaccessible terrain has Irrigation and weeding treatments were also fallen short. Researchers and practitioners applied across seeding treatments. Seeding should be encouraged to develop novel method was the only factor that limited techniques for situations where revegetation establishment with seedling density over equipment cannot be used. Th e use of sevenfold higher in the hand-seed compared domestic livestock as a method of designed to the drill-seeded treatments. Th e authors disturbance warrants further investigation. concluded that even modest improvements Soil surface depressions created by hoof in seeding technology might yield sub- prints can roughen the soil surface and stantial increases in seeding success (James cause a decrease in overland water fl ow and Svejcar, 2010). (Gutterman, 2003), thus accumulating and As discussed earlier in this chapter, retaining moisture for seeds and serving as a recommended seeding rates for revegetating 184 J.M. Mangold

invasive plant-infested sites can be two to Based on research that suggests low three times higher than standard diversity systems are more susceptible to recommendations (Sheley et al., 2008). invasion (Tilman, 1997; Levine, 2000; However, increasing seeding rate will only Dukes, 2002; Pokorny et al., 2005), selecting be useful if there are adequate safe sites species for revegetation should focus on available, created through designed dis- choosing species that are morphologically turbance, to accommodate all seeds. Th ere and functionally diverse. Th e rationale has been debate about whether seedling behind this is that species-rich communities recruitment is naturally limited by the will be less invasible because they are number of seeds and their dispersal occupying a larger proportion of available capabilities or the number of safe sites niches and maximizing resource uptake. (Satterthwaite, 2007). When designing Th erefore, resources are less available for use revegetation programs, theoretically, we by potentially invading species (Elton, should have complete control over site 1958). Studies that have actually quantifi ed availability; however, we have a fairly limited niche diff erentiation found invasion understanding of what constitutes a safe decreased when species richness and niche site and the relationship among safe site, occupation increased (Jacobs and Sheley, invasive plant propagules, and desired plant 1999; Carpinelli et al., 2004). Other research propagules (in other words, which species suggests that successfully establishing a will capture site availability) (Parker, 2000). species-rich seed mix that contains at least Further research into this question is one desired species of similar morphology to needed. the target invasive plant may limit reinvasion (Mangold et al., 2007a). Environmental conditions such as Species availability precipitation timing and amount, tem- perature, and solar radiation, during the Compared to advancements for designing year of seeding, may not match the ecological disturbance to address site availability, requirements of each individual species methods for addressing species availability selected for seeding (Wirth and Pyke, 2003). through species selection have experienced Th erefore, diverse seed mixes could improve greater advances and more novel approaches. seedling establishment by increasing the For example, the quantity and quality of probability that environmental conditions plant material for revegetation in the USA will match the requirements of at least one has increased substantially in the past decade species. Sheley and Half (2006) compared (Shaw and Pellant, 2010) and will likely seeding monocultures of six diff erent native continue to increase as researchers and forbs with seeding a mix of all six forbs practitioners continue to recognize the need under two diff erent watering regimes. for revegetation on invasive plant-infested Establishment of the mixture, as measured landscapes. While I will not go into a by density of seeded species, was similar to discussion about origin of plant material and the average density for individual species, maintenance of genetic integrity, researchers and the forb mixture was seven times more and practitioners are increasingly stressing competitive with the invasive forb spotted the importance of seed origin so that habitats knapweed than a single native forb was of restoration and donor sites can be when seeded alone. Because species diff er in appropriately matched and genetic integrity traits and ranges of tolerances, seeding can be maintained (Vander Mijnsbrugge et mixtures with greater species richness are al., 2010). I will also not address the use of more likely to contain a species within a native plant material versus introduced plant functional group that will germinate and material, but will instead present some novel emerge under varying and unpredictable approaches that integrate both native and environmental conditions (Tilman, 1994). introduced species into the revegetation Even though some research suggests low scheme. diversity systems are more susceptible to Using Current Technologies and Ecological Knowledge 185

invasion (see above), other studies suggest quantifying such traits for the broad array of native species diversity and invasive plant currently available plant materials may be diversity are positively correlated (Stohlgren very useful in designing seed mixes that are et al., 1999; Symstad, 2000; Bruno et al., both diverse and productive. In the future 2004). While this discrepancy may be due revegetation species will be described by to the scale upon which the diversity- their traits and requirements for establish- susceptibility to invasion question is ment, such as growth form, preferred soil examined, positive coupling between native type, minimum precipitation requirements, and invasive plant diversity can also be and seeding rate, but may include functional linked to productivity; that is, the highest traits such as relative growth rate and native plant diversity and invasion should specifi c leaf area as well. Such information occur on productive sites, or ‘hot spots’ of would be highly useful to practitioners who native diversity (Stohlgren et al., 1999). are striving to design seed mixes that are Highly productive sites likely meet plant invasion resistant. resource requirements and are extremely Just as designed disturbance that creates desirable for plant growth. Th erefore, such safe sites for desired species requires areas typically support many diff erent fl exibility across the landscape to refl ect species and may be especially vulnerable to variable conditions, there should also be invasion, especially if moderate disturbances fl exibility in choosing species appropriate free resources for invasive plants for various locations on the landscape. All (Radosevich et al., 2007). If this postulate is too often a species mix is created based on true, selecting species for revegetation that land management goals, and the same are highly productive may go further to species mix is then seeded across the entire prevent reinvasion than focusing on species- project area. Alternatively, species mixes rich seed mixes because highly productive could vary according to soil resource species will maximize resource use. Evidence availability and texture, topography, degree certainly exists to suggest this may be the of infestation severity, potential seed case. For example, crested wheatgrass, a dispersal routes (i.e. livestock and wildlife productive, introduced perennial grass, has trails, roads, waterways, etc.), and other been observed to resist invasion by factors commonly found in an ecological site knapweed and cheatgrass (Berube and description (ESD). It might also be useful to Myers, 1982; D’Antonio and Vitousek, identify areas on the landscape where 1992). James et al. (2008) compared native successful establishment of seeded species is bunchgrasses, perennial forbs, and annual critical, and then seed islands of chosen forbs in their soil nitrogen use and resistance species (Elstein, 2004). Species seeded in to annual grass invasion. Th ey found that islands could be chosen based on their native bunchgrasses, the dominant plant dispersal capabilities, with the intent of functional group at their study sites, choosing species that would have a high acquired the most nitrogen from all nitrogen likelihood of dispersing to non-seeded areas pools, accumulated the most biomass, and over time. Because areas seeded to islands was the only functional group that inhibited would be small relative to the entire annual grass establishment. landscape, they could be intensely managed Ideally, revegetation mixes will be diverse for successful establishment of seeded and productive. Recent and future species. In short, fl exibility and creativity in developments in the fi eld of plant materials selecting species and placing them on the will help this to become a reality. Because landscape should be encouraged. plant functional traits such as biomass If disturbances are designed to potentially production, canopy size, relative growth favor desired species, then the benefi t of rate, and specifi c leaf area have been shown high seeding rates of desired species to to be correlated with competitive per- overwhelm the pool of available propagules formance of a species (James and Drenovsky, (including those of invasive plants) and 2007; Isselin-Nondedeu et al., 2006), occupy the majority of safe sites can be 186 J.M. Mangold

realized. As stated earlier in this chapter, Dickson and Busby (2009) varied grass and seeding rates 5 and 25 times the forb seeding rates with the objective of recommended rate have been shown to fi nding the density of grass and forb seeds increase establishment of perennial grasses that would maximize forb abundance and in the short term in spotted knapweed- richness during prairie restoration in infested grasslands (2 years after seeding) Kansas, USA. Th ey found that higher seeding (Sheley et al., 1999). But did high seeding densities of grasses resulted in decreased rates result in favorable outcomes in the forb cover, biomass, and richness, and long-term? Six years after seeding, only the higher seeding densities of forbs increased highest seeding rate (12,500 seeds m−2) forb richness. Exotic species biomass produced wheatgrass density higher than dramatically declined at all seeding rate that of the non-seeded control (Sheley et al., combinations. Seeding islands of forb-rich 2005). A designed disturbance (tillage or mixes (Elstein, 2004) could eliminate herbicide application) combined with high unwanted competition between desired seeding rates doubled wheatgrass density forbs and grasses, would help to mimic the compared to seeding in the absence of natural patchiness of species usually found disturbance, and spotted knapweed biomass across natural landscapes (Dickson and tended to be lower where wheatgrass density Busby, 2009), and would allow for the and biomass was highest (Sheley et al., follow-up management of reinvading weedy 2005). Fifteen years after revegetation, the forbs with a broadleaf herbicide across the eff ect of high seeding rates on wheatgrass majority of the site (minus the forb-rich establishment was less pronounced as islands) without harming seeded, desired wheatgrass density and biomass at lower forbs. rates had increased to similar levels; how- It is vital that we continue to investigate ever, the eff ect of well-established wheat- and identify the key ecological processes at a grass on reducing reinvasion by spotted site that are contributing to propagule knapweed was more dramatic than at 2 or availability and establishment of desired 6 years, with an 86% reduction at the species. Seed and seedling herbivory is one highest seeding rate (Rinella et al., 2011). ecological process that has received some Th is long-term data set suggests high attention, and conclusions thus far suggest seeding rates can improve revegetation it may have far more signifi cant impacts on outcomes by reducing reinvasion. seedling establishment than previously Revegetation has been dominated by acknowledged (Hulme, 1994; Fenner and seeding relatively high rates of grasses in Th ompson, 2005). For example, Watts comparison to forbs. Th is is likely due to (2010) found that underground foraging by plant material availability and cost, the pocket gophers (Th omomys spp.) limited eff ectiveness of grasses to reduce soil erosion recruitment and establishment of native (Boyd, 1942; Jelinski and Kulakow, 1996), bunchgrasses in California, USA, grasslands. and the belief that highly productive grasses Bunchgrasses that attained a certain size can control weedy species and improve soil were able to survive the eff ects of predation, characteristics (Dickson and Busby, 2009). suggesting that short periods (3–5 years) of While grasses are important in reducing gopher exclusion could increase the chance reinvasion (e.g. James et al., 2008), other of survival for perennial bunchgrasses research suggests forbs may be very (Watts, 2010). In another study, grasshopper important for reducing reinvasion by weedy herbivory negatively impacted native plant forbs (Pokorny et al., 2005; Mangold et al., diversity in crested wheatgrass stands in 2007a). Seeded grasses can impede the North Dakota, USA; when abundant, establishment of seeded forbs (Kindscher grasshoppers and other invertebrate and Fraser, 2000; Parkinson, 2008). herbivores could impede native species Th erefore, it may be useful to alter the establishment, especially that of grasses proportion of desired grasses relative to (Branson and Sword, 2007). Accounting for desired forbs when considering seeding rate. processes, like herbivory, that present Using Current Technologies and Ecological Knowledge 187

obstacles to seedling establishment, and concentrations, high organic matter, and then adjusting seeding rates accordingly, well-developed and complex microbial will require a deeper understanding of how communities) for the establishment and such processes function. survival of native, late-seral species. Finally, Usually revegetation occurs by seeding a late-seral species would be seeded that meet mix of species that meet the long-term the long-term management objectives for management goals for the site. Th e seed mix the site. From a more practical perspective, is typically drill- or broadcast-seeded in fall, when revegetating invasive forb-infested and invasive-plant control measures are plant communities, it may be necessary to implemented shortly after (Sheley et al., fi rst establish grasses, while hindering the 2008). Because they are chosen to meet performance of the invasive forb with a long-term management objectives, the follow-up broadleaf herbicide application, seeded species are commonly slow-growing, then inter-seed with desired forbs and nutrient conservative, long-lived, late-seral shrubs during a second seeding phase. species. However, soil characteristics (e.g. Inter-seeding may allow the established nutrient concentrations, organic matter, grasses to facilitate the establishment of and microbiota) at the site may not be desired forbs and shrubs and reduce conducive to their establishment and growth interspecifi c competition between seedlings (Firn et al., 2009). For example, the existing of diff erent plant functional groups propagule pool likely contains invasive (Gunnell et al., 2010). plants, which are fast-growing and nutrient A similar approach has been proposed exploitative; the recent disturbance designed and tested for restoring plant communities to control invasive plant performance and dominated by annual grasses like cheatgrass increase site availability has likely resulted and medusahead (Taeniatherum caput- in an increase in soil resources and ambient medusae (L.) Nevski). In ‘assisted succession’ light, both of which are favorable for a site is moved from annual plant to colonizing species (i.e. weeds). perennial plant-domination by fi rst seeding Revegetating with multiple planting a competitive, aggressive perennial grass, phases using bridge species (cover crops) followed by creation of site availability and that transition the plant community from reinsertion of species native to the pre- one successional stage to the next (early-, invasion plant community (Cox and mid-, and late-seral) addresses the need to Anderson, 2004). An aggressive grass, such seed species into environmental situations as crested wheatgrass, has a higher likelihood that are more conducive to their growth as of establishing in a primarily monotypic described above. For example, initially fast- stand of invasive annual grass than do native growing, short-lived species could be seeded species. Over time the reestablishment of to provide immediate and direct com- perennial vegetation versus annual vege- petition with invasive plants that may be tation helps to restore ecological processes regenerating from the seed bank (i.e. nurse like nutrient cycling and fi re regimes to a or cover crops) (Vasquez et al., 2008; Perry condition more favorable for perennial et al., 2009). Th ese species may be seeded as native species. Further development and an individual phase or as a mix with mid- testing of this unique idea is needed. One seral species that are intermediate in their potential obstacle to its implementation is growth rate, nutrient use, and longevity. adequate control of the competitive, Th ese initial and mid-phase species could aggressive grass so that native species can very likely be exotic species (D’Antonio and ultimately be reestablished (Fansler and Meyerson, 2002). Th e fi rst one to two Mangold, 2011). If seeding methods are revegetation phases would function to going to include multiple seeding phases, return conditions at the site, such as soil care must be taken to ensure the bridge nutrient concentrations, organic matter species of choice do not become problematic and microbial communities, to a condition themselves (D’Antonio and Meyerson, more appropriate (e.g. low nutrient 2002). 188 J.M. Mangold

Because the seedling stage is arguably the management priority. In a like manner, most vulnerable stage in a plant’s life history populations of invasive plants may vary in (Otsus and Zobel, 2002; Clark and Wilson, their degree of ‘invasiveness’ based on the 2003), developing seeding methods that biology (e.g. seed production and vegetative assist a plant through this stage is critical. spread, recruitment, and growth rates) of One method may be to transplant juvenile the species and whether habitat require- plants instead of planting seeds. Trans- ments are being fully met for populations to planted seedlings have a higher rate of reach their maximum rates of growth and survival than seedlings germinating from spread. Managers should prioritize manag- seed in the fi eld (Page and Bork, 2005; ing the performance of more invasive Middleton et al., 2010). Transplanting populations that will increase in size and seedlings across large landscapes would be density at a faster rate than less invasive very labor and time intensive. Th erefore, populations, which may serve as a signifi cant similar to the island seeding concept source of propagules (Lehnhoff et al., 2008; (Elstein, 2004), transplanting could be Maxwell et al., 2009). applied by identifying areas on the landscape Controlling invasive plant performance where successful establishment of seeded will take time, and time should be readily species is critical and transplant small factored into revegetation plans. Th e length patches of desired species. Integrating of time since introduction of the invasive transplants along with seeding would ensure plant and the biology and life history of the that dispersal and establishment of certain species (i.e. prolifi c seed producer, annual species is guaranteed, thus improving the versus perennial species, and reproductive overall richness, diversity, and quality of strategy) can play a big role in the outcomes (Middleton et al., 2010). eff ectiveness and longevity of control measures, because they infl uence propagule availability and reinvasion potential Species performance (Reinhardt Adams and Galatowitsch, 2008). Th is, in turn, can determine the extent to As stated earlier, during revegetation weedy which seeded desirable species must species must be reduced or eliminated so compete with invasive species during their that desired species can establish and grow establishment phase. As seed production with minimal competition from undesired and seed longevity of an invasive species species. When considering revegetation increases, the more time it will take to strategies, it is necessary to carefully adequately control its performance. Because examine invasive and other nuisance weedy plant community composition is somewhat species in the context of disturbance, predictable based on seed bank composition succession, and ultimately restoration. (van der Valk and Pederson, 1989), sampling D’Antonio and Meyerson (2002) state that of the seed bank prior to designed dis- when setting priorities and goals for turbance will provide insight into whether revegetation, it is essential to understand invasive species are likely to remain domin- the potential transience of invasive and ant for some time (D’Antonio and Meyerson, weedy species at a site and the role they 2002), and if other weedy species are present might play in altering processes that that may increase following any type of soil infl uence the course of succession. Th ey also disturbance (Baskin and Baskin, 1998). If suggest that there may be some situations the proportion of invasive species propagules where a nuisance weedy species is short- to desired species propagules is very high, lived and successional to native species, thus then multiple years of control may be not requiring control in the post-disturbance necessary prior to reintroducing desired environment; in contrast, long-lived invasive species (Fansler and Mangold, 2011). While plants or tenacious invasive annuals that this may add time and expense to the are both good colonizers and persistent revegetation program, the probability for community members should be a top success may increase. Using Current Technologies and Ecological Knowledge 189

Following revegetation the threat of individuals who are interested not only in reinvasion is substantial due to the high implementing revegetation, but evaluating likelihood that a large number of invasive its eff ectiveness across short-, mid-, and species propagules still exist on the site. Th e long-term timeframes. Very few long-term likelihood of seedling mortality is also high. studies exist, but those that do suggest Monitoring to assess desired species initial trends may not be indicative of long- establish ment compared to reinvading term outcomes (Ferrell et al., 1998; Roche et species will allow the timely intervention al., 2008; Rinella et al., 2011). Rinella et al. and follow-up management to control (2011) re-sampled four projects that species performance. If grasses have been integrated various forms of invasive plant seeded fi rst in a multi-phase seeding process control and seeding in invasive forb-infested as described above, applying control grasslands in western Montana, USA. Th ey methods (e.g. herbicide application or measured seeded species and invasive forb grazing with sheep) to hinder the per- density and biomass 15 years after seeding. formance of reinvading weedy forbs will be Some seeded populations remained very necessary. If desired species seedlings small for 6 or more years but then became appear to be underperforming, intervention highly productive and greatly suppressed to improve their performance may be the invader (Fig. 10.3). Other populations necessary. Intervention may include sup- maintained high densities for 3 or more plemental irrigation, precision fertilization, years but then became exceedingly rare or or over-seeding (broadcast) to increase the extinct. Th e results suggest that seeding number of desired species propagules for sometimes provides appreciable long-term subsequent germination, emergence, and benefi ts, but at other times fails, and that establishment when conditions are more short-term data can both over- and under- amenable. While such activities will require estimate the lasting benefi ts of revegetation. more time and money, the benefi ts may be Additional long-term studies are needed to substantial. Furthermore, if critical areas for assess the likelihood of favorable seeding seedling establishment are identifi ed early outcomes, to identify good seeded species in the planning and revegetation process, traits, and refi ne ecologically based then strategies can be concentrated there strategies for controlling site availability, rather than being implemented over the species availability, and species performance entire area. Most ecological processes like during revegetation. nutrient and water cycling, decomposition rates, propagule dispersal patterns, and plant–plant interference, will not heal Conclusion within a few years, therefore management should be focused on directly infl uencing Severely degraded plant communities may their functionality. Evaluating plant com- require the modifi cation of site availability, munity composition (e.g. cover and species availability, and species per- frequency) and seed bank composition, formance to improve the success of sampling the soil for organic matter and revegetation and return plant communities nutrient concentrations, and noting to a desired state. Revegetation is a evidence of higher trophic level interactions challenging prospect that often results in (e.g. herbivory) over time are a few measures less than optimum outcomes. In a relatively that may be taken to assess whether the brief manner, this chapter has attempted plant community is transitioning in a to describe standard practices, identify desirable direction. limitations to those practices, and propose Monitoring and evaluating the success or new approaches for improving revegetation failure of our revegetation eff orts must be a in the future. Th e ideas and approaches long-term process and measured in the here are certainly not an exhaustive list, context of plant community successional but rather some ideas for which there is trajectories. It also takes dedicated supporting data. It is my hope that these 190 J.M. Mangold

(a) (b)

(c) Fig. 10.3. Spotted knapweed (Centaurea stoebe L.)-infested grasslands where no seeding (a) or seeding with bluebunch wheatgrass (Pseudoroegneria spicata (Pursh) A.Love) (b) or intermediate wheatgrass (Thinopyrum intermedium (Host) Barkworth & D.R. Dewey) (c) had occurred 15 years earlier. Notice persistence of spotted knapweed in the non-seeded plots and lack of spotted knapweed in the seeded plots. Photos courtesy of Jane Mangold, July 2010. Using Current Technologies and Ecological Knowledge 191

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abiotic processes and thresholds, exotic plants vegetation composition 94 altered hydrology description 93–94 hydraulic conductivity 48 inhibitory eff ects 94 irrigation 48 phytotoxic chemical 94 salt concentrations 48 root exudates 94 availability, soil nutrient shrubs 94–95 cycling and uptake 47 ‘soil sickness’ 94 fertilization 47 altered fi re regimes rate of recovery 46–47 description 24 synthetic fertilizers 46 grazing 24 distribution, soil nutrient sagebrush ecosystems 24 homogenization 46 shifting, plant community 24–25 ‘islands of fertility’ 46 stocking 24 soil structure and microtopography arbuscular mycorrhizal fungi (AMF) 90, 128, 130 compaction 45 plow pan 45 biological invasions 79 small-scale heterogeneity 45–46 biotic processes and thresholds, exotic plants acquisition resources vs. plant-soil resources altered competitive interactions canopies 155–156 Avena barbata 42, 43 grassland community 155 cultivation management practices 44 invasive and native species growth 155 life history traits 42 light 155–156 microbial biomass 42–44 limitations 155 mycorrhizal relationships 44 soil nutrients 156 description 40–41 water 157 recovery woody species 155 degradation 44 adaptive management foundation species 45 ecological site information 11, 12 repair strategies 44–45 site assessment and 109 seed limitation 44 allelochemicals 126–127 reproductive limitations allelopathy size and landscape position 41 and chemical interference wind pollination 41 antagonistic eff ects 94 seed limitations density-dependent factors 95 banking 41 separation 95 dispersal 41–42

197 198 Index

cogongrass 160 and STMs 8, 11, 113 cultivation see Land-use legacies ecosystem processes causes, succession 162 DAYMET US Data Center 115 disturbance regimes 163–164 diff use knapweed 157 EBIPM 162 domestic livestock invasive and desired species 163 grazing 26 ‘key-lock approach’ 162 herbivores 25 manipulating soil resources 164–166 seedlings 25–26 propagule pressure 164 stocking and mortality rates 25 ecosystems, invasibility woody plants 25 defi nition 160 disturbance 160–161 eastern US forest community 155 habitat and species diversity 161 EBIPM see Ecologically-based invasive plant propagule pressure 162 management (EBIPM) resource availability and open niches ecologically-based invasive plant management 161–162 (EBIPM) 162 El Niño Southern Oscillation (ENSO) 115, 116 ecological processes, revegetation exotic plants site availability abiotic processes and thresholds biotic and abiotic characteristics 182, 183 altered hydrology 48 designing disturbance 182 availability, soil nutrient 46–47 disturbance treatments, mosaic 182 distribution, soil nutrient 46 occurrence, invasive plant infested lands soil structure and microtopography 183 45–46 rangeland drills seeder 183 biotic processes and thresholds seeding rate 183–184 altered competitive interactions can favor soil legacies 182–183 exotic plants 42–44 solarization 183 reproductive limitations can favor exotic species availability plants 41 annual grass invasion 185 seed limitations can favor exotic plants cheatgrass and medusahead 187 41–42 cover crops 187 design seed mixes 185 foundation species 45, 49 desired species 185–186 diverse and productive 185 historical environmental conditions 184 fi re regimes 163 herbivory grasshopper 186–187 food regimes 163–164 islands, species seeded 185 homogenization 46 long-term management goals 187 native species diversity 185 invasive gorse 159 niche 184 invasive plant impacts perennial and annual vegetation 187 allelopathy 93–95 pocket gophers 186 amendments/fertilization 98 quantity and quality, plant material 184 ‘biological invasions’ 79 seeding density and rates 186 competitive ability 80 seedling stage 188 complex interactions 98 species mixes 185 description 79 species performance see Species performance direct interactions and linkages 81, 82 Ecological site description (ESD) ecosystem 79–80 MLRAs 7 exotic species, management/control 81 seeded species 185 interactions, direct and indirect Index 199

decomposition 82 invasive species management description 82, 83 applying ecological site information disturbance regimes 82 adaptive management cycle 11, 12 grazing and feeding 82–83 cattle 15 management 84 description 11 rhizosphere 83 honey mesquite 14–15 vegetation management 83 implementation phase 12 ‘invasive plant paradox’ 80 limestone soils 13–14 litter and decomposition prickly acacia 15 augmentation 85–86 redistribution, vegetation production 13 environment 85 sandy and loamy soils 14 management, natural systems 85 spatial heterogeneity 12 microbial mineralization, N 84 spatiotemporal heterogeneity 12 net primary productivity (NPP) 85 state-and-transition model 13 nitrogen 84 transition 12–13 pathways 84 in complex landscapes quantity and quality 84 agronomic and economic principles 4 management methods policy role 5–6 activated carbon (AC) 96–97 prediction approach 5 carbon addition and fertilization 95–96 quantitative information 4 description 95 remote sensing applications 4–5 microbial and invertebrate populations spatial and temporal patterns 3–4 97–98 spatiotemporal heterogeneity 6 microbial communities see Microbial time 4 communities vulnerability 5 nutrient cycling and biogeochemistry see weeds 4 Nutrient cycling description 3 physical properties, soil ecological sites aggregation 93 ‘climax’ vegetation 6–7 climates 93 description 7 description 92 ESD 7, 8 evapotranspiration 93 major land resource areas (MLRAs) 7 pH 93 practices and policies, weed 7 salinity 93 wildlands 6 vegetation 92–93 invaders 3 water salvage 93 patterns 15–16 potential management methods 98–99 STMs see State and transition models restoration 81 (STMs) soil invertebrates 91–92 invasive plants, reducing species performance key-lock approach 162 early-seral species 169 herbicides 169 land-use legacies historic grazing regimes 166 abiotic processes life cycle 166 recovery 48–49 linkages 166, 167–168 and thresholds 45–48 medusahead 166 anthropogenic vs. natural disturbance mowing 166 description 37 periodic burning , USA 169 duration 37–38 rapid germination and establishment 166 economic and social factors 37 sensitive riparian areas/grasslands 166 ecosystem properties 38 soil cultivation 166 intensity and severity 37 200 Index

land-use legacies continued growth rates 69–70 landscape positions, old fi elds 38, 39 nitrifi cation 69 life history and competitive, resident establishment and maintenance species 38 cottonwood and willow species 71 size and frequency 37 precipitation/fl ooding 70 biotic processes salt cedar and Russian olive 71 recovery 44–45 stream fl ow 70 and thresholds 40–44 sagebrush steppe ecosystem cultivation 27 cheatgrass biomass 69 degradation model 38, 40 description 68 description 36 grazing 69 exotic shrubs and forbs 36–37 maintenance 69 forage species potentials 68–69 grazing and fi re 28–29 soil water recharge 68 monocultures 28 microbial communities fragmented ownerships 28 AMF diversity 90 habitat fragmentation 27–28 canonical correlation analysis 90 intervention Chromolaena odorata and Fusarium 90 management techniques 49–50 description 89 secondary succession 50 mineralization 89 old fi elds 50–51 molecular techniques 91 secondary succession 36 nitrifi cation 89–90 structure and ownership pattern 27 nutrient cycling 89 Mojave and Sonoran deserts 161 management methods, invasive species mowing 67, 70, 126, 131, 166, 177 activated carbon (AC) allelopathic compounds 96 National Oceanic and Atmospheric description 96 Administration (NOAA) 115–116 and fertilizer 97 nutrient availability physical and chemical properties 96–97 discrepancies 143 plant-available N 97 evaluation 144–147 carbon addition and fertilization reconsideration 143–144 burning 96 soil N availability 143 conjunction 96 theoretical and empirical work 142–143 grazing 96 nutrient cycling nitrogen availability 95–96 and biogeochemistry removal, plant material 96 acquisition strategies 88 description 95 cheatgrass 87–88 microbial and invertebrate populations co-occurring processes 86 ecological interaction 97 generalizations, exotic species 86–87 electroshocking 97 management, weeds 88–89 fungal:bacterial ratios 97–98 Myrica faya 88 inoculation, N-fi xing bacteria 97 plant–soil N cycle 87 resource management 98 potassium and manganese 88 restoration shortcut 98 productivity and diversity 86 management scenarios, resource pool revegetation 89 management Russian olive 88 annual and perennial grass communities, symbiotic N fi xation 86 soil N wiener-leaf/saltlover 87 carbon addition 70 resource infl ux and effl ux, plant modifi cation cool-season 69 61 fl uctuations 70 soil fertility 47 Index 201

and soil water movement 38 limitations 155 nutrient environments 143, 149–150 soil nutrients 156 nutrient management, species performance water 157 advancement 147 woody species 155 generalist herbivore 148–149 adaptive qualities 154 potential resorption 148 and ecosystem processes, management see recycling 148 Ecosystem processes restoration outcomes 148 invasibility, ecosystems RGR species 147 defi nition 160 SLA species 147 disturbance 160–161 structural investments, leaf and root cells habitat and species diversity 161 148 propagule pressure 162 tissue construction strategy 147–148 resource availability and open niches tissue senescence 148 161–162 nutrient-poor systems 142, 147 management strategies early-seral species 169 plant–soil feedbacks herbicides 169 invaded state, invader’s 123–124 historic grazing regimes 166 mechanisms life cycle 166 allelochemicals 126–127 linkages 166, 167–168 challenges 132–134 medusahead 166 disturbance 132 mowing 166 litter 124, 126 periodic burning , USA 169 nitrogen 130–132 rapid germination and establishment 166 potential management tools 124, 125 sensitive riparian areas/grasslands 166 soil microbial community 128–130 soil cultivation 166 soil salinity 132 non-native plants, natural areas 154 restoration 135 phenotypic adaptation 154 shaping plant communities and invasions reproduction and dispersal 122–123 fundamental role 157–158 pollination 158 pollination 158 propagules propagule dispersal 159–160 dispersal 159–160 propagule pressure 158–159 pressure 158–159 seeds and propagules 158 and seeds 158 revegetation manipulations see Revegetation, ecological theory rangeland restoration reproduction monitoring and adaptive management 114 fundamental role 157–158 weather pollination 158 climate data resources 115–116 propagules see Propagules and long-term forecasts 114–115, 116 resource pool management state-and-transition probabilities see availability 58 Weather variability competitive interactions, plant communities successional processes and climate assembly and succession 65 109–112 changes 65 reducing invasive plant performance description 64 acquisition and plant-soil resource drought 64 interactions factors 64 canopies 155–156 germination 64–65 grassland community 155 trees/shrubs and grasses 65 invasive and native species growth 155 description 57 light 155–156 and fl uxes 202 Index

resource pool management continued germination and establishment, seeds 171 description 58, 59 heavily invaded 170 modulators 58 natural ecosystems 169 variation 59 plant communities water and nitrogen 59 site availability 177–179 grazing and mowing/burning 67 species availability 179–180 invasion 58 species performance 180–181 and invasive species seedbed preparation 171 disturbance 66 seeding 171 establishment 67 species rich seed mixes 170 exotic species, traits 66–67 transplant seedlings/young plants 170–171 light grazing 66 resource availability 65–66 long-term 67–68 secondary succession 36 modifi cation, species seed banking 41 large-scale infl uences 62 seed dispersal 41–42 nitrogen-enhancement 62 soil ecological knowledge (SEK) 182 nitrogen infl ux 61–62 soil invertebrates physical and chemical properties 62 bacteria and fungi 91 seedlings 61 decomposition 92 shrub removal 61 dominant species 92 nitrogen availability 67 fauna 91 plant modifi cation macrofauna 91 ecosystem engineers 59–60 mesofauna and microfauna 91–92 grazing and harvest 61 soil microbial community 128–130 infl ux and effl ux 61 soil micro-environment, rangeland restoration nutrient resources 61 climate and weather data resources 115–116 precipitation 60 ecological processes 109 transpiration 60–61 long-term weather forecasts 114–115 resistance 68 monitoring and adaptive management 114 scenarios see Management scenarios, planning cycle 109 resource pool management plant communities 107 species communities practices 109 diverse 62 precipitation 107 growth and maintenance 62–63 requirements 107 soil nitrogen 63–64 seasonal distribution, level 107, 108 solar fl ux 64 seedbed preparation and planting techniques water availability, layer models 62 108–109 stochastic events 58 successional processes, weather and climate revegetation, ecological theory seeding rate and optimization, species active and passive revegetation 170 availability 110 approaches site availability and optimization 109– site availability 182–184 110 species availability 184–188 species performance and competition, species performance 188–189 resources 110–112 ‘assisted succession’ 171 weather variability see Weather variability crested wheatgrass, USA 171 soil resources current strategies and application, desirable vs. invasive species 164–165 limitations 181–182 invasive annual grasses 165 desirable species 176–177 invasive perennial vine Cape ivy 165 functional diversity 170 microbial populations 166 Index 203

natural environments 165 management and patch-burn treatments nutrient availability 165 30 proportion, nitrogen 165 life histories strategies seed priming 165 fi re and grazing 23 yellow starthistle and jubatagrass 165 functional categories 23 soil salinity 132 resource availability and pathogen species availability community 23–24 annual grass invasion 185 traits/‘vital attributes’ 23 cheatgrass and medusahead 187 plot-level herbicide applications 31 cover crops 187 rangeland dynamics 19 design seed mixes 185 rates and patterns, species invasions 19 desired species 185–186 traits, life history 20 diverse and productive 185 species performance environmental conditions 184 control 188 herbivory grasshopper 186–187 determination 142 islands, species seeded 185 invasiveness 188 long-term management goals 187 low and high nutrient environments native species diversity 185 149–150 niche 184 monitoring and evaluating 189 perennial and annual vegetation 187 nutrient availability and invasion pocket gophers 186 discrepancies 143 quantity and quality, plant material 184 evaluation 144–147 seeding density and rates 186 reconsideration 143–144 seedling stage 188 soil N availability 143 species mixes 185 theoretical and empirical work species invasions 142–143 alterations, landscapes nutrient management 147–149 altered fi re regimes 24–25 reinvasion 189

C3 and CO2, atmospheric trace gases restoration practitioners 142 26 seedling establishment and populations climate-driven vegetation change 27 189 climatic infl uences 26–27 setting priorities and goals 188 domestic livestock 25–26 species-poor communities 155 land-use legacies 27–29 state and transition models (STMs) weather patterns, short-and long-term applications 10–11 26–27 description 8 disturbance processes, rangelands dynamic soil properties (DSPs) 8 biodiversity 23 and ESDs 11 complex landscape 22–23 limitation 11 conceptual model, path dynamics 22 plant attributes 8 defi nition 20–21 resistance 8 and diversity relationship 21–22 restoration 10 drought 21 sandy ecological site 8, 9 fi re and grazing 20, 21 states 8 ecological factors 20 thresholds 10 generalizations, species invasions 31 transitions 8, 10 grazing and fi re 29, 31 invasion processes 19–20 thresholds see Abiotic processes and Lespedeza cuneata thresholds, exotic plants; Biotic growth stages 30 processes and thresholds, exotic life history and traits 29 plants 204 Index

United States Department of Agriculture (USDA) cheatgrass 113 National Water and Climate Center 115 evaluation 113 Natural Resources Conservation Service 115 organizing and ranking historical data 113 USDA see United States Department of potential trajectories 113 Agriculture (USDA) sagebrush-bunchgrass rangeland 113 water corridors 164 vegetation states 112–113 weather variability wildland and long-term forecasts 116 detection, invasive species 4–5 state-and-transition probabilities and management 6 assessment predictive models 4