Multi-National Conservation of Alligator Lizards

Home , Abedus herberti, Abronia (lizard), Abronia campbelli, Abronia meledona, Abronia taeniata, Aechmea zebrina

MULTI-NATIONAL CONSERVATION OF ALLIGATOR LIZARDS:

APPLIED SOCIOECOLOGICAL LESSONS FROM A FLAGSHIP GROUP

ADAM G. CLAUSE

(Under the Direction of John Maerz)

ABSTRACT

The Anthropocene is defined by unprecedented human influence on the biosphere.

Integrative conservation recognizes this inextricable coupling of human and natural systems, and mobilizes multiple epistemologies to seek equitable, enduring solutions to complex socioecological issues. Although a central motivation of global conservation practice is to protect at-risk species, such organisms may be the subject of competing social perspectives that can impede robust interventions. Furthermore, imperiled species are often chronically understudied, which prevents the immediate application of data-driven quantitative modeling approaches in conservation decision making. Instead, real-world management goals are regularly prioritized on the basis of expert opinion. Here, I explore how an organismal natural history perspective, when grounded in a critique of established human judgements, can help resolve socioecological conflicts and contextualize perceived threats related to threatened species conservation and policy development. To achieve this, I leverage a multi-national system anchored by a diverse, enigmatic, and often endangered New World clade: alligator lizards. Using a threat analysis and status assessment, I show that one recent petition to list a California alligator lizard, Elgaria panamintina, under the US Endangered Species Act often contradicts the best available science. Building on this analysis, I also provide empirical evidence that the multi-species petition model under which Elgaria panamintina was proposed is problematic, thus corroborating claims made by the US Fish and Wildlife Service in their recent policy decision to ban such petitions. Shifting to Mesoamerica, I use global and regional status listings, distribution data, and a preliminary protected area gap analysis to show that an entire genus of alligator lizards (Abronia) is warranted for recognition as a flagship, despite conflicting social views toward the lizards. I supplement this finding with novel radio telemetry data on Abronia graminea, which reveals their strong arboreality, generalist forest habitat use, and adaptability to forest disturbance.

Finally, I provide a checklist and bilingual dichotomous key for all 29 species of Abronia, and offer best-practice solutions to broader species identification problems in understudied regions of the world. My findings challenge conventional wisdom in this system, and showcase lessons of broad relevance to applied conservation that account for social and biological linkages.

INDEX WORDS: Abronia, Elgaria, Batrachoseps, Biodiversity, California, Endangered

Species Act, Identification, Imperiled species, Integrative conservation,

Flagship species, Mesoamerica, Mexico, Museum science, Natural

resource management; Spatial ecology; Protected area gap analysis; Radio

telemetry, Transdisciplinary

MULTI-NATIONAL CONSERVATION OF ALLIGATOR LIZARDS:

APPLIED SOCIOECOLOGICAL LESSONS FROM A FLAGSHIP GROUP

ADAM G. CLAUSE

B.S., University of California, Davis, 2010

A Dissertation Submitted to the Graduate Faculty of The University of Georgia in Partial

Fulfillment of the Requirements for the Degree

DOCTOR OF PHILOSOPHY

ATHENS, GEORGIA

2018

Adam G. Clause

MULTI-NATIONAL CONSERVATION OF ALLIGATOR LIZARDS:

APPLIED SOCIOECOLOGICAL LESSONS FROM A FLAGSHIP GROUP

ADAM G. CLAUSE

Major Professor: John Maerz Committee: Byron Freeman Jeffrey Hepinstall-Cymerman Nik Heynen Fausto Sarmiento

Electronic Version Approved:

Suzanne Barbour Dean of the Graduate School The University of Georgia August 2018

DEDICATION

To Guy Clause, John Kinsey, and Joe Maher, who I wish could have seen this dissertation completed.

ACKNOWLEDGEMENTS

First to my parents, without whose encouragement this dissertation would not have been possible. Your patience with me, hands-on help with field work, and effort to keep me grounded were more important than you know. Thank you for trusting me to go off-the-grid for weeks at a stretch for my research, even though it caused worry sometimes.

My committee chair, Dr. John C. Maerz, was the single most influential force behind this dissertation, and I am forever grateful for his guidance and investment in me. John, you are my role model for how to be an effective scientist, scholar, teacher, leader, and friend. Thank you for everything, big and small, that you have done to propel my career. My committee members, Drs.

Byron J. Freeman, Jeffrey Hepinstall-Cymerman, Nik Heynen, and Fausto O. Sarmiento shared key feedback during the course of my dissertation research, for which I am grateful.

I thank Drs. Jana Johnson, H. Bradley Shaffer, Virginia Boucher, and Peter Lahanas for being my most important pre-grad school professional mentors. Despite my concerns to the contrary, you convinced me that I had what it took to be at least an average PhD student. My research was facilitated by the tireless staff at the UGA Science Library, who tracked down key literature in response to my countless interlibrary loan requests. A fantastic community of student colleagues shared much-needed companionship during my time at UGA. To Cyndi

Carter, Chris Cleveland, Becca Cozad, Brian Crawford, Jacob Daly, Jon Hallemeier, David

Haskins, Elizabeth Hincker, James Hunt, Connor Lake, Pearson McGovern, Rebecca McKee,

Chris Murphy, Todd Pierson, Sean Sterrett, Ben Thesing, and Ryan Unks: thanks for making days in Athens almost as good as days in Mexico or California or Fiji! Vanessa Kinney-Terrell

facilitated just about everything, and always with a smile and positive attitude. They are too numerous to name here, but many faculty, postdocs, and students in the ICON program at UGA enriched my education and contributed to my academic growth in exciting ways. Thanks to all for welcoming me into the program.

A veritable army of volunteers and colleagues helped me with Elgaria and Batrachoseps survey work in the mountains of eastern California: Brad Alexander, Amy Chandos, Justin

Clause, Linda Baeza, Nick Buckmaster, Nick Duenas, Jim Erdman, Bob Hansen, Nick Hubeek,

James Hunt, Stevie Kennedy-Gold, Jenn McKenzie, Erin Nordin, Ivan Parr, Dan Smith, Kayla

Smith, Jessie Terry, Ben Thesing, Erin Toffelmier, Cleo Tuday, Gary Wilson, and Tyler Wilson.

I hope that all of you enjoyed it as much as I did.

Special thanks to Levi Gray for bringing Abronia to my attention, for showing me first- hand the wonder, terror, and fun that comes with working in Mexico, and for convincing me that

I could make a dissertation of it. I profited immensely from the support and mentorship of Dr.

Adrián Nieto-Montes de Oca during my studies in Mexico, and Dr. Jonathan Campbell kindly encouraged me to enter the world of Abronia. Israel Solano-Zavaleta, Walter Schmidt-Ballardo,

Gustavo Jiménez-Velázquez, and Karlo Soto-Huerta were influential in the execution of my research in Mexico, and I feel privileged to call you my colleagues. Gracias, mis amigos. Levi

Gray, Carlos Pavón-Vázquez, Peter Scott, Chris Murphy, and Eric Schaad were a dream team on a whirlwind tour of southern Mexico that produced most of the data for Chapter 7 of this dissertation. That round-trip drive from Georgia to New Mexico to the Guatemala border and back was worth every minute. I am indebted to Malcolm Greeley for donating his time on two enjoyable, but Abronia-depauperate, Mexico ventures—someday, Malcolm, we’ll go back down there and see them in the wild!

It is non-hyperbolic to say that my ICON internship was the best internship ever. I am grateful to Robert Fisher, Kim Lovich, and Rajen Reddy for making advance arrangements on my behalf, and to Steve Anstey and the rest of my Likuliku and Malolo family for turning those months in Fiji into a truly life-changing experience. Vinaka vaka levu to all!

My graduate career, and the research presented in this dissertation, was made possible through generous funding provided by a University of Georgia Presidential Fellowship and

Warnell Graduate Assistantships.

And now, James and Connor…let’s go get that new species. It’s the final countdown.

vii TABLE OF CONTENTS

Page

ACKNOWLEDGEMENTS ...... v

CHAPTER

1 INTRODUCTION ...... 1

2 WHAT IS THE BEST AVAILABLE SCIENCE?: CONSERVATION STATUS OF

TWO CALIFORNIA DESERT VERTEBRATES ...... 19

3 RE-SHARPENING THE ENDANGERED SPECIES ACT: EMPIRICAL SUPPORT

FOR THE REGULATORY BAN ON MULTI-SPECIES PETITIONS ...... 85

4 WHOSE FLAGSHIP?: NATURAL HISTORY INFORMS CONSERVATION

PLANNING FOR AN IMPERILED MESOAMERICAN REPTILE CLADE...... 110

5 BREEDING-SEASON HABITAT USE, HABITAT SELECTION, AND

ADAPTABILITY TO DISTURBANCE IN ADULT ARBOREAL ALLIGATOR

LIZARDS, ABRONIA GRAMINEA...... 171

6 DICHOTOMOUS KEY AND CHECKLIST FOR THE ARBOREAL ALLIGATOR

LIZARDS (SQUAMATA: ANGUIDAE: ABRONIA) ...... 203

7 IDENTIFICATION UNCERTAINTY AND PROPOSED BEST-PRACTICES FOR

DOCUMENTING HERPETOFAUNAL GEOGRAPHIC DISTRIBUTIONS, WITH

APPLIED EXAMPLES FROM SOUTHERN MEXICO ...... 273

8 SYNTHESIS AND CONCLUSIONS ...... 329

APPENDICES

viii

A AHURA RESORTS IN-ROOM CONSERVATION COLLATERAL...... 342

CHAPTER 1

INTRODUCTION

The Anthropocene epoch is characterized by a level of human influence on the environment unparalleled in Earth’s history. Defining the precise onset of the Anthropocene is difficult (Crutzen and Steffen 2003, Smith and Zeder 2013), and there is ongoing debate whether this designation will become an official part of the geological timescale (Waters et al. 2014,

Zalasiewicz et al. 2017). Nonetheless, the significance of the Anthropocene on a biological timescale is apparent. More than ever before, humans and their environment are inextricably linked, with the strength of those interactions leading to serious consequences for ecological cycles and biodiversity (Steffen et al. 2007, Lennon 2015). Although ecosystem processes with minimal human influence do exist, and recognition of the Anthropocene need not undermine wilderness protection (Caro et al. 2012), conservation scientists cannot ignore the human element in their work (Corlett 2015, Lennon 2015).

Integrative conservation is a philosophical and practical approach that is ideally suited to conservation in the Anthropocene. As an interdisciplinary field, integrative conservation relies on multiple epistemologies from the social and biological sciences to provide an in-depth understanding of human-environment linkages (Hirsch and Brosius 2013). It embraces complexity and acknowledges that win-win conservation outcomes, although certainly possible, have been overemphasized in the conservation literature (Hirsch et al. 2010). In practice, stakeholder objectives and value systems are often conflicting, which can necessitate hard choices in which all parties do not necessarily benefit (McShane et al. 2011). Through its

1 interdisciplinarity, integrative conservation is a field well-placed to transparently illuminate these complexities, call attention to tradeoffs, and allow for more informed, realistic, and strategic decision-making. Without this plurality of perspectives, conservation interventions risk being one-sided, inequitable, and simplistic—and hence less likely to succeed.

An idealized conservation perspective is that decisions about which species warrant protection, which threats are important, and which actions are appropriate should be based on objective, data-rich science that can populate distribution projection models and population viability analyses. In reality, priority species are chosen for many reasons, including their economic importance to humans, their utility as a surrogate for maintaining ecosystem function, because they are iconic such as charismatic megafauna, because of political issues, or because they are valued as strange or unusual. These judgements are often based on select expert opinion or the will of stakeholders rather than data-rich analyses. This is typically a necessity and should not be viewed as a problem. Sufficient information to generate rigorous models and population viability analyses are unavailable for the vast majority of species of concern, yet management decisions must be made in real time, and action cannot be paralyzed by lack of extensive data.

Thus, at all stages, conservation is often a matter of social, cultural, political, and economic concern that depends on more qualitative expert judgements prone to human idiosyncrasies.

Failure to recognize this human element of applied conservation, much like failure to recognize human-environment linkages, will engender problems for practitioners. Crucially, this reality necessitates periodic interrogation of established expert judgement in light of new data and to investigate possible errors that may have been introduced to the literature. Transdisciplinary analyses and integrative approaches, and a sensitivity to political ecology, are needed in this realm.

In the absence of more extensive data, many decisions regarding imperiled species depend on expert judgement rooted in natural history. The study of natural history is concerned with the lifestyle, ecology, and behavior of organisms, often based on observational rather than experimental evidence. Natural history can thus frequently be grounded heavily in conventional wisdom and the opinions of authorities, rather than data-rich modeling projections. Devaluation of this line of inquiry has been identified in multiple forums (Greene and Losos 1988, Futuyma

1998, Tewksbury et al. 2014, Barrows et al. 2016), based on critiques accusing the field of being outdated and insufficiently rigorous. However, such critiques do not recognize that substantial conservation practice globally is driven by a desire to protect threatened species (Gibbons et al.

2000, Böhm et al. 2013) and forestall a sixth mass extinction (Ceballos et al. 2015, Ceballos et al. 2017). The widespread institutional emphasis placed on imperiled species lists, including the

IUCN Red List of Threatened Species, highlights this political focus on threatened taxa as conservation priorities (Rodrigues et al. 2006), and as surrogates to inform broader conservation planning goals (Hoffmann et al. 2008). For many species of concern, a failure to generate and use or the mischaracterization of baseline natural history information can misguide interventions on behalf of those species and their associated habitats. Although sometimes perceived as marginalized, a thorough understanding of natural history in these contexts thus remains vital

(Greene and Losos 1988, Greene 1994, Bury 2006), and it holds promise as a transdisciplinary way of emphasizing the connections between people and other organisms (Barrows et al. 2016).

Natural history, rather than being an isolated or irrelevant, is a foundational field that interdigitates with numerous disciplines including political ecology, genetics, and physiology, among others.

3 Given this reality, it is imperative that stakeholders concerned with protecting species of interest are clear about why and how they chose those species. It is also vital to rely on the best

available science, which likely necessitates consultation with experts and breaking the

researcher/practitioner divide (Arlettaz et al. 2012, Pietri et al. 2013). Too often, scientific data

relevant to conservation is “hidden” from interested parties in uncirculated government reports or

held by researchers in unpublished formats. This “gray literature” can present a problem for

decision makers, because it may not be easily available or searchable (Meek et al. 2015). In

systems where information is already limited, this data accessibility problem can be especially

harmful, because it can allow spurious perceptions to emerge and persist (Sutherland et al. 2004).

Language barriers can also impede the use of scientific information, particularly in international

systems where research output is often published in a language foreign to in-country

stakeholders. Such barriers create not only a power imbalance, but also can prevent on-the-

ground implementation by many actors who are perhaps most well-positioned to do so. To help

effect change in these areas, there have been repeated calls for improved student training and

skill development (Blickley et al. 2012, Lucas et al. 2017), and promotion of strategic

communication of results to bridge the research-implementation gap (Pietri et al. 2013).

Competing social perspectives toward organisms upheld as conservation priorities can

further complicate policy and forestall program implementation. Flagship species, which are

boundary objects linking social and biological components of conservation practice, are widely

used by non-governmental and governmental organizations as surrogates for broader

preservation agendas (Walpole and Leader-Williams 2002). However, the selection process by

which these actors identify flagships has been questioned (Home et al. 2009), and stakeholders

who interact with the species on-the-ground can have conflicting views regarding the flagships’

value (Douglas and Veríssimo 2013). For instance, deeply-held social fear and negative mythology associated with certain organisms can be a powerful inhibitor of their protection

(Ceríaco 2012). Alternatively, certain sectors of society may place high value on the flagship for consumption or other forms of extractive use, leading to negative outcomes such as poaching

(Auliya et al. 2016). Divergent social viewpoints can even influence the standards associated with scientific knowledge production itself. Perhaps nowhere is this more prominent than in field-based biodiversity science, which continues to rely on reasoned, regulated collection of museum specimens as a core methodology (Remsen 1995, Vonesh et al. 2010, Reynolds and

McDiarmid 2012, Das 2016). This practice is often regarded negatively by public audiences, even to the point of advocacy for physical violence against collectors (Rogers 2014, Johnson

2018). Importantly, museum collecting has also been critiqued by a small but vocal minority of scientists (Minteer et al. 2014, Henen 2016) as being in direct conflict with conservation goals— despite substantial, repeated effort to rebut those claims by broader segments of the scientific community (Krell and Wheeler 2014, Poe and Armijo 2014, Rocha et al. 2014, Hope et al.

2018). None of these conflicting social viewpoints can be simply brushed aside. Although some are arguably driven by ignorance, and hence education is the ultimate solution, others are incommensurate and conservation science practitioners must leave space for dissonance in negotiating these viewpoints in their work.

One group of at-risk species that exemplifies many of these issues, and which I leverage in this dissertation to highlight the continued necessity of devoting research attention to these problems, are the alligator lizards (Squamata: Anguidae: Gerrhonotinae). This ecologically diverse clade, with six recognized genera and 58 described species (The Reptile Database, http://www.reptile-database.org), is distributed from British Columbia, Canada (Lambertz and

Graba 2011) south through the western United States and Nuclear Central America to Chririquí,

Panamá (Lamar et al. 2015). They occupy a broad suite of habitats from arid deserts to coastal scrublands to mesic cloud forests, and they showcase a diversity of terrestrial, arboreal, and rock- dwelling forms. Multiple species and genera within this group are increasingly becoming the focus of conservation attention. These include the Panamint alligator lizard Elgaria panamintina, a species endemic to eastern California, USA that has long been on conservation “watch-lists”

(Jennings and Hayes 1994, Hammerson et al. 2005). This neglected species inhabits some of the most remote, arid landscapes in the western United States, and is enigmatic. Nonetheless, it has received much recent attention due to its proposed listing under one of the most powerful environmental laws in the world: the U.S. Endangered Species Act (Adkins Giese et al. 2012,

Service 2015). Yet, this attention appears somewhat divorced from scientific knowledge, and the literature contains many erroneous statements and uncirculated reports with limited availability to practitioners. Seemingly even more imperiled is the alligator lizard genus Abronia, which comprises 29 species that inhabit high-altitude forests in Mexico, Guatemala, El Salvador, and

Honduras (The Reptile Database, http://www.reptile-database.org). Despite their striking physical appearance, this genus is remarkably understudied. Several species are known to western science from just a single specimen and most species are range-restricted, being distributed across a single mountain range or even a single mountaintop (Campbell and Frost

1993). National and international groups are beginning to promote Abronia as flagships for forest conservation, but the suitability of the genus for such status has never been explicitly analyzed.

Importantly, this flagship narrative is contradicted by divergent social viewpoints and value systems: Abronia are the targets of black-market trafficking for the pet trade (Altherr 2014,

Anonymous 2014, Auliya et al. 2016), and local residents often consider the lizards dangerous

and kill them on sight (Campbell and Frost 1993, Ariano-Sánchez and Torres-Almazán 2010,

Martín-Regalado et al. 2012). Nonetheless, these impacts may be less important compared to other threats facing the genus, which center on the conversion, degradation, and fragmentation of their forest ecosystem (Ariano-Sánchez et al. 2011, Torres-Almazán and Urbina-Aguilar 2011,

Ponce-Reyes et al. 2012; Clause et al. 2016). Unifying these two geographic and taxonomic systems (Elgaria in arid-land California, and Abronia in mesic Central America) is their commonality in illustrating the potential for management progress despite problems that regularly arise in taxa for which idealized, model-driven decision making is not yet achievable.

In this dissertation, I leverage alligator lizards to demonstrate the socioecological opportunities and issues typical of data-poor imperiled species conservation foci, and how such systems differ from the idealized (yet infrequent) data-rich species conservation paradigm. As a thematic unifier, I ask: Why is it important to critique and re-visit expert opinion, and why must political ecology concerns be faced head-on in the context of species-focused conservation?

More specifically, I pursue the following overarching question: How can an organismal natural history perspective help to resolve incommensurate socioecological problems, and appraise perceived threats, related to imperiled species conservation? I sought to answer these questions in the sphere of imminent regulatory decisions and ongoing management interest related to

Elgaria panamintina and the genus Abronia. Does the literature currently overemphasize certain threats relative to the true drivers of these species’ imperilment, and if so, how? Furthermore, can natural history temper contradictory human narratives and value systems that are associated with certain species? As shown herein, I find that natural history does, indeed, advance our understanding of how best to contextualize the status of these species and prioritize limited resources toward their protection. Furthermore, this lesson can likely be extended to other

systems across the animal kingdom. Results pertinent to this theme are presented in Chapters 2, 4 and 5. More broadly, throughout my dissertation I shed light on the social realities of conservation programs devised in the face of imperfect data, and showcase the benefits to be gained from re-visiting longstanding conventional wisdom to improve practitioner behavior.

An additional theme unifying the diverse chapters that follow is strategic communication with non-academic audiences. Chapter 2 is framed in the context of reminding readers of basic scholarly standards when producing scientific work, and showing how conservation can be misdirected when these precepts are not followed. In this same chapter, I use metaphor to make the process of scientific knowledge production more concrete to this work’s intended audience of lawyers, policy makers, and resource managers. Chapter 6 builds further on this idea by considering not only how to strategically package a message, but also the need to pursue multilingual communication in certain cases. Throughout this dissertation, I present recommendations that are designed to translate my findings and stimulate action, while recognizing the complexities of multi-use resource management and competing value systems.

Further elaborating on this topic, I provide an appendix composed of informational materials I developed that contributed to a sustainably-minded Fijian resort being included in the National

Geographic Unique Lodges of the World collection—a somewhat unusual example of a true economic/environmental win-win.

Because human behavior and decision-making are an integral part of conservation, I also devote substantial effort toward exploring imperiled species-oriented policy development and scientific practice. In Chapter 3, I analyze the appropriateness of a recent policy change related to the administration of the US Endangered Species Act. In Chapter 7, I review trends in data production and documentation relating to reptile and amphibian geographic distributions in

southern Mexico—and advocate for a change in practitioner behavior. In particular, I encourage a return to the gold-standard of specimen-based vouchering, to minimize problems of species misidentification and maximize the scientific value of field-based biodiversity science.

Throughout this body of work, I demonstrate that alligator lizard conservation can and must proceed in the face of limited data, yet historical judgements should be viewed with a critical eye and tested or revised when possible. Furthermore, I show that natural history remains a vital lens for rigorously scrutinizing several issues within the sphere of alligator lizard conservation in California and Mesoamerica. Furthermore, strategic communication in this system helps to dissolve barriers between knowledge producers and knowledge consumers, and build consensus around best-practices for generating actionable information. Moreover, human behavior in the context of policy development and decision-making are integral to conservation, particularly that focused on species preservation. The lessons from this study system possess value that extends far beyond its taxonomic and geographic limits, and I expect that many conservation practitioners will find relevant applications to their work on a global stage.

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CHAPTER 2

WHAT IS THE BEST AVAILABLE SCIENCE?: CONSERVATION STATUS OF TWO

CALIFORNIA DESERT VERTEBRATES 1

1 Clause, A.G., Norment, C.J., Cunningham, L., Emmerich, K., Buckmaster, N.G., Nordin, E. and R.W. Hansen. Submitted to Journal of Fish and Wildlife Management, 7 June 2018.

19 Abstract

Scientific progress depends on evidence-based research, and reliance on accurate scholarship is essential when making management decisions for imperiled species. However,

erroneous claims are sometimes perpetuated in the scientific and technical literature, which can

complicate policy and regulatory judgments. The literature associated with two enigmatic

California desert vertebrates, the Panamint alligator lizard Elgaria panamintina and the Inyo

Mountains salamander Batrachoseps campi, exemplifies this problem. We produced a

comprehensive threat analysis and status assessment for these species, which are both under

review for possible listing under the US Endangered Species Act (ESA). Despite uncertainties

and limited data, we find that many sources contain factual errors about the status of these two

species, particularly the original petition that advocated for ESA listing. Although localized

declines may have gone undetected, no evidence exists of population declines, population

extirpation, or population-scale habitat conversion for E. panamintina. However, there is

evidence of recent flash flood damage to some occupied B. campi habitat, which has possibly led

to population declines at those localities. Contrary to inaccurate statements by some authors, all

known populations of both species occur exclusively on federal lands, and numerous populations

have likely benefited from recent federal management targeted at reducing known threats. Of the

12 threats that we identified for one or both species, only three currently appear to be serious:

water diversions, climate change, and flash floods. The remaining threats are neither widespread

nor severe, despite numerous contrary yet poorly supported statements in the literature. We thus

evaluate the contemporary conservation status of both species as relatively secure, although B.

campi is more at-risk compared to E. panamintina. This conclusion is independently supported

20 by a recent review. Nonetheless, ongoing stewardship of these species in a multi-use context by federal agencies remains vital, and we identify several priority management actions and research needs for both species. We also recommend updated determinations on the IUCN Red List, and the Species of Conservation Concern list of the Inyo National Forest. To maximize the quality and effectiveness of conservation planning, we urge government agencies, non-governmental organizations, and individual scientists to maintain high standards of scholarship and decision- making.

Introduction

Science is an incremental, evidence-based process whereby new research builds on earlier work. A common metaphor for this process is that individual works are like bricks, which are progressively assembled into structures that represent bodies of theory and descriptive knowledge (Forscher 1963; Courchamp and Bradshaw 2017). Building good “bricks” and

“structures” depends on proper interpretation of prior research, comprehensive review of relevant sources, synthetic data analysis, and placement of new findings in appropriate context. Failure to adhere to these scholarly principles can misdirect scientific progress. Such issues have motivated several recent reminders of author best-practices (Perry 2016; Anonymous 2017), and the life sciences have not escaped these problems (Grieneisen and Zhang 2012). For example, narratives on the impact of invasive species are sometimes affected by inaccurate or misinterpreted citations (Stromberg et al. 2009; Ricciardi and Ryan 2017), perspectives on predator-mediated trophic cascades may be overly simplistic and even misleading (Kauffman et al. 2010; Marshall et al. 2013; Marris 2014), and poorly documented claims of both species rediscovery and species extinction/extirpation are regularly falsified (Ladle et al. 2009; Roberts et al. 2010; Ladle et al.

2011; Scheffers et al. 2011; Caviedes-Solis et al. 2015; Clause et al. 2018).

Reliance on quality scholarship is especially important when making conservation decisions for imperiled species. It ensures the best possible justification for the decision

(Sutherland et al. 2004), and can increase the legitimacy of the decision among stakeholders

(Pullin and Knight 2009). This precept is codified in the decision-making process associated with what is, arguably, the most far-reaching piece of environmental legislation in the United States: the US Endangered Species Act (ESA 1973, as amended). Administered by the United States

22 Fish and Wildlife Service (USFWS) and the National Marine Fisheries Service, the ESA requires

both agencies to consider only the “best scientific and commercial data available” when making species-listing decisions (ESA 1973). However, the degree to which these agencies follow this standard is variable, due to numerous internal and external challenges (Lowell and Kelly 2016;

Murphy and Weiland 2016). Problems attributed to poorly supported ESA listing petitions have also motivated recent regulatory changes to the process for proposing new additions to the Lists of Endangered and Threatened Wildlife and Plants (USFWS et al. 2016).

In 2012, the US nonprofit Center for Biological Diversity submitted a multi-species petition to the USFWS advocating ESA listing for 53 amphibian and reptile taxa (Adkins Giese

et al. 2012). As required under the ESA, the USFWS subsequently released 90-day findings for

all 53 taxa, which represented the agency’s initial decision on whether the petitioner offered

substantial information in support of listing (USFWS 2015a, 2015b, 2015c; USFWS 2016a,

2016b, 2016c). These 90-day findings concluded that, for 17 of the 53 species, Adkins Giese et

al. (2012) did not present substantial information that the petitioned action (ESA listing) was

warranted. The remaining 36 taxa were advanced to the status review phase for more detailed

examination and public comment. Two of the taxa currently undergoing status review are the

Panamint alligator lizard Elgaria panamintina, and the Inyo Mountains salamander

Batrachoseps campi (Figure 2.1; USFWS 2015c).

These two species are endemic to eastern California, USA, where they are roughly codistributed in the arid mountain ranges of the western Great Basin and northern Mojave deserts

(Banta et al. 1996; Jockusch 2001). These mountains are among the most rugged and inaccessible landscapes in California. They support few paved roads, and their slopes are often incised by steep canyons with multiple waterfalls (Figure 2.2, Figure 2.3). Within the mountains,

E. panamintina and B. campi inhabit similar environs and sometimes occur in syntopy. Occupied microhabitats for both species include mesic riparian zones fed by perennial springs or creeks, and more arid talus slopes or limestone rock crevices far from standing water (Macey and

Papenfuss 1991a, 1991b). Due to their secretive behavior and remote habitats, little is known of these species’ biology and minimal literature has accrued since their discovery in 1954 and 1973, respectively (Stebbins 1958; Marlow et al. 1979). Nonetheless, both species are widely considered imperiled to some degree. The IUCN Red List of Threatened Species categorizes E. panamintina as Vulnerable (Hammerson 2007) and B. campi as Endangered (Hammerson

2004a). They have been designated as Species of Special Concern by the California Department of Fish and Wildlife (CDFW) for over 20 years (Jennings and Hayes 1994), and retained that status following a recent review (Thomson et al. 2016). However, some information that was incorporated into these determinations is inaccurate. Furthermore, many erroneous claims exist in the literature for both species, and substantial field survey data have accumulated since these listings were released. Identifying these errors, and accounting for new data, are especially important given the major regulatory decision that is pending for both species.

Here, we analyze the conservation status of E. panamintina and B. campi, using a dataset collated from white and gray literature, museum records, and contemporary field survey data.

Our objective is to build a comprehensive threat analysis, generate a status assessment, and contrast our findings against outmoded sources and factual inaccuracies in the literature. We conclude by presenting management recommendations and research needs for both E. panamintina and B. campi, and highlight the necessity of ensuring scholarly standards in both technical and peer-reviewed literature.

24 Methods

We reviewed the available literature on both E. panamintina and B. campi using their

common and scientific names (and all synonyms) as search terms in the ISI Web of Science and

Zoological Record databases. We also acquired copies of relevant uncirculated gray literature, in the form of reports prepared for resource management agencies. In addition, we included published sources such as species accounts from books (including field guides), the IUCN Red

List, and NatureServe within our concept of relevant literature despite their often less scholarly nature. Although these reports and other works are usually held in lower regard than peer- reviewed publications, they contain a large proportion of available technical knowledge relevant to our study, and many were cited in the Adkins Giese et al. (2012) listing petition for both species. As such, we consider it imperative to consider these sources in our analysis.

Concurrently with our literature review, we queried the VertNet online portal to create a

database of museum records, supplemented with data obtained directly from relevant museums

(California Academy of Sciences, CAS; Museum of Vertebrate Zoology, MVZ; Florida Museum

of Natural History, UF-Herpetology; and Natural History Museum of Los Angeles County,

LACM). To georeference literature/specimen locality data, evaluate land ownership, and

quantify the presence of roads, we used digital US Geological Survey 7.5-min 1:24,000

topographical maps published in 2012 and 2015, corroborated with the California Atlas &

Gazetteer™ (DeLorme 2015).

We complemented this literature review with field survey data that we collected in 2000–

2001 and 2009–2018. Our survey methodology primarily consisted of visual encounter surveys

completed on foot, but we also road cruised occasionally. We surveyed each locality 1–10+

times, with a total survey effort of 3–100+ person-hours per locality. When on foot, we surveyed riparian vegetation and talus habitats in an attempt to detect surface-active E. panamintina or B. campi, often supplemented by flip-and-replacement of cover objects such as rocks and logs.

During surveys we recorded all threats to either species, which we define as any anthropogenic or non-anthropogenic action or condition known or reasonably likely to negatively affect individuals or their habitat. Our definition of a locality corresponds to individual drainage basins or sub-basins, and every locality that we recognize is at least 1 airline km distant from the nearest portion of any other. At localities represented by point-source springs, we surveyed the length of the available habitat whenever possible. At localities represented by creeks or streams, impassable waterfalls or other barriers often prevented us from viewing habitat in upstream reaches. However, we consider our survey coverage of these long, linear localities sufficient to identify nearly all possible threats. Due to major access constraints higher in the remote, rugged reaches of many canyons, impacts from humans and their attendant infrastructure/animals are usually most intense near the canyon mouth (Figure 2.2). In keeping with these landscape-use patterns, we always covered the lower reaches of the creeks and canyons in our surveys. For all new localities and elevation records for E. panamintina and B. campi discovered during our surveys, we deposited vouchers at the LACM. These vouchers consisted of at least one of the following: whole-body specimen(s), genetic tissue sample(s), and digital photo(s).

After compiling this combined dataset, we first reviewed existing knowledge of the distribution and relevant natural history of each species, to provide appropriate context for evaluating threats. Next, we assessed all threats to these species that we identified during our field surveys or that were mentioned by Adkins Giese et al. (2012). We categorized these threats using the 5-factor analysis used by the USFWS for listing decisions. These are: (Factor A) the

26 present or threatened destruction, modification, or curtailment of the species’ habitat or range;

(Factor B) overutilization for commercial, recreational, scientific, or educational purposes;

(Factor C) disease or predation; (Factor D) the inadequacy of existing regulatory mechanisms; and (Factor E) other natural or manmade factors affecting the species’ continued existence (ESA

1973). Because of the broad nature of Factors A and E, we further divided them into seven and two sub-factors, respectively. In total, we thus identified 12 discrete threats to one or both focal species.

For each of these threats, we ranked its severity on a scale from 0 to 3, with definitions as follows: 0 = not currently affecting known localities, 1 = currently affecting <20% of known localities, 2 = currently affecting 20–50% of known localities, 3 = currently affecting >50% of known localities. We consider the divisions of this ranking scale fine enough to be informative, yet coarse enough to be resilient to changes in threat rankings following the acquisition of new survey data.

Results and Discussion

We surveyed 73% (24/33) of the documented localities for E. panamintina, and 81%

(17/21) of the documented localities for B. campi. These were generally the most logistically accessible localities. We also surveyed 16 additional localities with appropriate riparian/talus habitat that is suspected, but not known, to support one or both species (Figure 2.4). We did not detect either species’ presence at these additional sites, and instead only recorded the incidence of threats. Due to variable search effort and imperfect detection rates for these secretive species, additional surveys may show that they do occur at some of the 16 sites in which neither species was found. In all, we surveyed 63 localities (the numbers given above do not sum to 63 because

E. panamintina and B. campi co-occur at 7 localities). We provide threat scores for E. panamintina and B. campi across all known localities in Table 2.1. Below, we describe our results for both species in three separate sections: geographic distribution, natural history, and threats. In each section we compare our results against claims made in the literature, to clarify discrepancies.

Geographic Distribution

Across the western Great Basin and northern Mojave deserts of California, six named mountain ranges are known to support E. panamintina: the White, Inyo, Nelson, Coso, Argus, and Panamint mountains. In comparison, B. campi has a much more restricted distribution, and is known only from the Inyo Mountains (Figure 2.4). Banta (1965) predicted the occurrence of E. panamintina in three Nevada mountain ranges, Hammerson et al. (2005) included part of Nevada in their range map for the species, and Petersen et al. (2017) considered it “unconfirmed and potentially present” at two Nevada military installations. Nonetheless, no confirmed records exist for E. panamintina in Nevada, despite targeted survey effort in seemingly suitable habitat (J.

Jones, Nevada Department of Wildlife, personal communication). Additionally, range maps by

Hammerson (2004a, 2007) and Jockusch (2001) omit portions of the known distribution of E. panamintina and B. campi, respectively.

Elevation limits documented for E. panamintina range from 1,050 meters (m) (Surprise

Canyon, Panamint Mountains; LACM PC 1738) to 2,330 m (Silver Canyon, White Mountains;

LACM 187140). An imprecise record from Hunter Canyon, Inyo Mountains (LACM PC 2374) suggests that E. panamintina can occur below 600 m, but this remains unconfirmed. Elevation limits documented for B. campi are broader, ranging from 490 m (Hunter Canyon, MVZ

150363–66) to 2,625 m (Lead Canyon, LACM PC 2379). Although other authors present

28 different limits for one or both species (Behler and King 1979, Jockusch 2001, Stebbins 2003,

Mahrdt and Beaman 2009, Adkins Giese et al. 2012, Stebbins and McGinnis 2012), these

sources either rely on outdated citations or do not present any supporting data or vouchers.

Nevertheless, we predict that future surveys will expand the known elevation limits of both

species.

A total of 33 localities are reported for E. panamintina, but nine are unvouchered and

remain unverified (Figure 2.4). Of these 24 vouchered localities, we report eight here for the first

time. Across all 33 localities, ten are in the White Mountains (Dixon 1975; Stebbins 1985;

Cunningham and Emmerich 2001), nine in the Inyo Mountains (Banta 1963; Giuliani 1977;

Stebbins 1985; Macey and Papenfuss 1991b; Banta et al. 1996), one in the Nelson Mountains

(Banta 1963), one in the Coso Mountains (Giuliani 1993), five in the Argus Mountains (Phillips

Brandt Reddick Inc. 1983; Michael Brandman Associates Inc. 1988; LaBerteaux and Garlinger

1998; Morafka et al. 2001), and seven in the Panamint Mountains (Stebbins 1958; Anonymous

1982; Stebbins 1985; Banta et al. 1996; Cunningham and Emmerich 2001; Morafka et al. 2001).

Thirty localities are from Inyo County, and the remaining three are in Mono County. For B. campi, a total of 21 localities are reported, but two remain unvouchered and unverified (Figure

2.4). Of these 19 vouchered localities, we report two here for the first time. Fourteen localities are on the east slope Inyo Mountains, and the remaining seven are on the west slope (Giuliani

1977, 1988, 1990; Clause et al. 2014). All are from Inyo County.

Much confusion exists in the literature regarding the known locality-level distribution of both E. panamintina and B. campi. In the case of E. panamintina, some localities are considered independent by some authors, but actually are nested (e.g., Brewery Spring and Limekiln Spring lie within Surprise Canyon). Other localities are listed as separate, but actually represent a

spatially proximate group of records best represented as a single locality (e.g., the many records from the southwestern CA Highway 168 corridor). Synonymous localities are also treated as different (e.g., Batchelder Spring = Toll House Spring), and some are incorrectly spelled

(Westgard Pass misspelled as Westguard or West Guard), further adding to the confusion. These issues have led to repeated underreporting or overreporting of the true number of localities

(Banta et al. 1996; Hammerson et al. 2005; Hammerson 2007; Mahrdt and Beaman 2009).

Compounding problems are caused by authors overlooking gray literature or citing outdated sources, resulting in further underreporting of the distributions of one or both species (Jennings and Hayes 1994; Jockusch 2001; Hammerson 2004a, 2004b; Adkins Giese et al. 2012, Stebbins and McGinnis 2012). Mislabeled museum specimens have also contributed to one error for B. campi that we correct here. Specimens MVZ 150377–86, listed as originating from Pat Keyes

Canyon and accepted by Clause et al. (2014) as the sole substantiation of that locality, were in fact collected at McElvoy Canyon as shown by a careful reading of Giuliani (1977) and his unpublished 1976 field notes. Our examination of Kay Yanev’s field notes for these specimens also demonstrate that McElvoy Canyon is the correct locality, and that Giuliani was the original collector.

All reported localities for both E. panamintina and B. campi lie entirely on federal land.

Rangewide, E. panamintina occurs on lands managed by the USDA Forest Service, Inyo

National Forest (INF) (12 localities); U.S. Bureau of Land Management (BLM) (10 localities);

National Park Service, Death Valley National Park (DVNP) (8 localities, of which 2 are shared with the BLM); and China Lake Naval Air Weapons Station (CLNAWS) (5 localities). Three of these federal agencies also manage lands that support all reported localities for B. campi: BLM

(13 localities), INF (7 localities), and DVNP (1 locality). These lands are all under minimal or no

development pressure, and they retain their natural character. We are unaware of any private land with habitat that is potentially suitable for E. panamintina or B. campi. This reality contradicts a statement by Jennings and Hayes (1994) that “all except two of the known populations of

Panamint alligator lizard occur on private lands.” This statement was erroneous even in 1994, but has propagated widely across the literature (Hammerson 2007; Mahrdt and Beaman 2009;

Adkins Giese et al. 2012; Stebbins and McGinnis 2012; Thomson et al. 2016). Furthermore,

Adkins Giese et al. (2012) make an implicit claim, without supporting evidence, that at least one population of B. campi is on private land.

The known distribution of both E. panamintina and B. campi solely on public land directly relates to their population health and habitat quality, which have been misinterpreted in the literature. Jennings and Hayes (1994) state that populations of E. panamintina are experiencing “habitat loss,” Adkins Giese et al. (2012) indicate a “decline” in the species, and

Hammerson (2007) states that there is a “probably continuing decline” in both population size and extent and quality of habitat. Similarly, Hammerson (2004a) indicates a “continuing decline” in number of mature individuals and in extent and quality of habitat for B. campi, Evelyn and

Sweet (2012) claim that abundance of B. campi has “likely declined,” Papenfuss and Macey

(1986) assert that spring diversions “likely” led to extirpation of “some populations,” and Adkins

Giese et al. (2012) state that water diversion “causes extirpations” of this species. However, all of these claims are speculative and unsupported by data. Although it is possible that localized declines may have gone undetected, there is no evidence of population declines, population extirpation, or population-scale habitat conversion for E. panamintina anywhere in its range. For

B. campi our survey work indicates that five localities experienced recent habitat loss and

31 possible population declines due to flash floods, although the majority of B. campi localities that

we surveyed have maintained their habitat and consequently, populations.

Natural History

Habitat requirements for E. panamintina and B. campi are an oft-misunderstood aspect of

their natural history. Both species are typically considered narrow habitat specialists found only in microhabitats immediately adjacent to perennial surface water (Stebbins 1958; Marlow et al.

1979; Papenfuss and Macey 1986; Jennings and Hayes 1994; Stebbins 2003; Hammerson et al.

2005). However, this narrative ignores or minimizes both early (Banta 1963; Giuliani 1977) and more recent (Morrison and Hall 1999; Cunningham and Emmerich 2001) data showing much broader ecological tolerances for both species, although far more data exist for E. panamintina.

To date, 12 independent observations exist for E. panamintina in arid rocky habitat 2.7–

6.4 kilometers (km) from perennial surface water or riparian habitat. These observations are

spread across seven localities, and ten are supported by a museum voucher (MVZ 75918, MVZ

150327–29, MVZ 227761, LACM PC 1835, LACM PC 1849, LACM TC 4376–77, UF-

Herpetology 152976). The sightings include a mating pair (Morafka et al. 2001; LACM PC

1849) and a likely gravid adult female (LACM TC 4377), suggesting that at least some

observations reflect the existence of breeding populations in these areas. The recognition that E.

panamintina occupies habitats far from surface water is analogous to our understanding of

habitat use in the closely-related central peninsular alligator lizard, Elgaria velazquezi. Endemic

to the deserts of Baja California Sur, Mexico (Leavitt et al. 2017), E. velazquezi was

hypothesized to occur only at isolated oases (Grismer 1988; Grismer and McGuire 1993) but is

now known from multiple arid, rocky localities several kilometers from the nearest perennial

surface water or riparian zone (Grismer and Hollingsworth 2001).

For B. campi, four specimens (MVZ 190989–92) were collected from an “antifreeze pitfall trap set beside [a] mossy opening in limestone” atop a ridge about 0.4 km from the nearest riparian habitat (Giuliani 1977). Although five additional anecdotal accounts exist for B. campi individuals claimed to be found in nearly identical microhabitats “far from water” in moist ridgetop crevices (Giuliani 1977), these anecdotes are unvouchered and unsubstantiated by biologists. We did not survey the vouchered locality for verification, although we did survey some sites with similar microhabitat characteristics elsewhere but did not detect B. campi.

Despite the pitfall-trapped specimens demonstrating that B. campi can occupy habitat far from flowing water, occupancy rates in suitable, moist, non-riparian microhabitats remain unknown.

We encourage additional survey effort in these non-traditional areas to better resolve this situation, while recognizing that these surveys will likely be challenging because such habitat is rarely found at the surface. Nonetheless, well-documented vouchered records do show that neither E. panamintina nor B. campi is restricted solely to areas with flowing perennial water or riparian vegetation, contrary to earlier stereotypes and recent misstatements by some authors

(Adkins Giese et al. 2012; Stebbins and McGinnis 2012). Although we recognize that the presence of E. panamintina, and particularly B. campi, in these more arid habitats does not necessarily reflect the existence of self-sustaining populations with long-term viability, this uncertainty also applies to many mesic localities where these species occur. More data are needed on the metapopulation dynamics that might influence the suitability and population stability across these two habitat types for both species.

In addition to a lack of information on habitat preference, no rigorous, comprehensive estimates of population size or occupied habitat exist for E. panamintina or B. campi (but see

Larson et al. [1984]). As such, these metrics for evaluating the species’ imperilment remain

33 poorly quantified, and we instead use available data on the presence of suitable habitat, ongoing reproduction, and incidence of threats to infer population health elsewhere in this contribution.

Estimates for E. panamintina by Hammerson et al. (2005) and Hammerson (2007), which are derived from an arbitrary assumption that 50 or more adults exist in each of 20 populations, yielded a total adult population estimate of at least 1,000 individuals. Hammerson et al. (2005) and Hammerson (2007) also provided a similarly coarse estimate of total occupied habitat of less than 5 km2 for E. panamintina, based on the arbitrary assumption of dimensions of 2 km X 0.1 km for each of about 20 occupied habitat patches. For B. campi, Larson et al. (1984) estimated a total effective population size of 14,000 across 12 populations, based on allele frequencies derived from allozyme data first published by Yanev and Wake (1981). Using field survey data from 12 occupied localities, Giuliani (1977) categorized B. campi habitat into four separate bins, reporting 11.9 linear mi of “excellent” and “good” habitat, and 10.3 linear mi of “poor” and

“very poor” habitat. However, this analysis excluded substantial riparian habitat at those localities, due to access difficulties. Papenfuss and Macey (1986) subsequently produced estimates for 13 B. campi localities, 10 of which overlapped with those analyzed by Giuliani

(1977). In their study, Papenfuss and Macey estimated 14.82 ha of “ideal habitat” for B. campi, but they did not specify their criteria for diagnosing such habitat. Adkins Giese et al. (2012) subsequently misinterpreted these studies and did not acknowledge their limitations, citing those works as support for their statement that occupied habitat for B. campi “totals less than 20 ha.”

Threat Overview

Table 2.1 quantifies our assessment of 12 discrete threats of potential concern for E. panamintina and/or B. campi: water diversions, climate change, flash floods, grazing, roads and off-highway vehicles, invasive plants, illegal marijuana cultivation, mining, disease, renewable

energy development, overutilization, and inadequate regulatory mechanisms. Below, we also offer a narrative account of these 12 threats, summarizing available data and contrasting our analysis with assertions in the literature. We particularly concentrate on Adkins Giese et al.

(2012), because they present arguably the most liberal threat assessment for both species. We discuss the 12 threats in roughly decreasing order of severity.

Threat of Water Diversions.— The hydrology of surface flows in the mountain ranges inhabited by E. panamintina and B. campi is not well studied. However, these flows appear to be driven by precipitation, which feeds groundwater cells that discharge as perennial springs above the regional groundwater level (Patchick 1964; Jones 1965; Bedinger and Harrill 2012). We are unaware of any evidence to suggest that ongoing regional groundwater pumping, such as the highly-regulated groundwater withdrawal in the Owens Valley floor (Elmore et al. 2003), is decreasing surface flow at any site occupied by E. panamintina or B. campi.

The threat of water diversion or other anthropogenic change to hydrology is generally mentioned only in passing in the literature associated with E. panamintina (Jennings and Hayes

1994; Hammerson et al. 2005; Adkins Giese et al. 2012; Thomson et al. 2016), and only one source gives a specific example of this threat affecting an occupied locality (Anonymous 1982).

In contrast, Adkins Giese et al. (2012) feature this threat prominently in their discussion of B. campi, but cite Giuliani (1988) for support without recognizing the outdated nature of this source, similar to Hansen and Wake (2005). Giuliani (1988) described substantial degradation due to water diversions and grazing at Barrel Springs in the Inyo Mountains, a locality occupied by both E. panamintina and B. campi. However, our repeated surveys since 2013 indicate that the nearby mine is inactive, all water diversion infrastructure is defunct, grazing impacts are nonexistent, and the riparian zone has regenerated to at least half of the linear extent described

35 by Giuliani (1988). We have observed multiple individuals of both E. panamintina (including a likely gravid female) and B. campi (including juveniles and a gravid female) at this locality on several visits since 2013, indicating the presence of reproducing populations (A. G. Clause and

C. J. Norment, unpublished data).

Cumulatively, our surveys documented active diversion infrastructure at five localities in the White Mountains occupied by E. panamintina, and at five other localities possibly occupied by this species in the White, Argus, and Coso mountains. Stretches of riparian vegetation up to

150 m in length were killed off at two of these sites, seemingly due to lack of sufficient moisture.

Although we did not document active water diversions at any localities occupied by B. campi, we did observe defunct water diversion infrastructure at six B. campi localities (some syntopically occupied by E. panamintina).

Although available evidence indicates that active water diversions are not currently widespread among the known localities for either species, and have decreased in occurrence from historical levels in parallel with a decrease in regional mining activity (discussed subsequently), such diversions still pose a current threat to E. panamintina and a potential future threat to both species. We emphasize the need for ongoing management of this threat throughout the range of both E. panamintina and B. campi, particularly in the context of increased future water demand due to climate change (discussed next).

Threat of Climate Change.— Recent climate models for California generally predict a hotter, wetter future climate statewide, but the direction and magnitude of these predicted changes fluctuates broadly depending on the model (Polade et al. 2017). According to one recent forecast, by 2060 the region inhabited by E. panamintina and B. campi will experience a mean temperature increase of ca. 2–3°C, and a mean precipitation increase of ca. 10–60 mm (Wright et

al. 2016). However, future climate regimes in California could also bring more extreme droughts

(Cook et al. 2015; MacDonald et al. 2016; Swain et al. 2018) and reduced summer monsoon precipitation (Pascale et al. 2017). Although both E. panamintina and B. campi survived a prolonged regional mid-Holocene drought (LaMarche Jr. 1973), any climate-related loss of precipitation-fed riparian habitat would almost certainly be a stressor on populations occupying those habitats. In addition, droughts would likely create pressure to initiate new agricultural and municipal water diversions from these springs and creeks, potentially exacerbating the loss of riparian vegetation. Adkins Giese et al. (2012) briefly mention climate change as a threat to both

E. panamintina and B. campi, but do not discuss climate forecast variability. Few other authors mention climate change as a threat to either species, and most do so only in passing (Hammerson

2004b; Thomson et al. 2016).

In addition to large uncertainties surrounding California’s future climate, it remains unclear how severely the outcome of a hotter, wetter, yet more variable and thus drought-prone climate would affect E. panamintina and B. campi populations. Recent thermal modeling indicates that climate warming will likely depress the activity and energetics of arid-land lizards, but these studies predicted lower climate change-related extinction risk in anguids (which includes all alligator lizards) than most lizard families analyzed (Sinervo et al. 2010, 2017). In contrast, species-specific maximum entropy (Maxent) ecological niche models predict E. panamintina and B. campi to be at high and intermediate risk, respectively, of climate change creating conditions unsuitable for population persistence by 2050 (Wright et al. 2013).

Nonetheless, it is unclear if riparian-zone populations would become extirpated or instead persist at lower population sizes in more restricted patches of non-riparian habitat. It is also unclear if populations inhabiting rocky areas far from riparian zones or standing surface water will become

extirpated under those future climatic conditions, although again those conditions would likely be a strong stressor on those populations, particularly for B. campi.

Ultimately, we consider climate change to be perhaps the greatest potential threat to the long-term persistence of both species, both intrinsically and because it could worsen the stressors of water diversions (discussed previously) and flash floods (discussed next).

Threat of Flash Floods.—Beaty (1963) suggested that flash floods occur regularly in the White

Mountains, caused primarily by localized summer thunderstorms. Dramatic re-sculpturing of canyon topography and severe destruction to riparian vegetation are typical results. Across the ranges of E. panamintina and B. campi, forecasted wetter climate regimes in California could exacerbate the frequency and/or severity of these flash floods in the future (Modrick and

Georgakakos 2015; Polade et al. 2017; Swain et al. 2018), but see Pascale et al. (2017) for alternative predictions. Hansen and Wake (2005) and Adkins Giese et al. (2012) discuss the threat of flash floods to B. campi, but only Cunningham (2010) mentions this threat for E. panamintina. Based on our surveys and several published sources (Giuliani 1990; Hansen and

Wake 2005; Cunningham 2010), over the last 30 years at least seven thunderstorms have caused flash floods across eight Inyo Mountains localities occupied by E. panamintina and/or B. campi, plus an additional locality occupied by E. panamintina in the Panamint Mountains. Additional flash floods, throughout the range of both species, likely remained undocumented during that period. However, based on our surveys and those of Giuliani (1996), following a documented flash flood E. panamintina has persisted at every locality and B. campi has persisted at most.

There are three localities at which B. campi has not been detected in post-flood resurveys: the south fork of Union Wash (although E. panamintina has persisted there), Waucoba Canyon, and the middle fork of Willow Creek (C. J. Norment, unpublished data). Although recent flash floods

have damaged or destroyed known habitat for B. campi and E. panamintina, our surveys also suggest that many known localities of both species are likely insulated from this threat because occupied habitat lies in side-canyon drainages too small to capture enough rainfall for scouring to occur. Furthermore, because heavy rainfall is a known behavioral cue for many organisms, including stream abandonment behavior in a few invertebrate taxa (Lytle 1999), E. panamintina and B. campi could possess behavioral mechanisms to help them escape flash floods. Regardless of possible mechanisms, ultimately flash floods represent a natural disturbance regime that both species have withstood for millennia. However, flooding could certainly act as a driver of local extinctions in a metapopulation dynamic, and we hypothesize that more frequent or extreme flash floods might exceed the recolonization or demographic capacity of both species to respond to this stressor in the future.

Threat of Grazing.—Feral burros and feral horses have populated much of California’s desert wildlands for decades, primarily a legacy of abandoned stock associated with historic settlers and miners (Weaver 1974). The negative effects and widespread distribution of feral burros in DVNP were discussed by Sanchez (1974), and Giuliani (1977) subsequently documented extensive damage by feral burros to riparian zones at multiple B. campi localities in the adjacent east slope

Inyo Mountains. Surveys at the E. panamintina-occupied Haiwee Spring in the Coso Mountains reported it as suffering “heavy” and “concentrated” use by feral burros (Woodward and

McDonald 1979), and Giuliani (1993) later reported “over-grazing” by cattle and continued presence of feral burros at this locality. A review by Kauffman and Krueger (1984) demonstrated that intense grazing by non-native ungulates typically causes direct loss of riparian vegetation cover due to browsing, breaking, and trampling, accompanied by compaction and erosion of soils. Jones (1981) correlated these structural habitat changes with reduced lizard community

abundance and diversity in Arizona. Reinsche (2008) subsequently reviewed additional studies that variously resolved both positive and negative effects of grazing on several lizard assemblages in arid and semi-arid landscapes. Although none of these studies involved alligator lizards or salamanders, we consider it reasonable that heavy grazing pressure is likely not beneficial to either E. panamintina or B. campi due to negative effects such as reduced vegetative cover, disturbance of microsites, and contamination of water sources.

Importantly, contemporary grazing severity at most localities for both species is reduced from historical levels, due to major removal efforts by federal land managers. From 1979–1981, the BLM removed over 1,500 feral burros from the east-slope Inyo Mountains (Papenfuss and

Macey 1986). From the 1980s to 2005, the Navy removed 9,500 feral burros and 3,280 feral horses from CLNAWS lands. Navy removals are ongoing, to fulfill the CLNAWS

Comprehensive Land Use Management Plan objectives of eliminating feral burros and maintaining a cumulative feral horse herd of 170 animals (U.S. Navy and Bureau of Land

Management 2005). The BLM cooperates with the Navy in this effort, and has removed hundreds of additional feral ungulates from adjacent BLM lands known to support E. panamintina. Moreover, DVNP has engaged in control of feral ungulates since 1939 (Sanchez

1974). The Park Service has removed hundreds of burros from within DVNP; cooperatively implements burro control on adjacent BLM lands; has a long-term management goal of zero burros within the park; and plans to retire cattle from the Hunter Mountain allotment, which supports a known E. panamintina locality (National Park Service 2002).

The effects of these control efforts have been dramatic in many areas, although feral ungulates are far from being completely eradicated from the range of E. panamintina or B. campi. Our surveys documented grazing damage to riparian habitat at only two E. panamintina

localities: one each in the Argus and Nelson mountains. Elsewhere in the Argus Mountains, on land managed by the CLNAWS and BLM, surveys by LaBerteaux and Garlinger (1998) indicated “low” or “moderate” feral burro grazing impacts at four additional E. panamintina localities that we did not survey. Nonetheless, anecdotal evidence suggests that grazing impacts could remain high in parts of the Argus, Nelson, Coso and Panamint mountains, which were comparatively under-represented in our recent surveys for E. panamintina. Ongoing removal efforts in these three ranges might be below annual recruitment rates, suggesting that populations of feral burros and perhaps feral horses could be on the rise in these areas (Tom Campbell,

CLNAWS, personal communication). Moreover, funding constraints and deep-seated political controversy (e.g., Animal Welfare Institute [2012]) complicate the long-term management or eradication of feral ungulates (Crowley et al. 2017). Ultimately, current data on grazing severity are unavailable for many localities, particularly for E. panamintina. Nonetheless, our surveys found no evidence of feral or domestic ungulate grazing at any B. campi locality, and we consider it unlikely that this threat would cover a large portion of the species’ range due to the many inaccessible locations it occupies.

Adkins Giese et al. (2012) largely overlook data that indicate recent but variable reductions in non-native grazing animals on rangelands. Instead, they cite outdated secondary sources (Papenfuss and Macey 1986; Jennings and Hayes 1994) to support their claims that grazing is a major contemporary threat to both E. panamintina and B. campi. Adkins Giese et al.

(2012) also incorrectly cite a third source (Mahrdt and Beaman 2002) by claiming that overgrazing “is” a threat to E. panamintina when Mahrdt and Beaman (2002) indicate only that it

“could” be a threat.

Threat of Roads and Off-Highway Vehicles.— Neither roads nor off-highway vehicles (OHV) are mentioned in the literature as a threat to B. campi, save for an unsubstantiated claim by

Evelyn and Sweet (2012) that “road construction” is a likely contributor to declines. However,

Adkins Giese et al. (2012) make several erroneous statements about the threat these factors pose to E. panamintina. They overlook contrary evidence to claim that OHV use “has increased significantly” in the Panamint Mountains, and that road “construction” threatens the species. For both claims, they cite only Mahrdt and Beaman (2002) for support, despite that source’s outdated nature and lack of supporting documentation. Adkins Giese et al. (2012) also mischaracterize a statement by Mahrdt and Beaman (2002), claiming that vehicular traffic “threatens lizard populations” when their source says only that it “could threaten lizard populations.”

Available evidence indicates that roads do pose an ongoing threat to E. panamintina, but no threat to B. campi. A two-lane paved road parallels or bisects occupied habitat at four known

E. panamintina localities. At one of these localities, multiple road-killed E. panamintina have been documented (Morrison and Hall 1999; Cunningham and Emmerich 2001; specimens UF-

Herpetology 152976 and LACM 189186). However, the sole patch of riparian vegetation (which the road bisects) and nearby roadside talus still consistently yield detections of this species 43 years after their discovery there (Dixon 1975). Furthermore, although data are limited, there is no indication of a decline in detection probability; our annual surveys of the spring-fed riparian habitat since 2013 have documented over two dozen individual lizards, about one-quarter of which were juveniles (A. G. Clause, unpublished data). Elsewhere in the range of E. panamintina, dirt access roads regularly approach the mouths of occupied canyons, but only at seven localities do dirt roads parallel and/or bisect riparian or talus habitat. Although grading and widening of three of these dirt roads in 2012 damaged riparian plants (Klingler 2015), our

42 surveys indicate that much of the vegetation has since recovered. For B. campi, no paved road exists within 3 km of occupied habitat. Furthermore, only at four localities does a dirt road

approach within 2 km of occupied habitat, and those roads never reach riparian zones inhabited

by B. campi. At one locality (Barrel Springs), an old dirt road that closely approached occupied

riparian habitat is now completely impassable to vehicles due to intentional placement of

boulders in the roadcut.

The related threat of OHV use is even less consequential to E. panamintina and again a

non-threat to B. campi. Many canyons where these species occur have multiple steep, bedrock

waterfalls (Giuliani 1977) that restrict OHV passage (Figure 2.3). Except for one locality in the

White Mountains (Redding Canyon), our surveys did not document evidence of unauthorized

OHV use in or along riparian habitats occupied by E. panamintina. Contrary to statements made

by Adkins-Giese et al. (2012), the severity of this threat has been much reduced from historical

levels, due to targeted efforts by federal land managers. Over 15 years ago, the BLM prohibited

all vehicular travel at the E. panamintina type locality in the Panamint Mountains (BLM 2001).

Our surveys show that this canyon’s riparian zone has regenerated substantially in the absence of

vehicular traffic, reclaiming much of the former dirt road that was a popular site for OHV

enthusiasts. For B. campi, we are unaware of any OHV use at a known locality.

Threat of Invasive Plants.— The only non-native plant mentioned in the literature, or that we

identified during our surveys, as a threat to E. panamintina or B. campi is saltcedar or tamarisk,

Tamarix spp. These shrubs or small trees can form dense monoculture stands in invaded riparian

areas (Di Tomaso 1998); they have variable, but sometimes high, evapotranspiration rates that

can potentially reduce surface water availability (Cleverly 2013; Nagler and Glenn 2013); and

they are often correlated with elevated salinity levels in soil and groundwater, although causation

has rarely been demonstrated (Ohrtman and Lair 2013). Research into the effect of Tamarix on lizard communities in the arid southwestern U.S. was reviewed by Bateman et al. (2013), and although no study involves alligator lizards, available research generally reveals a pattern of reduced lizard diversity and abundance in Tamarix stands relative to uninvaded riparian habitat.

We are unaware of any studies exploring the effect of Tamarix on salamanders, but we infer that reduced surface water availability and elevated salinity levels would likely negatively affect B. campi.

LaBerteaux and Garlinger (1998), documented Tamarix at two known E. panamintina localities in the Argus Mountains, but noted that the plants were highly localized across the riparian habitat. DeDecker (1991) indicated a “widespread infestation” of Tamarix in low- elevation reaches of the west slope White and Inyo Mountains, where the plant had become a

“serious threat to springs and seeps.” Adkins Giese et al. (2012) considered Tamarix a threat to

E. panamintina, and cited DeDecker (1991) and Mahrdt and Beaman (2002) to support their position. However, Mahrdt and Beaman (2002) only paraphrase statements from DeDecker

(1991) for support and present no novel data. In a subsequent work, Mahrdt and Beaman (2009) again mentioned invasive plants and Tamarix as a possible threat to E. panamintina without offering supporting evidence. In contrast, to our knowledge no published source identifies invasive plants or Tamarix as a possible threat to B. campi.

Although it was likely more abundant in the region historically as indicated by DeDecker

(1991), our survey data indicate that Tamarix is currently neither a widespread nor severe threat to E. panamintina or B. campi, although it is a greater threat to the latter species. Our surveys documented Tamarix at ten localities in the Inyo Mountains, of which three were occupied by E. panamintina and seven occupied by B. campi. Of these ten localities, four support < 20 plants

and appear to be in an early stage of colonization, three support established populations that were recently treated mechanically and chemically by the BLM with some success, and two support plants only at the canyon mouth far from habitat occupied by either species. Elsewhere within the range of E. panamintina, our surveys documented Tamarix at only one additional locality, where the plants were present low in the canyon far from occupied habitat. Cumulatively, there is little evidence that Tamarix or other invasive plants currently pose a substantial threat to E. panamintina or B. campi. This reality is attributable, in large part, to decades of Tamarix control efforts by multiple federal agencies. Nonetheless, without concerted management this threat could worsen in the near future given the capacity of Tamarix to colonize and spread.

Threat of Illegal Marijuana Cultivation.—No literature source mentions marijuana grows as a threat to either E. panamintina or B. campi. However, since 2014 our surveys revealed three recently destroyed or abandoned marijuana grows in remote canyons: one at an E. panamintina locality in the east slope Argus Mountains, one at a B. campi locality in the east slope Inyo

Mountains, and one in the Inyo Mountains at a locality that could support one or both species. At the grow site in the Argus Mountains, we observed chopping damage to mature willows, terracing of the slopes immediately adjacent to the riparian zone, compaction of leaf litter, and defunct water diversion driplines, with these impacts covering a 2-hectare area. Additional negative effects, such as other forms of streamflow diversion (Bauer et al. 2015) along with water and soil contamination from pesticide/herbicide application, are also probable. Installation of similar grows elsewhere is a future threat to the riparian habitat of E. panamintina and B. campi, particularly in isolated canyons otherwise exposed to minimal direct human activities, as has been found elsewhere in California wildlands (Butsic and Brenner 2016). Although we caution that clandestine activities such as illegal marijuana cultivation are inherently challenging

45 to quantify, which complicates any assessment of their prevalence or severity, this threat

warrants ongoing management attention.

Threat of Mining.— Knopf (1912) described a widespread decline in mining activity across the

Inyo and White mountains beginning in the late 19th century. A review of Inyo Mountains mineral resources by McKee et al. (1985) indicated the general continuation of this pattern, albeit

with periodic spikes in mining activity corresponding to rises in gold prices. Papenfuss and

Macey (1986) subsequently reported 361 mining claims in the Inyo Mountains “filed in and

around 13 canyons where [B. campi] is found,” some of which also support E. panamintina.

However, these authors did not define the phrase “in and around,” nor did they indicate which mining claims were active, inactive, or not yet acted upon. Adkins Giese et al. (2012) list mining as a threat to both E. panamintina and B. campi, but they provide no examples to support their assertions nor do they acknowledge the limitations of the Papenfuss and Macey (1986) source, which they cite in their discussion of B. campi. Other authors (Hammerson 2004b, Hansen and

Wake 2005) cite Papenfuss and Macey (1986) in a similar fashion, overlooking its outdated

nature, particularly in the case of B. campi, because all known populations of the species now

occur either within DVNP or the Inyo Mountains Wilderness, which was created in 1994.

Although wilderness designation offers some protection for at-risk species and their habitats,

valid mining claims existing prior to 1 January 1984 can legally be exploited; permitted activities

include “where essential the use of mechanized ground…equipment” and “use of land

for…waterlines” (Legal Information Institute undated).

In keeping with the general decline in mining noted by Knopf (1912) and McKee et al.

(1985), our surveys suggest that mining has continued to decline across the known range of both

species, and is not currently affecting the habitat of either. Although the large footprint of the

active Briggs gold mine in the Panamint Mountains lies adjacent to riparian habitat that might be occupied by E. panamintina, our surveys revealed no active mines within 0.8 km of known E. panamintina or B. campi habitat—only abandoned ones. We regularly documented old mining- related debris among riparian habitat occupied by both species but we never observed water flowing from abandoned mines or mine tailings, suggesting minimal water pollution by mining- related contaminants. Nonetheless, legacy effects of mining in the region have not been well- studied. Despite the apparent regional decline in mining activity, it has not ceased completely and economic shifts in the supply/demand of gold, silver, and other minerals could increase regional mining pressures in the future. For instance, an application to re-open the Robbie Hoyt

Memorial Mine at a known E. panamintina locality in the White Mountains, which proposes widening a dirt road that currently impinges on occupied riparian habitat, was recently submitted for review (Inyo National Forest 2017a). Furthermore, a controversial application for a large gold mine on the Conglomerate Mesa, Inyo Mountains, which was later formally withdrawn

(Timberline Resources Corporation 2008), has also been recently re-opened (Silver Standard

Resources Inc. 2016). Ongoing vigilance by land managers against potential future mining threats remains essential.

Threat of Disease.—No authors claim that disease threatens E. panamintina or B. campi, and no documentation exists of wild individuals of either species showing outward signs of ill health.

Nevertheless, the future and possible current threat posed to B. campi from the disease chytridiomycosis, which is caused by the fungi Batrachochytrium salamandrivorans (Bsal), and

B. dendrobatidis (Bd), warrants consideration.

Due to its recent discovery (Martel et al. 2013), Bsal remains poorly studied. Although it has not yet been documented in North America, Bsal has devastated salamander populations in

47 northern Europe (Yap et al. 2017). In the region where B. campi occurs, spatial models predict

low to moderate habitat suitability for Bsal (Yap et al. 2015), and low to moderate salamander

vulnerability to Bsal (Richgels et al. 2016), although the latter result could be an artifact of low

salamander species diversity in the region. In comparison, Bd is better studied and has been

correlated with enigmatic declines in terrestrial plethodontid salamanders (Cheng et al. 2011).

Nonetheless, field and laboratory studies indicate highly variable Bd infection rates among this

group of salamanders, to which B. campi belongs (Van Rooij et al. 2011; Moffitt et al. 2015;

Mendoza-Almeralla et al. 2016). One study of Batrachoseps attenuatus revealed evidence of

mixed susceptibility to Bd and no evidence of measurable declines in wild populations

(Weinstein 2009), while a retrospective analysis of three species of Batrachoseps from insular

populations revealed consistently low prevalence of Bd infection (Yap et al. 2016). Conversely, a

retrospective study of B. attenuatus negatively correlated modern-day population persistence

with time to first detection of Bd infection (Sette et al. 2015). The sister species to B. campi (B.

wrighti; Jockusch et al., 2015) is known to be capable of infection based on a single Bd-positive

specimen (Weinstein, 2009), but no other information relating to chytridiomycosis susceptibility

exists for the Plethopsis subgenus of Batrachoseps, which includes B. campi. The deep

evolutionary divergence of Plethopsis (ca. 40 MYA; Shen et al. 2016) coupled with the

ecological extremes inhabited by its component species (Jockusch and Wake 2002) could limit

accurate inference about the effects of chytridiomycosis on B. campi populations using data from

non-Plethopsis congeners. Thus, Plethopsis-specific chytridiomycosis research is needed to

evaluate the threat this disease might pose to B. campi.

Threat of Renewable Energy Development.—Although the literature for E. panamintina and B.

campi does not mention renewable energy development as a stressor, we consider it worthy of

management attention for E. panamintina. Utility-scale solar projects have been proposed by the

Los Angeles Department of Water and Power in the Owens Valley at the base of the Inyo

Mountains, and although these proposals were later withdrawn or cancelled (Manzanar

Committee 2015), if reopened in the future such projects could impact the lower edge of E. panamintina habitat on upper alluvial fans at canyon mouths. Furthermore, a major geothermal energy development project in the Coso Mountains is less than 10 km from a known E. panamintina locality, and potentially suitable talus habitat exists within the project footprint.

However, this habitat has never been surveyed for the species, and the draft Environmental

Impact Statement did not consider possible impacts to E. panamintina (BLM 2012). Despite a lack of evidence that E. panamintina is currently being affected by this or any other energy infrastructure, renewable energy is a growth industry in the California deserts (CA Senate Bill

No. 2 2011; California Energy Commission 2014) and thus warrants ongoing attention as a potential future threat.

Threat of Overutilization.—The issue of overutilization has received attention in the literature for

E. panamintina and B. campi, but this attention has been speculative. For E. panamintina,

Mahrdt and Beaman (2002, 2009) indicated that illegal collecting “may” threaten populations, and Adkins Giese et al. (2012) cited the former source as the sole support for their claim that illegal collection “likely threatens populations” of the species. Giuliani (1977, 1988, 1990) expressed concern about collector-driven disturbance to B. campi populations, but he did not specify the basis for those concerns.

To our knowledge, no major hobbyist market exists for any species of Batrachoseps or

Elgaria. We are also unaware of any evidence that overutilization is a population-level threat to

E. panamintina or B. campi. Furthermore, both species are likely inherently resistant to

overutilization due to their secretive life histories, generally low detection rates (Giuliani 1977,

1996; C. J. Norment, unpublished data), and occupancy of remote, rugged habitats (Figure 2.1,

Figure 2.2). Reported scientific whole-body collection of E. panamintina amounts to fewer than

50 specimens spread across 16 localities over a period of 60+ years. Similarly, reported whole- body collection of B. campi sums to fewer than 200 individuals across 17 localities over a period of 40+ years. Collecting at these levels over such extended time periods is unlikely to have an appreciable effect on population persistence (Dubois and Nemésio 2007; Krell and Wheeler

2014; Poe and Armijo 2014; Rocha et al. 2014; Hope et al. 2018). Furthermore, current lethal scientific collection of both species is strictly regulated, and generally allowed only when documenting a new locality (L. Patterson, CDFW, personal communication). Nonetheless, quantifying the magnitude of legal and illegal wildlife trade is challenging, and can be prone to underestimation (Salzberg 1996; Schlaepfer et al. 2005).

Our surveys documented no evidence of collector-driven habitat disturbance, and we encountered possible collecting equipment only twice: two plywood boards (one since removed by unknown person[s]) at an occupied E. panamintina locality, and wood roofing shingles at a locality occupied by both E. panamintina and B. campi. Importantly, none of these cover objects showed signs of recent disturbance during our repeated surveys, suggesting a lack of regular visitation. Furthermore, these two localities represent the most easily-accessible sites for these species. For this reason, these sites would potentially be those most strongly affected by illegal collecting pressure; yet, they support populations that repeatedly yielded captures of multiple individuals, including juveniles, during our surveys (A. G. Clause and C. J. Norment, unpublished data).

50 Threat of Inadequate Regulatory Mechanisms.—Several sources mention the existing state and

federal regulatory mechanisms that cover E. panamintina and B. campi in California, in

generally positive terms (e.g., Hansen and Wake 2005; Thomson et al. 2016). However, Adkins

Giese et al. (2012) downplay these government protections, characterizing them as “insufficient” and “inadequate.” Importantly, Adkins Giese et al. (2012) do not fully acknowledge the role

played by these regulations in past management actions implemented specifically to mitigate

several threats to E. panamintina and B. campi. As discussed previously, existing legal

protections have directly motivated beneficial interventions on behalf of both species in recent

decades, resulting in decreased levels of grazing, OHV use, and invasive Tamarix spp. in

occupied habitat.

At the state level, under California Code of Regulations Title 14 Sections 5.05 and 5.60,

there is a zero bag limit for E. panamintina and B. campi under sportfishing regulations, making

it illegal to collect either species for recreational purposes. A recreational collecting moratorium

also exists for all Batrachoseps salamanders in Inyo County. Moreover, it is illegal to

commercially collect either E. panamintina or B. campi in California unless a biological supply

house obtains specific authorization from the California Department of Fish and Wildlife

(CDFW) to collect them for sale to bona fide scientific and educational facilities—an unlikely

scenario for these species (L. Patterson, CDFW, personal communication). For over 20 years, E.

panamintina and B. campi have also been administratively designated as Species of Special

Concern by CDFW. This designation is intended to direct research and management toward

enigmatic but likely imperiled species as a means to prevent more stringent future listing, but it

does not directly regulate destruction of habitats or individuals (Jennings and Hayes 1994;

Thomson et al. 2016). Also, the State of California has some jurisdiction over water diversions

on federal land through the California Environmental Quality Act (CEQA). Under this statute, surface water cannot be legally diverted without a state permit; applications to divert are made though the California Water Resources Control Board, which would require a CEQA analysis (S.

Parmenter, CDFW, personal communication). This process could provide an additional level of regulatory protection for E. panamintina and B. campi, as long as the State of California has the political will to prevent water diversions in sensitive habitat.

At the federal level, both species are currently designated as Sensitive by the BLM (BLM

2006), and Species of Conservation Concern by the USDA Forest Service (under FSM 2670).

The BLM is mandated to manage Sensitive species and their habitat in a multi-use context, by minimizing threats affecting the species and improving habitat, where applicable (BLM 2008).

The USDA Forest Service is mandated to develop and implement management objectives for

Species of Conservation Concern and their habitat. This management is designed to ensure that the species maintain viable populations on Forest Service lands, and do not become threatened or endangered due to Forest Service actions (see FSM 2670). However, in their ongoing Forest Plan revision, the Inyo National Forest proposed to exclude E. panamintina from their Species of

Conservation Concern list (Inyo National Forest 2017b), which would decrease management attention for over one-third of the species’ known populations.

Legal protection of E. panamintina, B. campi, and their habitats would be strengthened by ESA listing and would help compensate for possible lax enforcement or even the repeal of existing protections in the future. Nevertheless, historical regulatory protections have clearly improved the status of both species, and there is no signal to suggest that similarly beneficial management will cease in the near future.

52 Conclusion and Recommendations

Although scientific information on E. panamintina and B. campi is limited, our literature review and threat analysis showsthat available information often contradicts or highlights uncertainty associated with historical and contemporary literature for both species. Although many knowledge gaps remain and additional data are needed, we consider the current conservation status of E. panamintina and B. campi to be relatively secure, although B. campi is comparatively more imperiled due to flash food impacts on habitat and populations. It is true that both species are California endemics known from comparatively few localities, which places them at increased risk for local stochastic extinction; however, all known populations occur exclusively on federally managed land that retains or is recovering its natural character under existing regulatory mechanisms and associated management efforts. Although it is possible that localized declines may have gone undetected, no evidence exists of population declines, population extirpation, or large-scale habitat conversion for E. panamintina, and habitat loss and possible population declines are documented at only five sites for B. campi, all due to flash floods.

Of the 12 threats to E. panamintina and/or B. campi that we identified in our review, only three appear to be currently important: water diversions, climate change, and flash floods.

Available data indicate that water diversions actively threaten multiple populations of E. panamintina and formerly threatened some B. campi populations. Pressure to initiate new diversions might increase in the future under predicted climate change scenarios, or if federal regulations are relaxed regarding wilderness area protection or mining claim development.

Shrinking riparian zones under a hotter, more drought-prone predicted climate is also a concern.

Additionally, if climate change causes more severe and/or frequent summer thunderstorms, the

resulting destructive flash floods could exceed the adaptive capacity of populations of both species to deal with this stressor, and flash floods seem to have caused recent declines at some B. campi localities. However, substantial uncertainty exists in the regional climate forecasts across the range of both species. Available evidence (albeit more substantial for E. panamintina than B. campi) suggests that populations can persist in more arid, rocky habitat far from canyon-bottom riparian zones, although such non-riparian habitat may be rarely occupied by B. campi and may support smaller, less stable populations compared to riparian habitat.

Based on available data, the nine remaining threats to both species (grazing, roads and

OHV use, invasive plants, illegal marijuana cultivation, mining, disease, renewable energy development, overutilization, and insufficient regulatory protection) are neither widespread nor severe at this time. Threats due to mining have declined from historical levels, and threats due to grazing, OHV use, and invasive Tamarix spp. have been reduced from historical levels through targeted action by federal land managers. However, we emphasize that these nine threats could become more severe in the future, in part due to potential changes in the political and regulatory landscape. Ongoing stewardship by resource managers is necessary to appropriately safeguard populations of E. panamintina and B. campi.

Broadly, the results of our threat assessment are independently supported. Using a rigorous, transparent eight-metric risk assessment framework, Thomson et al. (2016) identified and scored 45 imperiled yet non-ESA-listed amphibian and reptile taxa across their entire known distribution in the United States (all in California). The cumulative threat scores for E. panamintina and B. campi ranked low (28th and 19th, respectively) within this cohort of taxa, corroborating our assessment of the comparatively secure conservation status of these two species. Remarkably, of the 15 most threatened taxa identified by Thomson et al. (2016), only

54 four were included in the Adkins Giese et al. (2012) ESA petition. This discrepancy alludes to what we perceive as a disconnect between scientific knowledge and the species selection process used by Adkins Giese et al. (2012). Although Thomson et al. (2016) was published four years after Adkins Giese et al. (2012), most of the same data and sources were available to and used by both sets of authors.

A similar, although less pronounced, disconnect also exists between current scientific knowledge and the IUCN Red List categorizations for E. panamintina and B. campi. The categorization for both species hinges on judgments that in some cases are weakly supported by contemporary data, an unsurprising result given that these species accounts were generated over a decade ago (Hammerson 2004a, 2007). For E. panamintina, the Vulnerable categorization rests on the assumption that there is continuing decline in the number of mature individuals, and that no subpopulation exceeds 1,000 mature individuals, fulfilling criterion C2a(i) (Hammerson

2007). The categorization of B. campi as Endangered rests on the assumption that there are continuing declines in habitat extent, habitat quality, and number of mature individuals, fulfilling criteria B1b(iii,v) and B2b(iii,v) (Hammerson 2004a). Appropriately, Hammerson (2004a, 2007) qualifies his judgments throughout, and makes it clear that they are inferences or projections.

However, no rigorous, comprehensive estimates of historical or contemporary population size, or number of mature individuals, exist for either species. Nor is there clear evidence of widespread, critical declines in habitat extent or quality, particularly for E. panamintina. Furthermore, based on the available evidence, inferences or projections that widespread population declines have occurred are weak, for three reasons. First, all contemporary resurveys of historical localities have produced detections (with three exceptions for B. campi), a result that would be unexpected if widespread declines had occurred. Second, although available data are limited, there is no

strong signal of decreasing detections at any locality for E. panamintina, although B. campi detections have declined at 5 of 14 localities for which we have resurvey data (including the three that lack detections during resurveys). Third, at localities with perhaps the most severe human impacts (Barrel Springs and CA Highway 168), repeated detection of juveniles indicates successful ongoing reproduction in these populations. As such, despite lacking comprehensive rangewide data, available information suggests that populations of both species are likely relatively stable overall, albeit with possible population declines in some B. campi populations for which we have resurvey data, due to recent flash floods. We thus re-evaluate both E. panamintina and B. campi as Near Threatened on the IUCN Red List (IUCN Red List Criteria

Version 3.1), downlistings that we consider a positive development (see Mallon and Jackson

[2017]).

Independent of their IUCN listings, we expect that the existing status of both E. panamintina and B. campi as CDFW Species of Special Concern and BLM Sensitive Species will motivate continued attention to their protection. Available evidence indicates that these status listings are warranted, and we suggest that they remain unchanged. We similarly advocate for the continued inclusion of B. campi on the INF’s Species of Conservation Concern list, and strongly recommend reconsideration of the proposed exclusion of E. panamintina from that list

(Inyo National Forest 2017b). A recent analysis of the conservation status of E. panamintina on

INF lands evaluated the species as being of “high concern” or “some concern” across all eight threat categories assessed (Evelyn and Sweet 2012). Our surveys indicate that the 12 known localities for E. panamintina on INF lands include some of the most at-risk populations rangewide. Half are affected by roads that bisect riparian or talus habitat, almost half are affected by ongoing water diversions, and two localities (Barrel Springs and CA Highway 168) are

perhaps the most strongly human-altered of any known site. Furthermore, the localities that drain into the Owens and Chalfant valleys are among the most vulnerable to pressures from increased water diversions, due to the agricultural, ranching, and housing development in those valleys.

For these reasons, we consider it imperative that E. panamintina be included on the INF’s

Sensitive Species list, to promote continued management efforts by INF that will help forestall any need for more stringent listing status in the future.

All federal agencies that support populations of both E. panamintina and B. campi face challenges managing for these species in the context of multi-purpose land use and a changing political environment (Norment, in press). This reality can lead to unavoidable complexity and tradeoffs for many management actions, problems that are widely recognized among conservation practitioners (Hirsch et al. 2010; Roe and Walpole 2010; McShane et al. 2011). For instance, regional agricultural and municipal water needs are likely to increasingly conflict with those of E. panamintina and B. campi, control of feral horses and burros often contradicts deep- seated value systems held by some stakeholders, and mining-related economic development can clash with protection of sensitive riparian habitat. Furthermore, limited resources and competing goals may complicate implementation of management interventions, especially given the remote landscapes occupied by E. panamintina and B. campi. Scientific uncertainty, which we regularly identified in our threat assessment, will also necessitate adaptive management of these species

(Runge 2011). While recognizing these challenges, we nonetheless offer four management recommendations designed to promote the long-term population viability of both species (Figure

2.5). Our recommendations are targeted toward preservation of sensitive riparian habitats that are critical not only to E. panamintina and B. campi, but also to co-occurring species of regulatory interest including the Inyo California towhee, Melozone crissalis eremophilus (federally

57 Threatened, state Endangered) and desert bighorn sheep, Ovis canadensis nelsoni (BLM

Sensitive and INF Species of Conservation Concern). We hope that this alignment with broader

resource protection goals will increase the relevance of our recommendations in prioritizing future conservation action. First, we recommend that existing water withdrawals on federal lands be carefully enumerated and tracked, and that any proposal to initiate new water withdrawals be vetted using a detailed environmental impact assessment. Second, we recommend the continued reduction of feral burro and horse populations on federal lands, and the drawdown of permitted animal unit months in the Hunter Mountain cattle grazing allotment within DVNP. Third, we recommend continued control of Tamarix spp. on federal lands using appropriate control

methods, with particular emphasis on localities where Tamarix eradication is feasible due to low

plant abundance, while considering potential impacts to native species and habitat. Fourth, we

recommend that any new proposal for mineral resource extraction on federal lands be vetted

using a detailed environmental impact assessment, and that any mining-related destruction or

degradation of riparian zones be carefully controlled. Attention to the other threats identified in

our assessment is also important for conserving E. panamintina, B. campi, and their habitats. But

we argue that focusing limited available resources on the management of water withdrawals,

grazing, invasive Tamarix, and mining will likely maximize return on investment, and minimize the need for more strict regulation to protect these species.

Through our status assessment and threat analysis, we also identified several research needs of immediate management relevance to both E. panamintina and B. campi (Figure 2.5).

These include: (1) updated species distribution models, to inform targeted surveys of potential

new localities and more rigorously evaluate the possibility of private-land populations; (2)

rangewide GIS analysis of all riparian zones in the mountains occupied by these species, to

produce a baseline assessment of habitat extent in the face of climate change; (3) comprehensive multi-day survey expeditions at suspected localities to verify and voucher species’ occupancy, ground-truth current threats, and evaluate habitat quality; (4) conservation genomics of known populations, to evaluate genetic diversity and estimate rates of gene flow among and between localities; and (5) field and/or laboratory studies of the susceptibility and prevalence of chytridiomycosis in B. campi populations from both Bd and Bsal. Some of these research needs have been advocated elsewhere (Thomson et al. 2016). Moreover, several goals (e.g., 3 and 4) are inherently linked and can be efficiently pursued simultaneously. We especially advise field workers in the Inyo Mountains to consider both E. panamintina and B. campi in their study aims, because these species likely co-occur at most riparian zones in that range. We encourage agency funding to support research on these topics, and emphasize that the results of such studies might lead to revisions of our threat assessment. Although our recent survey coverage included the majority of the known localities for both species, gaps do exist, particularly on lands managed by

DVNP and CLNAWS.

Species-focused recommendations aside, our work additionally revealed the presence of several recurring scholarly problems that are of general interest to the broader scientific community, given the reliance of species status assessments on the best available science. The factual errors that we identified in the literature with respect to E. panamintina and B. campi are attributable to several causes, including: overlooking or selective use of available data, limited availability of some gray literature, the use of and failure to contextualize older data in light of more recent findings, misinterpretation or misrepresentation of data, and perpetuation of pre- existing literature errors. These problems are not unique, and have been identified elsewhere

(Rubel and Arora 2008; Stromberg et al. 2009). In the E. panamintina and B. campi literature,

the high frequency of errors could be a consequence of the relatively limited amount of data available in a peer-reviewed format, and limited accessibility of original data found only in uncirculated agency reports. For this reason, we encourage biologists and resource managers to prioritize the release of novel scientific data in a publicly accessible, peer-reviewed format whenever possible. Furthermore, we invite those who produce any scientific literature to strive for the following scholarly standards: (1) provide supporting evidence/citations for claims or statements; (2) reference original source literature, or scholarly review papers, when citing evidence for claims or statements; (3) cite references accurately, in a way that does not misrepresent the work of earlier authors; (4) cite older references with caution, and indicate when these might not reflect contemporary reality, and (5) comprehensively review the available literature on the topic of interest (Figure 2.5). Achieving these standards is a labor-intensive process, and no publication is ever perfect. Yet by striving to fulfill these scholarly guidelines

(Figure 2.5), researchers will promote the best available science and help agencies tasked with resource protection to best prioritize their limited time and budgets. Furthermore, these recommendations will help to ensure that authors in any discipline will maximize the accuracy, value, and utility of their work, thereby assuring the integrity of the scientific community’s

“bricks.”

Acknowledgments.—Author order follows the “sequence-determines-credit” approach as defined by Tscharntke et al. (2007). We thank B. Alexander, L. Bell, S. Brown, J. Calvin, T. Campbell,

M. Dickes, J. Dittli, T. Hibbitts, S. Hillard, E. L. Jockusch, J. D. Hunt, C. Klingler, O. Klingler-

Brown, J. Mattern, G. Milano, M. Murphy, D. Nielsen, M. T. Norment, S. P. Parker, S.

Parmenter, K. Rademacher, D. Silverman, R. Smith, B. J. Thesing, D. York, and the late D.

Giuliani for field assistance and data sharing. D. Pritchett at the University of California White

Mountain Research Center, and C. Fidler at the University of California Berkeley Museum of

Vertebrate Zoology Archives, graciously provided copies of the unpublished field notes of D.

Giuliani and K. Yanev, respectively. We are grateful to N. Camacho and G. B. Pauly (LACM) for accessioning vouchers. D. J. House, M. R. Lambert, J. C. Maerz, and several members of the

Maerz lab offered helpful comments on prior versions of this manuscript. D. LaBerteaux and C.

Carter shared key literature. Financial support was provided through a University of Georgia

Presidential Fellowship to AGC; a sabbatical leave granted to CJN by the College at Brockport,

SUNY; and a contract awarded to CJN by the U.S. Fish and Wildlife Service. These funding sources had no role in influencing this research or its publication. Field work was authorized under California Department of Wildlife Scientific Collecting Permit #011663, Bureau of Land

Management Permits #1110 (CA-170.31) P, #1110 (CA-6601.26) P, and #6500 (CA-170.31) P,

USDA Inyo National Forest Permits #WMD15002 and #2300, USDI National Park Service

Permit #DEVA-2017-SCI-0036, and University of Georgia IACUC AUP #A2012 10-004-Y1-

A0, #A2016 02-001-Y1-A0, and #A2016 02-001-Y2-A0. A. Ellsworth issued a CDFW letter of authorization for survey work at the Indian Joe Springs Ecological Reserve. We dedicate this paper to the late David J. Morafka and Derham Giuliani.

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Tables and Figures

Table 2.1. Rangewide threat scores for Elgaria panamintina and Batrachoseps campi. Score values correspond to the percentage of occupied localities currently known to be affected by the threat, as follows: 0 = 0%, 1 = less than 20%, 2 = 20–50%, 3 = over 50%. Scores marked with an asterisk (*) denote a prediction for the future; their current score is 0.

USFWS Elgaria panamintina Batrachoseps campi

Threat ESA Listing Factor Score Score

Water diversions Factor A 2 0

Grazing by feral/domestic livestock Factor A 2 0

Mining Factor A 0 1

Roads and off- highway vehicles Factor A 2 0

Invasive plants

(Tamarix spp.) Factor A 1 2

Marijuana cultivation Factor A 1 1

Renewable energy development Factor A 0 0

Overutilization Factor B 0 0

Disease Factor C 0 0

Inadequate regulatory mechanisms Factor D 0 0

Climate change Factor E 3* 3*

Flash floods Factor E 1 2

Figure 2.1. Panamint alligator lizard, Elgaria panamintina (left), and Inyo Mountains

Salamander, Batrachoseps campi (right). Photographs by Adam G. Clause.

Figure 2.2. Representative landscapes for Elgaria panamintina and/or Batrachoseps campi showing the undeveloped character of the mountainous regions they inhabit. Left to right: Union

Wash, Inyo Mountains; Piute Creek, White Mountains; Surprise Canyon, Panamint Mountains.

Photos taken from the approximate vantage point of the nearest paved road; a high-clearance dirt access road is visible in the middle photo. Photographs by Adam G. Clause.

Figure 2.3. Habitat photos for Elgaria panamintina and/or Batrachoseps campi showing the rugged, rocky terrain and bedrock waterfalls that are common in occupied canyon-bottom microhabitat. Left to right: Water Canyon, Argus Mountains; unnamed canyon between Union

Wash and Reward Mine, Inyo Mountains; French Spring, Inyo Mountains. Photographs by

Adam G. Clause.

Figure 2.4. Rangewide locality-level distribution and survey coverage for E. panamintina and B. campi. Solid symbols show localities surveyed for this work, hollow symbols show historically surveyed localities.

Priority Management Actions for Elgaria panamintina & Batrachoseps campi v Strictly regulate montane water withdrawals v Remove non-native ungulate grazers from wildlands v Prohibit mining-related degradation of riparian habitat v Control Tamarix spp. using appropriate methods

Research Needs for E. panamintina & B. campi q Updated species distribution models q GIS quantification of riparian habitat extent q Comprehensive field surveys q Conservation genomics q Susceptibility of B. campi to chytridiomycosis (Bd & Bsal)

Scholarly Publishing Standards Ø Support claims with data/citation(s) Ø Read and cite original source literature Ø Cite references accurately and objectively Ø Cite older references with caution Ø Comprehensively review relevant literature Ø Prioritize peer-review outlets for novel datasets

Figure 2.5. Recommendations for management and research relating to Elgaria panamintina and

Batrachoseps campi, and reminder of scholarly scientific standards.

CHAPTER 3

RE-SHARPENING THE ENDANGERED SPECIES ACT: EMPIRICAL SUPPORT FOR THE

REGULATORY BAN ON MULTI-SPECIES PETITIONS 1

1 Clause, A.G. and J.C. Maerz. To be submitted to Endangered Species Research.

Abstract

One of the most powerful environmental laws in the United States is the federal

Endangered Species Act (ESA), but many challenges impede its effectiveness. How species are petitioned for protection under the ESA is one such challenge, and multi-species petitions are of particular concern. This type of petition was the recent subject of a controversial regulatory ban, the validity of which remains unstudied. Here, we offer the first empirical data germane to this policy change. Our analysis of 81 amphibian and reptile taxa recently proposed for ESA listing under multi-species petitions suggests that many were unwarranted for such protection. Of the petitioned taxa endemic to the United States, only 50% are categorized as threatened on the

International Union for the Conservation of Nature Red List of Threatened Species. Similarly, although over 90% of the petitioned taxa are considered threatened according to the NatureServe classification system, this pattern is largely driven by the inclusion of 40% of the taxa in the

Vulnerable category, which generally does not indicate a serious threat of extinction.

Furthermore, most of the petitioned taxa were supported by comparatively little direct evidence of imperilment provided in the multi-species petition, and authoritative independent observers identified serious errors of data omission, misinterpretation, and reliance on outdated data for multiple species. To date, 27% of all petitioned taxa have been found unwarranted for listing by the U.S. Fish and Wildlife Service, and this number is likely to grow. Cumulatively, our results suggest a lack of petitioner expertise and an accompanying disconnect between scientific knowledge and the petitioners’ species selection process. We conclude that the ban on multi- species petitions was likely justified as a means of re-sharpening the ESA, and will help prevent further poorly-supported taxa from clogging the listing pipeline in the future. Similar to other authors, we encourage close collaboration between petitioners, species experts, and state wildlife

management agencies during all aspects of petition development, to maximize the data-driven selection of petitioned species and thereby further improve the effectiveness of the ESA.

Introduction

The federal Endangered Species Act (ESA) is one of the premier pieces of environmental legislation in the United States. This law tasks the United States Fish and Wildlife Service

(USFWS) and the National Marine Fisheries Service (NMFS) with protecting at-risk species against extinction. First enacted in 1973, the ESA today extends legal protection to over 1,600 domestic and almost 700 foreign plant and animal taxa and their habitats (USFWS, 2018a).

Many new taxa continue to be listed annually, and the cumulative listing curve shows no sign of leveling off (Puckett et al., 2016). As such, the assessment process for ESA listing continues to orchestrate changes to species-specific conservation policy at a national scale. However, the effective implementation of the ESA listing process is beset by many challenges, including taxonomic biases (Wilcove et al., 1993; Metrick and Weitzman, 1998; Gratwicke et al., 2012), political pressure (Lieben, 1997; Ando, 1999; Greenwald et al., 2006), and insufficient budgetary appropriations (Gibbs and Currie, 2012; Puckett et al., 2016). The synergistic challenges associated with expertise, data quality, and efficiency in species petitions remain unstudied, however.

Lack of ESA petitioner expertise can result in inappropriate species selection, and the use of incomplete and misinterpreted data to justify petitions. For the rare and little-known species that are often proposed for ESA listing (MacCracken et al., 2013; Weijerman et al., 2014), significant unpublished data is frequently held by wildlife management agencies or scholarly researchers, or available only in uncirculated government reports, and is effectively “hidden” from other actors (Meek et al., 2015). More readily available published data can thus be outdated, and inaccurately reflect current on-the-ground realities. Furthermore, the literature itself can contain key errors that might go unrecognized without a deep knowledge of the species

or system in question. Such errors can be severe, including the attribution of a species’ type locality to the wrong country (Gibbons and Watkins, 1982), and erroneous claims of species extirpation that can be sustained for decades (Clause et al., 2018). Because the USFWS and

NMFS are legally mandated to make listing decisions “on the basis of the best scientific and commercial data available” (ESA 1973, as amended), data quality is ultimately a major driver of petition success, but see Doremus (2004) and Murphy and Weiland (2016) for discussion of complicating factors.

Efficiency of the ESA listing process has long been undermined by a lengthy backlog of petitioned species, resulting in prolonged listing delays. As reported by Puckett et al. (2016), from 1973 to 2014 the median time to listing was 12.1 years across all taxa, although they perhaps overestimated the length of the delay in some cases (Walls et al., 2016). Regardless of their exact duration, these delays can allow population declines in petitioned species to worsen

(Wilcove et al., 1993), leading to reduced chances of species recovery upon eventual listing and thereby decreasing the likelihood of ESA success (Neel et al., 2012). Dozens of petitioned or candidate species have even become extinct while awaiting ESA listing (Suckling et al., 2004;

CBD, 2017a; USFWS, 2017b), but some caution may be warranted given the prematurity of many extinction claims (Scheffers et al., 2011; Caviedes-Solis et al., 2015; Clause et al., 2018).

Despite insufficient government funding being perhaps the primary cause of listing delays

(Stokstad, 2005; Puckett et al., 2016), the effect of funding deficits could be compounded if poorly-supported petitions overwhelm the assessment pipeline.

Until recently, a commonly-used approach in ESA petitions was to combine dozens of species into one multi-species petition, or “mega-petition” (Brown and Wolf, 2009; CBD, 2009,

2010, 2012; Curry et al., 2008; Adkins Giese et al., 2012) . By defining shared ecosystems, life

history traits, or geographic distributions across the petitioned taxa, multi-species petitioners can theoretically avoid unnecessary repetition and streamline listing and recovery efficiency

(USFWS 1992, 1994). However, in practice this consolidation appears to sometimes result in superficial species accounts and a detailed, but generalized, petition narrative with unclear applicability to the component species. In addition, multi-species petitions can, in practice, occasionally be motivated by a “shotgun” approach, in which petitioners attempt to misappropriate the ESA as a tool to protect geographic regions of conservation interest, not species (Wilcox and Elderd, 2003).

Due to these seemingly entrenched problems in multi-species petitions, and the related problems of poor data quality and a preponderance of unwarranted species petitions, the USFWS and NMFS recently abolished this practice (USFWS et al., 2016). Only one species may now be included in a petition for ESA listing. In the same final rule announcing this change, more detailed protocols for petition formatting and content were also established. The USFWS and

NMFS justified these regulatory changes as a means to boost the quality of scientific content on a per-species basis, thus ultimately improving the “efficiency and effectiveness” of the listing process (USFWS et al., 2016). These agencies cited direct experience with petitions containing incomplete, misleading, and outdated content as support for their decision. However, the scientific literature is bereft of studies to back up those claims with respect to multi-species petitions, and the issue remains “a controversy” (Walls et al., 2016). While some recent authors seem implicitly supportive of the ban (Walls et al., 2016), others suggest that multi-species petitions are in fact efficient (Puckett et al., 2016), and the new regulations have even been provocatively characterized as “premised on right-wing myths, not facts” (CBD, 2016). The question thus remains: was the prohibition on multi-species petitions truly justified as a tool to

re-sharpen the listing process of the ESA? In this contribution, we harvest data from two multi- species petitions to provide the first empirical data germane to this policy question.

Materials and Methods

We used a taxonomic approach for our analysis, harvesting data on all amphibian and non-avian reptile taxa from two multi-species petitions submitted by the US national non-profit

Center for Biological Diversity (hereafter, CBD). Our analysis included a total of 81 taxa, including 28 taxa petitioned by CBD (2010), and 53 taxa petitioned by Adkins Giese et al.

(2012). This sample comprised 46 amphibian taxa and 36 reptile taxa, with representation from all major lineages: 38 salamanders, 21 lizards and snakes, 15 turtles, and 7 frogs and toads. This set of 81 taxa is national in scope, with a cumulative geographic distribution that includes 44 US

States—albeit with a concentration of species in the Southeast and California.

To quantify petition rigor for these 81 taxa on a per-taxon basis, we measured the volume of text and number of citations specifically devoted to the case for listing each taxon. We measured volume of text in pages, rounded up to the nearest quarter-page, including all content except for language specifying the taxon name, its non-ESA status listings, and literature cited.

All text was in 12-point single-spaced Times New Roman font, with a full line between paragraphs, and with document borders of 1”. We tallied citations verbatim from the taxon- specific References sections in the original petitions, with no filtering.

To corroborate the species-selection process by the petitioners, we compared the 81 petitioned species against two highly-regarded classification systems for assessing species’ conservation status: the International Union for the Conservation of Nature Red List of

Threatened Species (hereafter, IUCN), and NatureServe. For the IUCN, we were only able

record the evaluation category for 68 of the 81 taxa, because 13 taxa are considered subspecies or as-yet unnamed population clusters and thus are precluded from this classification system.

For NatureServe, we evaluated 80 of the 81 taxa; Lampropeltis getula meansi has not been assessed by NatureServe.

Finally, using the Environmental Conservation Online System (ECOS) maintained by the

USFWS, we also tracked the progress of all 81 petitioned taxa through the USFWS assessment process, current as of 1 June 2018. For species that were advanced to a public comment period following a positive 90-day finding by the USFWS, we reviewed all public comments to identify those that were submitted by species experts. We considered commenters to be species experts if they had authored or co-authored one or more peer-reviewed papers on the species or genus in question, or if they presented substantial unpublished data on the species in question within their comments.

Results

The level of detail for the taxa-specific sections of the petitions were variable, but often relatively limited. The average number of pages devoted to each taxon, rounded to the nearest quarter-page, was 3.25 (range: 0.75–16.75). The average number of citations devoted to each taxon, rounded to the nearest whole number, was 24 (range: 1–103). However, a few high-effort examples skew those averages; when the 12 most well-populated petitions are excluded, the averages fall to 2.5 pages and 17 citations. Overall, the individual cases for listing over one-third of the taxa comprised less than 2 pages of text and/or were justified by fewer than 15 citations.

The taxon-selection process by the petitioners did not closely reflect the imperilment rankings of the IUCN Red List. Of the taxa endemic to the United States and for which an IUCN

assessment existed at the time the taxon was petitioned, only 50% (26 of 52) were evaluated as threatened, which includes the IUCN categories of Vulnerable, Endangered, and Critically

Endangered. Furthermore, less than 10% (5 of 52) of this subset of taxa were ranked higher than a category of Vulnerable by the IUCN, and none were Critically Endangered.

There was much stronger congruence between the NatureServe classification system and the petitioned taxa. Over 90% (73 of 80) of the petitioned taxa are categorized as Vulnerable,

Imperiled, Critically Imperiled, or Possibly Extinct by NatureServe. Once again, however, fewer of these taxa fell into the most at-risk NatureServe categories, with only 50% (40 of 80) ranked higher than Vulnerable and only 19% (15 of 80) ranked as Critically Imperiled or Possibly

Extinct. This discrepancy between the IUCN and NatureServe classifications appears to be due to a more liberal set of ranking metrics in the latter system. For instance, of the taxa ranked as

Vulnerable by NatureServe that have been evaluated by the IUCN, 65% (20 of 31) are ranked only as Near Threatened or Least Concern by the IUCN, and some authors generally do not consider NatureServe Vulnerable taxa to be threatened (Wilcove and Master, 2005). Regardless of the incongruity between the IUCN and NatureServe classifications, both nonetheless suggest that the conservation status of many taxa, at least at the time they were petitioned, was considered relatively secure across their entire geographic range.

Of the 81 petitioned taxa, only one has been listed under the ESA as of 1 June 2018

(Necturus alabamensis, Endangered; USFWS 2018). Conversely, 19 taxa were excluded from further review by the USFWS after 90-day findings concluded that the petitioner did not present substantial information that the petitioned action (ESA listing) was warranted. An additional three taxa that passed the 90-day finding benchmark were subsequently excluded from further review by the USFWS after a 12-month review found them not warranted for listing. The

petition content associated with each of these 22 excluded taxa was consistently among the most limited, averaging 2 pages and 16 citations per taxon. This appears to indicate that the USFWS’ decision not to list them was legitimate, being based not on political interference but rather on lack of information presented by the petitioner (but see CBD [2017b] for one possible exception). Of the remaining 61 taxa, we identified seven (11%) for which species experts, during the public comment period following the 90-day finding, identified serious problems of incomplete data, misinterpreted data, or reliance on outdated data in the original petition (Clarke,

2015; Clause et al., 2015; Clause and Hansen, 2015; Hoyer, 2015; Jones, 2015; Rabe, 2015;

Sweet, 2015; Sweet and Evelyn, 2015). These comments imply that these taxa also have a high probability of being excluded from consideration by the USFWS in the future, as their assessments progress. Cumulatively, the available data thus seems to indicate that 36% (29 of

81) of the petitioned taxa were likely inappropriate for listing consideration. The remaining 58 taxa that still await a final decision by the USFWS are undergoing, or slated to undergo, a detailed Species Status Assessment (see Smith et al. 2018).

Discussion

Our analysis empirically corroborates the claims made by the USFWS in support of their regulatory ban on multi-species petitions. For the 81 petitioned taxa we surveyed, the text volume of the taxon-specific narratives and the number of citations supporting those narratives was often low, suggesting that many petitioned species were not well justified for ESA listing.

Authoritative public commenters identified serious errors of omission, misinterpretation, and reliance on outdated data in the taxon-specific petition language for 11% of the 61 taxa opened to public comment, suggesting lack of petitioner expertise for the taxa in question. Furthermore,

only 50% of the US-endemic taxa assessed by the IUCN were categorized as threatened, and less than 10% were ranked higher than Vulnerable. Although over 90% of the petitioned taxa were categorized by NatureServe as Vulnerable or higher, this pattern is driven by a high proportion

(40%) of Vulnerable-ranked taxa, a category that often includes taxa of limited extinction risk.

These realities have already translated into a problem of petition inefficiency: USFWS determinations to date show that the ESA listing success rate for these 81 taxa will be 73% at best, and likely much lower. Although it is possible that these results are an artifact of the taxonomic group we analyzed, and do not reflect the characteristics of multi-species petitions more broadly, our dataset nonetheless reveals patterns that are consistent across both petitions.

The general brevity of the taxon-specific content in the petitions has two possible causes, which could be interacting synergistically. First, very little information on the taxa may actually exist, which is a widespread problem with imperiled species (Bland and Böhm, 2016; Meiri et al., 2018) and unavoidably prevents more detailed listing narratives. Second, petitioners may have overlooked pertinent references and data. For any given petitioned taxon, it can be challenging to confirm which of these factors is at play. However, neither reflects well on the expertise or strategy of the petitioners. If minimal relevant data is available for a taxon, building a strong case that ESA listing is warranted will likely be onerous. Similarly, if relevant data is excluded from a petition, the comprehensive nature of the USFWS review process is almost certain to eventually reveal such omissions—but only after the time-intensive 12-month review and public comment process is triggered. Thus, USFWS recognition of an inaccurate petition is difficult prior to the expenditure of substantial resources, which could have been better invested in truly warranted petitioned species.

Although popular and well regarded, neither the IUCN Red List nor NatureServe are perfect systems for evaluating species’ conservation status (Regan et al., 2004; Regan et al.,

2005; Hoffmann et al., 2008; Bland and Böhm, 2016). Nor are their evaluation criteria and imperilment rankings exactly analogous to each other (O’Grady et al., 2004; Regan et al., 2005), nor to that of the ESA (Laband and Nieswiadomy, 2006; USFWS, 2011). Nevertheless, in the most recent reviews of the topic, Harris et al. (2012) and Gratwicke et al. (2012) found that the

IUCN Red List and NatureServe, respectively, identify numerous imperiled U.S. amphibians and reptiles that are not extended protection under the ESA, suggesting a useful starting point for selecting taxa deserving of petitioner attention. In particular, some 80–82% of the IUCN-listed

(Harris et al., 2012) and NatureServe-listed (Gratwicke et al., 2012) threatened amphibian species in the U.S. remain unlisted under the ESA. Thus, the high proportion (50%) of IUCN- assessed species petitioned by Adkins Giese et al. (2012) and CBD (2010) that were not categorized as threatened by the IUCN is surprising. Together, these two listing systems imply at least some degree of disconnect between scientific knowledge and the petitioners’ species- selection process.

An additional line of evidence suggesting a disconnect between scientific knowledge and petition development, specifically relating to Adkins Giese et al. (2012), is offered by Thomson et al. (2016). Using an eight-metric risk assessment framework, Thomson et al. (2016) identified and scored 45 imperiled yet non ESA-listed amphibian and reptile taxa across their entire distribution in the western United States (all in California). Of these 45 taxa, Adkins Giese et al.

(2012) included only 11 in their multi-species petition; CBD (2010) intentionally did not include

California taxa in the scope of their petition. Remarkably, of the 10 most threatened taxa identified by Thomson et al. (2016), only three were included in the Adkins Giese et al. (2012)

petition, and none of them ranked among the top five. Different data availability does not explain this disparity; both sets of authors had almost identical information available (or potentially available) to them despite Thomson et al. (2016) being published four years after Adkins Giese et al. (2012).

It has been claimed that petitioners are unlikely to propose species that probably do not meet the criteria for ESA listing (Brosi and Biber, 2012). Studies have also shown that ESA- listed taxa originally petitioned by citizens are equally if not more deserving of ESA protections compared to listed taxa originally proposed by the USFWS (Biber and Brosi, 2010; Brosi and

Biber, 2012) Furthermore, work by Puckett et al. (2016) indicated similar process time to eventual listing across both single-species and multi-species petitions. However, these studies did little to explore the overall efficiency or effectiveness of citizen petitions, i.e., the proportion of petitioned species that were eventually listed. As we have shown, three independent assessment frameworks indicate the repeated, perhaps even widespread inclusion of poorly- justified species within both multi-species petitions we analyzed. Moreover, 23% (19 of 81) of the petitioned taxa were excluded from listing consideration at the earliest possible stage of the

USFWS assessment process (USFWS, 2011, 2015a, 2015b, 2015c, 2016a, 2016b, 2016c).

Although citizen petitioners are undoubtedly capable of identifying species that warrant ESA listing, and many of the petitioned species we analyzed are likely strong candidates for ESA protection, our analysis nevertheless also offers clear evidence that citizen petitioners often lack the expertise to distinguish species that likely warrant listing from those that do not. Arguably, this constitutes misuse of the multi-species petition process.

We consider it self-evident that submission of such poorly supported ESA petitions unnecessarily burdens for-profit, non-profit, private, state, and federal stakeholders, and

encumbers USFWS and NMFS rulemaking infrastructure. This can lead to waste of taxpayer money, and could engender public disillusionment with the ESA itself—which has long been under critical scrutiny (Shea, 1977; Holden, 1982; Gordon et al., 1997; Stokstad, 2005; Corn et al., 2013) despite its notable successes (Schwartz, 2008). Importantly, saturating the listing pipeline with unwarranted petitions can worsen delays in the listing of species that legitimately deserve such status. Although many factors contribute to the dramatic backlog of listing decisions (Stokstad, 2005; Puckett et al., 2016), our results suggest that the recent federal ban on multi-species petitions was a necessary step to help alleviate the backlog, as implied by other recent authors (Walls et al., 2016).

However, our results do not condemn the multi-species approach as an inherently undesirable technique for pursuing species listings under the ESA. To the contrary, multi-species internal proposals by the USFWS, and multi-species citizen petitions, have historically streamlined the successful listing of several closely-related and/or geographically cohesive groups of taxa. These include all 41 species of O’ahu tree snails in the genus Achatinella

(USFWS, 1980, 1981), three species of Branchinecta fairy shrimp endemic to California

(USFWS, 1991, 1994), and the Appalachian salamanders Plethodon nettingii and P. shenandoah

(USFWS, 1988, 1989), to name a few. For over 20 years, these benefits motivated USFWS policy supportive of multi-species petitions. Unfortunately, the seemingly well-intentioned misuse of the multi-species petition has now precipitated the loss of this tool for kickstarting the efficient addition of certain taxa to the ESA. Two likely consequences of this misuse are contemporary delays in listing of deserving species, and future delays in listing groups of species that could have been more efficiently considered together rather than separate. The latter will likely become particularly true for invertebrates, which constitute the most underrepresented

taxonomic group on the ESA proportionate to their diversity (Wilcove and Master, 2005). In coming decades, invertebrates could begin to dominate the ESA petition landscape as larger, more charismatic taxa become increasingly well-represented under the ESA.

Regardless of the future ESA petition landscape, our findings highlight the necessity of a best-practice for improving species selection and petition development. Echoing similar recommendations made elsewhere (Hanson et al., 2016; Walls et al., 2016), we encourage all

ESA petitioners to collaborate closely with recognized species experts and state wildlife agencies. This recommendation is symmetrical in that we also encourage academic and agency scientists to prioritize such collaboration, as a means of bridging the research/implementation gap that is an ongoing issue for conservation practitioners (Arlettaz et al., 2012). Adoption of this approach will help to ensure that multiple forms of relevant data, including difficult-to-access content in “gray literature” and unpublished formats (Meek et al., 2015), are incorporated into petitioner decisions. Without such communication with appropriate experts, petitioners open themselves to error when deciding whether or not ESA listing might be realistic, and scientifically justified, for a species of interest (see Sutherland et al. 2004). Although language requiring petitioner consultation with state wildlife agencies was ultimately removed from recent proposed regulations (USFWS and NMFS, 2015), the benefit to be gained from such consultations remains clear. An important distinction must be made, however: petition input, and possible petition co-authorship, is beneficial only if sought from biologists who possess legitimate expertise with the species under consideration. Reliance on close dialogue with such knowledgeable, authoritative partners will maximize petition effectiveness by facilitating data- driven species selection and inclusion of strong, comprehensive science in support of the listing.

In turn, such re-sharpening of petition development will help to strengthen and legitimize one of

the world’s most powerful environmental laws. This cannot come too soon for the numerous truly imperiled taxa, many of them on the brink of extinction, that still await ESA protection.

Acknowledgements

We thank members of the Maerz lab and M. R. Lambert for comments that improved earlier versions of this manuscript. Funding provided by a University of Georgia Presidential

Fellowship.

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CHAPTER 4

WHOSE FLAGSHIP?: NATURAL HISTORY INFORMS CONSERVATION PLANNING

FOR AN IMPERILED MESOAMERICAN REPTILE CLADE 1

1 Clause, A.G. To be submitted to Biodiversity and Conservation.

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Abstract

The search for robust, lasting solutions to complex socioecological problems is a central challenge for conservation scientists. Integrative conservation draws from multiple epistomologies to address such problems, but one perspective that has been criticized and sometimes overlooked is natural history. Nonetheless, a natural history perspective remains crucial to a major motivator of global conservation practice: safeguarding imperiled species in light of a sixth mass extinction. In this contribution, I demonstrate the value of natural history to the protection of arboreal alligator lizards in the genus Abronia, an at-risk clade of

Mesoamerican highland forest-dwelling reptiles. Using a review of the available literature and museum specimens, coupled with a protected area gap analysis, I find that Abronia are perhaps the most imperiled reptile group in Mesoamerica and are justified for recognition as a conservation flagship. Given the comparatively understudied distribution of Abronia and their predicted presence in many protected areas, additional field surveys are urgently needed and natural history can help to maximize the success of such work. Furthermore, because numerous

Abronia species are under-represented in the existing protected area network, establishment of additional community-based reserves and increased local-level collaboration to legitimize existing government reserves are warranted. Although illegal black-market collecting and fear- based killing of Abronia by indigenous residents complicate the flagship narrative, the lizards’ arboreal life history likely makes most populations resilient to those threats. Widespread ongoing habitat destruction from tree-cutting and climate change, although less tractable to address, is a far greater threat not only to Abronia but to broader landscape-level processes. Ultimately, a viewpoint grounded in natural history helps to illustrate how Abronia are a potentially useful, but

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imperfect vehicle for promoting holistic, equitable conservation of their montane forest habitat, which is part of a globally recognized biodiversity hotspot.

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Introduction

A central challenge in conservation science is to find durable, equitable solutions to complex socioecological problems. Often, these solutions necessitate decision-making in the face of uncertainty, and unavoidable tradeoffs (Hirsch et al. 2010, Roe and Walpole 2010, Milner-

Gulland and Shea 2017). Integrative conservation engages multiple viewpoints to address such problems, drawing from both commensurate and incommensurate perspectives to enrich planning and practice (Hirsch et al. 2010, McShane et al. 2011). However, certain lenses useful for understanding many conservation systems are sometimes overlooked.

One such lens is the study of natural history—the lifestyle, ecology, and behavior of organisms. Devaluation of this line of inquiry has been repeatedly identified (Greene and Losos

1988, Futuyma 1998, Tewksbury et al. 2014, Barrows et al. 2016), with critiques labeling the field as outdated and insufficiently rigorous. Yet, because much conservation practice on a global scale is motivated by protecting threatened species (Gibbons et al. 2000, Böhm et al.

2013) and forestalling a sixth mass extinction (Ceballos et al. 2015, Ceballos et al. 2017), a grounding in organismal natural history remains critically relevant in this context (Greene and

Losos 1988, Greene 1994, Bury 2006) and many others (Futuyma 1998, Greene 2005, Schwenk et al. 2009, Tewksbury et al. 2014, Barrows et al. 2016). Reliance on imperiled species lists, such the International Union for the Conservation of Nature Red List of Threatened Species, is widespread and epitomizes the institutional emphasis placed on threatened taxa as conservation priorities (Rodrigues et al. 2006), or at least as surrogates for systematic conservation planning

(Hoffmann et al. 2008).

Flagship species, which are boundary objects that link social and ecological aspects of conservation, represent another widespread application of the species-focused approach.

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Numerous governmental and non-governmental actors rely on conservation flagships as programmatic foci, advertising them to non-scientist audiences for both fund-raising and awareness gains (Walpole and Leader-Williams 2002). Nonetheless, the selection process for flagship species has been accused of a lack of rigor in some cases (Home et al. 2009).

Furthermore, many flagships are poorly understood ecologically, and may be viewed negatively by certain communities (Douglas and Veríssimo 2013). The effectiveness of conservation interventions motivated by flagship species, and their socioecological appropriateness, can thus be questionable. One taxon that is emerging as an international conservation flagship, but whose natural history and public perception are problematic, is the tropical lizard genus Abronia.

Arboreal alligator lizards in the genus Abronia, with 29 described species, represent the most species-rich of 11 recognized genera within the New World squamate family Anguidae

(The Reptile Database, 2018; http://www.reptile-database.org). These visually striking lizards

(Figure 4.1) occur primarily in mesic highland forests across Mesoamerica, from Mexico to

Honduras. Most species are comparatively poorly studied, and several are known to science from just a single individual. Tihen (1949) and Good (1988) were the first to assemble comprehensive treatises on the gerrhonotine anguid lizards, a group that includes the genus Abronia. Campbell and Frost (1993) later produced a revisionary monograph focused solely on the genus Abronia, but noted that their study was not “burdened with an overabundance of comparative material.”

Although Abronia remain enigmatic, in the 25 years since Campbell and Frost’s landmark treatment over 70 scientific contributions on the genus have been published. These contributions include novel information on natural history, distributional and elevational range extensions, and new species descriptions. Review of this body of literature is needed to set a new baseline for our understanding of the diversity and biogeography of this genus, which is

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considered among the most imperiled clades of mainland vertebrates in Mesoamerica. A groundswell of global support for in-situ protection of Abronia has begun to promote the genus as a linchpin for safeguarding montane forests in a global biodiversity hotspot (Myers et al.

2000, Myers 2003, Mittermeier et al. 2004), further emphasizing the practical import of a new analytical review. A major complication for this conservation narrative, however, is that many human communities that co-habit forests with Abronia view the lizards with fear and antipathy, often believing them to be dangerously venomous (Campbell and Frost 1993, Ariano-Sánchez and Torres-Almazán 2010, Ariano-Sánchez et al. 2011, Martín-Regalado et al. 2012). An additional problem is that Abronia are highly prized by hobbyists and are illegally trafficked on an international black market (Anonymous 2009, Altherr 2014, Anonymous 2015, Auliya et al.

2016), creating additional challenges for conservation practice.

Here, drawing on a comprehensive literature review, a global database of museum records, and an analysis of Mesoamerican protected areas, I attempt to answer the following questions. First, does a natural history perspective improve conservation practice for Abronia, by informing how best to evaluate threats and maximize the success of interventions? Second, does a preliminary gap analysis of protected areas support the recognition of Abronia as a flagship genus for highland forest conservation? And finally, are national and international threatened species classification systems congruent with respect to their categorization of Abronia? I conclude by discussing the need for strategic, holistic conservation programs for Abronia and their habitat, without which enduring conservation outcomes will prove elusive. My goal is to demonstrate how an approach rooted in the natural history of a little known, threatened taxonomic group can be a useful lens for practitioners, with applied relevance to integrative conservation in a broader socioecological context.

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Methods

I compiled available literature on the genus Abronia using the ISI Web of Science and

Zoological Record databases, using genus/subgenus names and all synonyms as search terms. I also included scholarly books, species accounts in the IUCN Red List, and unpublished theses and dissertations despite their generally less rigorous nature compared to peer-reviewed publications. Important knowledge relevant to the genus is available only in these types of sources, and I consider it necessary to account for these data. I also supplemented my review with unpublished personal field work observations from Mexico.

To complement the literature, I assembled a comprehensive global database of museum records for Abronia. I populated this database with records obtained directly from institutional curators at a total of 40 museums: American Museum of Natural History (AMNH); Academy of

Natural Sciences, Drexel University (ANSP); Monte L. Bean Life Science Museum, Brigham

Young University (BYU); California Academy of Sciences (CAS); Carnegie Museum of Natural

History (CM); Colección Nacional de Anfibios y Reptiles, Instituto de Biología, Universidad

Nacional Autónoma de México (CNAR); Cornell University Museum of Vertebrates (CUMV);

Facultad de Ciencias Biológicas, Benemérita Universidad Autónoma de Puebla (EBUAP);

Universidad Autónoma del Estado de Morelos (EBUM); El Colegio de la Frontera Sur, San

Cristóbal de Las Casas (ECO-SCH); Escuela Nacional de Ciencias Biológicas, Instituto

Politécnico Nacional (ENCB-IPN); Field Museum of Natural History (FMNH); Instituto de

Historia Natural de Chiapas (IHN); University of Kansas Biodiversity Institute (KU); Natural

History Museum of Los Angeles County (LACM); Louisiana State University Museum of

Natural Science (LSUMZ); Milwaukee Public Museum, Vertebrate Zoology (MPM); Museum of

Comparative Zoology, Harvard University (MCZ); Muséum National d’Histoire Naturelle, Paris

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(MNHN); Musuem of Southwestern Biology, University of New Mexico (MSB); Museum of

Vertebrate Zoology, UC Berkeley (MVZ); Museo de Zoologia “Alfonso L. Herrera,” Facultad de

Ciencias, Universidad Nacional Autónoma de México (MZFC-HE); Natural History Museum,

London (NHMUK); Naturhistorisches Museum (NMW); James R. Slater Museum of Natural

History (PSM); San Diego Natural History Museum (SDNHM); Senckenberg Forschungsinstitut und Naturmuseum (SMF); Staatliches Museum für Naturkunde, Stuttgart (SMNS); Texas A&M

University Biodiversity Research and Teaching Collections (TCWC); Texas Natural History

Collection (TNHC); Universidad Autónoma del Estado de Hidalgo (UAEH); Universidad

Autónoma de Nuevo León (UANL); University of Colorado Museum of Natural History (UCM);

University of Illinois Museum of Natural History (UIMNH); University of Michigan Museum of

Zoology (UMMZ); National Museum of Natural History, Smithsonian Institution (USNM),

University of Texas at Arlington, Department of Biology (UTA); Yale University, Peabody

Museum of Natural History (YPM); Museum für Naturkunde (ZMB); Zoologische

Staatssammlung München (ZSM). Queries directed to the curators at an extensive list of other

Mesoamerican museum collections did not yield additional non-duplicative Abronia specimen records.

From this combined literature and specimen database, I georeferenced all published and/or vouchered Abronia localities using the Mapa Digital de México V6.3.0

(http://www.inegi.org.mx/geo/contenidos/mapadigital/) and the Pueblos America

(http://mexico.pueblosamerica.com) platforms. I define a locality as being ≥ 1 km distant from any other locality, and treated records < 1 km apart as a single locality. I excluded potential localities from consideration under the following circumstances: (1) existing locality data too imprecise, available only at the municipality level or coarser; (2) an irreconcilable contradiction

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exists between the locality’s narrative description and GPS coordinates; or (3) the locality name is absent from existing maps and online platforms, and lacks associated GPS coordinates. For cases where a specimen or literature record appeared to be misidentified, I obtained high-quality photographs or personally examined the specimen to confirm the species. In rare cases where the taxonomic status of a record was uncertain, I typically followed the published literature by tentatively attributing it to a described species with no comment.

To determine the representation of Abronia in existing protected areas, I queried the

World Database on Protected Areas (WDPA; http://www.protectedplanet.net) and corroborated those results using data on Mexican protected areas from the Comisión Nacional de Áreas

Naturales Protegidas (CONANP; http://sig.conanp.gob.mx/website/pagsig/datos_anp.htm), and data on Guatemalan protected areas from the Consejo Nacional de Áreas Protegidas (CONAP; http://conap.gob.gt/AreasProtegidas.aspx). To identify protected areas that possibly support

Abronia, but which remain unconfirmed, I selected those that met two criteria: (1) inclusion of a mountain or massif with at least one Abronia record, and (2) inclusion of forest at an elevational range within that known for the Abronia species in question. I filtered all protected area results using the IUCN Protected Area Categories System, including in my analysis only those protected areas that met one of the six categories of this system (Dudley 2008, Stolton et al. 2013). I complemented my analysis by accounting for private or local-level reserves known to support

Abronia, based on their mention in the literature or discovery during my personal field work.

Because no lists of private reserves exist for Mexico, Honduras, or El Salvador, my analysis likely underestimates their number and importance. In rare cases where multiple protected areas designations overlap, such as private reserves nested within a federal reserve, I account only for the larger, more encompassing protected area.

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I consulted five different threatened-species classification systems to evaluate how scientists, practitioners, and governments have assessed the imperilment of Abronia. Two classification systems are international in scope, yet have no legal weight: the International

Union for the Conservation of Nature Red List of Threatened Species (hereafter, IUCN Red

List), and the Environmental Vulnerability Score (hereafter, EVS). The IUCN Red List is generally considered the leading global species imperilment assessment mechanism (Rodrigues et al. 2006, Hoffmann et al. 2008), although its application to Mesoamerican reptiles has been critiqued (Wilson and Townsend 2010, Johnson et al. 2017). The EVS system was first developed by Wilson and McCranie (2004b) for the reptiles and amphibians of Honduras, and has since been revised and applied to all Mesoamerican countries as a rapid, easily updated alternative to the IUCN Red List (Johnson et al. 2015). The third classification system is the

Convention on the International Trade in Endangered Species of Wild Fauna and Flora

(hereafter, CITES), which administers legal regulatory power over species imports/exports among signatory countries. The remaining two classification systems are national in scope, and have legal influence in their respective countries: the Norma Oficial Mexicana list of the

Mexican federal government (NOM-059-SEMARNAT-2010), and the Lista de Especies

Amenazadas of the Guatemalan federal government (CONAP 2009).

Several phylogenies have resolved the genus Abronia as paraphyletic with respect to the genus Mesaspis (Solano-Zavaleta 2011, Pyron et al. 2013, García-Vásquez et al. 2018).

However, these studies have included only a few genetic loci and a small proportion of the described species of both genera. Pending more comprehensive phylogenies, and due to the morphological and ecological divergence of Abronia and Mesaspis, I here tentatively treat them as independent lineages and do not discuss the latter genus further. Within the genus Abronia, six

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clades are traditionally recognized and denoted as subgenera: Auriculabronia, with 12 described species; Abronia, 8 species; Lissabronia, 3 species; Scopaeabronia, 3 species; Abaculabronia, 2 species; and Aenigmabronia, 1 species. These clades have been generally stable over the past 25 years and are morphologically and molecularly diagnosable (Campbell and Frost 1993, Campbell et al. 1998, Chippindale et al. 1998). I occasionally reference these clades in this contribution to better interpret and categorize the diversity within this comparatively speciose genus, while recognizing that their phylogenetic relationships could change in the near future. To avoid confusion, hereafter I distinguish between the two groups named Abronia by referring to the smaller, nested group as the “Abronia clade,” and the larger, inclusive group as the “genus

Abronia” or simply “Abronia.”

Results

Although over 200 literature sources offer information on Abronia, and nearly 1,100 preserved specimens exist in museums worldwide, many knowledge gaps remain. Excluding original species descriptions, only a handful of refereed publications focus on Abronia; most contributions mention the genus only in passing, or report serendipitously acquired data.

Furthermore, over half of available museum specimens belong to just three species (A. graminea,

A. taeniata, and A. mixteca), and 15 species are represented by fewer than ten individuals known to science. Despite this uncertainty, clear patterns are identifiable in the available data. Below, I present the pertinent results of my review in four sections: distribution, natural history, preliminary protected area gap analysis, and threatened species classification systems.

Distribution

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Members of the genus Abronia range from central Mexico south through Guatemala and into northern El Salvador and southwestern Honduras (Figures 4.2, 4.3). The northernmost and easternmost populations lie in Mexico, and are referred to A. taeniata in Tamaulipas (Terán-

Juárez et al. 2015) and A. deppii in Michoacán (Alvarado-Díaz et al. 2013), respectively.

Alvarado-Díaz et al. (2013) claim that A. deppii occurs even farther east in the Mexican state of

Jalisco, but their assertion is unsubstantiated. The most southern and western populations of

Abronia are those of A. salvadorensis in Honduras (Hidalgo 1983, Dueñas et al. 2000). Mexico supports 19 species, Guatemala 10 species (two of them shared with Mexico), Honduras 2 species, and El Salvador 1 species (shared with Honduras) (Table 4.1). Twelve Mexican states

(Chiapas, Guerrero, Hidalgo, Michoacán, Morelos, México, Oaxaca, Puebla, Querétaro, San Luis

Potosí, Tamaulipas, and Veracruz), ten Guatemalan departments (Alta Verapaz, Baja Verapaz,

Escuintla, Huehuetenango, Jalapa, Quetzaltenango, Quiché, Sacatepéquez, San Marcos, and

Sololá), three Honduran departments (Copán, Intibucá, and La Paz), and one El Salvadoran department (Santa Ana) are known to support at least one verified Abronia record (Table 4.1).

Most Abronia species are documented from only a handful of localities, largely due to their restricted ranges. Although 5 species (all members of the Abronia clade: A. deppii, A. graminea,

A. mixteca, A. oaxacae, and A. taeniata) are reported from twenty or more localities, over half of the species in the genus (15 of 29) are known from three or fewer localities (Figures 4.2, 4.3).

The distribution of Abronia is highly discontinuous across Mesoamerica, being restricted to the heterogeneous mountain ranges scattered across the region (Figures 4.2, 4.3). This discontinuity mirrors the patchy geographic distribution of the mesic highland forests they inhabit (Ponce-Reyes et al. 2012). Most of the literature has emphasized the “nearly invariable” allopatry (Campbell et al. 2016) between species of Abronia. Although allopatry remains the

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dominant pattern among members of the genus, growing evidence now indicates that sympatry

(which I define here as two species occurring within 10 km of each other on the same mountain or massif) is more widespread than previously understood. Over one-third of the 29 described species are now known to be sympatric with a congener. These six sympatric species pairs, as reviewed by Clause et al. (2018), are: A. bogerti and A. ornelasi, A. chiszari and A. reidi, A. graminea and A. taeniata, A. mixteca and A. oaxacae, A. frosti and A. lythrochila, and A. gaiophantasma and A. fimbriata. Furthermore, A. fuscolabialis and A. mitchelli might be sympatric on Cerro Pelón, Oaxaca, Mexico (Good and Schwenk 1985), and A. cuetzpali and A. oaxacae could prove sympatric in parts of the Sierra Madre del Sur of Oaxaca (Campbell et al.

2016). Speculation also exists that A. lythrochila might occur in sympatry with A. leurolepis and/or A. ochoterenai in Chiapas, Mexico or adjacent Guatemala, although much doubt exists due to the imprecise type locality for the latter two species (Campbell and Frost 1993, Peterson and Nieto-Montes de Oca 1996, Casas-Andreu and Smith 1991). However, in these known and suspected cases of sympatric species pairs, the region of overlap usually represents only a fraction of their geographic distribution, such that one or both component species primarily occurs in allopatry respective to all congeners. Of biogeographical interest is that four of the known or suspected sympatric species pairs (A. bogerti/A. ornelasi, A. chiszari/A. reidi, A. frosti/A. lythrochila, and A. fuscolabialis/A. mitchelli) are not close relatives, belonging to different clades within the genus.

Despite sympatry among Abronia becoming recognized as increasingly common, syntopy has been reported in only two species pairs: A. gaiophantasma and A. fimbriata (Campbell and

Frost 1993), and A. mixteca and A. oaxacae (Aldape-López and Santos-Moreno 2016).

Congeneric occupancy of individual trees is documented for both. Syntopy has never been

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documented for other species pairs, and thus it remains possible that those congeners do not occur side by side in the same habitat. Rather, they might be isolated to different forest types in the mountains they co-habit (Campbell and Frost 1993, Torres et al. 2013). Montane forests in

Mesoamerica are complex, with dramatic changes in topographical relief, soil type, and precipitation that often produce a wide array of different forest types on the same mountain

(Clause et al. 2016b, Williams-Linera et al. 2013, Williams-Linera and Vizcaíno-Bravo 2016), thus creating a habitat mosaic that could facilitate niche segregation among sympatric species.

These caveats aside, the available evidence indicates that Abronia can no longer be considered to have diversified in near-perfect allopatry across Mesoamerica. This biogeographical reality has important implications for conservation and knowledge production.

Prioritization of limited conservation resources could be improved by targeting programs in areas that support two co-occurring species of Abronia, instead of just one. Furthermore, documentation of new populations of Abronia must be supported by sufficient physical evidence to diagnose the species in question. Relying solely on geography to clinch an identification might lead to erroneous conclusions, especially given that many Abronia species can be challenging to identify (see Chapter 6).

Natural History

Elevation

In the quarter-century since Campbell and Frost (1993) summarized the elevational limits of Abronia, new data has revised the known range for about two-thirds of the species they treated. Consistent with the common stereotype of Abronia as high-elevation specialists, over three-quarters of the described species have elevation ranges that lie within 1300–2900 m (Table

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4.1). There is little signal of clade-level differentiation in occupied elevation, save for the three species that comprise the Scopaeabronia clade (A. bogerti, A. chiszari, and A. ramirezi).

Although understudied and cumulatively documented from fewer than 10 localities (Clause et al.

2016b), Scopaeabronia inhabit an unusually low elevation range of ca. 360–1540 m that includes lowland tropical rainforest habitats. They are also morphologically among the most divergent members of the genus (Campbell and Frost 1993, Clause et al. 2016b). However, the minimum elevation known for the genus Abronia corresponds to the distantly related A. taeniata, which is recorded from 125–2970 m and has the most expansive geographic and elevational distribution within the genus (Figure 4.2), yet may represent multiple cryptic species (Campbell and Brodie

Jr. 1999). This minimum elevation of 125 m corresponds to a locality nestled in the foothills of the Sierra Madre Oriental in extreme northeastern Hidalgo (Mendoza-Paz and Fernández-Badillo

2018), and is atypical. Nonetheless, this area supports mesic broadleaf forest laden with epiphytic growth, which is considered a classic microhabitat for the species and genus (Campbell and Frost 1993).

Due to these and other recent low elevation records for Abronia (A. taeniata at ca. 300 m

[Lemos-Espinal and Dixon 2016; Mendoza-Paz and Fernándo-Badillo 2018], A. graminea at

1170 m [Clause et al. 2018], and A. reidi at 1000–1200 m [Thesing et al. 2017]), I caution against overreliance on elevation to diagnose forests suitable for the genus. Moreover, elevation itself has no direct ecological relevance save for atmospheric oxygen levels, instead being merely a proxy for precipitation and temperature. I encourage field biologists and conservation planners to be attentive to the possibility of Abronia inhabiting forests well below 1000 m, particularly those supporting dense epiphytic growth and situated in mesic or thermally sheltered sites.

However, as presented in Chapter 5, even dense epiphytic growth is not always a required habitat

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element for Abronia.

Microhabitat Use

Abronia appear to be highly arboreal, and arboreality is a consistent trait across the clades within the genus. Campbell and Frost (1993) drew this conclusion after an exhaustive review of the then-available evidence. Since their work, additional data has accrued. Vega-Trejo (2010) and Cruz-Ruiz et al. (2012) found multiple individuals of A. oaxacae on pine tree trunks,

Aldape-López and Santos-Moreno (2016) reported an adult male of this species over 1 m high on the trunk of a pine tree, and A. oaxacae specimen LACM 163902 was taken 5 m high in an oak tree. Aldape-López and Santos-Moreno (2016, 2018) reported mating pairs of A. mixteca perched 1 and 1.7 m high on pine trunks. Ariano-Sánchez and Torres-Almazán (2010) implied that all 35 A. campbelli they captured were found on oak trees, and Abronia meledona has been seen repeatedly in trees up to about 12 m high (Campbell and Brodie Jr. 1999). One A. frosti was found about 2.5 m high on the trunk of an unidentified tree (Campbell et al. 1998), and five

Abronia bogerti were reported from 3.5–8 m high in oak trees (Clause et al. 2016b). The only existing targeted data that rigorously analyzes Abronia microhabitat use is presented in Chapter

5, and those results from A. graminea strongly support the arboreality narrative for the genus.

However, like most arboreal squamates Abronia are often found on the ground, as noted by Campbell and Frost (1993). Many of these observations seem directly attributable to tree- felling operations or natural tree falls (Bogert and Porter, 1967; Campbell & Brodie 1999;

Campbell et al. 1998; Franzen and Haft, 1999; Ariano-Sanchez et al., 2011; Torres et al. 2013).

In other cases, the lizards appear to have inadvertently fallen from the forest canopy (Smith and

Álvarez del Toro 1963, Smith and Smith 1981, Flores-Villela and Vogt 1992). Single specimens of A. anzuetoi (AMNH 102177) and A. deppii (MZFC-HE 22673) were collected after they fell

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from trees, and I have personally observed five A. graminea fall to the ground from the forest canopy. In other cases, it is possible that Abronia descended to the ground naturally. Wagner

(2010) reports finding unspecified Abronia species “on the ground and in rotting wood.”

Multiple adult A. gaiophantasma have been encountered on the forest floor (Franzen and Haft

1999, Eisermann and Acevedo 2016), and Campbell (2000) indicated that this species and A. fimbriata “were generally taken on or near the ground.” One A. meledona was encountered while basking in the early morning at the margin of a small gravel road (Campbell and Brodie Jr.

1999), and single A. matudai were captured among rocks and alive on a road, respectively (UTA

R-40643 and R-40659. I have observed both A. graminea and A. taeniata resting upon or crossing paved and unpaved roads, and roadkill has been documented in A. oaxacae (MZFC HE-

3409 and 24434). Vega-Trejo (2010) found A. oaxacae on the ground, and Canseco-Márquez and Gutiérrez-Mayen (2010) suggested that this species descends to the forest floor to forage in leaf litter. One A. mixteca (UTA R-4711) was collected on the ground in pine-oak forest, two additional specimens (UTA R-6246 and 6248) were captured while basking on rocks in pine-oak forest, and Martín-Regalado et al. (2012) reported finding single A. mixteca individuals at the base of a pine tree and on the forest floor, respectively. Two of the three known individuals of A. cuetzpali were found on the forest floor (Campbell et al. 2016), and one A. martindelcampoi

(UTA R-4451) was taken on the ground near a pine tree. Only twice have Abronia been discovered while turning cover objects: a juvenile A. fuscolabialis found “beneath a large rock in a cleared area” (Campbell and Frost 1993), and an A. matudai (UTA 34280) collected from under a rock.

Ultimately, the large number of ground-level Abronia observations should not be interpreted to mean that Abronia spend much of their time on the ground. Instead, it probably

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means that the ground is where most herpetologists are lucky enough to find them, and this perspective is supported in Chapter 5. The strongly prehensile tail documented in all Abronia species for while live specimens have been studied further corroborates their highly arboreal lifestyle. Based on this natural history knowledge, those who survey for Abronia should be attentive to the possibility that these lizards can appear almost anywhere in or adjacent to occupied forest habitat, but scanning accessible parts of trees will likely maximize detection rates in most cases. This arboreal life history trait also allows for biologically realistic interpretation of threats posed by actors who remove Abronia from the wild, either for the pet trade or due to fear- based killing. Because only a minor proportion of a given Abronia population is likely subject to such impacts at any one time, they may not represent serious threats to their long-term survival.

Venom

In addition to natural history tempering concerns about the population-level severity of threats posed by illegal collecting and fear-based killing of Abronia, it also helps to contextualize the latter viewpoint. All anguid lizards are deeply phylogenetically nested within the Toxicofera or “venom clade” (Vidal and Hedges 2005, Fry et al. 2006), which is widely recognized among contemporary herpetologists and evolutionary biologists (Pyron et al. 2013, Reeder et al. 2015,

Koludarov et al. 2017, Streicher and Wiens 2017). Although a few authors continue to question the validity of the Toxicofera (Sweet 2016, Mackessy and Saviola 2017), their arguments have been rebutted (Koludarov et al. 2017; Streicher and Wiens 2017). The Toxicofera clade suggests a single origin of venom that has been repeatedly lost and elaborated among its constituent lineages, and it has motivated a surge of research into the oral secretions of many squamates previously believed non-venomous. The only such study on Abronia involves A. graminea

(Koludarov et al. 2012). These authors found that the mandibular glands on A. graminea are

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well-developed and serous, and they recovered several secreted peptides homologous with those present in venomous helodermatid lizards. Although Koludarov et al. (2012) emphasized that these peptides constitute insufficient evidence to consider A. graminea dangerously venomous to humans, they suggested that the species’ venom is relevant both biologically and ecologically.

Furthermore, they did not rule out the possibility of human allergic responses to Abronia oral secretions, although no clinical symptoms from Abronia bites have ever been reported in the literature. In fact, limited anecdotal evidence indicates that Abronia bites pose no serious health risk to humans (Smith 1941, Wagner 2010), a conclusion supported by my personal lack of ill effects following intentionally inflicted bites from wild A. bogerti, A. graminea, and A. taeniata.

Despite its limitations, available data appears to suggest that the widespread antipathy directed toward Abronia by indigenous residents is overly stringent.

Activity Patterns and Reproductive Timing

All available observations indicate that Abronia are strictly diurnal (e.g., Campbell and

Frost 1993, Campbell and Brodie Jr 1999, Carabias Lillo et al. 2000), with no published account of any Abronia exhibiting crepuscular or nocturnal activity. In general, Abronia appear to be most vagile and visible to human observers on warm, sunny days. Campbell and Brodie (1999) indicated as such for A. meledona, and suggested that it was true for the genus more generally.

Campbell et al. (1998) also reported a failure to find A. frosti on cold, rainy, overcast days at the type locality, compared to a survey yield of five lizards with similar search effort on warm, sunny days. However, I have observed several A. graminea active and awake in trees during cold, overcast, misty days, and Chapter 5 describes movements by A. graminea during similar conditions. A gravid adult female A. matudai (UTA R-40643) was found on a cloudy and rainy day among the rocks of a roadside wall, at an ambient temperature of 17°C. These observations

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are unsurprising, given that two unpublished theses (Villamar Duque 1998, Fierro Estrada 2013) indicate that A. graminea, A. mixteca and A. taeniata are eurythermic as is the related alligator lizard Elgaria multicarinata in California, USA (Kingsbury 1994). Physiologically, as a genus

Abronia are thus likely much more tolerant of inclement conditions than most other lizards.

Across all alligator lizards, much of the scant knowledge available on reproduction comes from animals kept in captivity, and Abronia are no exception. Much of the available reproduction-related data for Abronia was summarized by Schmidt-Ballardo et al. (2015), with additional information reported by other authors (Davis and Dixon 1961, Carpenter 1978,

Campbell 1984, Ariano-Sánchez and Melendez 2009, Vega-Trejo 2010, Lemos-Espinal and

Dixon 2016, Aldape-López and Santos-Moreno 2018). Despite over half of all Abronia species lacking any published data on their reproductive biology, the information that is available is consistent. Members of the genus are live-bearers, breed in the summer/fall, and give birth in the spring.

These data suggest that the best diel/seasonal conditions to survey for wild Abronia are warm, sunny days during the summer/fall breeding season. Such conditions, although by no means required for successful searches, will likely maximize detection rates given that the lizards will be most active and hence more easily visible to human observers.

Diet, Predation, and Parasites

Trophic dynamics of Abronia remain mysterious, and much of the available literature on this topic is hypothetical in nature. Save for my personal observation of a subadult A. taeniata catching and eating an unidentified species of cricket from a shallow cavity on a mossy tree trunk near La Mojonera, Hidalgo, I am unaware of any dietary data for wild Abronia.

Speculation abounds, however. Koludarov et al. (2012) stated that A. graminea “is a generalist

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predator of both invertebrates and small vertebrates such as skinks,” but they provide no data or citations to support this claim. Bogert and Porter (1967) surmised that epiphyte-dwelling invertebrates form a large component of the diet of wild Abronia, an idea implicitly supported by

Smith (1941). Multiple authors (Gadow 1908, Canseco-Márquez and Gutiérrez-Mayén 2010,

Carabias Lillo et al. 2000) cumulatively indicate that wild A. fuscolabialis, A. graminea, A. mixteca, and A. oaxacae eat insects, but do not substantiate their assertions. Without providing supporting evidence, Marschang et al. (2002) characterized the genus, probably erroneously, as

“omnivorous.” As with the topic of reproduction, observations from captive lizards provide a clearer picture. Several authors indicate that grasshoppers and other orthopterans seem to be especially favored by captive Abronia (Ariano-Sánchez and Melendez 2009, Dixon and Lemos-

Espinal 2010, Wagner 2010). However, these and other authors also report that captive Abronia accept virtually every type of feeder insect commonly available in the hobbyist trade: crickets, cockroaches, mealworms, and waxworms (Formanowicz et al. 1990, Marschang et al. 2002,

Langner 2007, Wagner 2010, Schmidt-Ballardo et al. 2015, Clause et al. 2016c). Wagner (2010) indicates that Abronia will also take sphingid moth caterpillars, soldier fly larvae, spiders, and snails, while young locusts and “meadow plankton” are reported as prey taken by A. graminea specifically (Langner 2007). Cumulatively, this data appears to suggest that Abronia are generalist consumers of arthropods. However, the inclusion of small-bodied vertebrate prey in their diet remains possible, as does a potential preference for certain type(s) of prey. There is no evidence available that Abronia regularly consume highly venomous or toxic prey like centipedes or scorpions, nor that they sequester ingested toxins in their body tissues. However,

Abronia likely could consume such prey, and this may contribute in some way to the common perspective among co-occurring human communities that the lizards are themselves dangerously

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venomous.

Given the arboreality of Abronia, birds likely represent a major guild of predators as suggested for the alligator lizard Coloptychon rhombifer, a similarly arboreal species (Lamar et al. 2015). A female resplendent quetzal, Pharomachrus mocinno was documented carrying an adult Abronia to feed her nestlings (Eisermann 2013), and a male P. mocinno was photographed with a dead adult Abronia in its bill (Ariano-Sánchez et al. 2009). In both cases, the lizards cannot be confidently identified to species but were either A. gaiophantasma or A. fimbriata. The stomach contents of an adult male great black hawk, Buteogallus urubitinga (specimen MVZ

120962) from Chiapas, Mexico near the Guatemala border contained a lizard identified as A. lythrochila. Both P. mocinno and B. urubitinga were previously known to include other lizards in their diet (Wheelwright 1983, Gerhardt et al. 1993, Ávila et al. 1996). Other such lizard-eating birds that occur in Abronia habitat include numerous raptors such as kites, hawks, falcons, and forest-falcons; additional species of trogons in the genus Trogon; various species of motmots; the bright-rumped attila, Attila spadiceus; masked and black-crowned tityras, Tityra semifasciata and T. inquisitor; and jays in the genera Aphelocoma, Cyanocorax, Cyanocitta, Cyanolyca, and

Psilorhinus (Howell and Webb 1995). Terrestrial and semi-arboreal snakes are also known to eat

Abronia. A single A. smithi was recovered from the stomach of an adult Godman’s montane pitviper, Cerrophidion godmani along with parts of both a rodent and a hylid frog (Campbell and

Frost 1993). Flores-Villela and Sanchez-H. (2003) reported finding an A. martindelcampoi in the stomach of a Mexican horned pitviper, Ophryacus undulatus. Both of these viper species are documented predators of other lizard species, and O. undulatus is reported to climb up to 4 m into trees (Campbell and Lamar 2004). I expect other lizard-eating snake taxa that co-occur with

Abronia will eventually be revealed as natural predators, including neotropical racers,

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Drymobius spp.; milksnakes, Lampropeltis spp.; parrot snakes, Leptophis spp.; the tropical ratsnake, Pseudelaphe flavirufa; jumping pitvipers, Atropoides spp.; palm pitvipers, Bothriechis spp.; and additional species of both Cerrophidion and Ophryacus (Campbell and Lamar 2004,

Köhler 2008). Although no observations yet exist, I consider it probable that some mammals occasionally eat Abronia as well, especially those considered arboreal or semi-arboreal and/or that take lizards as prey. Such mammals include opossums in the genera Caluromys, Metachirus, and Philander; tayra, Eira barbara; jaguarundi, Herpailurus yagouaroundi; margay, Leopardus wiedii; ring-tailed cat, Bassariscus astutus; cacomistle, B. sumichrasti; and perhaps the woolly false vampire bat, Chrotopterus auritus and great false vampire bat, Vampyrum spectrum (Reid

2009, Ceballos 2014).

With the exception of “heavy infestations” of an unidentified trombiculid mite along the lateral fold of all known individuals of A. ornelasi (Campbell 1984), no external or internal parasites have yet been documented on wild or captive Abronia. This remains an area of much- needed research in this genus. Given the unusual arboreal habits of these lizards, and their discontinuous distribution across Mesoamerica, they could be hosts to new, undescribed species of parasites (Davis et al. 2012), which would further emphasize the conservation importance of

Abronia.

Preliminary Protected Area Gap Analysis

The existing protected area network is incomplete in its coverage of Abronia taxonomic diversity (Table 4.2), and also in its coverage of major mountain ranges. Over one-third (13/29) of Abronia species are neither known nor expected to occur in existing protected areas, and I consider them “gap species” of high conservation priority (Rodrigues et al. 2004). These include

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five Guatemalan endemics and eight Mexican endemics. At the other extreme, a single protected area encompasses the entire known and expected range of three species: A. chiszari, A. ramirezi, and A. reidi, all Mexican endemics. The Sierra de los Tuxtlas and Sierra Madre de Chiapas mountain ranges in Mexico, and the Sierra de las Minas of Guatemala, are entirely or largely encompassed by Reservas de las Biosferas (Biosphere Reserves), indicating their excellent coverage at least on paper. In Guatemala, the eastern massifs of the Central Plateau lack recognized protected areas and the expansive Sierra de los Cuchumatanes supports only a single protected area, the newly established Yal Unin Yul Witz Ecological Preserve (Vásquez-Almazán

2016). In Mexico, the Chimalapas highlands, Sierra Mixe, Sierra de Miahuatlán, Sierra de

Quatro Venados, Mixteca Alta, and Sierra Madre del Sur of Guerrero all lack recognized protected areas.

Of the Abronia species that are known to occur in at least one protected area, those reserves provide coverage that is likely sufficient to support viable populations well into the future, so long as habitat within the reserves remains truly protected from deforestation. Most montane reserves in Mesoamerica are centered on one or more major peaks or ridgelines, and hence they provide extensive rather than peripheral coverage of high-elevation forest habitat for

Abronia—both currently and when considered under the reality of future altitudinal shifts due to climate change. Only one of the 28 confirmed protected areas with Abronia that I identified

(Reserva de la Biósfera Tehuacán-Cuicatlán, which supports the Mexican endemic A. oaxacae) just barely covers habitat suitable for the species. Even the eight comparatively small state, municipal, and private protected areas 60–1,000 ha in size confirmed to harbor Abronia could presumably support viable populations in the long term, assuming that the comparatively low vagility and adaptability to forest fragmentation documented in Chapter 5 for A. graminea hold

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true for the entire genus. These eight reserves all lie entirely, or almost entirely, within the known elevational range of the Abronia species they protect.

For three Abronia species (A. deppii of Mexico, and A. fimbriata and A. gaiophantasma of Guatemala), the number of protected areas in which they are known to occur is exceeded by the number in which they are expected to occur (Table 4.2). This showcases the opportunity to promote Abronia-focused biodiversity surveys in those areas. Five additional species are expected from as many protected areas as they are currently known to occupy, further emphasizing the need for additional search effort. Few protected areas in Mexico and Guatemala have been thoroughly surveyed by biologists, and the flagship status of Abronia could help stimulate funding for much-needed biodiversity inventories in these areas (see Griffin 2015 for an example). Existing Abronia detections within protected areas usually pre-date the establishment and recognition of those protected areas. Of the 28 protected areas that include one or more known Abronia populations, only 10 yielded their first Abronia detection after official designation of protected status. As such, any potential bias toward Abronia survey effort inside protected areas likely has only minimally influenced our current understanding of their locality- level distribution, and it implies that the majority of coverage gaps are not an artifact.

The quality and effectiveness of many reserves in the existing protected area network is questionable. For over one-third (6/16) of Abronia species known or suspected to occur in a protected area, the parcels in question are categorized as Biosphere Reserves (Reservas de las

Biosferas) or National Parks (Parques Nacionales) (Table 4.2). These federal designations often represent little more than paper parks, with minimal government enforcement or buy-in from local communities. Their boundaries regularly include entire towns, whose members generally were not consulted during the federal designation process and many of whom may consider the

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government designation illegitimate. In Mexico, where land ownership regimes are dominated by communally-owned land known as ejidos, federal protected areas are especially problematic because most or all of the land is neither owned nor managed by the government (Figueroa and

Sánchez-Cordero 2008, Halffter 2011). Widespread, ongoing forest conversion is documented in many of these protected areas across Mesoamerica, including those known to support Abronia

(Wilson and McCranie 2004a, Figueroa and Sánchez-Cordero 2008, Toledo-Aceves et al. 2011,

McCranie and Köhler 2015). For these reasons, the de-facto protected area landscape as it relates to Abronia is less meaningful than it might appear at first glance.

Protected areas are immediately relevant to Abronia in light of the diverse threats to the integrity of their montane forest habitat (Toledo-Aceves et al. 2011). Destruction and fragmentation of forests inhabited by Abronia are attributable to clearing for firewood (Ariano-

Sánchez et al. 2011, Pope et al. 2015), charcoal production (Clause et al. 2016b), food and ornamental crop production (Torres-Almazán and Urbina-Aguilar 2011, Pope et al. 2015), and cattle grazing (Ariano-Sánchez and Torres-Almazán 2010). Furthermore, elevational contraction of forest due to the effects of climate change (Ponce-Reyes et al. 2012, Rojas-Soto et al. 2012), along with anthropogenic forest fires (Asbjornsen et al. 2005, Lamoreux et al. 2015) and the harvest of epiphytes for sale as ornamentals (Chaparro and Ticktin 2011, Aranda-Coello et al.

2012) are all additional vectors for habitat loss in areas that support Abronia. If these ecosystem- scale threats are to be mitigated, they must receive priority attention by practitioners.

Threatened Species Classification Systems

Although most classification systems would benefit from substantial updates, the various status listings applied to Abronia reveal a consistent assessment of the genus as highly at-risk,

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whether considered independently or together(Table 4.2). The IUCN Red List categorizes the majority (19/29) of Abronia species as threatened (either Vulnerable, Endangered, or Critically

Endangered). An additional seven species are classified as Data Deficient. Two species (A. lythrochila and A. smithi) are considered Least Concern, and A. cuetzpali remains Not Evaluated due to its recent recognition as a new species. Many of these assessments need revision, given that 16 species were evaluated over a decade ago and many assessments do not account for contemporary data. The average Environmental Vulnerability Score (EVS) for the entire genus is

16.9 out of 20, with individual species scores ranging from 15–18. Under the EVS system, a score of 14 or higher is considered to represent “high vulnerability.” All scores were produced between 2015 and 2017 (Johnson et al. 2015, Johnson et al. 2017). In early 2017, five

Guatemalan species of Abronia (Auriculabronia: A. anzuetoi, A. campbelli, A. fimbriata, A. meledona; Lissabronia: A. frosti) were listed under the most stringent CITES category, Appendix

I, in which all commercial trade is prohibited. The remaining members of the genus were placed on Appendix II, under which trade is allowed only in accordance with a CITES export permit.

However, there is a zero export quota for wild specimens of five Appendix II species from

Guatemala, Honduras, and El Salvador (Auriculabronia: A. aurita, A. gaiophantasma, A. vasconcelosii; Lissabronia: A. montecristoi, A. salvadorensis). In Mexico, the federal Secretaría de Medio Ambiente y Recursos Naturales (SEMARNAT) agency lists all but four of the 19

Mexican Abronia species on its Norma Oficial Mexicana special status list (NOM-059-

SEMARNAT-2010). However, substantial revision to this list is needed (García-Aguilar et al.

2017). Imperilment severity categorizations assigned by SEMARNAT to many Abronia species are idiosyncratic and unexplained, and three species (A. leurolepis, A. ramirezi, and A. smithi) were erroneously excluded from consideration. In Guatemala, the federal Consejo Nacional de

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Áreas Protegidas (CONAP) includes all but three of the 10 Guatemalan Abronia species on its

Lista de Especies Amenazadas (LEA; CONAP, 2009). This national list warrants an update to add the three currently absent Guatemalan species (A. lythrochila, A. meledona, and A. vasconcelosii).

Cumulatively, these conservation status listings consistently rank Abronia as arguably the most threatened squamate genus in mainland Mesoamerica. They are rivaled only by spiny-tailed iguanas in the genus Ctenosaura, a group of 18 recognized species that occur in low- to mid- elevation forested and scrubby habitats from Mexico to Colombia (Buckley et al. 2016). The majority of Ctenosaura species (10 of 18) are categorized as threatened on the IUCN Red List and the genus has an average EVS score of 16.9 just like Abronia (Johnson et al. 2015, Johnson et al. 2017), but only four species are listed under CITES (all Appendix II). Notwithstanding the substantial uncertainty regarding the true status of many species due to lack of data, the emergence of Abronia as a regional flagship genus appears sound because of their broad inclusion in multiple national and international listing classifications, and is bolstered by their eye-catching physical appearance (Ariano-Sánchez and Melendez 2009; Figure 4.1).

Discussion

Despite limited data availability, my review indicates that the genus Abronia deserves recognition as one of the most threatened vertebrate clades in Mesoamerica. Their generally restricted geographic ranges, ongoing anthropogenic conversion of their montane forest habitat, forecasted contraction of this habitat due to climate change, and removal of animals from natural populations due to illegal wildlife trafficking and fear-based killing, all contribute to the severe imperilment of Abronia. In recognition of these threats to their ongoing survival, the five

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endangered species classification systems that I reviewed are remarkably consistent in their assessment of the genus as a highly at-risk taxon, although updates to these listings are needed.

These factors, coupled with the striking morphology and bright coloration of most Abronia, justify their emerging status as conservation flagships for the preservation of mesic high- elevation forests.

A major complication in the promotion of Abronia as flagships, however, is the antipathy often directed toward them by coexisting human communities that consider the lizards potently venomous. This reality cannot be ignored, and educational campaigns to change those behaviors will be critically important to the success of any conservation program with Abronia as a linchpin. One Mesoamerican reptile species, the Guatemalan beaded lizard Heloderma charlesbogerti, is promoted as a surrogate for forest conservation (Ariano-Sánchez and Salazar

2015) despite its highly venomous bite (Ariano-Sánchez 2008)—emphasizing that even potentially dangerous species can be useful for advancing conservation awareness. Although aversion toward objectively harmless native reptiles by local communities is documented for numerous other genera in Mesoamerica and globally (Ceríaco 2012, Ceríaco and Marques 2013), it remains unclear if Abronia possess venom that is clinically harmful to humans. Mandibular secretions homologous to those of Heloderma are documented in the one Abronia species studied to date (A. graminea; Koludarov et al. 2012). Although those authors considered the species “technically” venomous and suggest that its venom is both biologically and ecologically relevant, they caution that their results do not imply a danger to humans. Additional research is needed to determine the true biological effects of Abronia bites, especially given the intraspecific variability in oral gland secretions documented in another anguimorph lizard genus (Varanus;

Koludarov et al. 2017). In particular, the possibility of human allergic reactions to otherwise

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harmless Abronia oral secretions warrants attention. Nonetheless, there is no scientifically verified, published evidence that Abronia bites produce clinical effects in humans or other animals. Important steps are already being taken to strategically communicate this information to indigenous communities and other stakeholders (Sánchez-Herrera et al. 2017), but more work is needed. Based on my personal experience with a resident community at one field site (discussed in detail in Chapter 5), such educational efforts can successfully lead to behavior change. After multiple demonstrations in which I allowed adult Abronia graminea to bite the soft tissue of my hands during my 10-week stay in the community, many residents appeared to accept that the lizards did not pose a serious threat to humans or livestock. By the end of my visit several residents, including an influential landowner with high standing in the community, began to voluntarily advocate among their peers for the cessation of lethal action when these lizards were encountered.

A second social incommensurability for the Abronia conservation flagship narrative is their valuation in the black-market hobbyist trade. Despite limited available data on the severity of this trade in Abronia, it drove the recent listing of the entire genus under Appendix I and II of

CITES. Nonetheless, although ongoing illicit traffic and possession of wild-caught Abronia for the hobbyist market is problematic and reprehensible, it remains unclear if the wildlife trade is a biologically-relevant stressor to the genus as a whole. Complicating this judgment are inherent challenges in attempts to quantify black-market clandestine activities, a process that is prone to underestimation (Salzberg 1996, Schlaepfer et al. 2005). Nonetheless, a perspective rooted in natural history suggests that even when coupled with misguided killing of lizards, poaching is likely not driving the endangerment of most Abronia species. Their secretive and strongly arboreal habits make the majority of any given population consistently unavailable to humans

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(see Chapter 5). Clear exceptions include those few species that exist only in small fragments of declining habitat on a single, relatively accessible and populated mountaintop (A. campbelli, A. meledona, and possibly A. frosti), for which collection pressure and killing are probably meaningful contributors to their extinction risk. Furthermore, arguments for withholding exact locality data for marketable taxa like Abronia from publications (Stuart et al. 2006, Lindenmayer and Scheele 2017) are well supported, involve little downside, and deserve consideration by the academic Abronia community. Yet despite a clear need in Mesoamerica for better enforcement of wildlife trade laws and more severe penalties to dissuade potential violators (Auliya et al.

2016), such actions alone appear unlikely to safeguard any Abronia in the long term. The rapid rate of forest destruction, and the forecasted contraction of these forests due to climate change, are far more important and widespread threats—albeit more difficult and intractable to address.

Targeted survey effort to document new Abronia populations remains an urgent priority, particularly in the protected areas identified here as potentially supporting one or more species.

To maximize the efficiency and success of such surveys, careful attention should be given to the natural history of the genus, including diel and seasonal weather conditions. To promote rigorous scientific investigation of the diversity within this genus, acquisition of additional physical material and especially DNA samples also remains a priority (Clause et al. 2018), similar to many other Mesoamerican herpetofauna (Clause et al. 2016a). Although it has been suggested that scientific collecting and conservation can be at odds (Minteer et al. 2014, Henen 2016), these arguments are largely speculative and have been repeatedly refuted by a broad spectrum of the scientific community (Krell and Wheeler 2014, Poe and Armijo 2014, Rocha et al. 2014,

Hope et al. 2018). I join these latter scientists in encouraging educators, academics, and particularly resource managers to embrace the enduring value of limited scientific collecting.

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This is especially true for squamates of conservation concern such as Abronia, for which knowledge deficiency continues to be a serious challenge (Böhm et al. 2013, Andrade-Díaz et al.

2017, Meiri et al. 2018).

There is a great need to expand the existing protected area network in montane regions of

Mexico and Guatemala. This conclusion, and the priority areas identified here, are largely congruent with other prioritization studies (Ariano-Sánchez 2010, Urbina-Cardona and Flores-

Villela 2010, Toledo-Aceves et al. 2011, Ochoa-Ochoa et al. 2017). It remains equally important to strengthen government-private partnerships in existing protected areas, to boost the on-the- ground meaningfulness of Biosphere Reserves and National Parks (Halffter 2011). Furthermore, given the institutional dominance of communal and private lands in Mesoamerica and particularly in Mexico, establishment of community and/or private reserves likely represents a productive avenue for advancing conservation practice in the region (Ochoa-Ochoa et al. 2009).

Given the patchy distribution of Abronia and their declining montane forest habitats, a small- scale local approach toward land preservation and management could be particularly effective.

However, significant geopolitical and institutional challenges to this process remain, emphasizing the complexity of conservation in the developing world (Barrett et al. 2001, Ochoa-

Ochoa et al. 2009, Pope et al. 2015). For instance, ongoing civil unrest in Mexico’s Chimalapas highlands, particularly along the contested border between Chiapas and Oaxaca, complicate any conservation initiatives for the two Abronia (A. ornelasi and A. bogerti) that are endemic or nearly endemic to that area (Lamoreux et al. 2015, Clause et al. 2016b). Ongoing violence related to drug trafficking and clandestine marijuana and opium poppy cultivation in the mountains of Guerrero, Michoacán, and Morelos in Mexico also forestalls conservation work in those areas, which support A. deppii, A. martindelcampoi, and A. mixteca. However, in some

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cases geopolitical conflict may offer de-facto protection, such as the land mines on Volcán de

Agua in Guatemala that limit incursion into the only known habitat for A. anzuetoi (Ariano-

Sánchez et al. 2014).

Ultimately, the preservation of Abronia is best approached as a means for promoting the conservation of their declining montane forest habitat, in the context of its status as a biodiversity hotspot that supports numerous other sympatric species of conservation concern. Such co- occurring threatened taxa include other squamates (Mesaspis spp., Exiliboa placata, Ungaliophis continentalis, Bothriechis spp., Ophryacus spp.) and numerous amphibians (Dendrotriton spp.,

Isthmura spp., Ixalotriton spp., Ecnomiohyla spp., Sarcohyla spp., Pletrohyla spp.), along with several imperiled birds (e.g., Pharomacrus mocinno, Oreophasus derbianus, Harpia harpyja) and certain high-profile mammals (e.g., Tapirus bairdi and Panthera onca) (Howell and Webb

1995, Köhler 2008, 2011, Ceballos 2014). Nonetheless, this biological viewpoint represents just one lens for analyzing and promoting the complex process of robust, lasting conservation in montane Mesoamerica. The highly heterogeneous linguistic, cultural, and sociopolitical dynamics of this region form an inextricable, and equally important, component of the biological diversity of such landscapes (Wilder et al. 2016). Much additional research remains necessary to better understand the institutions, cultural practices, belief systems, and possible leverage points useful for promoting equitable conservation programs over the long term. In reviewing the current state of scientific knowledge associated with Abronia here, I demonstrate the applied value of a natural history perspective for properly contextualizing threats, evaluating the suitability of the genus as a flagship, and maximizing the success and impact of much-needed future work. Through this taxonomically-focused survey of an imperiled clade both challenges

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and opportunities have emerged, which I hope will encourage pluralistic, strategic ongoing study and conservation practice in Mesoamerican highland forests.

Acknowledgements

I thank the following people and museums for sharing data on their Abronia holdings, and for allowing me to report on that data: David Dickey (AMNH), Ned Gilmore (ANSP), Jack

Sites and Wesley Skidmore (BYU), Jens Vindum (CAS), Stephen P. Rogers (CM), Víctor Hugo

Reynoso-Rosales (CNAR), John Friel (CUMV), Héctor Eliosa León (EBUAP), Rubén Castro

Franco (EBUM), Antonio Muñoz Alonso (ECO-SCH), Juan Carlos López Vidal (ENCB-IPN),

Alan Resetar (FMNH), Froilán Esquinca Cano and Roberto Luna Reyes (IHN), Rafe M. Brown and Richard E. Glor (KU), Gregory B. Pauly and Neftali Camacho (LACM), Christopher C.

Austin and Catherine E. Newman (LSUMZ), Jose Rosado (MCZ), Annemarie Ohler (MNHN),

Julia Colby (MPM), Tom Giermakowski (MSB), Carol L. Spencer (MVZ), Adrián Nieto-Montes de Oca and Edmundo Pérez-Ramos (MZFC), Patrick Campbell (NHMUK), Heinz Grillitsch

(NMW), Gary Shugart (PSM), Bradford D. Hollingsworth and Laura Kabes (SDSNH), Gunther

Köhler and Linda Acker (SMF), Andreas Schlüter and Alexander Kupfer (SMNS), Toby Hibbitts

(TCWC), Travis J. LaDuc (TNHC), Irene Goyenechea (UAEH), David Lazcano (UANL), Emily

M. Braker (UCM), Christopher A. Phillips and Christine A. Mayer (UIMNH), Greg Schneider

(UMMZ), Jeremy F. Jacobs (USNM), Carl J. Franklin (UTA), Gregory Watkins-Colwell (YPM), and Mark-Oliver Rödel and Frank Tillack (ZMB). I am indebted to Adrián Nieto-Montes de Oca,

Israel Solano-Zavaleta, Walter Schmidt-Ballardo, and Gustavo Jiménez-Velázquez for providing vital input and aid in the field and lab, without which this work would have been impossible. I thank César T. Aldape-López, Leonardo Fernández-Badillo, and Nelson Martín Cerón de la Luz

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for sharing key information. Funding support provided by a University of Georgia Presidential

Fellowship.

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