Endangered and Threatened Wildlife and Plants; Endangered Species Status for Beardless Chinchweed With Designation of Critical Habitat, and Threatened Species Status for Bartram's Stonecrop With Section 4(d) Rule

DEPARTMENT OF THE INTERIOR

Fish and Wildlife Service

50 CFR Part 17

[Docket No. FWS-R2-ES-2018-0104; 4500030113]

RIN 1018-BD35

Endangered and Threatened Wildlife and Plants; Endangered Species Status for Beardless Chinchweed With Designation of Critical Habitat, and Threatened Species Status for Bartram's Stonecrop With Section 4(d) Rule

AGENCY:

Fish and Wildlife Service, Interior.

ACTION:

Proposed rule.

SUMMARY:

We, the U.S. Fish and Wildlife Service (Service), propose to list

Pectis imberbis

(beardless chinchweed), a plant species from southern Arizona and northern Mexico, as an endangered species and to designate critical habitat for Beardless chinchweed under the Endangered Species Act of 1973 (Act), as amended. In total, we propose to designate approximately 10,604 acres (4,291 hectares) in southern Arizona as critical habitat for this plant. We also announce the availability of a draft economic analysis of the proposed designation of critical habitat for beardless chinchweed.

In addition, we propose to list

Graptopetalum bartramii

(Bartram's stonecrop), a plant species from southern Arizona and northern Mexico, as a threatened species under the Act and to issue a rule under section 4(d) of the Act to provide for the conservation of Bartram's stonecrop. We are not proposing to designate critical habitat for Bartram's stonecrop because we find that a designation is not prudent. If we make this rule final as proposed, it would extend the Act's protections to both of these species and to beardless chinchweed's critical habitat.

DATES:

We will accept comments received or postmarked on or before February 4, 2020. Comments submitted electronically using the Federal eRulemaking Portal (see

ADDRESSES

, below) must be received by 11:59 p.m. Eastern Time on the closing date. We must receive requests for a public hearing, in writing, at the address shown in

FOR FURTHER INFORMATION CONTACT

by January 21, 2020.

ADDRESSES:

Written comments:

You may submit comments by one of the following methods:

(1)

Electronically:

Go to the Federal eRulemaking Portal:

http://www.regulations.gov.

In the Search box, enter FWS-R2-ES-2018-0104, which is the docket number for this rulemaking. Then, click on the Search button. On the resulting page, in the Search panel on the left side of the screen, under the Document Type heading, click on the Proposed Rule box to locate this document. You may submit a comment by clicking on “Comment Now!”

(2)

By hard copy:

Submit by U.S. mail or hand-delivery to: Public Comments Processing, Attn: FWS-R2-ES-2018-0104; U.S. Fish and Wildlife Service, MS: BPHC, 5275 Leesburg Pike, Falls Church, VA 22041-3803.

We request that you send comments only by the methods described above. We will post all comments on

http://www.regulations.gov.

This generally means that we will post any personal information you provide us (see

Public Comments,

below, for more information).

Document availability:

The draft economic analysis is available at

http://www.fws.gov/southwest/es/arizona/Docs_Species.htm,

at

http://www.regulations.gov

at Docket No. FWS-R2-ES-2018-0104, and at the Arizona Ecological Services Field Office (see

FOR FURTHER INFORMATION CONTACT

).

The coordinates or plot points or both from which the map is generated are included in the administrative record for this critical habitat designation and are available at

https://www.fws.gov/southwest/es/arizona/Docs_Species.htm,

at

http://www.regulations.gov

at Docket No. FWS-R2-ES-2018-0104, and at the Arizona Ecological Services Field Office (see

FOR FURTHER INFORMATION CONTACT

). Any additional tools or supporting information that we may develop for this critical habitat designation will also be available at the Fish and Wildlife Service website and Field Office set out above, and may also be included in the preamble and/or at

http://www.regulations.gov.

FOR FURTHER INFORMATION CONTACT:

Jeff Humphrey, Field Supervisor, U.S. Fish and Wildlife Service, Arizona Ecological Services Field Office, 9828 North 31st Avenue, #C3, Phoenix, AZ 85051-2517; telephone 602-242-0210. Persons who use a telecommunications device for the deaf (TDD) may call the Federal Relay Service at 800-877-8339.

SUPPLEMENTARY INFORMATION:

Executive Summary

Why we need to publish a rule.

Under the Act, if a species is determined to be an endangered or threatened species throughout all or a significant portion of its range, we are required to promptly publish a proposal in the

Federal Register

and make a determination on our proposal within 1 year. Under section 4(d) of the Act, the Secretary of the Interior has the discretion to issue such regulations as he deems necessary and advisable to provide for the conservation of threatened species. Critical habitat shall be designated, to the maximum extent prudent and determinable, for any species determined to be an endangered or threatened species under the Act. Listing a species as an endangered or threatened species, adopting provisions under section 4(d) of the Act for a threatened species, and designations and revisions of critical habitat can only be completed by issuing a rule.

What this document does.

We propose to list beardless chinchweed as an endangered species and Bartram's stonecrop as a threatened species. This proposed rule assesses all available information regarding status of and stressors to beardless chinchweed and Bartram's stonecrop. We also propose a rule issued under section 4(d) of the Act to provide for the conservation of Bartram's stonecrop. In addition, we propose to designate critical habitat for beardless chinchweed. We are not proposing critical habitat for Bartram's stonecrop as we have determined that the designation of critical habitat for this species is not prudent.

The basis for our action.

Under the Act, we can determine that a species is an endangered or threatened species based on any of five factors: (A) The present or threatened destruction, modification, or curtailment of its habitat or range; (B) overutilization for commercial, recreational, scientific, or educational purposes; (C) disease or predation; (D) the inadequacy of existing regulatory mechanisms; or (E) other natural or manmade factors affecting its continued existence.

For beardless chinchweed, we have determined that the key factors supporting the proposed endangered finding are: Loss of habitat due to invasion by nonnative species (Factor A); altered fire regime exacerbated by nonnative invasion (Factors A and E); altered precipitation, drought, and temperature (Factors A and E); road and trail maintenance, mining, livestock, wildlife, and post-wildfire runoff (Factors A and E); grazing from wildlife and livestock (Factor C); and small population size exacerbating all other stressors (Factor E). The existing regulatory mechanisms are not adequate to address these factors such that the species does not meet the definition of

an endangered or threatened species (Factor D).

For Bartram's stonecrop, we have determined the key factors supporting the proposed threatened finding are: Reduction in water availability (Factors A and E); erosion, sedimentation, and burial (Factors A and E); trampling (Factor E); altered fire regime (Factors A and E); loss of shade (Factors A and E); altered flooding regime (Factors A and E); drought (Factors A and E); predation of individuals and shade trees (Factors A, C, and E); illegal collection (Factor B); and small population size (Factor E). The existing regulatory mechanisms are not adequate to address these factors such that the species does not meet the definition of an endangered or threatened species (Factor D).

Under the Act, any species that is determined to be an endangered or a threatened species shall, to the maximum extent prudent and determinable, have habitat designated that is considered to be critical habitat. Section 4(b)(2) of the Act states that the Secretary shall designate and make revisions to critical habitat on the basis of the best available scientific data after taking into consideration the economic impact, the impact on national security, and any other relevant impact of specifying any particular area as critical habitat. The Secretary may exclude an area from critical habitat if he determines that the benefits of such exclusion outweigh the benefits of specifying such area as part of the critical habitat, unless he determines, based on the best scientific data available, that the failure to designate such area as critical habitat will result in the extinction of the species. Under section 4(d) of the Act, the Secretary of the Interior has the discretion to issue such regulations as he deems necessary and advisable to provide for the conservation of threatened species.

We prepared an economic analysis of the proposed designation of critical habitat.

In order to consider economic impacts, we prepared an analysis of the economic impacts of the proposed critical habitat designation. We hereby announce the availability of the draft economic analysis and seek public review and comment.

Peer review.

In accordance with our joint policy on peer review published in the

Federal Register

on July 1, 1994 (59 FR 34270), we have sought the expert opinions of three appropriate and independent specialists regarding the scientific information in the species status assessment upon which this proposed rule is based. The purpose of peer review is to ensure that our listing determinations and critical habitat designation are based on scientifically sound data, assumptions, and analyses. The peer reviewers have expertise with beardless chinchweed's or Bartram's stonecrop's biology, habitat, physical or biological factors, or stressors. Species status assessment reports for beardless chinchweed and Bartram's stonecrop were developed (Service 2018a and 2018b, entire), which represent a compilation of the best scientific and commercial data available concerning the status of the species, including the past, present, and future stressors to the species. We requested peer review of each species status assessment report from three independent specialists, with expertise with the species, to ensure that we based our determinations on scientifically sound data, assumptions, and analyses. The peer reviewers' comments have been considered and incorporated where appropriate in the species status assessment reports (Service 2018a and 2018b, entire), which are available at

https://www.fws.gov/southwest/es/arizona/Docs_Species.htm,

and at

http://www.regulations.gov

at Docket No. FWS-R2-ES-2018-0104. The peer review comments will be available along with other public comments in the docket for this proposed rule on

http://www.regulations.gov

(Docket No. FWS-R2-ES-2018-0104).

Information Requested

Public Comments

We intend that any final action resulting from this proposed rule will be based on the best scientific and commercial data available and be as accurate and as effective as possible. Therefore, we request comments or information from other concerned governmental agencies, Native American tribes, the scientific community, industry, or any other interested parties concerning this proposed rule. We particularly seek comments concerning:

(1) Beardless chinchweed and Bartram's stonecrop biology, range, and population trends, including:

(a) Biological or ecological requirements of these species, including habitat requirements for germination, growth, and reproduction;

(b) Genetics and taxonomy;

(c) Historical and current range, including distribution in Mexico;

(d) Historical and current population levels, and current and projected trends; and

(e) Past and ongoing conservation measures for these species, their habitats, or both.

(2) Factors that may affect the continued existence of these species, which may include habitat modification or destruction, overutilization, disease, predation, the inadequacy of existing regulatory mechanisms, or other natural or manmade factors.

(3) Biological, commercial trade, or other relevant data concerning any stressors (or lack thereof) to these species and existing regulations that may be addressing those stressors.

(4) Additional information concerning the historical and current status, range, distribution, and population size of these species, including the locations of any additional populations of these species.

(5) Information related to climate change within the range these species and how it may affect these species' habitats.

(6) Information on regulations that are necessary and advisable to provide for the conservation of these species and that the Service can consider in developing a 4(d) rule for the species. In particular, information concerning the extent to which we should include any of the section 9 prohibitions in the 4(d) rule or whether any other forms of take should be excepted from the prohibitions in the 4(d) rule.

(7) The reasons why areas should or should not be designated as critical habitat as provided by section 4 of the Act (16 U.S.C. 1531

et seq.

) including information to inform the following factors such that a designation of critical habitat may be determined to be not prudent:

(a) The species is threatened by taking or other human activity and identification of critical habitat can be expected to increase the degree of such threat to the species;

(b) The present or threatened destruction, modification, or curtailment of a species' habitat or range is not a threat to the species, or threats to the species' habitat stem solely from causes that cannot be addressed through management actions resulting from consultations under section 7(a)(2) of the Act;

(c) Areas within the jurisdiction of the United States provide no more than negligible conservation value, if any, for a species occurring primarily outside the jurisdiction of the United States;

(d) No areas meet the definition of critical habitat.

(8) The following specific information on:

(a) The amount and distribution of habitat;

(b) What areas, that are currently occupied and that contain the physical and biological features essential to the conservation of these species, should be

included in a critical habitat designation and why;

(c) Special management considerations or protection that may be needed for the essential features in potential critical habitat areas, including managing for the potential effects of climate change; and

(d) What areas not occupied at the time of listing are essential for the conservation of the species. We particularly seek comments regarding:

(i) Whether occupied areas are inadequate for the conservation of the species; and,

(ii) Specific information that supports the determination that unoccupied areas will, with reasonable certainty, contribute to the conservation of the species and, contain at least one physical or biological feature essential to the conservation of the species.

(9) Land use designations and current or planned activities in the subject areas and their possible impacts on proposed critical habitat.

(10) Any probable economic, national security, or other relevant impacts of designating any area that may be included in the final designation, and the benefits of including or excluding areas that may be impacted.

(11) Information on the extent to which the description of probable economic impacts in the draft economic analysis is a reasonable estimate of the likely economic impacts.

(12) Whether any specific areas we are proposing for critical habitat designation should be considered for exclusion under section 4(b)(2) of the Act, and whether the benefits of potentially excluding any specific area outweigh the benefits of including that area under section 4(b)(2) of the Act.

(13) The likelihood of adverse social reactions to the designation of critical habitat, as discussed in the associated documents of the draft economic analysis, and how the consequences of such reactions, if likely to occur, would relate to the conservation and regulatory benefits of the proposed critical habitat designation.

(14) Whether we could improve or modify our approach to designating critical habitat in any way to provide for greater public participation and understanding, or to better accommodate public concerns and comments.

(15) Additional guidance and methods that the Service could provide or use, respectively, to streamline the implementation of the proposed 4(d) rule for Bartram's stonecrop.

Please include sufficient information with your submission (such as scientific journal articles or other publications) to allow us to verify any scientific or commercial information you include.

Please note that submissions merely stating support for or opposition to the action under consideration without providing supporting information, although noted, will not be considered in making a determination, as section 4(b)(1)(A) of the Act directs that determinations as to whether any species is an endangered or a threatened species must be made “solely on the basis of the best scientific and commercial data available.”

You may submit your comments and materials concerning this proposed rule by one of the methods listed in

ADDRESSES

. We request that you send comments only by the methods described in

ADDRESSES

.

If you submit information via

http://www.regulations.gov,

your entire submission—including any personal identifying information—will be posted on the website. If your submission is made via a hardcopy that includes personal identifying information, you may request at the top of your document that we withhold this information from public review. However, we cannot guarantee that we will be able to do so. We will post all hardcopy submissions on

http://www.regulations.gov.

Comments and materials we receive, as well as supporting documentation we used in preparing this proposed rule, will be available for public inspection on

http://www.regulations.gov,

or by appointment, during normal business hours, at the U.S. Fish and Wildlife Service, Arizona Ecological Services Field Office (see

FOR FURTHER INFORMATION CONTACT

).

Public Hearing

Section 4(b)(5) of the Act provides for a public hearing on this proposal, if requested. Requests must be received within 45 days after the date of publication of this proposed rule in the

Federal Register

(see

DATES

, above). Such requests must be sent to the address shown in

FOR FURTHER INFORMATION CONTACT

. We will schedule a public hearing on this proposal, if requested, and announce the date, time, and place of the hearing, as well as how to obtain reasonable accommodations, in the

Federal Register

and local newspapers at least 15 days before the hearing.

Previous Federal Actions

Beardless Chinchweed

Beardless chinchweed was a candidate for listing from 1980 to 1996. It was first a Category 1 candidate species, as identified in our December 15, 1980, notice of review (45 FR 82480). Category 1 is a term no longer in use, having been replaced by the term “candidate species.” A candidate species is a species for which the Service has on file sufficient information on biological vulnerability and threat(s) to support issuance of a proposed rule to list, but issuance of the proposed rule is precluded by higher priority actions to amend the Lists of Endangered and Threatened Wildlife and Plants. In 1983, beardless chinchweed was reclassified as a Category 2 species (48 FR 53640; November 28, 1983). A Category 2 species referred to a species for which the Service had some indication that listing as endangered or threatened might be warranted, but there were insufficient data available to justify a proposal to list. The species remained so designated in subsequent annual candidate notices of review (50 FR 39526, September 27, 1985; 55 FR 6184, February 21, 1990; 58 FR 51144; September 30, 1993). In 1996, the Service eliminated Category 2 species; consequently, this species dropped off the candidate list. The Service received a petition in July 2010 to list beardless chinchweed and designate critical habitat under the Act (Center for Biological Diversity 2010, entire). The Service published a 90-day finding on August 8, 2012 (77 FR 47352), concluding that the petition presented substantial scientific or commercial information indicating that listing of the species may be warranted.

Bartram's Stonecrop

Bartram's stonecrop was a candidate for listing from 1980 to 1996. It was first a Category 1 candidate species, as identified in our December 15, 1980, notice of review (45 FR 82480), and then in 1983, it was reclassified as a Category 2 species (48 FR 53640; November 28, 1983). The species remained so designated in subsequent annual candidate notices of review (50 FR 39526, September 27, 1985; 55 FR 6184, February 21, 1990; 58 FR 51144; September 30, 1993). In 1996, the Service eliminated Category 2 species; consequently, this species dropped off the candidate list. The Service received a petition in July 2010 to list Bartram's stonecrop and designate critical habitat under the Act (Center for Biological Diversity 2010, entire). The Service published a 90-day finding on August 8, 2012 (77 FR 47352), concluding that the petition presented substantial scientific or commercial information indicating that listing of the species may be warranted.

I. Proposed Listings

Background

To provide the necessary and most up-to-date information and background on which to base our determination, we completed a species status assessment (SSA) report for beardless chinchweed (Service 2018a, entire), and an SSA report for Bartram's stonecrop (Service 2018b, entire), which are available online at

http://www.regulations.gov,

under Docket No. FWS-R2-ES-2018-0104. The SSA reports document the results of the comprehensive biological status review for each species, and each provides an account of the applicable species' overall viability through the forecasting of the condition of populations into the future. We generally define viability as the ability of the species to persist over the long term and, conversely, to avoid extinction (Service 2016, entire). In the SSA reports, we summarize the relevant biological data; describe the past, present, and likely future risk factors (causes and effects); and conduct an analysis of the viability of the species. The SSA reports provide the scientific basis that informs our regulatory decision regarding whether these species should be listed under the Act. This decision involves the application of standards within the Act, its implementing regulations, and Service policies (see Determination, below). Further, these SSA reports contain the risk analysis on which this determination is based, and the following discussion is a summary of the results and conclusions from these SSA reports. Species experts and appropriate agencies provided input into the development of these SSA reports.

Beardless Chinchweed

Beardless chinchweed is plant of the Asteraceae, or sunflower, family. Beardless chinchweed was first collected by Charles Wright in the early 1850s in Sonora, Mexico (now part of Santa Cruz County, Arizona), and was described by Asa Gray in 1853 (Phillips

et al.

1982, p. 1; Keil 1978, p. 135). The name has remained unchanged since that time, and there are no known synonyms. Based on this information as the best available scientific and commercial data, we accept the characterization of beardless chinchweed as a valid species.

Beardless chinchweed is an erect, many-branched, perennial herb growing 3 to 12 decimeters (1 to 4 feet (ft)) from a slender, woody, taprooted caudex (stem base) (Keil 1978, p. 143; Phillips

et al.

1982, p. 2; Keil 2017, pers. comm.). The glabrous (without hairs) leaves are 1 to 5 centimeters (cm) (0.4 to 2 inches (in)) in length and 1 to 2 millimeters (mm) (0.04 to 0.08 in) wide with pointed tips (Phillips

et al.

1982, p. 2). Daisy-like flower heads containing yellow ray and disk flowers are solitary or in open, flat-topped clusters at the tips of the branches (Phillips

et al.

1982, p. 2). In fruit, the heads have red to purple drying phyllaries (bracts around the flower head of a composite plant) and have small (<5 mm (0.2 in) long), spreading, awned black achenes (simple dry fruit) (Fishbein and Warren 1994, p. 19). Although we do not know exactly how long individual beardless chinchweed live, experts estimate 5 to 10 years (Keil 2017, pers. comm.).

Young beardless chinchweed plants have been noted in April (Dahlby 2017, pers. comm.), and are still present in November (Westland 2010, p. 10). Flowering occurs from August to October, when the plants are more than 0.5 meters (m) (1.6 ft) in height (Kearney and Peebles 1951, p. 935; Phillips

et al.

1982, p. 8). There have been no reports of the plant from winter months, when beardless chinchweed is presumed to die back to the ground. It is unknown how long flowers remain open. In one measurement of the number of flowers per stem, these range from 0 to 55, with an average of 28.3 per stem (Service 2015, p. 1). It was estimated that there were 6 to 8 seeds per head, resulting in a potential of roughly 832 seeds per plant, although seed loss to grazing, desiccation, and abortion were not accounted for. Germination and establishment may be sporadic or require specific conditions for success (Keil 1978, p. 144). There is no information available on the seedbank longevity of the species; however, we are aware that within populations, a variety of age classes are represented (Phillips

et al.

1982, p. 7; Service 2011, p. 4; Service 2014a, p. 2; Service 2015, p. 1; Sebesta 2017, pers. comm.). Therefore, we believe viable seeds are being produced and reproduction is occurring.

The species has been reported to reproduce both by seed and rhizomes (Westland 2010, p. 10), although there is no evidence that the species is rhizomatous (Keil 2017, pers. comm.). It is not known whether plants are able to pollinate themselves or require the pollen of another plant. However, it is likely that the plant requires pollinators. The pollinators of beardless chinchweed are not known, but other

Pectis

species are reported to be pollinated by bees and flies (Cockerell 1897, pp. 148-149; Cockerell 1911, pp. 136-137, 141-142; Simpson and Neff 1987, p. 434; Phillip

et al.

2006, pp. 532, 535-536, 538), and both an

Acmaeodera

beetle and a

Diadasia

bee were noted visiting beardless chinchweed plants (Sebesta 2017, pers. comm.). Butterflies may also use this species, as showy yellow heads containing both ray and disk flowers serve as landing platforms and are easily accessible to a variety of low energy pollinators such as butterflies (Schmitt 1980, p. 935; Keil 2017, pers. comm.).

Beardless chinchweed is typically found in oak woodlands at higher elevations, and desert grasslands and oak savannas at lower elevations (McLaughlin

et al.

2001, pp. 119, 121). However, it has also been found on disturbed road cuts, arroyo cuts, and unstable rocky slopes, where it has little competition for sunlight and nutrients (Phillips

et al.

1982, pp. 4, 6; Fishbein and Warren 1994, p. 19). It is found at elevations from 1,158-1,737 m (3,799-5,699 ft) (SEINet 2017, entire). Plants are typically noted to occur on steep, south-facing, sunny to partially shaded hillslopes, with eroding bedrock and open areas with little competition from other plants. The nonstable substrate, which could be moved through gravity, erosion, or impact, reduces competition with other vegetation, favoring beardless chinchweed. It is presumed to be a poor competitor due to its preferred open habitat and inability to find the species under dense vegetation conditions.

Beardless chinchweed requires a lack of competition from other plants. The different shaped and sized canopy and root systems of associated plant species within healthy grasslands, savannas, and woodlands create heterogeneity of form, height, and open patches needed by beardless chinchweed. Open patches are created and maintained through a variety of abiotic and biotic mechanisms (Porensky

et al.

2013, p. 591), including natural erosion (from things like precipitation events, gravity, and animals); the grazing and browsing of native animals, such as black-tailed prairie dogs (

Cynomys ludovicianus

) and pronghorn antelope (

Antilocapra americana

) (BANWR 2012, entire; Bahre 1995, p. 231; McPherson and Weltzin 2000, p. 4); and low severity, frequent wildfires (Hoffmeister 1986, pp. 194-195; McPherson and Weltzin 2000, p. 5; Brooks and Pyke 2002, p. 6; McDonald and McPherson 2011a, p. 385; Fryer and Leunsmann 2012, entire). The desert grasslands, oak savannas, and oak woodlands of southern Arizona historically had large-scale, low severity fire roughly every 10 to 20 years and following periods of adequate moisture (McPherson and Weltzin 2000, p. 5; Brooks and Pyke 2002, p. 6; McDonald and McPherson 2011a, p. 385; Fryer and

Leunsmann 2012, entire). Precipitation within the mountain ranges is bimodal, with dormant season snow and rain, and growing season monsoon rain. Data are lacking to indicate how beardless chinchweed uses dormant season versus growing season precipitation; however, we believe that dormant season precipitation is more important because this is needed for seed germination and growth.

The historical range of beardless chinchweed was larger than the current range, with a greater number of populations than persist today in southeastern Arizona and northern Sonora and Chihuahua Mexico. The historical distribution included 21 separate beardless chinchweed populations within the Atascosa-Pajarito, Huachuca, Patagonia, and Santa Rita Mountains and Canelo Hills of Cochise, Pima, and Santa Cruz Counties, Arizona, as well as in northern Chihuahua and Sonora Mexico (see Table 1, below). We define a population of beardless chinchweed as one or more subpopulations that occur within 1 kilometer (km) (0.62 miles (mi)) of other beardless chinchweed individuals allowing for gene flow and movement through cross-pollination. Because many bees and butterflies can travel a distance of 1 km (0.62 mi), we believe plants within this distance to be a single population. Subpopulations within a population are separated by between 300 and 999 m (984.3 and 3,278 ft). Of the 21 populations, 15 were in Arizona and 6 were in Mexico. The number of individuals seen historically in Mexico is not available, and no beardless chinchweed have been reported from Mexico since 1940. Nine populations and one subpopulation in Arizona have become extirpated since 1962.

Table 1—Current Status of Beardless Chinchweed Populations

Mountain range/country

Population name

Population status

Subpopulation name *

Subpopulation status

Atascosa-Pajarito Mountains, USA

Pena Blanca Lake

Extirpated

N/A

Extirpated.

Ruby Road

Extant

N/A

Extant.

Summit Motorway

Extirpated

N/A

Extirpated.

Canelo Hills, USA

Audubon Research Ranch

Extant

Post Canyon

Extirpated.

Tributary of O'Donnell Canyon

Extant.

Copper Mountain

Extirpated

N/A

Extirpated.

Harshaw Creek

Extirpated

N/A

Extirpated.

Lampshire Well

Extirpated

N/A

Extirpated.

Huachuca Mountains, USA

Scotia Canyon

Extant

N/A

Extant.

Coronado National Memorial

Extant

State of Texas Mine

Extant.

Visitor Center

Extant.

Joe's Canyon Trail

Extirpated

N/A

Extirpated.

Patagonia Mountains, USA

Flux Canyon

Extirpated

N/A

Extirpated.

Washington Camp

Extirpated

N/A

Extirpated.

Santa Rita Mountains, USA

Box Canyon Road

Extirpated

N/A

Extirpated.

McCleary Canyon—Gunsight Pass

Extant

N/A

Extant.

McCleary Canyon—Wasp Canyon

Extant

N/A

Extant.

Chihuahua, Mexico

Batopililas

Unknown; presume extant

N/A

Unknown; presume extant.

Guasaremos

Unknown; presume extant

N/A

Unknown; presume extant.

Sonora, Mexico

Canon de la Petaquilla

Unknown; presume extant

N/A

Unknown; presume extant.

Canyon Estrella

Unknown; presume extant

N/A

Unknown; presume extant.

Horconcitos

Unknown; presume extant

N/A

Unknown; presume extant.

Los Conejos

Unknown; presume extant

N/A

Unknown; presume extant.

* In this column of the table, N/A means “not applicable.”

Currently, there are 12 populations in Arizona and Mexico. In Arizona, there are currently 387 individual beardless chinchweed spread across less than 2 hectares (ha) (5 acres (ac)) within six extant populations spread across the following four mountain ranges: The Atascosa-Pajarito, Huachuca, Santa Rita mountain ranges, and the Canelo Hills (see Table 1, above). Five of the six populations in Arizona contain fewer than 50 individuals. Most of the mountain ranges in the United States have been surveyed for beardless chinchweed, and it is unlikely that any large populations remain unaccounted for therein. In addition, there are six populations in northern Mexico for which we have no current information. Inquiries between February 17 and December 12, 2017, with 11 researchers familiar with the flora of Chihuahua and Sonora revealed no information on the status of the species in Mexico. We believe these populations are extant, but with few individuals and with poor habitat condition (similar to the smallest extant populations in the United States), because much of the grasslands in beardless chinchweed' historical range in Mexico have been invaded by nonnative species (Romo

et al.,

2012, entire; Arriaga

et al.,

2004, entire).

For beardless chinchweed to maintain viability, its populations or some representative portion thereof must be resilient. Resiliency describes the ability of populations to withstand stochastic events (arising from random factors). We can measure resiliency based on metrics of population health (for example, germination versus death rates and population size). Highly resilient populations are better able to withstand disturbances such as random fluctuations in germination rates (demographic stochasticity), variations in rainfall (environmental stochasticity), or the effects of anthropogenic activities. A beardless chinchweed population with high resiliency is one in which abundance is high, the number of subpopulations is high and spatially dispersed, seed production is high,

recruitment is such that the population remains stable or increases, and the population is able to withstand stochastic events or recover to current or better condition from stochastic events from seed bank. Population resiliency categories for beardless chinchweed are described in section 3.2 of the SSA report (Service 2018a).

In addition to the above demographic needs, populations also need habitat elements for resiliency. Based on where the species has typically been found, a resilient population needs eroding granite or limestone soils or rock outcrops with native-dominated habitat, on sunny to partly shaded southern exposures. Beardless chinchweed plants are also often associated with active disturbances from frequent, low severity wildfire; grazing and browsing of native animals; and natural erosion of nonstable substrates, thus reducing competition for beardless chinchweed. In addition, resilient populations need soil moisture for seed germination, growth, and reproduction in the form of dormant season (October through March) precipitation. The minimum amount of precipitation needed for individual survival is unknown. We believe that deviation from the timing and amount of precipitation would impact the resiliency of a population, because soil moisture would be impacted. This would lead to decreased seed germination, reduced growth, reduced flowering, and decreased seed production. Further, the presence of pollinators is needed for effective fertilization, out-crossing, and seed production in beardless chinchweed. Habitat resiliency categories for beardless chinchweed are described in Table 2, below, and in section 3.2 of the SSA report (Service 2018a).

Table 2—Population Resiliency Category Definitions for Beardless Chinchweed

Condition category

Subpopulations

Abundance

Native-dominated

habitat

Dormant season (October through March)

precipitation

High (3)

Three or more subpopulations per population

Number of adults in each population is >300 individuals

No nonnative plants

More than 12 inches of winter rain on average during the past 5 years as recorded at the nearest weather station.

Moderate (2)

Two subpopulations per population

Number of individuals in each population is 100 to 300 individuals

Native plants dominate

Between 6.1 and 12 inches of winter rain on average during the past 5 years as recorded at the nearest weather station.

Low (1)

One subpopulation per population

Number of individuals in each population is <100 individuals

Mix of nonnative and native plants, where there is not a clear dominance of either

6 or fewer inches of winter rain on average during the past 5 years as recorded at the nearest weather station.

Ø

No subpopulations; population is extirpated

No individuals are found during surveys

Nonnative plants dominate the habitat

6 or fewer inches of winter rain on average during the past 5 years as recorded at the nearest weather station.

Maintaining representation in the form of genetic or ecological diversity is important to maintain the capacity of beardless chinchweed to adapt to future environmental changes. Representation describes the ability of a species to adapt to changing environmental conditions. Representation can be measured by the breadth of genetic or ecological diversity within and among populations. The more representation, or diversity a species has, the more it is capable of adapting to changes (natural or human-caused) in its environment. In the absence of species-specific genetic and ecological diversity information, we evaluate representation based on the extent and variability of habitat characteristics across the geographical range.

Genetic analysis of beardless chinchweed has not been conducted within or among populations or mountain ranges. However, populations on different mountain ranges are widely separated, making cross-pollination highly unlikely, and most of the populations contain small numbers of individuals. Therefore, there is the potential for genetic diversity among mountain ranges. However, these populations are isolated and contain small numbers of individuals. Small, isolated populations are susceptible to the loss of genetic diversity, genetic drift, and inbreeding. This could mean that between-population genetic diversity may be greater than within-population diversity (Smith and Wayne 1996, p. 333; Lindenmayer and Peakall 2000, p. 200). It is possible that there has been a loss of genetic diversity in the species due to the fact that multiple populations are already extirpated. Currently, there are six extant populations across four widely separated mountain ranges in the United States, and six populations in northern Mexico that are presumed extant.

Beardless chinchweed has been reported from both decomposing granite and limestone substrates. This variability of substrate preference may be important in maintaining environmental and genetic diversity. Similarly, the species is found over a relatively wide range of elevations of 1,158 to 1,737 m (3,799 to 5,699 ft) and vegetation communities (oak woodlands at higher elevations, and grasslands and oak savannas at lower elevations), which could be important in terms of representation. The precise genetic and ecological diversity needed is unknown, but given the loss of populations, the low number of individuals in the majority of the populations, and the distance among populations, it is likely that some diversity has been lost. Consequently, at a minimum, we likely need to retain populations throughout the range of the species to maintain the overall potential genetic and life-history attributes that can buffer the species' response to environmental changes over time.

Beardless chinchweed needs to have multiple resilient populations distributed throughout its range to provide for redundancy. Redundancy describes the ability of a species to withstand catastrophic events, measured by the number of populations, and their resiliency, distribution, and connectivity. The more populations, and the wider the distribution of those populations, the more redundancy the species will exhibit. Redundancy reduces the risk that a large portion of the species' range will be negatively affected by a catastrophic natural or anthropogenic event at a given point in time. Species that are well-distributed across their historical range are considered less susceptible to extinction and more likely to be viable than species confined to a small portion of their range (Carroll

et al.

2010, entire).

With the known six extant populations being separated by as much as 35 km (21.8 mi) in southern Arizona and even farther with the six populations believed to be extant in northern Mexico, a localized stressor such as grazing during flowering would impact only those groups of plants nearby the activity. Conversely, such distance among populations reduces connectivity among populations and mountain ranges, which may be important for genetic exchange and recolonization. Nonnative plant invasion and repeated, large-scale, moderate and high severity fires have impacted and will continue to impact many populations throughout the plant's range. The minimum number of populations needed to provide for sufficient redundancy is unknown. However, based on the number of populations now extirpated and the wide-ranging impacts from nonnatives and wildfire, the species likely needs to retain its existing population redundancy across multiple mountain ranges throughout the range to minimize impacts from catastrophic events.

Bartram's Stonecrop

Bartram's stonecrop is a plant of the Crassulaceae or stonecrop family (Phillips

et al.

1982, p. 2; Moran 1994, p. 192). Acevedo

et al.

(2004, entire) investigated the phylogenetic relationship of

Graptopetalum

and other genera of Crassulaceae. Their work clearly separates Bartram's stonecrop from other species (Acevedo

et al.

2004, p. 1101). The Flora of North America (2008, p. 227) recognizes

Graptopetalum

and

Dudleya

as distinct, and recognizes this species as Bartram's stonecrop in the genus

Graptopetalum.

Based on this information as the best available scientific and commercial data, the Service accepts this taxonomy.

Bartram's stonecrop is a small, succulent (fleshy), acaulescent (without a stem) perennial plant (Phillips

et al.

1982, p. 2; Moran 1994, p. 192). Bartram's stonecrop has a basal rosette that is 7 to 16 centimeters (cm) (2.75 to 6.3 in) wide comprised of 20 or more flat to concave, smooth, blue-green leaves (Rose 1926, p. 2; Phillips

et al.

1982, p. 2; Moran 1994, p. 192). One to seven showy inflorescences (includes stems, stalks, bracts, and flowers) up to 30.5 cm (12 in) in height are produced in equilateral panicles (pyramidal loosely branched flower cluster). The branches of the panicles produce one to six (usually three) flowers each (Rose 1926, p. 2). The fruits are follicles (capsule that splits along one side to release seeds), with minute seeds (0.5 to 0.9 mm (0.02 to 0.04 in) in length)) having little or no endosperm (tissue surrounding the embryo that provides nutrition; Shohet 1999, pp. 3, 48). The lifespan of Bartram's stonecrop is thought to be approximately 5 years (Ferguson, 2017b, tables 1-3; Ferguson 2017, pers. comm.).

The inflorescence stalks of Bartram's stonecrop grow for 30 to 40 days, around July and August, before coming to their full height, with the flowers then opening primarily between September and November (Kearney and Peebles 1951, p. 361; Phillips

et al.

1982, pp. 2, 7; Shohet 1999, p. 25). Individual flowers produce both male and female parts, but the timing of male and female flower stages differs. Individual flowers open in succession, such that the length of time each flower remains open overlaps, allowing for various stages of flowering and fruiting to be simultaneous within an individual plant for a month or more. The two stages of floral growth may reduce the probability of self-pollination, though it likely does still occur (Ferguson 2017, pers. comm.). Flowering is triggered by fall rains and does not occur during periods of water stress (Shohet 1999, pp. 22, 25, 36, 39).

Bartram's stonecrop requires pollination for reproduction. The major pollinators of Bartram's stonecrop are

Sarcophaga

spp. (true flies) and

Musca

spp. (house flies), although

Apis mellifera

(honey bee) may also play a role in pollination. Other species noted on Bartram's stonecrop include wasps, butterflies, and Tachinidae and Bombyliidae flies (Shohet 1999, p. 41; Ferguson 2014, p. 26; Ferguson 2017b, p. 13). Fertilization success is greatest in earliest opening flowers, possibly due to more pollinators being available earlier in the season, but having a long period of flowering increases overall chance of pollination (Shohet 1999, p. 57). Of the seeds produced, approximately 20 percent are viable under optimal conditions (Shohet 1999, p. 48). Because seedlings (plants less than 1.5 cm [0.6 in] in diameter) have been located in most populations, we believe pollinator availability is not a limiting factor for this species. Given their geographic location in the landscape (

i.e.,

in canyons with springs and streams), it is possible that seeds are transported by water and that populations may have been founded by a single individual plant or seed (Shohet 1999, p. 58). Seeds may also be dispersed via gravity and wind.

There is little information available regarding the seedbank of Bartram's stonecrop. In general, a seed that is very tiny has evolved a requirement of sunlight for germination, as they cannot successfully emerge from deep burial (Venable and Brown 1987, p. 360). Similarly, it is thought that Bartram's stonecrop seeds reside at the soil surface beneath the litter (Shohet 1999, p. 48). It is possible that because the seed is so small, with little endosperm, mycorrhizae (the symbiotic association of a fungus with the roots of plants) may be required for seedling establishment and growth, but this has not been studied (Felger 2017, pers. comm.). Researchers at the Desert Botanical Gardens have attempted to grow Bartram's stonecrop from seed. They had no difficulty with seed germination; however, they have experienced high seedling mortality, perhaps related to a requirement for mycorrhizae for seedling establishment.

The species typically occurs on rocky outcrops with erodible soils in deep, narrow canyons in heavy cover of litter and shade within Madrean woodlands at elevations ranging from 1,067 to 2,042 m (3,500 to 6,700 ft). Madrean woodlands are a forested community dominated by evergreen oaks, but also containing junipers and pine trees, and characterized by mild winters and warm wet summers (Brown 1982, p. 59). Madrean evergreen woodland is typically bounded by semi-desert grasslands and savanna at warmer, drier sites in the lower elevations, and by evergreen and broadleaf forests on more mesic and cooler sites at higher elevation, at north aspect, or near riparian areas. Bartram's stonecrop root into crevices on rock ledges and cliffs on slopes of various aspects (Shohet 1999, p. 22; Ferguson 2014, p. 41; NPS 2016, p. 7). In addition, Bartram's stonecrop are almost always located near water sources (springs, seeps, or intermittent streams), but above the floodline (Phillips

et al.

1982, p. 4; Shohet 1999, p. 22; NPS 2014, p. 2). Plants are typically within 10 m (32.8 ft) from a streambed in the bottom of canyons on rocky outcrops, but can be much farther on occasion (Shohet 1999, p. 5; Ferguson 2014, p. 41; NPS 2014, p. 2; Ferguson 2016a, p. 14). Based on microhabitats in which the species is typically found, the species' needs include crevices (with or without soil) for seeds to lodge and germinate, shade and deep leaf litter to help maintain soil moisture, and a humid microhabitat in this arid environment. Proximity to water may provide humidity for the plant's microclimate. The deep, narrow canyons and associated overstory species provide shade during a portion of the day, creating a cooler temperature and aiding in maintaining a humid microenvironment. In addition, the vegetation litter provides retention of

soil moisture, further promoting the humid microenvironment. The specific substrate component does not seem to be critical. In addition, for reestablishment, moist soil for seedbank may be important for this species following extended periods of drought.

Madrean evergreen woodlands of the sky island mountain ranges have evolved with frequent, low-severity fire and have warm wet summers and mild winters. The maximum interval between the relatively widespread fires typically ranged from about 10 to 30 years in the pine-dominant forests (Swetnam

et al.

2001, p. 4). Precipitation within the sky island mountain ranges is bimodal, with winter snow and rain, and summer monsoon rain. Mean annual precipitation in the Madrean woodland habitat of southern Arizona is 250 to 450 mm (10 to 17 in), with more than 50 percent occurring in summer. The winter snow and rain coincide with Bartram's stonecrop seed germination and growth. Winter precipitation is needed for Bartram's stonecrop germination (although some germination likely occurs following summer rains), and both summer (July and August) and fall precipitation (captured partially in the October and November “winter” data) is needed for Bartram's stonecrop flower production.

Bartram's stonecrop is known to have historically occurred in 33 separate populations within 13 isolated sky island mountain ranges, 10 in southern Arizona and 3 in northern Mexico. While the overall range of the species is likely unchanged, the number and size of populations has been reduced. Four populations have become extirpated in the United States in recent years, and a fifth population has contracted in size. In three instances, extirpation was associated with the drying of habitat, which rendered it no longer suitable for the species to persist; we do not know the cause of extirpation in the fourth instance. In addition, there have been many changes in the southeastern Arizona landscape since the 1890s due to intensive cattle grazing, water development, and fire suppression (

e.g.,

Bahre 1991, entire). These impacts may have reduced the range or number of populations and individuals.

We define a population as occurring within the same water course (

i.e.,

stream) in a sky island range and within the distance pollinators can travel. A population may consist of one or more subpopulations of Bartram's stonecrop. These subpopulations are separated by up to 8 km (5 mi). Within each subpopulation are groupings of plants. Groupings are separated by up to 1.7 km (1 mi).

As of 2017, when the SSA analysis was completed, there were 29 extant populations across 12 mountain ranges in the United States and Mexico: 26 extant populations from 9 mountain ranges in southern Arizona and 3 presumed extant populations from 3 mountain ranges in northern Mexico (see Table 3, below). Within these 29 populations, there are approximately 3,756 individuals within about 2 ha (5 ac).

In 2018, four additional populations were located in the United States in the Rincon Mountains, one additional population was located in Mexico, and a known population in Mexico, which we did not have recent data for, was confirmed. The new populations in the United States included the Upper Rincon Creek population with 38 individuals (including “many” seedlings), Turkey Creek population with 4 individuals (seedlings not differentiated, but photos look like adult rosettes and flowering), Deer Creek population with 10 individuals (adult rosettes and flowering), and Chiminea Tributary population with 13 plants (seedlings not differentiated). In Sonora, Mexico, a new population (Mesa Tres Rios population) with 80 living and 28 dead plants was found in Mesa Tres Rios. In the Río Piedras Verdes near Colonia Pacheo area of Chihuahua, seven individuals were located, confirming the presence of an extant population “near Colonia Pacheco”; it is unknown if this is the exact historical location. Seedlings were not differentiated in either of the Mexico surveys. In total, only 145 new individuals were found, including seedlings, with 65 from the United States and 80 from Mexico. All but one population (Mesa Tres Rios) are small populations with fewer than 150 individuals. The number of extant populations as of 2018 is 34 across 13 mountain ranges in the United States and Mexico.

Table 3—Current Status of Bartram's Stonecrop Populations

Mountain ranges

Population

Population

status

Subpopulation

Subpopulation status

UNITED STATES

Baboquivari Mountains

Brown Canyon

Extant

Brown Canyon

Extant.

Thomas Canyon

Extant

Thomas Canyon

Extant.

Chiricahua Mountains

Echo Canyon

Extant

Echo Canyon

Rhyolite Canyon

Sugarloaf Mountain

Extant.

Extant.

Extant.

Indian Creek

Extirpated

Indian Creek Canyon

Extirpated.

Dragoon Mountains

Carlink Canyon

Extirpated

Carlink Canyon

Extirpated.

Jordan Canyon

Extant

Jordan Canyon

Extant.

Sheepshead

Extant

Sheepshead

Extant.

Slavin Gulch

Extant

Lower Slavin Gulch

Extant.

Stronghold Canyon East

Extant

Cochise Spring

Park Canyon

Extant.

Extant.

Stronghold Canyon West

Extant

Rockfellow Dome Trail

Stronghold Canyon West

Stronghold Canyon—hanging canyon drainage

Extant.

Extant.

Extant.

Empire Mountains

Empire Mountains

Extirpated

Empire Mountains

Extirpated.

Mule Mountains

Juniper Flat

Extant

Juniper Flat and vicinity

Extant.

Pajarito/Atascosa Mountains

Alamo Canyon

Extant

Alamo Canyon

Extant.

Holden Canyon

Extant

Holden Canyon

Extant.

Sycamore Canyon

Extant

Montana Peak Vicinity

Montana Canyon

Mule Ridge

Extant.

Extant.

Extant.

Penasco Canyon; below dam

Summit Motorway

Sycamore Canyon

Extant.

Extant.

Extant.

Warsaw Canyon

Extant

Warsaw/Old Glory Canyons

Extant.

Patagonia Mountains

Alum Gulch

Extant

Alum Gulch

Flux Canyon

Extant.

Extant.

Rincon Mountains

Chimenea-Madrona Canyons

Extant

Chimenea Canyon + Manning Camp Trail + Madrona Canyon

Extant.

Happy Valley North

Extirpated

Happy Valley North

Extirpated.

Happy Valley South

Extant

Happy Valley South

Extant.

Upper Rincon Creek

Extant

Upper Rincon Creek

Extant.

Turkey Creek

Extant

Turkey Creek

Extant.

Deer Creek

Extant

Deer Creek

Extant.

Chiminea Tributary

Extant

Chiminea Tributary

Extant.

Santa Rita Mountains

Adobe Canyon

Extant

Adobe Canyon

Extant.

Gardner Canyon

Extant

Cave Creek Canyon

Gardner Canyon

Sawmill Canyon

Extant.

Extant.

Extant.

Josephine Canyon

Extant

Bond Canyon

Josephine Canyon

Extant.

Extant.

Madera Canyon

Extant

Madera Canyon

Extant.

Squaw Gulch

Extant

Squaw Gulch

Extant.

Sycamore Canyon

Extant

Sycamore Canyon

Extant.

Temporal Gulch

Extant

Temporal Gulch

Upper Jones Canyon

Extant.

Extant.

Walker Canyon

Extant

Big Casa Blanca Canyon

Walker Canyon Basin

Extant.

Extant.

Whetstone Mountains

Death Trap Canyon

Extant

Death Trap Springs

Extant.

French Joe Canyon

Extant

French Joe Canyon

Extant.

MEXICO

Sierra Las Avispas, Sonora

Sierra Las Avispas, Sonora

Presumed Extant

Sierra Las Avispas, (Nogales County)

Presumed Extant.

Sierra La Escuadra, Chihuahua

Sierra La Escuadra, Chihuahua

Extant

Near Colonia Pacheco (in the Municipio Nuevo Casas Grandes)

Extant.

Sierra La Estancia, Chihuahua

Sierra La Estancia, Chihuahua

Presumed Extant

Cuarenta Casas (northwest of Las Varas, Municipio Madera)

Presumed Extant.

Sierra Los Mojones

Mesa Tres Rios

Extant

Mesa Tres Rios

Extant.

The number of populations within each sky island mountain ranges from one population (

e.g.,

Mule Mountains) to as many as eight populations (

e.g.,

Santa Rita Mountains). Each of these populations contains from one to eight subpopulations, which can be separated by up to 8 km (5 mi). Within each subpopulation, plants grow in groups or clusters of one to eight groups, which are separated by up to 1.7 km (1 mi). Within each subpopulation, plants grow across an area of 1 to 140 m (3.3 to 459 ft) (Ferguson 2014, entire; Ferguson 2016a, p. 14).

Bartram's stonecrop typically occurs in small populations with limited numbers of individuals. Most populations contain fewer than 100 plants (Ferguson 2014, entire; Ferguson 2016a, entire), but occasionally hundreds of plants can be found within a single population. The number of individuals in a given population can vary greatly from year to year and from season to season, depending on weather and stressors present (Ferguson 2017b, pp. 8, 15).

For Bartram's stonecrop to maintain viability, its populations or some representative portion thereof must be resilient. Resiliency describes the ability of populations to withstand stochastic events (arising from random factors). We can measure resiliency based on metrics of population health (for example, germination versus death rates and population size). Highly resilient populations are better able to withstand disturbances such as random fluctuations in germination rates (demographic stochasticity), variations in rainfall (environmental stochasticity), or the effects of anthropogenic activities. Resilient Bartram's stonecrop populations must be large enough that stochastic events do not eliminate the entire population. A highly resilient population of Bartram's stonecrop consists of multiple subpopulations, with a large number of individuals in each subpopulation. Highly resilient Bartram's stonecrop populations must also produce and disperse seeds, establish seedlings that survive, and maintain mature reproductive individuals in the population; recruitment should exceed or be equal to mortality. This allows for shared pollinators and seed dispersal between subpopulations and groups within the population, which can allow the population to recover from disturbance events and maintain or increase genetic diversity. Population resiliency categories for Bartram's stonecrop are described in section 3.2 of the SSA report (Service 2018b, entire).

In addition to the above demographic needs, populations also need habitat elements for resiliency. Based on where the species has typically been found, a resilient population needs riparian characteristics (

i.e.,

proximity to water and associated vegetation),

precipitation, shade, and bedrock or soil pockets in rock ledges and cliffs. Precipitation is needed to maintain soil moisture, cooler temperatures, and humidity in the microenvironment; shade from trees, canyon walls, and leaf litter aid in moisture retention. Small population size has the potential to decrease Bartram's stonecrop's population resiliency, as all stressors are exacerbated in populations with only a small number of individuals. Area of occupied habitat, abundance, number of subpopulations, and recruitment all affect population resiliency. Habitat resiliency categories for Bartram's stonecrop are described in Table 4, below, and in section 3.2 of the SSA report (Service 2018b).

Table 4—Population Resiliency Category Definitions for Bartram's Stonecrop

Condition

category

Population factors

Subpopulations

Abundance

Recruitment

Habitat factors

Riparian elements

Winter (October through March) precipitation

Shade

High (3)

Three or more subpopulations of plants/population

Number of adults in each population is >300 individuals

Populations contain more seedlings (<1.5 cm [0.6 in]) than dying individuals

Water is within 10 m from individuals or riparian vegetation present indicating subsurface water nearby

More than 12 inches of winter rain on average during the past 5 years as recorded at the nearest weather station

Overstory cover of

Juniperus, Quercus, Pinus

or other is >80%.

Moderate (2)

Two subpopulations of plants/population

Number of individuals in each population is 150 to 300 individuals

Populations contain an equal number of seedlings (<1.5 cm [0.6 in]) to dying individuals

Water at or near the surface (riparian vegetation present indicating subsurface water) is within 10-20 m from individuals

Between 6.1 and 12 inches of winter rain on average during the past 5 years as recorded at the nearest weather station

Overstory cover of

Juniperus, Quercus, Pinus

or other is between 50 and 80%.

Low (1)

One subpopulation of plants/population

Number of individuals in each population is <150 individuals

Populations contain fewer seedlings (<1.5 cm [0.6 in]) than dying individuals

Water at or near the surface (riparian vegetation present indicating subsurface water) is within 20-30 m from individuals

6 or fewer inches of winter rain on average during the past 5 years as recorded at the nearest weather station

Overstory cover of

Juniperus, Quercus, Pinus

or other is between 20 and 50%.

Ø

No subpopulations

No individuals are found during surveys in appropriate microhabitat

Population is made up primarily of dead and dying individuals that do not produce seed or no individuals found

Streambed near plants is dry and invaded by non-riparian plant species indicating shift of vegetation community and complete loss of suitable habitat

6 or fewer inches of winter rain on average during the past 5 years as recorded at the nearest weather station

Overstory cover has been removed.

Maintaining representation in the form of genetic or ecological diversity is important to maintain the capacity of Bartram's stonecrop to adapt to future environmental changes. Representation describes the ability of a species to adapt to changing environmental conditions. Representation can be measured by the breadth of genetic or ecological diversity within and among populations. The more representation, or diversity, a species has, the more it is capable of adapting to changes (natural or human-caused) in its environment. In the absence of species-specific genetic and ecological diversity information, we evaluate representation based on the extent and variability of habitat characteristics across the geographical range.

Genetic analysis of Bartram's stonecrop has not been conducted within or among populations or mountain ranges. However, populations on different mountain ranges are widely separated (ranging from roughly 14 to 42 km (8.7 to 26 mi) apart), making cross-pollination highly unlikely, and most of the populations contain small numbers of individuals. Therefore, there is the potential for genetic diversity among mountain ranges. Because multiple populations have been extirpated, it is possible that there has been a loss of genetic diversity. There may be genetic diversity between populations within and among the sky island mountain ranges due to response to elevational and other environmental differences between locations. As such, maintaining representation in the form of genetic diversity across multiple populations and sky island mountain ranges may be important to the capacity of Bartram's stonecrop to adapt to future environmental change.

The species is found over a relatively wide range of elevations of 1,067 to 2,042 m (3,500 to 6,700 ft) and vegetation communities (oak woodlands at higher elevations, and grasslands and oak savannas at lower elevations), which could be important in terms of representation. Such variability in elevation could aid in survival of future environmental changes, such as warming temperatures or decreased precipitation from climate change. At a minimum, we likely need to retain populations throughout the geographic and elevational ranges of the species to maintain the overall potential genetic and environmental diversity that can maximize the species' response to environmental changes over time.

Bartram's stonecrop needs to have multiple resilient populations distributed throughout its range to provide for redundancy such that a catastrophic event will not result in the loss of all populations. Redundancy describes the ability of a species to withstand catastrophic events, measured by the number of populations, and their resiliency, distribution, and connectivity. The more populations, and the wider the distribution of those populations, the more redundancy the species will exhibit. Redundancy reduces the risk that a large portion of the species' range will be negatively affected by a catastrophic natural or anthropogenic event at a given point in time. Species that are well-distributed across their historical range are considered less susceptible to extinction and more likely to be viable than species confined to a small portion of their range (Carroll

et al.

2010, entire). There is little connectivity potential between the sky island mountain ranges (separated from roughly 14 to 42 km (8.7 to 26 mi) apart); therefore, a localized stressor such as dewatering from a mine or a high-severity wildfire would impact only those populations near the activity. Regional drought and altered fire regime could impact many populations throughout the plant's range. There are 34 populations spread throughout the range of the species, many with multiple subpopulations. Conversely, such distance among populations reduces connectivity among populations and mountain ranges, which may be important for genetic exchange and recolonization. At a minimum, the species likely requires retaining population redundancy across multiple sky island mountain ranges throughout

the species' range to minimize impacts from catastrophic events.

Summary of Biological Status and Stressors

Section 4 of the Act (16 U.S.C. 1533) and its implementing regulations (50 CFR part 424) set forth the procedures for determining whether a species is an “endangered species” or a “threatened species.” The Act defines an endangered species as a species that is “in danger of extinction throughout all or a significant portion of its range,” and a threatened species as a species that is “likely to become an endangered species within the foreseeable future throughout all or a significant portion of its range.” The Act requires that we determine whether any species is an “endangered species” or a “threatened species” because of any of the following factors:

(A) The present or threatened destruction, modification, or curtailment of its habitat or range;

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

(C) Disease or predation;

(D) The inadequacy of existing regulatory mechanisms; or

(E) Other natural or manmade factors affecting its continued existence.

These factors represent broad categories of natural or human-caused actions or conditions that could have an effect on a species' continued existence. In evaluating these actions and conditions, we look for those that may have a negative effect on individuals of the species, as well as other actions or conditions that may ameliorate any negative effects or may have positive effects.

We use the term “threat” to refer in general to actions or conditions that are known to or are reasonably likely to negatively affect individuals of a species. The term “threat” includes actions or conditions that have a direct impact on individuals (direct impacts), as well as those that affect individuals through alteration of their habitat or required resources (stressors). The term “threat” may encompass—either together or separately—the source of the action or condition or the action or condition itself.

However, the mere identification of any threat(s) does not necessarily mean that the species meets the statutory definition of an “endangered species” or a “threatened species.” In determining whether a species meets either definition, we must evaluate all identified threats by considering the expected response by the species, and the effects of the threats—in light of those actions and conditions that will ameliorate the threats—on an individual, population, and species level. We evaluate each threat and its expected effects on the species, then analyze the cumulative effect of all of the threats on the species as a whole. We also consider the cumulative effect of the threats in light of those actions and conditions that will have positive effects on the species—such as any existing regulatory mechanisms or conservation efforts. The Secretary determines whether the species meets the definition of an “endangered species” or a “threatened species” only after conducting this cumulative analysis and describing the expected effect on the species now and in the foreseeable future.

The Act does not define the term “foreseeable future,” which appears in the statutory definition of “threatened species.” Our implementing regulations at 50 CFR 424.11(d) set forth a framework for evaluating the foreseeable future on a case-by-case basis. The term foreseeable future extends only so far into the future as the Services can reasonably determine that both the future threats and the species' responses to those threats are likely. In other words, the foreseeable future is the period of time in which we can make reliable predictions. “Reliable” does not mean “certain”; it means sufficient to provide a reasonable degree of confidence in the prediction. Thus, a prediction is reliable if it is reasonable to depend on it when making decisions.

It is not always possible or necessary to define foreseeable future as a particular number of years. Analysis of the foreseeable future uses the best scientific and commercial data available and should consider the timeframes applicable to the relevant threats and to the species' likely responses to those threats in view of its life-history characteristics. Data that are typically relevant to assessing the species' biological response include species-specific factors such as lifespan, reproductive rates or productivity, certain behaviors, and other demographic factors.

We completed a comprehensive assessment of the biological status of beardless chinchweed and Bartram's stonecrop, and prepared an SSA report for each species (Service 2018a and 2018b, entire), which provides a thorough account of the species' overall viability. We define viability here as the ability of the species to persist over the long term and, conversely, to avoid extinction. In the following discussion, we summarize the conclusions of the SSA reports, which can be accessed at Docket FWS-R2-ES-2018-0104 on

http://www.regulations.gov

and at

https://www.fws.gov/southwest/es/arizona/Docs_Species.htm.

Beardless Chinchweed

Several stressors influence whether beardless chinchweed populations will grow to maximize habitat occupancy, which increases the resiliency of a population to stochastic events. We evaluated the past, current, and future stressors (

i.e.,

negative changes in the resources needed by beardless chinchweed) that are affecting what beardless chinchweed needs for viability. These stressors are described in detail in chapter 4 of the SSA report (Service 2018a). Stressors that have the potential to affect beardless chinchweed population resiliency include:

• Loss of habitat due to invasion by nonnative species;

• Altered fire regime exacerbated by invasion by nonnative species;

• Altered precipitation, drought, and temperature;

• Erosion, sedimentation, and burial from road and trail maintenance, mining, livestock, wildlife, and post-wildfire runoff;

• Grazing from wildlife and livestock; and

• Small population size exacerbating all other stressors.

The stressors that pose the largest risk to future viability of the species are: (1) Loss of habitat caused by the invasion of nonnative grasses that compete for space, water, light, and nutrients and that alter wildfire regimes; and (2) small population size (fewer than 50 individuals), which potentially causes other stressors to seriously damage or extirpate populations. The size of fewer than 50 individuals as a small population was determined by assessing the range of known population sizes. Much of the historical range of beardless chinchweed in both the United States and Mexico has been altered by an invasion of nonnative grasses and herbaceous plants. Although there are many nonnative plant species growing in historical beardless chinchweed habitats in both the United States and Mexico, two species in particular are most problematic to beardless chinchweed at this time: Lehman's lovegrass (

Eragrostis lehmanniana

) and rose natal (

Melinis repens

). Both of these species are strong competitors on southern exposures where beardless chinchweed occurs.

Habitat Loss Caused by Nonnative Grasses

Lehman's lovegrass, a nonnative grass from South Africa, has numerous

competative advantages over native grasses in southern Arizona. Lehman's lovegrass resprouts from roots and tiller nodes not killed by hot fire, is not hampered by the reduction in mycorrhizae associated with fire and erosion, is able to respond to winter precipitation when natives grasses are dormant, is able to produce copious seed earlier than native grasses, maintains larger seed banks than native grasses, and has higher seedling survival and establishment than native grasses during periods of drought (Anable 1990, p. 49; Anable

et al.

1992, p. 182; Robinett 1992, p. 101; Fernandez and Reynolds 2000, pp. 94-95; Crimmins and Comrie 2004, p. 464; Geiger and McPherson 2005, p. 896; Schussman

et al.

2006, p. 589; O'Dea 2007, p. 149; Archer and Predick 2008, p. 26; Mathias

et al.

2013, entire). This species outcompetes native grasses for water, light, and nutrients, forming nonnative-dominated grasslands that reduce structural, species, and spatial diversity and that produce two to four times the biomass of native grasslands (D'Antonio and Vitousek 1992, p. 70; McPherson 1995, pp. 136-137; VanDevender

et al.

1997, p. 4; Huang

et al.

2009, pp. 903-904;). This change in vegetation structure results in a higher fuel load that is highly lignified (long-lasting through slow decomposition) and results in more frequent fires that have longer flames, faster rates of spread, and higher severity and frequency than historical low-intensity burns of native desert grasslands (Anable

et al.

1992, p. 186; Dennet

et al.

2000, pp. 22-23; Williams and Baruch 2000, p. 128; Crimmins and Comrie 2004, p. 464). In addition, Lehman's lovegrass-dominated grasslands recover quickly from fire, as fires scarify the ample seeds and remove canopy, allowing for high seedling emergence (Cable 1965, p. 328; Anable 1990, p. 15; Roundy

et al.

1992, p. 81; McPherson 1995, p. 137; Biedenbender and Roundy 1996, p. 160).

Rose natal, a native of Africa and Madagascar, is invasive in many locations, including southern Arizona and northern New Mexico (Stevens and Fehmi 2009, p. 379; Romo

et al.

2012, p. 34). Similar to Lehman's lovegrass, rose natal is capable of growing in low moisture situations and has many advantages to outcompete native grasses of southern Arizona, such as prolific seed production and culms that root from the nodes (Stokes

et al.

2011, p. 527). This aggressive grass displaces native vegetation in shrublands and oak stands, and increases fire frequency (Romo

et al.

2012, p. 35; Center for Agriculture and Biosciences International 2017, entire).

In addition, several other African grasses (

e.g., Eragrostis cilianensis

[stinkgrass],

Eragrostis curvula

[Boer lovegrass],

Eragrostis echinochloidea

[African lovegrass], and

Dichanthium annulatum

[Kleberg's bluestem]) have been documented in southern Arizona and northern Mexico (Van Devender and Reina 2005, p. 160; NatureServe, entire; Fire Effects Information System, entire; SEINet, entire), as has the Asian grass,

Bothriochloa ischaemum

(yellow bluestem). Studies of other nonnative grasses in Mexico show rapid expansion and degradation of native communities, with the potential to invade large areas of northern Mexico (Arriaga

et al.

2004, p. 1504). There are no beardless chinchweed populations in the United States that are more than 1 km (0.6 mi), and no beardless chinchweed populations in Mexico that are more than 27 km (16.8 mi), from documented nonnative grasses (SEINet, entire; Heitholt 2017, pers. comm.). Because we have seen nonnative infestations in the field in locations not shown in SEINet, we believe only a small portion of nonnative plants are reported into the SEINet system in either country. Based on the above information, we believe that it is unlikely any beardless chinchweed population is free of nonnative plants. This encroachment of nonnatives has reduced beardless chinchweed population numbers and habitat, and as nonnatives continue to encroach on beardless chinchweed populations, the number of individuals and available habitat will continue to decrease.

Altered Fire Regime

The desert grasslands, oak savannas, and oak woodlands of southern Arizona historically had large-scale, low-severity fire roughly every 10 to 20 years and following periods of adequate moisture (McPherson and Weltzin 2000, p. 5; Brooks and Pyke 2002, p. 6; McDonald and McPherson 2011a, p. 385; Fryer and Leunsmann 2012, entire). Fires now are more frequent and intense due to the unnaturally dense and evenly spaced canopies of nonnative-dominated communities (as compared to more open and heterogeneous native-dominated grasslands), coupled with more frequent fire starts from recreationist and cross-border violators (Anable

et al.

1992, p. 186; D'Antonio and Vitousek1992, p. 75; Dennet

et al.

2000, pp. 22-23; Williams and Baruch 2000, p. 128; Crimmins and Comrie 2004, p. 464; Emerson 2010, pp. 15, 17; United States Government Accountability Office 2011, p. 1; Wildland Fire Lesson's Learned Center 2011, entire). Nonnative grasses have higher seed output and large seed banks, earlier green-up in the spring, and greater biomass production than native grasses; all of these characteristics help to perpetuate a grass-fire cycle (

e.g.,

D'Antonio and Vitousek 1992, p. 73; Zouhar

et al.

2008, pp. 17, 21; Steidl

et al.

2013, p. 529).

In many locations in southern Arizona in recent decades, repeat fires have occurred within short periods of time, aided by the dominance of nonnative grasses in the landscape. For example, in the Pajarito and Atascosa Mountains area, multiple fires burned the landscape between 2008 and 2016 (Figure 4.4 in Service 2018a). This landscape is now dominated by both nonnative Lehman's lovegrass and rose natal (Service 2014c, entire; Heitholt 2017, entire), and many historically documented locations that supported beardless chinchweed have not been found again (Service 2014c, entire; Fernandez 2017, pers. comm.; Haskins and Murray 2017, p. 4). High-severity wildfires burn hotter than fires that beardless chinchweed evolved with; consequently, we believe the plant is not capable of surviving high-severity fires.

Altered Precipitation, Drought, and Temperature

Altered precipitation timing and form (snow versus rain), as well as reduced winter and spring precipitation and prolonged drought, are currently occurring and projected to increase or be altered from normal in the Southwest (Garfin

et al.

2014, entire). Recently there has been a decrease in the amount of snowpack, earlier snowmelt, and increased drought severity in the Southwest (Garfin

et al.

2013, entire; Garfin 2013b p. 465). Further, more wintertime precipitation is falling as rain rather than snow in the western United States (IPCC 2013, p. 204; Garfin 2013b p. 465). This means that the amount of runoff in the spring when snow melts is reduced, as is soil moisture. Precipitation is bimodal with the mountain ranges where beardless chinchweed occurs, with dormant season snow and rain, and growing season monsoon rains (CLIMAS 2014, entire). We believe that precipitation during October through March is important for beardless chinchweed germination and growth. In addition, beardless chinchweed does not flower until it reaches a height of more than 0.5 m (1.6 ft) tall; without sufficient precipitation, beardless chinchweed may be unable to attain adequate size for reproduction (Phillips

et al.

1982, p.

8). Further, reduced precipitation, change in the timing and type of precipitation, and prolonged drought impact soil and ambient moisture availability for beardless chinchweed germination, growth, and flowering. In addition, due to increased nonnative competition during times of reduced precipitation and drought, impacts from these stressors to beardless chinchweed would be exacerbated (Anable 1990, p. 49; Robinett 1992, p. 101; Fernandez and Reynolds 2000, pp. 94-95; Geiger and McPherson 2005, p. 896; Schussman

et al.

2006, p. 589; Archer and Predick 2008, p. 26; Mathias

et al.

2013, entire).

Under a continuation of A2-high emissions scenario, reduced winter and spring precipitation is consistently projected for the southern part of the Southwest by 2100, as part of the general global precipitation reduction in subtropical areas (Garfin

et al.

2014, p. 465). Analyses of the southwestern United States indicate future drying, primarily due to a decrease in winter precipitation under both the RCP 4.5 and 8.5 scenarios (IPCC 2013, p. 1080). The annual projected changes in precipitation for 2025 to 2049 under the RCP 4.5 and 8.5 scenarios range from an increase of 1.3 cm/month (0.5 in/month) to a decrease of 1.5 cm/month (0.5 in/month), with a an annual average of no change compared to 1981 to 2010 (USGS 2019, entire). However, winter and spring precipitation under both emission scenarios is projected to decrease from −0.3 to −1 cm (−0.1 to −0.4 inches) (MACA 2019) or a decrease up to 10 percent for 2016-2035 relative to 1986-2005 under RCP 4.5 (IPCC 2013, p. 985). The decrease in winter and spring precipitation would likely be greater under the RCP 8.5 scenario. There is some evidence from comparing observations with simulations of the recent past that climate models might be underestimating the magnitude of changes in precipitation in many regions (IPCC 2013, p. 986). The climate-model-projected simulations indicate that a high degree of variability of annual precipitation will continue during the coming century, for both low and high emission scenarios (Garfin 2013, p. 110). This suggests that the Southwest will remain susceptible to unusually wet spells and, on the other hand, will remain prone to occasional drought episodes (Garfin 2013, p. 110). However, decrease in soil moisture across much of the Southwest is projected under both scenarios by mid-century, due to increased evaporation (IPCC 2013, p. 1259). Late winter-spring mountain snowpack in the Southwest is predicted to continue to decline over the 21st century under the high emission scenario (A2), mostly because of projected increased temperature (Garfin

et al.

2013, p. 6). Reduced rain and snow, earlier snowmelt, and drying tendencies cause a reduction in late-spring and summer runoff. Together these effects, along with increases in evaporation, result in lower soil moisture by early summer (Gafrin 2013, p. 117).

Climatic events such as snowpack, earlier snowmelt, and increased drought are regional and will impact all populations of beardless chinchweed. Precipitation timing and amount impacts the germination, growth, and flowering of beardless chinchweed, resulting in the loss of individuals and recruitment, and overall reducing the population size.

In the Southwest, temperatures increased 2.7 degrees Celcius (°C) (1.6 degrees Fahrenheit (°F)) plus or minus 0.9 °C (0.5 °F), between 1901 and 2010, and more heat waves occurred over the Southwest during 2001-2010 compared to average occurrences in the 20th century. In the future, under RCP 4.5, the annual maximum temperature is projected to increase by 5 °C (2.7 °F) for 2025-2049 and 7.3 °C (4 °F) for 2050-2074, and 5 °C (2.7 °F) for 2025-2049 and 10.4 °C (5.7 °F) for 2050-2074 under RCP 8.5, all relative to 1981-2010 (USGS 2019, entire). When temperatures rise, as has been occurring in recent decades and as is projected to continue into the future, evapotranspiration rates also increase and soil moisture decreases. Along with projected warming and increased evapotranspiration, it is highly likely that droughts will become more severe (Garfin 2013, pp. 137-138). A decrease of up to 4 percent soil moisture is projected under RCP 4.5 scenario for 2016-2035, relative to 1986-2005. The decrease in soil moisture would likely be greater under the RCP 8.5 scenario. Further, the evaporation deficient increases under RCP 4.5 and increases more in RCP 8.5 in 2025 to 2049, relative to 1981 to 2010. Based on the high emissions scenario, the current 100-year drought will become commonplace in the second half of this century and future droughts will be much more severe than those previously recorded (Garfin 2013, p. 138). This projection of intensified drought conditions on the Colorado River is not due to changes in precipitation, but rather due directly to warming and its effect on reducing soil moisture (Garfin 2013, p. 138). Physiological effects of CO

2

may involve both the stomatal response, which acts to restrict transpiration, and an increase in plant growth and leaf area, which acts to increase evapotranspiration (IPCC 2013, p. 986). An increase in evapotranspiration results in water loss from the plant and increases stress on the plant. This increase in stress impacts photosynthesis, respiration, transpiration, water use efficiency, leaf conductance, growth rate, vigor, and gas exchange. These impacts result in reduced growth, flowering, and seed production and, therefore, in reduced overall recruitment and population numbers.

Although rare species in the southwestern United States evolved with drought, recent changes in temperature, and rainfall patterns present stressful conditions of increased magnitude greater than what the species faced historically and raise the question of whether the species, can persist. Some species may shift their distributions in response to warming of the climate (McLaughlin

et al.

2002, p. 6070). However, it is highly unlikely that beardless chinchweed would be able to naturally shift its range to keep up with current and high projected rates of climate change, due to its overall population decline and inability to maintain current populations. Since plants are not mobile, expanding the distribution of this species is dependent on seed dispersal. Further, extant populations are small, which limit the amount of seed production for dispersal. It is highly unlikely that under elevated environmental stress associated with climate change, the species would be able to both maintain populations and also colonize new areas with more suitable climate conditions. Thus localized extirpations over portions of the beardless chinchweed range could result (lower elevations), and, in other portions of its distribution, the occupied range (higher elevation) may expand, depending upon habitat availability.

Erosion, Sedimentation, and Burial

General road maintenance and widening could disturb populations along road cuts and create erosion (Phillips

et al.

1982, p. 8). Of the six extant U.S. populations, the Ruby Road and Scotia Canyon populations, and the State of Texas Mine subpopulation of Coronado National Memorial occur along roadcuts; similarly, the Visitor Center subpopulation of the Coronado National Memorial population contains some plants that occur along a maintained trail. These plants could be damaged or removed by road or trail maintenance. Impacts from such stressors could be profound for

populations with fewer than 50 individuals. In addition, nonnative plant introduction and spread often occur in areas of disturbance, such as along roadways, along trails, in mining sites, and in areas of recreational use (Gelbard and Belnap 2003, p. 421; Brooks 2007, pp. 153-154; Anderson

et al.

2015, p. 1).

The McCleary Canyon—Gunsight Pass population is in the path of a proposed alignment of a secondary access road for the proposed Rosemont Mine (Westland 2010, p. iv), and the McCleary Canyon—Wasp Canyon population is within the processing facility portion of the proposed Rosemont Mine (Westland 2017, entire). Collectively, these plants represent approximately 33 percent of the total beardless chinchweed populations known across the U.S. range and 16 percent of all known individuals. The proposed road alignment would eliminate these populations.

Dust from mining operations or recreational travel can impact beardless chinchweed populations along dirt roadways. Dust may negatively affect plant growth and vigor as a result of changes in physiological and biochemical processes (

e.g.,

photosynthesis, respiration, transpiration, water use efficiency, leaf conductance, growth rate, vigor, and gas exchange) and reduced pollination (Phillips

et al.

1982, pp. 9-10; Chibuike and Obiora 2014, p. 1; Waser

et al.

2017, p. 90). These impacts could affect those populations within 30 meters (98 feet) of roads and mine sites (Waser

et al.

2017, p. 90). This stressor could impact four of the six populations in the United States.

Grazing

There are two different perspectives on the influence of grazing on beardless chinchweed:

(1) Wildfire historically maintained native open habitat where beardless chinchweed occurred, but with fire suppression, overgrazing may have alternatively provided native open habitats for this species to expand its range in the early 1900s, even without frequent fire (Schmalzel 2015, p. 2), due to open space being created and maintained by cattle; and

(2) Grazing pressure may have contributed to the species' rareness (Keil 1982, entire) due to reduced reproduction and alteration in habitat.

Regardless, grazing that occurs in small populations (fewer than 50 individuals) of beardless chinchweed would have a negative population-level impact through the reduction of flowers and seeds, and possibly individuals. Beardless chinchweed does not flower until it reaches a height of more than 0.5 m (1.6 ft) tall, suggesting that grazing in summer or fall when the plant is growing and flowering could reduce seed production and recruitment.

Small Populations

Small population size has the potential to affect beardless chinchweed' population resiliency, as all stressors are exacerbated in populations with only a small number of individuals (fewer than 50). Known population sizes of beardless chinchweed were used to quantify the size of a small population. Small populations are less able to recover from losses caused by random environmental changes (Shaffer and Stein 2000, pp. 308-310), such as fluctuations in reproduction (demographic stochasticity), variations in rainfall (environmental stochasticity), or changes in the frequency or severity of disturbances, such as wildfires. Five of the six extant beardless chinchweed populations in the United States contain fewer than 50 individuals. Based on populations in the United States, which are mostly small and occur in habitat dominated by nonnatives, we believe that the six populations in Mexico are of similar size but may be in worse condition, because of limited native habitat management, similar climate change impacts, equally frequent wildfires, and likely more impacts from grazing. Loss due to mining, erosion, road and trail maintenance, trampling, grazing, or other stressors mentioned above are exacerbated in small populations, and have the potential to seriously damage or completely remove these small populations. Synergistic interactions among wildfire, nonnative grasses, decreased precipitation, and increased temperatures cumulatively and cyclically impact beardless chinchweed, and all stressors are exacerbated in small populations.

Current Condition of Beardless Chinchweed

Since 1962, we are aware of nine populations and one subpopulation of beardless chinchweed in the United States that have become extirpated. Currently, six extant beardless chinchweed populations are spread across four mountain ranges in southern Arizona: The Atascosa-Pajarito, Huachuca, Santa Rita, and the Canelo Hills. These six populations consist of 387 individuals spread across less than 2 ha (5 ac). Additionally, six populations have been reported from northern Mexico, but this information is from 1940 or earlier.

Population Resiliency of Beardless Chinchweed

To help determine current condition, we assessed each population in terms of its resiliency. Our analysis of the past, current, and future stressors on the resources that beardless chinchweed needs for long-term viability revealed that there are a number of stressors impacting this species. All beardless chinchweed populations likely contain nonnative grasses. Further, altered fire regime has the potential to affect all populations. This altered fire regime enhances the spread of nonnatives, and all populations of beardless chinchweed contain nonnatives. Consequently, fire will aid in the spread of nonnatives, and is currently a risk to all populations of beardless chinchweed and will be further exacerbated by nonnative grasses in the near future (approximately 10 years). Altered precipitation, increased temperatures, increased evapotranspiration, decreased soil moisture, and decreased winter and spring precipitation are current and ongoing regional actions that are impacting all populations of beardless chinchweed. These environmental conditions exacerbate an altered fire regime, which in turn further drives the spread of nonnatives. In addition, nonnative grasses have competitive advantage over native grasses during periods of drought.

Road maintenance is likely resulting in the direct killing of individuals in three populations (Ruby Road, Scotia Canyon, and Coronado National Memorial). In addition, all individuals in these three populations are currently being impacted by dust from the road. These three populations are already of low resiliency. Two additional populations (McCleary Canyon—Gunsight Pass and McCleary Canyon—Wasp Canyon) will be impacted by Rosemont mining operations and dust in the near future (approximately 10 years; Westland 2010, p. iv). One of these populations is already of low resiliency, and the other is of moderate resiliency. Eleven of the 12 populations (92 percent) are small population (fewer than 50 individuals). Synergistic interactions among wildfire, nonnative grasses, decreased precipitation, and increased temperatures cumulatively and cyclically impact beardless chinchweed, and all stressors are exacerbated in small populations. Of the six extant populations, two are moderately resilient and four are in low resiliency (Table 5, below). Population resiliency categories are described in Table 2, above, and in the SSA report (Service 2018a).

Table 5—Beardless Chinchweed Current Population Condition

Mountain range/country

Population

Number of

individuals

Current

condition

Atascosa-Pajarito Mountains, USA

Pena Blanca Lake

0

Extirpated.

Ruby Road

10

Low.

Summit Motorway

0

Extirpated.

Canelo Hills, USA

Audubon Research Ranch

37

Low.

Copper Mountain

0

Extirpated.

Harshaw Creek

0

Extirpated.

Lampshire Well

0

Extirpated.

Huachuca Mountains, USA

Scotia Canyon

40

Low.

Coronado National Memorial

241

Low.

Joe's Canyon Trail

0

Extirpated.

Patagonia Mountains, USA

Flux Canyon

0

Extirpated.

Washington Camp

0

Extirpated.

Santa Rita Mountains, USA

Box Canyon

0

Extirpated.

McCleary Canyon—Gunsight Pass

32

Moderate.

McCleary Canyon—Wasp Canyon

32

Low.

Chihuahua, Mexico

Batopililas, Rio Mayo

~10

Low.

Guasaremos, Rio Mayo

~10

Low.

Sonora, Mexico

Canon de la Petaquilla

~10

Low.

North of Horconcitos

~10

Low.

Canyon Estrella, Sierra de los Cendros; southeast of Tesopaco

~10

Low.

Los Conejos, Rio Mayo

~10

Low.

Beardless Chinchweed Representation

No genetic studies have been conducted within or between the 21 historical populations of beardless chinchweed in southern Arizona and Mexico. Mountain ranges that have only one or two populations, or have only have one subpopulation per population, or low numbers of individuals per population with several miles between mountain ranges, may not be as genetically diverse because pollination or transport of seeds between populations may be very limited or nonexistent. Five of the six extant U.S. populations do not have multiple subpopulations. The Coronado National Memorial population has two subpopulations. The six extant U.S. populations are separated geographically into the Atascosa-Pajarito, Huachuca, and Santa Rita Mountains, and the Canelo Hills, which are separated by 16 to 61 km (9.9 to 37.9 mi). There is likely genetic diversity among mountain ranges, but reduced genetic diversity within populations. Further, overall genetic diversity is likely reduced given that some populations are extirpated.

The 15 historical beardless chinchweed populations in the United States range in elevation from 1,158 m (3,799 ft) to 1,737 m (5,699 ft). Of these, eight (about 53 percent) fall below 457 m (1,500 ft) elevation. Of these eight, six have become extirpated in recent decades. This essentially indicates a loss at this lower elevational range and possibly loss of some local adaptation to warmer or dryer environments and genetic differentiation among populations.

In the Ruby Road, Scotia Canyon, and Coronado National Memorial populations, plants have been reported over many decades, indicating that these populations may have the genetic and environmental diversity needed to adapt to changing conditions. Note, however, that both the Ruby Road and Scotia Canyon populations have been reduced in size in the past 30 years, and we have no previous count data at Coronado National Memorial for comparison.

Beardless Chinchweed Redundancy

The beardless chinchweed populations in the United States and Mexico are naturally fragmented between mountain ranges. Currently, six extant beardless chinchweed U.S. populations are spread across Atascosa-Pajarito, Huachuca, and Santa Rita Mountains and the Canelo Hills. The Atascosa-Pajarito Mountains and the Canelo Hills have only one extant population each, while the Santa Rita and Huachuca Mountains have two extant populations each. These mountain ranges are separated from each other by 16 to 61 km (9.9 to 37.9 mi), so natural gene exchange or re-establishment following extirpation is very unlikely. In addition, six historical populations of beardless chinchweed are distributed across two general areas in northern Chihuahua and Sonora, Mexico. Their status is unknown, but we believe they are small populations with poor habitat based on populations in the United States, which are small and dominated by nonnative species. Although this may imply some level of redundancy across the range of beardless chinchweed, note that five of the six extant populations in the United States contain fewer than 50 individual plants. Further, nine populations and one subpopulation have been extirpated in recent decades, largely from the lower elevations of the species' range, and several populations have been reduced in size in recent decades.

Future Condition of Beardless Chinchweed

We also assessed the future condition of beardless chinchweed under several plausible scenarios in our SSA report (Service 2018a, entire). We present a summary of the relevant information here; the detailed future condition analysis is available in the SSA report.

We developed four scenarios incorporating the stressors that are ongoing or will occur in the future to consider the range of possible future conditions. For each scenario, we describe the level of impact from the identified stressors that would occur in each population. All of the scenarios involve some degree of uncertainty; however, they present a range of realistic and plausible future conditions (Table 6). All scenarios consider impacts from nonnative invasion, altered wildfire regime, and drought because there is no likely future scenario where these stressors would not affect the species. In addition, effects on individual plants (small population size) from multiple stressors are assessed, including cross-border

violator traffic, mining, trampling, erosion, road and trail maintenance, and grazing. We projected the likelihood of each scenario occurring at 40-years. We chose 40 years because this is within the range of available hydrological and climate change model forecasts, is within the time period of the Rosemont Mine effects, and it represents four generations of the plant.

Below is a summary of the four scenarios. For more detail, see Chapter 6 of the SSA (Service 2018a, entire).

Table 6—Future Scenarios for Beardless Chinchweed

Risks

Mining activity

Altered fire

regime *

Climate

Individual

effects

Conservation

Risk described

• Burial

• Removal

• Dust

• Lightning

• Nonnative plants

• Cross border violators

• Recreation

• Reduction in available water **

• Seedling desiccation

• Flowering halt

• Grazing

• Trampling

• Trail and road maintenance

• Erosion

Conservation actions implemented.

Scenario 1 Continuation continuing into the future

Rosemont mine implemented with indirect and direct impacts

Number of wildfires annually increases at the same rate as the last 10 years

Available water and drought continue at the same level as in the past 10 years, emissions 4.5

Applied to populations <50 individuals

No new individuals, subpopulations or populations found. No augmentation of existing populations, little seed preservation, nonnatives not controlled, some woodland areas thinned.

Scenario 2 Conservation

Rosemont mine implemented with indirect and direct impacts; with mitigation

Number of wildfires does not increase from current rate

Available water remains stable, emissions 4.5

Applied to populations <50 individuals

Sites revisited and additional plants are located, sites are augmented, or new sites are established, some nonnatives are controlled, and additional woodland areas are thinned.

Scenario 3 Moderate increase in negative effects

Rosemont mine implemented with direct impacts and additional mines implemented with indirect impacts

Number of wildfires increases

Available water is reduced per 4.5 emissions scenario

Applied to populations <50 individuals

No new individuals, subpopulations or populations found. No augmentation of existing populations, little seed preservation, nonnatives not controlled, some woodland areas thinned.

Scenario 4 Major increase in negative effects

Rosemont mine implemented and additional mines implemented with direct impacts

Number of wildfires increases

Available water is reduced per 8.5 emissions scenario

Applied to populations <50 individuals

No new individuals, subpopulations or populations found. No augmentation of existing populations, little seed preservation, nonnatives not controlled, some woodland areas thinned.

The “continuation” scenario evaluates the condition of beardless chinchweed if there is no increase in risk of stressors to the populations relative to what exists today. The other scenarios evaluate the response of the species to changes in those risks. The “conservation” scenario takes into account realistically possible additional protective measures, which may or may not happen. The “moderate effects” scenario is an increase in the risk of stressors to populations. The “major effects” scenario is a further increase in risk of stressors to populations.

We examined the resiliency, representation, and redundancy of beardless chinchweed under each of these plausible scenarios (see table 6.7 in the SSA report). The overall resiliency categories are the same as those used for current condition. We expect the six extant beardless chinchweed populations to experience changes to aspects of their habitat in different ways under the different scenarios. We projected the expected future resiliency, representation, and redundancy of beardless chinchweed based on the risk of stressors that would occur under each scenario (see Table 7). Under the “continuation” scenario, we would expect the viability of beardless chinchweed to be characterized by a loss of resiliency, representation, and redundancy at the level that is currently occurring. Under the “conservation” scenario, we would expect the viability of beardless chinchweed to be characterized by higher levels of resiliency, representation, and redundancy than it exhibits under the current condition. Under the “moderate effects” scenario, we would expect the viability of beardless chinchweed to be characterized by lower levels of resiliency, representation, and redundancy than it has in the “continuation” scenario. Under the “major effects” scenario, we would expect all populations of beardless chinchweed to be extirpated at the 40-year time step.

Table 7—Beardless Chinchweed Population Conditions Under the Current Condition and All Future Scenarios

Mountain range

Population name

Current condition

Continuation

scenario

Conservation

scenario

Moderate effects scenario

Major effects

scenario

Atascosa-Pajarito

Pena Blanca Lake

Extirpated

Extirpated

Extirpated

Extirpated

Extirpated.

Ruby Road

Low

Extirpated

Extirpated

Extirpated

Extirpated.

Summit Motorway

Extirpated

Extirpated

Extirpated

Extirpated

Extirpated.

Canelo Hills

Audubon Research Ranch

Low

Low

Low

Extirpated

Extirpated.

Copper Mountain

Extirpated

Extirpated

Extirpated

Extirpated

Extirpated.

Harshaw Creek

Extirpated

Extirpated

Extirpated

Extirpated

Extirpated.

Lampshire Well

Extirpated

Extirpated

Extirpated

Extirpated

Extirpated.

Huachuca

Scotia Canyon

Low

Low

Low

Extirpated

Extirpated.

Coronado National Memorial

Low

Low

Low

Low

Extirpated.

Joe's Canyon Trail

Extirpated

Extirpated

Extirpated

Extirpated

Extirpated.

Patagonia

Flux Canyon

Extirpated

Extirpated

Extirpated

Extirpated

Extirpated.

Washington Camp

Extirpated

Extirpated

Extirpated

Extirpated

Extirpated.

Santa Rita

Box Canyon Road

Extirpated

Extirpated

Extirpated

Extirpated

Extirpated.

McCleary Canyon—Gunsight Pass

Moderate

Low

Low

Extirpated

Extirpated.

McCleary Canyon—Wasp Canyon

Low

Extirpated

Extirpated

Extirpated

Extirpated.

Chihuahua, MX

Batopililas

Low

Extirpated

Extirpated

Extirpated

Extirpated.

Guasaremos

Low

Extirpated

Extirpated

Extirpated

Extirpated.

Sonora, MX

Canon de la Petaquilla

Low

Extirpated

Extirpated

Extirpated

Extirpated.

Canyon Estrella

Low

Extirpated

Extirpated

Extirpated

Extirpated.

Horconcitos

Low

Extirpated

Extirpated

Extirpated

Extirpated.

Los Conejos

Low

Extirpated

Extirpated

Extirpated

Extirpated.

Bartram's Stonecrop

Several factors influence whether Bartram's stonecrop populations will grow to increase habitat occupancy, which increases the resiliency of a population to stochastic events. We evaluated the past, current, and future stressors that are affecting what Bartram's stonecrop needs for viability. These stressors are described in detail in the chapter 4 of the SSA report (Service 2018b, entire). Stressors that have the potential to affect Bartram's stonecrop population resiliency include:

• Loss of water in nearby drainages from mining and drought;

• Erosion, sedimentation, and burial from mining, livestock, wildlife, recreation trails and roads, cross-border violators, and post-wildfire runoff;

• Trampling from humans, wildlife, and livestock, and predation;

• Altered fire regime resulting from fires ignited by recreationists, cross-border violators, and lightning;

• Illegal collection;

• Altered precipitation, drought, flooding, and freezing regime from current and future climate change, resulting in loss of seedling, immature, and adult plants, and in loss of reproduction; and

• Small population size exacerbating all other stressors.

The stressors that pose the largest risk to future viability of the species, which are related to habitat changes, include:

(1) Groundwater extraction and prolonged drought that may reduce nearby water levels and humidity within Bartram's stonecrop habitat; and

(2) Altered fire regimes leading to erosion of Bartram's stonecrop habitat, sedimentation that could cover individuals, and loss of overstory shade trees. These stressors play a large role in the future viability of Bartram's stonecrop, especially for smaller populations. These stressors may reduce nearby water levels, shade, and humidity within Bartram's stonecrop habitat and may directly impact individuals.

Loss of Water

Dewatering of streams from mining operations may lead to overstory canopy losses and resulting loss of shade, as well as reduction in spring and stream flow and humidity in nearby Bartram's stonecrop populations. The Rosemont Mine Final Environmental Impact Statement states that no Bartram's stonecrop were found in the project area or the footprint of the connected actions; however, individuals growing in the analysis area could experience indirect impacts from groundwater drawdown (USFS 2013a, p. 676). According to the Rosemont Mine Final Environmental Impact Statement (USFS 2013a, p. 339), the proposed mine pit would create a permanent drawdown of the water table, and groundwater would flow toward the pit and be lost to evaporation. The water would be perpetually replenished in part by groundwater from the regional aquifer, and the pit would act as a hydraulic sink. Given that Bartram's stonecrop is consistently found in locations with nearby springs or other water sources, the loss of groundwater at the nearby unmapped spring in Box Canyon/Sycamore Canyon confluence, between Ruelas Spring and the Singing Valley Road residences, could significantly impact these Bartram's stonecrop plants. In the range of Bartram's stonecrop, there are many mining claims, trenching and exploration drilling activities, and a few active and proposed mines. Many currently undeveloped areas of locatable mineral deposits may be explored and/or mined in the future. We do not know the extent of future mine activity within the range of Bartram's stonecrop; however, a number of proposed mines are identified for development within Bartram's stonecrop habitat. The range of current and projected mining activities varies from 1 to 10 per sky island mountain range containing Bartram's stonecrop (USFS 2012, entire). The loss of water in any Bartram's stonecrop population could lead to extirpation of that population.

Erosion, Sedimentation, and Burial

Bartram's stonecrop typically occurs on steep slopes with erodible soils and areas susceptible to rock fall, making the plant particularly vulnerable to physical damage to its environment (Phillips

et al.

1982, p. 10; Shohet 1999, p. 50; Ferguson 2014, p. 42; Ferguson 2016a, pp. 15, 26). Soil erosion can result in burying plants, eroding the soil the plant is growing in, or dislodging plants. While displaced plants may re-root (Shohet 1999, pp. 50-51, 60), it is more likely that these plants will not survive (Ferguson 2015, p. 2). The potential of soil disturbance and erosion within or above Bartram's stonecrop habitat or the trampling of individual Bartram's stonecrop plants may occur from a variety of activities, including livestock and wildlife movement; the placement and maintenance of infrastructure, trails, and roads; and recreationists or cross-border violators traveling along established trails or cross country (Phillips

et al.

1982, p. 10; Shohet 1999, p. 60; Ferguson 2014, p. 42; NPS 2015, p. 4; Ferguson 2016a, p. 26).

Direct removal of Bartram's stonecrop individuals and substrate due to erosion, or burial of individuals, may occur due to the placement of mineral extraction sites and debris piles. These impacts could severely impact small Bartram's stonecrop populations. Erosion from test pits (an excavation made to examine the subsurface conditions of a potential mine site) has been documented to remove portions of habitat occupied by Bartram's stonecrop in Flux Canyon (Phillips

et al.

1982, pp. 9-10).

Trampling

The trampling of individual Bartram's stonecrop plants may occur from a variety of activities, including livestock and wildlife movement; the placement and maintenance of infrastructure, trails, and roads; and recreationists or cross-border violators traveling along established trails or cross country (Phillips

et al.

1982, p. 10; Shohet 1999, p. 60; Ferguson 2014, p. 42; NPS 2015, p. 4; Ferguson 2016a, p. 26). Given the potential for these stressors, those populations with fewer than 50 individuals may be heavily impacted during periods of unusual recreational use. This stressor is considered in our analysis of future viability only when it may impact a population with fewer than 50 individuals.

Altered Fire Regime

Since the mid-1980s, wildfire frequency in western forests has nearly quadrupled compared to the average of the period 1970 to 1986 (Westerling

et

al.

2006, p. 941). The timing, frequency, extent, and destructiveness of wildfires are likely to continue to increase (Westerling

et al.

2006, p. 943), especially given historical land management actions, an increase in fire starts from cross-border violators and recreationists (

e.g.,

from campfires, cigarettes, target shooting), nonnative plant invasion, and continuing drought conditions (Westerling

et al.

2006, p. 940; FireScape 2016, entire; Fire Management Information System 2016, p. 2; Tersey 2017, pers. comm.). Altered fire regimes can have direct and indirect impacts to Bartram's stonecrop and its habitat. Direct impacts include burning of individual Bartram's stonecrop plants, resulting in injury, reduction in reproductive structures, or death. Indirect impacts of fire on Bartram's stonecrop may include increased runoff of floodwaters, post-fire flooding, deposition of debris and sediment originating in the burned area, erosion, changes in vegetation community composition and structure, increased presence of nonnative plants, alterations in the hydrologic and nutrient cycles, and loss of overstory canopy shade essential for maintaining Bartram's stonecrop microhabitat (Griffis

et al.

2000, p. 243; Crawford

et al.

2001, p. 265; Hart

et al.

2005, p. 167; Smithwick

et al.

2005, p. 165; Stephens

et al.

2014, p. 42; Ferguson 2014, p. 43; Ferguson 2016a, p. 26).

We are aware of 11 wildfires (Alamo, Brown, Elkhorn, Hog, Horseshoe II, La Sierra, Lizard, Mule Ridge, Murphy, Soldier Basin, and Spring) that have occurred in known Bartram's stonecrop sites in the past decade that killed some Bartram's stonecrop individuals and removed shade in some instances. When looking at the number of acres burned per sky island mountain range in comparison to the number of adult individuals known from that range, the two largest populations occur in sky island mountain ranges that have had the fewest acres burned in the past 10 years. It is not known if this is coincidence or is of significance, as we do not have pre-fire population counts in any population to address this question. Wildfires have burned in all nine sky island mountain ranges of southern Arizona that support Bartram's stonecrop during this time period. Fires did not burn through Bartram's stonecrop populations in all cases, but fire could occur in any population within this 10-year timeframe. Wildfire could potentially cause extirpation of small Bartram's stonecrop populations throughout the range of the species and have negative impacts on larger populations. In addition, because it is thought that Bartram's stonecrop seeds reside at the soil surface and the seeds are very tiny (Shohet 1999, p.48), it is likely that the seeds would not survive a wildfire.

The nonnative plants in the uplands and within Bartram's stonecrop populations include nonnative grass species such as Lehman's lovegrass and rose natal, both of which have numerous advantages over native grasses. Lehman's lovegrass resprouts from roots and tiller nodes not killed by hot fire, is not hampered by the reduction in mycorrhizae associated with fire and erosion, responds to winter precipitation when natives grasses are dormant, produces copious seed earlier than native grasses, maintains larger seedbanks than native grasses, and has higher seedling survival and establishment than native grasses during periods of drought (Anable 1990, p. 49; Anable

et al.

1992, p. 182; Robinett 1992, p. 101; Fernandez and Reynolds 2000, pp. 94-95; Crimmins and Comrie 2004, p. 464; Geiger and McPherson 2005, p. 896; Schussman

et al.

2006, p. 589; O'Dea 2007, p. 149; Archer and Predick 2008, p. 26; Mathias

et al.

2013, entire). Rose natal is capable of growing in low moisture situations, has prolific seed production, and culms that root from the nodes (Stokes

et al.

2011, p. 527). Both species outcompete native plants, reduce structural and spacial diversity of habitats, and increased biomass and fuel loads, increasing the fire frequency. Nonnative grasses have been reported with Bartram's stonecrop individuals in two instances, at French Joe Canyon and Juniper Flat populations, increasing the likelihood of fire occurrence and subsequent impacts to these two populations (Heritage Database Management System, E.O. ID 55; Simpson 2017, pers. comm.). Nonnative plant species increase the frequency and severity of wildfires, such wildfires can directly and indirectly impact individuals and populations.

Illegal Collection

Bartram's stonecrop is an attractive small plant that can be easily collected by gardeners and succulent enthusiasts. Tagged individuals were uprooted and taken from two sites in the Santa Rita Mountains, one near a campsite (Shohet 1999, p. 60). In a 2016 on-line Google search for Bartram's stonecrop for sale, an advertisement from a collector in Texas offered to pay cash for Bartram's stonecrop seedlings or rooted cuttings. One website notes that the similar southern Arizona occurring species,

G. rusbyi,

is cultivated and legally available for sale from cactus nurseries; however, Bartram's stonecrop is not (because it is more difficult to propagate and maintain in captivity) and is therefore vulnerable to collection. Small populations may not be able to recover from collection, especially if the mature, reproductive plants are removed. The removal of mature plants reduces the overall reproductive effort of the population, thereby reducing the overall resilience of the population.

Altered Precipitation, Drought, Flooding, and Freezing Regimes

Precipitation within the sky island mountain ranges is bimodal, with winter snow and rain, and summer monsoon rain (CLIMAS 2014, entire). Fall and winter (October through March) precipitation is needed for Bartram's stonecrop germination, and both summer (July and August) and fall precipitation (October and November) is needed for Bartram's stonecrop flower production. Flowering is triggered by fall rains and does not occur during periods of water stress (Shohet 1999, pp. 22, 25, 36, 39). Altered precipitation timing and form (

i.e.,

snow versus rain), as well as reduced precipitation in the winter and spring and prolonged drought, are important considerations in the analysis of the future stressors to Bartram's stonecrop due to increased nonnative competition during times of reduced precipitation and drought, which exacerbate impacts from stressors (Anable 1990, p. 49; Robinett 1992, p. 101; Fernandez and Reynolds 2000, pp. 94-95; Geiger and McPherson 2005, p. 896; Schussman

et al.

2006, p. 589; Archer and Predick 2008, p. 26; Mathias

et al.

2013, entire). In addition, reduced precipitation in the winter and spring and drought will also impact moisture availability for Bartram's stonecrop's germination, growth, and flowering.

Altered precipitation timing and form (snow versus rain), as well as reduced winter and spring precipitation and prolonged drought, are currently occurring and projected to increase or be altered from normal in the Southwest (Garfin

et al.

2014, entire). Recently there has been a decrease in the amount of snowpack, earlier snowmelt, and increased drought severity in the Southwest (Garfin

et al.

2013, entire; Garfin 2013b, p. 465). Further, more wintertime precipitation is falling as rain rather than snow in the western United States (IPCC 2013, p. 204; Garfin 2013b p. 465). This means that the amount of runoff in the spring when snow melts is reduced, as is soil moisture.

Under a continuation A2-high emissions scenario, reduced winter and

spring precipitation is consistently projected for the southern part of the Southwest by 2100, as part of the general global precipitation reduction in subtropical areas (Garfin

et al.

2014, p. 465). Analyses of the southwestern United States indicate future drying, primarily due to a decrease in winter precipitation under both the RCP 4.5 and 8.5 scenarios (IPCC 2013, p. 1080). The annual projected changes in precipitation for 2025 to 2049 under RCP 4.5 and 8.5 scenarios ranges from an increase of 1.3 cm/mo (0.5 to a decrease of 0.5 in/mo), with an annual average of no change compared to 1981 to 2010 (USGS 2019, entire). However, winter and spring precipitation under both emission scenarios is projected to decrease from −0.3 to −1 cm (−0.1 to −0.4 in) (MACA 2019) or a decrease up to 10 percent for 2016-2035 relative to 1986-2005 under RCP 4.5 (IPCC 2013, p. 985). The decrease in winter and spring precipitation would likely be greater under the RCP 8.5 scenario. There is some evidence from comparing observations with simulations of the recent past that climate models might be underestimating the magnitude of changes in precipitation in many regions (IPCC 2013, p. 986). The climate-model-projected simulations indicate that a high degree of variability of annual precipitation will continue during the coming century, for both low and high emission scenarios (Garfin 2013, p. 110). This suggests that the Southwest will remain susceptible to unusually wet spells and, on the other hand, will remain prone to occasional drought episodes (Garfin 2013, p. 110). However, decrease in soil moisture across much of the Southwest is projected under both scenarios by mid-century, due to increased evaporation (IPCC 2013 p. 1259). Late winter-spring mountain snowpack in the Southwest is predicted to continue to decline over the 21st century under the high emission scenario (A2), mostly because of projected increased temperature (Garfin

et al.

2013, p. 6). Reduced rain and snow, earlier snowmelt, and drying tendencies cause a reduction in late-spring and summer runoff. Together these effects, along with increases in evaporation, result in lower soil moisture by early summer (Gafrin 2013, p. 117).

Precipitation timing and amount impacts the germination, growth, and flowering of Bartram's stonecrop, resulting in the loss of individuals and recruitment, and overall reducing the population size.

In the Southwest, temperatures increased 2.7°C (1.6 °F) plus or minus 0.9 °C (0.5 °F), between 1901 and 2010, and more heat waves occurred over the Southwest during 2001-2010 compared to average occurrences in the 20th century. In the future, under RCP 4.5, the annual maximum temperature is projected to increase by 5°C (2.7°F) for 2025-2049 and 7.3 °C (4°F) for 2050-2074, and 5 °C (2.7°F) for 2025-2049 and 10.4 °C (5.7°F) for 2050-2074 under RCP 8.5, all relative to 1981-2010 (USGS 2019, entire). When temperatures rise, as has been occurring in recent decades and as is projected to continue into the future, evapotranspiration rates also increase and soil moisture decreases. Along with projected warming and increased evapotranspiration, it is highly likely that droughts will become more severe (Garfin 2013, pp. 137-138). A decrease of up to 4 percent soil moisture is projected under RCP 4.5 for 2016-2035, relative to 1986-2005. The decrease in soil moisture would likely be greater under RCP 8.5. Further, the evaporation deficient increases under RCP 4.5 and increases more in RCP 8.5 in 2025 to 2049, relative to 1981 to 2010. Based on the high emissions scenario, the current 100-year drought will become commonplace in the second half of this century and future droughts will be much more severe than those previously recorded (Garfin 2013, p. 138). This projection of intensified drought conditions on the Colorado River is not due to changes in precipitation, but rather due directly to warming and its effect on reducing soil moisture (Garfin 2013, p. 138). Physiological effects of CO

2

may involve both the stomatal response, which acts to restrict transpiration, and an increase in plant growth and leaf area, which acts to increase evapotranspiration (IPCC 2013 p. 986). An increase in evapotranspiration results in water loss from the plant and increases stress on the plant. This increase in stress impacts photosynthesis, respiration, transpiration, water use efficiency, leaf conductance, growth rate, vigor, and gas exchange. These impacts result in reduced growth, flowering, and seed production, and, therefore, reduces overall recruitment and population numbers.

Although rare species in the southwestern United States evolved with drought, recent changes in temperature and rainfall patterns present stressful conditions of increased magnitude above what the species faced historically and raise the question of whether the species in this rule can persist. Some species will shift their distributions in response to warming of the climate (McLaughlin

et al.

2002, p. 6070). It is highly unlikely that Bartram's stonecrop would be able to naturally shift its range to keep up with current and high projected rates of climate change due to its general state of population decline, lack of suitable intervening habitat, and abundant nonnative competitors. Thus, localized extinctions over portions of Bartram's stonecrop's range could result.

Small Populations

Stressors are exacerbated in populations with only a small number (

e.g.,

fewer than 50) of individuals. Small populations are less able to recover from losses caused by random environmental changes (Shaffer and Stein 2000, pp. 308-310), such as fluctuations in reproduction (demographic stochasticity), variations in rainfall (environmental stochasticity), or changes in the frequency or severity of wildfires. Approximately half of the extant Bartram's stonecrop populations contain 50 or fewer individuals. Loss due to erosion, trampling, collection, predation, fire, severe frost, or other stressors have the potential to seriously damage or completely remove these small populations.

In summary, the stressors that pose the largest risk to future species viability are primarily related to habitat changes: Groundwater extraction from mining, long-term drought, and alteration in wildfire regime. These stressors may reduce nearby water levels, shade, and humidity within Bartram's stonecrop habitat and may directly impact individuals. Other important stressors include erosion or trampling from livestock, wildlife, or human activities; illegal collection; predation of Bartram's stonecrop or their shade trees by wildlife and insects; abnormal freezing or flooding events; or other stressors that have the potential to seriously damage or completely remove small populations. Synergistic interactions among wildfire, drought, altered precipitation, and increased temperatures cumulatively and cyclically impact Bartram's stonecrop, and all stressors are exacerbated in small populations.

Current Condition of Bartram's Stonecrop

Historically, we know of 33 populations spread across 13 mountain ranges. Four populations have been extirpated in the United States in recent years, and a fifth population has likely contracted in size. In addition, the southeastern Arizona landscape has experienced many changes since the 1890s, resulting from intensive cattle grazing, water development, and fire

suppression (

e.g.,

Bahre 1991, entire). These impacts may have reduced the range or number of populations and individuals. Currently, 29 extant populations occur across 12 mountain ranges in the United States and Mexico: 9 in southern Arizona and 3 in northern Mexico. The U.S. populations total 3,726 individuals within occupied habitats that total about 2 ha (5 ac). Data are lacking for the Mexico populations; however, based on populations in the United States, which are mostly small, we believe that the three populations in Mexico are of similar size to U.S. populations but may be in worse condition, because of limited native habitat management, similar climate change impacts, equally frequent wildfires, and likely more livestock impacts (Romo

et al.

2012, entire; Arriaga

et al.

2004, entire; Fishbein and Warren 1994, p. 20).

Population Resiliency for Bartram's Stonecrop

To help determine current condition, we assessed each population in terms of its resiliency and assessed the species' representation and redundancy. Our analysis of the past, current, and future stressors on the resources that Bartram's stonecrop needs for long-term viability revealed a number of stressors to this species. All Bartram's stonecrop populations likely contain nonnative grasses. Further, altered fire regime has the potential to affect all populations. This altered fire regime enhances the spread of nonnatives. Consequently, all populations of Bartram's stonecrop will be further impacted by nonnative grasses in the near future. Altered precipitation, increased temperatures, and decreased annual precipitation are current and ongoing regional conditions that are impacting all populations of Bartram's stonecrop. These environmental conditions exacerbate an altered fire regime, which, in turn, further drives the spread of nonnatives. In addition, nonnative grasses have competitive advantage over native grasses during periods of drought. Many currently undeveloped areas of locatable mineral deposits may be explored or mined in the future. We do not know the extent of future mine activity within the range of Bartram's stonecrop; however, there are 12 mining projects currently ongoing or proposed within 8 km (5 mi) of Bartram's stonecrop populations in Arizona. The range of current and projected mining activities varies from 1 to 10 per sky island mountain range containing Bartram's stonecrop (USFS 2012, entire). One population, Sycamore Canyon (115 adult individuals), would be affected by groundwater drawdown due to the Rosemont Mine. Sycamore Canyon is currently in moderate condition. Further, this species is collected and sold. Synergistic interactions among wildfire, nonnative grasses, decreased precipitation, and increased temperatures cumulatively and cyclically impact Bartram's stonecrop, and all stressors are exacerbated in small populations. In addition, because approximately 41 percent (12 populations) of the extant Bartram's stonecrop populations contain 50 or fewer individuals, loss due to erosion, trampling, collection, predation, fire, severe frost, or other stressors have the potential to seriously damage or completely remove these small populations. Of the 29 extant populations, 1 population (3 percent) is in high condition, 21 populations (72 percent) are in moderate condition, and 7 populations (24 percent) are in low condition (Table 8, below). Population resiliency categories are described in Table 4, above, and in the SSA report (Service 2018b).

Table 8—Bartram's Stonecrop Current Population Condition

Sky Island

Population

Number of

individuals

Current condition

Baboquivari

Brown Canyon

112

Moderate.

Thomas Canyon

5

Low.

Chiricahua

Echo Canyon

186

Moderate.

Indian Creek

0

Extirpated.

Dragoon

Carlink Canyon

0

Extirpated.

Jordan Canyon

415

Moderate.

Sheephead

45

Moderate.

Slavin Gulch

9

Moderate.

Stronghold Canyon East

188

Moderate.

Stronghold Canyon West

533

High.

Empire

Empire Mountains

0

Extirpated.

Mule

Juniper Flat

798

Moderate.

Pajarito-Atascosa

Alamo Canyon

134

Moderate.

Holden Canyon

7

Moderate.

Sycamore Canyon

298

Moderate.

Warsaw Canyon

13

Moderate.

Patagonia

Alum Gulch

123

Moderate.

Rincon

Chimenea-Madrona Canyon

9

Moderate.

Happy Valley North

0

Extirpated.

Happy Valley South

14

Moderate.

Santa Rita

Adobe Canyon

82

Moderate.

Gardner Canyon

14

Moderate.

Josephine Canyon

71

Moderate.

Madera Canyon

76

Moderate.

Squaw Gulch

5

Low.

Sycamore Canyon

115

Moderate.

Temporal Gulch

7

Moderate.

Walker Canyon

3

Moderate.

Whetstone

Deathtrap Canyon

135

Low.

French Joe Canyon

87

Low.

Sierra Las Avispas, Sonora

Sierra Las Avispas

10

Low.

Sierra La Escuadra, Chihuahua

Near Colonia Pacheco

10

Low.

Sierra La Estancia, Chihuahua

Cuarenta Casas

10

Low.

Bartram's Stonecrop Representation

No genetic studies have been conducted within or between the 33 historical populations of Bartram's stonecrop in southern Arizona and Mexico. However, we assessed representation for Bartram's stonecrop in the form of its geographic distribution across the range. Some genetic exchange likely occurs within populations containing many subpopulations or many plants per subpopulation. Sky island populations on different mountain ranges are widely separated (ranging from roughly 14 to 42 km (8.7 to 26 mi) apart), making cross-pollination across sky islands highly unlikely. Mountain ranges that have only one or two populations, have only one subpopulation per population, or have low numbers of individuals per population with several miles between mountain ranges may not be as genetically diverse because pollination or transport of seeds between populations may be very limited. However, there may be genetic diversity between populations within and between the sky island mountain ranges in response to elevational and other environmental differences between locations. Due to the loss of four populations, it is possible that there has been a loss of genetic diversity. However, because the species occurs across 29 populations in 12 mountain ranges, it is likely some genetic diversity exists among mountain ranges.

In addition, because the plant occurs on multiple substrate types and at a range of elevations (1,067 to 2,042 m (3,500 to 6,700 ft)), there is likely some local adaptation and genetic differentiation among populations. This range in elevation provides a variety of climatic conditions for the species to inhabit. Lastly, in at least three locations (Flux Canyon, Sycamore Canyon (Pajarito-Atascosa Mountains), and Gardner Canyon populations), Bartram's stonecrop have been reported over many decades, indicating that these populations may have the genetic and environmental diversity to adapt to changing conditions.

Bartram's Stonecrop Redundancy

The Bartram's stonecrop populations in the United States and Mexico are naturally fragmented between mountain ranges. Currently, 29 extant Bartram's stonecrop populations are spread across 12 different mountain ranges in southern Arizona and northern Mexico. Although these numbers may imply redundancy across its range, note that 24 of the 29 extant populations contain fewer than 150 total individual plants. Further, 14 of the 29 populations have 50 individuals or less, and 4 populations have been extirpated over recent (approximately 10) years. Five mountain ranges (Baboquivari, Chiricahua, Mule, Whetstone, and Patagonia Mountains) have only one or two populations each or have only have one subpopulation per population, and low numbers of individuals per population. These sky island mountain ranges are several miles away from the other sky island mountain ranges, so natural gene exchange or re-establishment following extirpation is unlikely. In addition, the Mule Mountains contain large number of Bartram's stonecrop individuals, but there is only one population and it is approximately 38 km (23.6 mi) away from the nearest population, making natural re-establishment of populations unlikely. In addition, this population is known to be contracting in size due to drying of habitat (The Nature Conservancy 1987, p. 2).

Future Condition of the Bartram's Stonecrop

We now consider the species' future condition of population resiliency and the species' representation and redundancy are likely to be. The future viability of Bartram's stonecrop depends on maintaining multiple resilient populations over time. The resiliency of Bartram's stonecrop populations depends on moisture in their microenvironment maintained by shade from overstory vegetation, spring and winter precipitation, proximity to water, and vegetation litter. We expect the 29 extant Bartram's stonecrop populations to experience changes to all of these aspects of their habitat, although it may be in different ways under the different conditions. In addition, direct impacts to Bartram's stonecrop through being dislodged, buried, or collected will continue to impact the species.

Given our uncertainty regarding the scope of the stressors manifesting and the species' response, we forecasted future conditions of Bartram's stonecrop under four plausible future scenarios (see chapter 6 of the SSA report; Service 2018b). We developed these scenarios to span a range of potential stressors that are ongoing or will occur in the future that we believe will influence the future status of the species. We chose 10 years to evaluate the current condition, as well as future projections out to 40 years because this is within the range of predictions of available hydrological and climate change model forecasts and is within the time period of the Rosemont Mine effects. This time frame represents eight generations of the Bartram's stonecrop, which allows us to assess reproductive effects on the species and allows the species opportunities to rebound after poor water years. The ten-year time step also represents a reasonable timeframe to judge the species' current vulnerability to threats as they are manifested now, without projecting changes to threats that longer timeframes would provide. Thus, the future scenarios forecast the viability of Bartram's stonecrop over the next 40 years. See table 9 below for a summary of the four scenarios. For more detail, see Chapter 6 of the SSA report (Service 2018b, entire).

Table 9—Future Scenarios for Bartram's Stonecrop

Risks

Mining activity

Altered fire regime

Climate

Climate

Individual effects

Conservation

Risk described

Water extraction, Excavation, Burial, Shade reduction

Lightning Recreation Cross border violators Nonnative plants

Reduction in available water * and/or shade

Dislodging from flooding events, Seedling desiccation, Flowering halt, Shade removal

Livestock Recreation Trampling Predation Collection

Conservation actions implemented.

Scenario 1. Continuation continuing into the future

Ongoing or planned mining activities as of 2012 (~20)

Number of wildfires annually increases at the same rate as the last 10 years

Available water and drought continue at the same level as in the past 10 years. Emissions 8.5

Number and severity of flooding events continues at the past 10 years. Emissions <4.5

Applied to populations <50 individuals

No new individuals, subpopulations or populations found. No augmentation of existing populations, no seed preservation, nonnatives controlled, and forest thinned.

Scenario 2. Conservation

Number of mining activities does not increase from current condition

Number of wildfires does not increase from current rate

Available water remains stable. Emissions 4.5

Flooding events do not increase. Emissions <4.5

Applied to populations <50 individuals

Sites revisited and additional plants are located, sites are augmented, or new sites are established, nonnatives controlled, and forest thinned.

Scenario 3. Moderate increase in negative effects

1-3 new mining activities (above the 2012 number) are implemented and/or existing mines expand

Number of wildfires increases in uplands

Available water is reduced per 8.5 emissions scenario

Increases in flash flooding per 4.5 emissions scenario

Applied to populations <50 individuals

No new individuals, subpopulations, or populations found, and no augmentation of existing populations, nonnatives controlled, and forest thinned.

Scenario 4. Major increase in negative effects

>3 new mining activities are implemented and/or existing mines expand

Number of wildfires increases in uplands

Available water is reduced per 8.5 emissions scenario

Increases in flash flooding per 8.5 emissions scenario

Applied to populations <50 individuals

No new individuals, subpopulations or populations found, and no augmentation of existing populations, nonnatives controlled, and forest thinned.

* Available water includes precipitation, soil moisture, humidity, surface water, aquifer recharge, reduction in riparian vegetation, and increased number of days without water.

All scenarios consider impacts from mining, wildfire, and climate. In addition, effects on individual plants from multiple stressors are assessed, including livestock, recreation, trampling, predation, and collection. The “continuation” scenario evaluates the condition of Bartram's stonecrop if there is no increase in risks to the populations relative to what exists today, while the other scenarios evaluate the response of the species to changes in those risks. The “conservation” scenario takes into account realistically possible additional protective measures which may or may not happen. The “moderate effects” scenario is an increase in the risks to populations with changes in climate as projected in a lower (8.5) emissions scenario along with increases in other stressors. The “major effects” scenario is a further increase in risks to populations, with changes in climate projected at a higher (8.5) emissions scenario, and with additional increases in other stressors. These are described in more detail in chapter 6 of the SSA report (Service 2018b).

The most likely scenario is the “moderate effects” scenario, with impacts to the species occurring around the 40-year time step. Under the “moderate effects” scenario, water flow reduction due to drought and groundwater extraction continues to reduce the humid microhabitat for this species. Cross-border violator traffic continues, and risk of catastrophic wildfire is high due to dry conditions; invasion of nonnatives in the uplands; and increased risk of fire starts from illegal activity, recreation, and natural causes. Mining impacts individuals in the Patagonia and Santa Rita Mountains. Collection, trampling, freezing, predation, and human impacts also continue at current or increased levels. The full analyses of all scenarios are available in the SSA report (Service 2018b, chapter 6); however, we are only presenting the full results of the “moderate effects” scenario here because it gives the most realistic projection of the future condition of the species.

Under the “moderate effects” scenario, within the 40-year timeframe, we expect Bartram's stonecrop's viability to be characterized by lower levels of resiliency, representation, and redundancy than it has currently, which is already reduced as described above. Under the “moderate effects” scenario, no populations would be in high condition, 4 populations (12 percent) would remain in moderate condition, 16 populations (52 percent) would be in low condition, and 13 populations (36 percent) would be extirpated, further reducing population redundancy and connectivity (see table 6.6 in the SSA report; Service 2018b). Under the “moderate effects” scenario, because of the intensity of stressors discussed above, 22 populations would be reduced from their current condition (see Table 10, and see figure 6.3 and table 6.6 in the SSA report (Service 2018b)). We further believed that in the “moderate effects” scenario, one of the three small populations in Mexico becomes extirpated due to the amount of nonnatives contributing to fire, reduction in precipitation, increase in drought, and low resiliency of a small population.

Table 10—Bartram's Stonecrop Population Conditions Under the “Moderate Effects” Scenario

Sky Island

Population

Condition under the

“moderate effects” scenario

Baboquivari

Brown Canyon

Low.

Thomas Canyon

Low.

Chiricahua

Echo Canyon

Low.

Indian Creek

Extirpated.

Dragoon

Carlink Canyon

Extirpated.

Jordan Canyon

Moderate.

Sheephead

Low.

Slavin Gulch

Low.

Stronghold Canyon East

Moderate.

Stronghold Canyon West

Moderate.

Empire

Empire Mountains

Extirpated.

Mule

Juniper Flat

Low.

Pajarito-Atascosa

Alamo Canyon

Low.

Holden Canyon

Extirpated.

Sycamore Canyon

Moderate.

Warsaw Canyon

Extirpated.

Patagonia

Alum Canyon

Extirpated.

Rincon

Chimenea-Madrona Canyon

Low.

Happy Valley North

Extirpated.

Happy Valley South

Low.

Santa Rita

Adobe Canyon

Low.

Gardner Canyon

Low.

Josephine Canyon

Low.

Madera Canyon

Extirpated.

Squaw Gulch

Extirpated.

Sycamore Canyon

Extirpated.

Temporal Gulch

Low.

Walker Canyon

Extirpated.

Whetstone

Deathtrap Canyon

Low.

French Joe Canyon

Extirpated.

Sierra Las Avispas, Sonora

Sierra Las Avispas

Low.

Sierra La Escuadra, Chihuahua

Near Colonia Pacheco

Extirpated.

Sierra La Estancia, Chihuahua

Cuarenta Casas

Low.

Determination

Section 4 of the Act (16 U.S.C. 1533), and its implementing regulations at 50 CFR part 424, set forth the procedures for adding species to the Federal Lists of Endangered and Threatened Wildlife and Plants. Under section 4(a)(1) of the Act, we may list a species based on (A) The present or threatened destruction, modification, or curtailment of its habitat or range; (B) overutilization for commercial, recreational, scientific, or educational purposes; (C) disease or predation; (D) the inadequacy of existing regulatory mechanisms; or (E) other natural or manmade factors affecting its continued existence.

We have carefully assessed the best scientific and commercial information available regarding the past, present, and future stressors to beardless chinchweed and Bartram's stonecrop.

The Act defines an endangered species as any species that is “in danger of extinction throughout all or a significant portion of its range” and a threatened species as any species “that is likely to become endangered throughout all or a significant portion of its range within the foreseeable future.”

Therefore, on the basis of the best available scientific and commercial information, we propose listing beardless chinchweed as endangered in accordance with sections 3(6) and 4(a)(1) of the Act and Bartram's stonecrop as threatened in accordance with sections 3(20) and 4(a)(1) of the Act.

Beardless Chinchweed

Historically there were 21 populations. Nine populations have been extirpated, leaving 12 extant populations (six in the United States and six in Mexico). The six populations in the United States consist of 387 individuals spread across less than 2 ha (5 ac). The six populations have been reported from northern Mexico, but this information is from 1940 or earlier.

The proliferation of invasive nonnative grasses throughout most of the beardless chinchweed's range has greatly affected this species through increased competition and altered fire regimes. Many of these historical locations no longer support beardless chinchweed due to this alteration of habitat (National Park Service 2014, pp. 3-4; Service 2014b, pp. 1-2; Service 2014c, entire; Service 2014d, pp. 1-2).

All beardless chinchweed populations likely contain nonnative grasses, resulting in habitat loss (Factor A). Further, altered fire regime (Factors A and E), which is currently or in the near future impacting all populations, drives the spread of nonnatives (Factor A), exacerbating the encroachment of nonnative grasses. Consequently, all remaining populations of beardless chinchweed are impacted by nonnative grasses now or will be in the near future. Altered precipitation (Factors A and E), increased temperatures (Factors A and E), and decreased annual precipitation (Factors A and E) are current and ongoing regional conditions that are impacting all populations of beardless chinchweed. These environmental conditions exacerbate an altered fire regime, which, in turn, drives the spread of nonnatives. In addition, nonnative grasses have competitive advantage over native grasses during periods of drought. Road and trail maintenance (Factors A and E) is altering habitat and likely resulting in the direct killing of individuals in three populations (Ruby Road, Scotia Canyon, and Coronado National Memorial). In addition, all individuals in these three populations are being impacted by dust (Factor E) from the road. These three populations are already of low resiliency. Two additional populations (McCleary Canyon—Gunsight Pass and McCleary Canyon—Wasp Canyon) will be impacted by roads (Factor A) related to mining operations in the near future (Westland 2010, p. iv). All individuals of these two populations will also be impacted by dust (Factor E). One of these populations is already of low resiliency and the other is of moderate resiliency. Of the 12 populations, 11 (92 percent) are small populations (fewer than 50 individuals). Synergistic interactions among wildfire, nonnative grasses, decreased precipitation, and increased temperatures cumulatively and cyclically impact beardless chinchweed, and all stressors are exacerbated in small populations (Factor E). No conservation efforts have been implemented for this species.

We consider beardless chinchweed to have poor representation in the form of potential genetic diversity (Factor E). All but one population has fewer than 50 individuals. Small populations are susceptible to the loss of genetic diversity, genetic drift, and inbreeding. There are currently six populations spread across four mountain ranges in the United States and six populations in northern Mexico that are presumed extant. Five of the six extant U.S. populations do not have multiple subpopulations (all but the Coronado National Memorial population, which has two subpopulations). Mountain ranges that have only one or two populations, have only have one subpopulation per population, or have low numbers of individuals per population with several miles (16 to 61 km (9.9 to 37.9 mi)) between mountain ranges, may not be genetically diverse because pollination or transport of seeds between populations may be very limited. This could mean that between-population genetic diversity may be greater than within-population diversity (Smith and Wayne 1996, p. 333; Lindenmayer and Peakall 2000, p. 200). Further, nine populations are extirpated, and it is possible that there has been a loss of genetic diversity.

Beardless chinchweed populations in the United States range in elevation from 1,158 m (3,799 ft) to 1,737 m (5,699 ft) in elevation. Of the 15 historical U.S. populations, 8

(approximately 53 percent) fall below 457 m (1,500 ft) elevation. Of these eight, six have become extirpated in recent decades. This essentially indicates a loss at this lower elevational range and possibly loss of some local adaptation to warmer or dryer environments and genetic differentiation among populations (Factor E).

Beardless chinchweed needs to have multiple resilient populations distributed throughout its range to provide for redundancy. Beardless chinchweed needs multiple resilient populations spread over their range that are distributed in such a way that a catastrophic event will not result in the loss of all populations. With the known extant populations being separated by as much as 35 km (21.8 mi) in southern Arizona and even farther in northern Mexico, there is little connection potential between known disjunct populations. Therefore, a localized stressor such as grazing during flowering would impact only those groups of plants nearby the activity. However, repeated, large-scale, moderate- and high-severity fires, nonnative plant invasion, and climatic changes occur across the region and could impact all populations now or in the near future. The distance among populations reduces connectivity among populations and mountain ranges, making it unlikely that a site that is extirpated can be naturally recolonized by another population (Factor E).

We find that beardless chinchweed is presently in danger of extinction throughout its entire range based on the severity and immediacy of stressors currently impacting the species. The overall range has been significantly reduced (nine populations extirpated), and the remaining habitat and populations are threatened by a variety of factors acting in combination to reduce the overall viability of the species. The risk of extinction is high because the remaining populations are small, isolated, and have limited potential for natural recolonization. Therefore, on the basis of the best available scientific and commercial information, we propose listing beardless chinchweed as endangered in accordance with sections 3(6) and 4(a)(1) of the Act. We find that a threatened species status is not appropriate for beardless chinchweed because of the species's current precarious condition due to its contracted range, because the stressors are severe and occurring rangewide, and because the stressors are ongoing and expected to continue into the future.

Under the Act and our implementing regulations, a species may warrant listing if it is endangered or threatened throughout all or a significant portion of its range. Because we have determined that beardless chinchweed is endangered throughout all of its range, we find it unnecessary to proceed to an evaluation of potentially significant portions of the range. Where the best available information allows the Service to determine a status for the species rangewide, that determination should be given conclusive weight because a rangewide determination of status more accurately reflects the species' degree of imperilment and better promotes the purposes of the statute. Under this reading, we should first consider whether listing is appropriate based on a rangewide analysis and proceed to conduct a “significant portion of its range” analysis if, and only if, a species does not qualify for listing as either endangered or threatened according to the “all” language. We note that the court in

Desert Survivors

v.

Department of the Interior,

No. 16-cv-01165-JCS, 2018 WL 4053447 (N.D. Cal. Aug. 24, 2018), did not address this issue, and our conclusion is therefore consistent with the opinion in that case.

Therefore, on the basis of the best available scientific and commercial information, we propose to list beardless chinchweed as an endangered species across its entire range in accordance with sections 3(6) and 4(a)(1) of the Act.

Bartram's Stonecrop

Bartram's stonecrop has experienced population declines and four populations have been lost entirely. Currently, there are 29 extant populations. All Bartram's stonecrop populations contain or are near nonnative grasses resulting in habitat loss in the future (Factor A). Further, altered fire regime (Factors A and E), which is currently and in the future impacting all populations, drives the spread of nonnatives (Factor A), exacerbating the encroachment of nonnative grasses. Consequently, all populations of Bartram's stonecrop will be impacted by nonnative grasses in the future. Altered precipitation (Factors A and E), increased temperatures (Factors A and E), and decreased annual precipitation (Factors A and E) are current and ongoing regional conditions that are impacting all populations of Bartram's stonecrop. These environmental conditions exacerbate an altered fire regime, which, in turn, drives the spread of nonnatives. In addition, nonnative grasses have competitive advantage over native grasses during periods of drought. Many currently undeveloped areas of locatable mineral deposits may be explored or mined in the future (Factors A and E). The range of current and projected mining activities varies from 1 to 10 per sky island mountain range containing Bartram's stonecrop (USFS 2012, entire). One population, Sycamore Canyon (115 adult individuals), will be affected by groundwater drawdown due to the Rosemont Mine, which will impact the shade and moist microclimate this species needs (Factor A). This species is known to be collected and sold (Factor B), and plants in close proximity to

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