Encelia farinosa, white brittlebush
Brahmsteadt, Taryn R. 2024. Encelia farinosa, white brittlebush. In: Fire Effects Information System, [Online]. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Missoula Fire Sciences Laboratory (Producer). Available: https://research.fs.usda.gov/feis/species-reviews/encfar
| Abbreviation | Common Name | Scientific Name | Classification | Status |
|---|---|---|---|---|
| Plants | ||||
| ENCFAR | white brittlebush | Encelia farinosa | Life Form: Plants/Shrub, Plants/Subshrub Kingdom: Plantae Class: Dicot Order: Asterales Family: Asteraceae Genus: Encelia | Fed. Protected: No Nativity: Native Invasiveness: Noninvasive |
This review summarizes the information that was available in the scientific literature as of 2024 on the biology, ecology, and effects of fire on white brittlebush in North America.
White brittlebush is a shrub that can grow to 1 m tall. It is native to the southwestern United States and northern Mexico. It is most common in North American warm desert scrub, chaparral, and coastal sage scrub. It can form monospecific stands throughout its range, especially after disturbances, and often dominates frequently disturbed desert scrub and coastal sage scrub. It occurs most often on rocky slopes, hillsides, and upper bajadas. It is drought tolerant and frost intolerant, and its range is limited by cold temperatures.
White brittlebush establishes by seed, and seedlings readily establish in the postfire environment. It can also resprout after top-kill from fire. It is usually a weak resprouter, although the degree of resprouting depends on plant, site, and fire characteristics. Resprouting plants contribute to postfire recovery to varying degrees, from providing most postfire regeneration to providing very little.
White brittlebush is easily damaged by fire and is often killed or top-killed, but populations typically recover quickly, often becoming more abundant in recently burned than long unburned areas. However, populations may take many years to recover on sites lacking on-site seed sources following very frequent, intense, and/or large fires. Populations may decrease after fire if buffelgrass or other nonnative invasive plants preclude plants from establishing. Alternatively, white brittlebush may dominate in burned communities for decades or longer.
Historically, desert scrub communities with white brittlebush did not have enough fine fuel to carry fire in most years. However, in some areas, nonnative invasive annual and perennial grasses form a more persistent, continuous layer of fine fuels than native species, resulting in more frequent fires. Fuel and fire regime characteristics in these areas have likely shifted outside the range of historical variation, primarily due to the introduction and spread of invasive grasses and human ignitions. In some chaparral and coastal sage scrub communities, grass invasions have led to more frequent fire and conversion to invasive-dominated grasslands with few remaining shrubs other than white brittlebush. Thus, the primary management in white brittlebush communities usually involves fuels reduction, particularly of invasive species, and full suppression of wildfires and desert fires.
Climate models suggest that North American warm deserts may become warmer and drier over the next century, possibly reducing the number of germination and establishment events for white brittlebush.
Figure 1—A stand dominated by white brittlebush in southern California.
Taxonomy
The scientific name of white brittlebush is Encelia farinosa A. Gray ex Torr. (Asteraceae) [20,94,119,121,265]. Flora of North America (2024) [94] does not recognize any lower taxa, but several sources recognize different varieties (see Synonyms).
White brittlebush hybridizes with other Encelia species, including button brittlebush [94,137], California brittlebush, Vizcaino encelia, Central Desert encelia, Palmer encelia [226], and plains encelia [63,85], as well as at least one desertsunflower species—hairy desertsunflower [136]. Hybrids within the same genus are fertile in the wild and in cultivation, and hybrids from different genera are typically sterile [63].
Common names are used throughout this Species Review. See the Appendix for scientific names of plants and animals mentioned in this review.
Synonyms
- Encelia farinosa A. Gray ex Torr. var. farinosa [97,105,137]
- Encelia farinosa A. Gray ex Torr. var. phenicodonta (S. F. Blake) I. M. Johnston [94,97,105,137,204]
- Encelia farinosa A. Gray ex Torr. var. radians Brandegee ex S.F. Blake [97,105,204]
- Encelia radians Brandegee [97,105,204]
Other Common Names
brittlebush, incienso, goldenhills, hierba del bazo
General Distribution
Figure 2— Native distribution of white brittlebush.
White brittlebush is native to the Mojave and Sonoran deserts [53,100,221,247,250], as well as interior valleys of California’s South Coast [69,75,108,173] and the northwesternmost Chihuahuan Desert [138]. It occurs in Arizona, southern California, southern Nevada, and southwestern Utah in the United States [121,255,257,265], and extends south into Baja California Sur and Sinaloa in Mexico [94]. It is one of the most prevalent shrub species in the Sonoran Desert [221].
White brittlebush is naturalized at relatively low elevations on dry sites in Maui, Hawai’i [97], where it was first collected in 1909 [231,261]. It has also been introduced to Saudi Arabia [242].
White brittlebush is commonly planted for restoration of disturbed sites, especially along roadsides [20,27,35,45,48,100], and may be expanding its range in Arizona, California, and Nevada as a result [133].
States and Provinces
- United States: AZ, CA, HI, NV, UT [97,255]
- Mexico: BCN, BCS, Sin, Son [94]
Site Characteristics
White brittlebush grows in arid and semi-arid climates [228]. Annual precipitation ranges from 50 to 400 mm and varies greatly from year to year. Precipitation patterns vary across its range and are characterized by long dry periods. White brittlebush occurs in all three North American warm deserts, all of which exhibit seasonal rainfall bimodality to some degree. Precipitation peaks in the winter in the Mojave Desert, in the summer in the Chihuahuan Desert, and in either winter or summer in the Sonoran Desert, depending on location. Summer precipitation occurs during late summer monsoons, when localized, high-intensity thunderstorms deliver sometimes substantial amounts of rainfall [72,147,250,279]. In coastal California, precipitation peaks in the winter [124].
Across its range, maximum summer temperatures commonly exceed 32 °C. Although winters are generally mild [129,217], temperatures sometimes drop below freezing [147], but freezing rarely occurs in most communities where white brittlebush is a significant component [108]. White brittlebush is frost intolerant, and its distribution is limited in the north and at high elevation by cold weather extremes, including hard freezes or frequent frosts [29,246,262]. Frost damage can cause complete defoliation at -5 °C [228].
| Area | Elevational range (m) |
|---|---|
| Southwest | 900–1,000 [40] |
| Sonoran Desert | <1,500 [249] |
| Arizona | <1,100 [121,240] |
| California | <1,500 [20,67,168,245] |
| Nevada | 200–1,200 [119] |
| Utah | 850–1,280 [257,265] |
White brittlebush is often common or dominant in communities on rocky slopes, hillsides, and upper bajadas [35,178,197,200,228]. It is drought tolerant [202,213], has an affinity for warm, dry sites [108,122,130,143,250], and occurs on soils with low infiltration rates [53] or on rocky, shallow soils [86]. It frequently occurs on disturbed sites [33,142,161,166,178].
Populations are often relatively dense on south-facing slopes [58,79,154,177,178,194,238], but it is also common or dominant on north-facing slopes, depending on elevation, latitude, and landform [139,182]. In California, white brittlebush cover was eight times higher on warm, dry south-facing slopes than north-facing slopes at 820 m, but at 520 m it was six times higher on warm, dry north-facing slopes than south-facing slopes [182]. White brittlebush occurs on the upper portions of bajadas in its southern extent and on lower portions in its northern extent [44,53,74].
White brittlebush generally grows in coarse-textured soils [228]. It grows in colluvium, alluvium, and other unconsolidated material [111,243], including gravelly bajada substrates [95,148,250]. It often grows in soils with granitic, basaltic, or other volcanic parent materials [53,143,146,194,221,243,270]. Soils are often shallow, rocky, and poorly developed [197]. White brittlebush may grow in areas where root-restricting soil layers, such as bedrock or caliche, occur near the soil surface [58,85], although in some cases a caliche layer may prevent it from becoming abundant [228]. In the Lake Mead National Recreation Area, the root-restricting layer averaged 21 cm deep in white brittlebush communities, which was the shallowest among 17 plant communities [3].
Although white brittlebush is flood-intolerant, it may be present to dominant in wash communities [3,69]. It is most prevalent in transition zones where flash flooding no longer occurs because it generally does not survive breakage of stems when flooding does occur [84].
Plant Communities
Figure 3—A stand dominated by white brittlebush in southern California.
White brittlebush is an important component in plant communities in the mountains, foothills, bajadas, and plains of the southwestern United States and northern Mexico [3,27,69,146,221,243,250,272], and often forms nearly monospecific stands throughout much of its range [86,151,214], particularly following disturbance [69] (see Successional Status).
White brittlebush is a common component in desert scrub [178,250,273], coastal sage scrub [13,130,162,270], chaparral [69,126,173], and frequently disturbed plant communities such as those in washes and arroyos [122,161,173]. More information about these communities is provided in the following sections, which are organized by NatureServe (2024) [173] and USNVC (2021) [93] groups. They are described by NatureServe (2024) [173] and others (as cited) and correspond to Biophysical Settings (BpSs) described by LANDFIRE (2020) [138]. White brittlebush also occurs in semidesert grasslands [29,118,272] and oak woodlands [29].
Mojave-Sonoran Bajada & Valley Desert Scrub Group
BpS: Sonora-Mojave Creosotebush – White Bursage Desert Scrub (10870)
White brittlebush is among several species of shrubs that may be present to codominant with creosotebush and/or white bursage (a.k.a. burrobush) in valleys, bajadas, plains, and hills in the Mojave and Sonoran deserts [53,106,108,109,122,173,197,217,247].
White brittlebush is dominant in the following alliances in this group [173]:
- Brittlebush Desert Scrub Alliance
- Creosotebush – Ocotillo Upper Bajada & Rock Outcrop Desert Scrub Alliance
It is present but not dominant in three other alliances in this group:
- Creosotebush – Burrobush Bajada & Valley Desert Scrub Alliance
- Burrobush Desert Dwarf Scrub Alliance
- Teddy-bear Cholla Cacti Scrub Alliance
Sonoran Paloverde-Mixed Cacti Desert Scrub Group
Figure 4—White brittlebush (with yellow flowers) dominates the understory of a Sonoran paloverde-mixed cacti desert scrub community in Arizona.
BpS: Sonoran Paloverde-Mixed Cacti Desert Scrub (11090)
In the Sonoran Desert, white brittlebush is a common understory component of mixed cacti desert scrub communities [80,178,192,217,233], particularly on south-facing slopes [143,177,230]. Saguaro often occurs as a sparse emergent layer with or without a short tree or succulent layer dominated by yellow paloverde [173]. This group includes scrub dominated by desert ironwood, triangle bur ragweed, and teddybear cholla [173]. Mesquite, crucifixion thorn, jojoba, and ocotillo are common components, among others [29,74,118,135,143,176,177,221,248]. White brittlebush is common to dominant on upper bajadas, plains, and hillslopes in these communities [106,166,221,228]. These communities are susceptible to invasion by the nonnative perennial grass, buffelgrass [5,187,274].
White brittlebush is a named dominant in one provisional alliance in this group [173]:
- Desert-ironwood – Brittlebush Desert Scrub Alliance
It is common to subdominant in the shrub layer of two other alliances in this group:
- Saguaro – Yellow Paloverde – Velvet Mesquite Desert Scrub Alliance
- Jojoba – Crucifixion-thorn – Eastern Mojave Buckwheat Desert Scrub Alliance
Warm Semi-desert Shrub & Herb Dry Wash & Colluvial Slope Group
BpS: North American Warm Desert Riparian Systems – Stringers (11552)
White brittlebush is one of a wide variety of woody species that dominate intermittently flooded washes and arroyos that intersect bajadas, mesas, plains, and basin floors throughout the warm deserts of the western United States and northwestern Mexico [173].
White brittlebush is present to codominant in four alliances in this group [173]:
- Catclaw Acacia – Desert-lavender – Beloperone Desert Wash Scrub Alliance
- Mojave Rabbitbrush Mojave Desert Wash Scrub Alliance
- Burrobrush – Sweetbush Mojave-Sonoran Desert Wash Scrub Alliance
- Blue Paloverde – Desert-ironwood Desert Wash Scrub Alliance
Mojave Mid-Elevation Mixed Desert Scrub Group
BpS: Mojave Mid-Elevation Mixed Desert Scrub (10820)
At mid-elevations in the eastern and central Mojave Desert, this group represents the desert scrub communities in the zone between creosotebush ─ white bursage desert scrub and lower montane woodlands. Vegetation is dominated by drought-adapted evergreen shrubs, including Joshua tree, Mojave yucca, rough jointfir, Eastern Mojave buckwheat, and blackbrush. White brittlebush does not typically dominate communities in this group. It is present to common in five alliances [173]:
- Rough Joint-fir Scrub Alliance
- Parry’s Bear-grass – Sacahuista Scrub Alliance
- Mojave Yucca Scrub Alliance
- Eastern Mojave Buckwheat – Parish’s Goldeneye Desert Scrub Alliance
- Buckhorn Cholla/Big Galleta Scrub Alliance
North American Warm Semi-desert Cliff, Scree & Pavement Sparse Vegetation Group
This group occurs across the southwestern United States and northern Mexico and consists of barren and sparsely vegetated substrates from a variety of landscapes in the Chihuahuan, Sonoran, and Mojave deserts. White brittlebush occurs most often in rocky Mojave Desert stands in this group. It is present to common in three alliances in this group [173]:
- Desert-holly Scrub Shrubland Alliance
- Warm Semi-desert Ephemerally Vegetated Pavement Alliance
- Schott’s Pygmy-cedar – Bush Arrowleaf Shrubland Alliance
Central & Southern Californian Coastal Sage Scrub Group
BpS: Southern California Coastal Scrub (10920)
White brittlebush is a characteristic species in dry stands of coastal sage scrub [13,75,130,268], especially on steep, south-facing slopes and intermittently flooded channels [162,173]. It may form monospecific stands on east-, west-, and south-facing slopes [151]. This community occurs from central California to Baja California, Mexico, and is dominated by drought-deciduous shrubs but can have characteristic resprouting, deep-rooted, leathery-leaved shrubs [173]. Coastal sagebrush (a.k.a. California sagebrush) and Eastern Mojave buckwheat frequently occur in these stands [130]. White brittlebush occurs most frequently in inland or “Riversidean” coastal sage scrub [108,211].
Increasing fire frequency in this group is causing type conversion to nonnative and invasive annual grasslands [62,124]. Species capable of resprouting after multiple fires, such as laurel sumac, lemonade sumac, and coastal pricklypear, are increasing in abundance.
White brittlebush is present to subdominant in two alliances in this group [173]:
- Coastal Sagebrush – San Luis Purple Sage Mesic Scrub Alliance
- Xeric Coastal Sage Scrub Alliance
Californian Xeric Chaparral Group
BpSs: California Maritime Chaparral (10960); Southern California Dry-Mesic Chaparral (11100)
This group is characterized by drought- and fire-adapted shrublands occurring inland of coastal chaparral communities from Baja California to southern Oregon. This group is primarily composed of obligate seeders, such as chamise, ceanothus, manzanita, California flannelbush, and others. White brittlebush is codominant in one alliance in this group [173]:
- White Sage – Black Sage – Chamise Chaparral Alliance
Sonoran Sarcocaulescent Desert Scrub Group
BpS: Sonoran Granite Outcrop Desert Scrub (10900)
Vegetation communities in this group align with the Central Gulf Coast subdivision of the Sonoran Desert [173], where white brittlebush is one of the most abundant deciduous shrubs [221]. Stands are dominated by pachycereus, organpipe cactus, physicnut, and elephant tree [173,250]. White brittlebush is associated with stands of this group especially where it transitions between upper bajada and mountain slopes. White brittlebush is an associated species in the only alliance described in this group [173]:
- Elephant-tree – Physicnut – Organ Pipe Cactus Desert Scrub Alliance
Botanical Description
This description covers characteristics that may be relevant to fire ecology and is not meant for identification. Identification keys are available (e.g., [20,70,119,121,249]).
Aboveground Description
White brittlebush is a rounded, many-branched shrub. Brittle, aromatic stems arise from one or several gray, linear, woody trunks [20,67,70,168,257,261,265], and form an “umbrella”-like canopy [54]. Tomentose young stems become smooth and woody as they mature [20]. It generally grows from 0.3 to 1.0 m tall [119,121,257,265], but robust individuals may reach 1.5 m [20,67,70,76,133,249].
Leaves are silvery gray or white tomentose and borne in dense clusters on short petioles at branch tips [20,168,265]. Under moist conditions, white brittlebush produces drought-deciduous, relatively water inefficient leaves (“winter” leaves), which desiccate from lower portions of the branches up to the tips as summer drought progresses [72,234]. Under dry conditions, it produces long-lasting, drought-tolerant leaves [82,186], which generally persist through the hot, dry summer months (“summer” leaves) [58,71,72,179,218,219]. Under sufficiently hot and dry conditions, even the summer leaves are shed [140] (see Seasonal Development).
White brittlebush plants produce perfect flowers [196,225] that develop on tall, loose panicles [133] that extend above the leaf canopy [175]. Fertile disk flowers produce achenes when fertilized and are surrounded by 11 to 21 infertile ray florets 8 to 12 mm long [208]. Achenes lack pappi for long-distance dispersal [94].
Belowground Description
Figure 5—Taproot of a juvenile white brittlebush plant.
White brittlebush has a relatively shallow root system [3,159] comprised of a taproot, lateral roots, and filamentous roots [58]. The taproot may extend at least 90 cm deep, but its depth is often limited by impassable soil layers, such as caliche [58], which are common in white brittlebush communities [3] (see Site Characteristics). Lateral roots can extend 1.5 m or more from the taproot but are often shorter [58,181]. For example, in the northwestern Sonoran Desert, white brittlebush with canopies 29 cm across had maximum root depths from 24 to 36 cm and mean root depths from 10 to 13 cm. Lateral roots radiated up to 34 cm from the taproot. Mean root depth increased in plants growing next to competing plant species (n = 4 plants) relative to isolated plants [181]. At Tumamoc Hill in Arizona, the tap root of a 90-cm tall plant extended 55 cm vertically until it reached an impassable soil layer then extended an additional 50 cm or more horizontally. Lateral roots occurred only within the upper 30 cm of the soil surface, extending horizontally about 1.5 m from the taproot [58]. In Saddle Peak Hills, California, maximum rooting depth of white brittlebush on recently eroded wash edges was 80 cm [84].
Stand and Age Structure
White brittlebush forms monospecific stands throughout much of its range, including in the Mojave Desert [86], the Sonoran Desert [221], and coastal sage scrub [151]. White brittlebush communities typically have a sparse to open canopy layer [172], although they typically have higher densities of woody species than communities dominated by species of larger shrubs or small trees, such as creosotebush or paloverde [221,236]. White brittlebush plants often have a clumped distribution [83,214]. Two statistical measures of aggregation applied to white brittlebush populations in Arizona’s Estrella Mountains indicated high and moderate degrees of aggregation (clumping) among individual plants on the landscape. In contrast, white brittlebush had a random distribution at a site in the northwestern Sonoran Desert [79].
Establishment and mortality are episodic in many white brittlebush communities [32,36,86], which may result in relatively even-aged stands. Establishment is typically correlated with higher than average precipitation from cool-season or late summer rain events, especially when El Niño conditions occur after seedlings have germinated [32,41,86] (see Seedling Establishment and Mortality). Mortality is also episodic and appears to be associated with the transition between the warm phase of the Pacific Decadal Oscillation to the cold phase [86] (see Plant Growth and Mortality).
Raunkiaer Life Form
- Phanerophyte [71,205]
Seasonal Development
White brittlebush germinates most often after cool-season (October–April) or late summer (August–September) rains [25,133], from approximately August to April, depending on location [36,41,267]. For example, in the Sonoran Desert near Tumamoc Hill, seeds germinated after cool-season rain events from October to April [36]. In the Mojave Desert, seeds germinated after late summer monsoons in September and October [267]. In Joshua Tree National Park, white brittlebush seeds germinated in August and September after sporadic heavy summer rain [266].
White brittlebush grows new leaves throughout the cool season. Early in the cool season, when soil moisture is high and temperatures are low, it produces large winter leaves [58], which are capable of rapid photosynthesis [140]. As soil moisture decreases and temperatures rise, it produces summer leaves, which are smaller than winter leaves and have a greater density of reflective hairs, which reduces their photosynthetic efficiency and increases their water use efficiency. The large, water-inefficient winter leaves fall off as weather gets hotter and drier. The smaller, water-efficient summer leaves generally remain on the plant unless conditions become too hot and dry, at which point even they fall off and the plant goes dormant [87,140].
While white brittlebush can flower any month of the year (table 2), it typically flowers in winter or spring [74] after cool-season rains [39], but also in summer after enough monsoonal rains [35,58,86,267]. However, over 6 years at Tumamoc Hill it never flowered in summer, suggesting high temperatures may inhibit flowering. It flowered mainly from February through May, and also between October and January when there was enough cool-season rain [39]. Peak flowering typically occurs between the beginning of March and end of April in Arizona [39,157]. The span of the flowering season may be lengthened [157] or shortened [39] because of high or low soil moisture, respectively. Flowering timing and length of the flowering season may also be altered by temperature. For example, white brittlebush planted in grassy areas of the Phoenix metropolitan area (e.g., schoolyards with grass, herbaceous, and tree cover) flowered later, longer, and at a higher percentage than those planted in surrounding desert areas. Although commercial urban land cover types are usually hotter (≈2 °C warmer) than surrounding desert areas, urban grassy areas in Phoenix have higher plant cover and are about 2 °C cooler than surrounding desert areas [175].
For plants that flower in late winter to spring, seeds mature [164] and disperse between May and July but do not germinate until after enough cool-season precipitation occurs [39]. At Tumamoc Hill, seeds were dispersed in May but did not germinate until the following February, March, and April following rain events [36].
| Area | Event |
|---|---|
| Intermountain West | Flowers mostly March to May, sometimes December to February at southern extent [70] |
| Sonoran and Mojave deserts | Flowers March to May [133]. Fruiting May to July [133,163]. |
| Sonoran Desert | Flowers mostly February to May. May flower October to January, and rarely as early as August [249]. |
| Arizona | Flowers November to May [121] |
| California | Flowers January to June and August to September [20]. Other publications state it flowers March to May [67,168] or April to May [245]. |
| California, Sonoran Desert | Leaf growth January to February. Peak number of leaves March to April. Loss of deciduous (winter) leaves in May [180]. |
| Nevada | Flowers March to May [119] |
| Nevada, Nevada Test Site | Flowers April [27] |
| Utah | Flowers February to May and August to September [257] |
Regeneration Processes
White brittlebush primarily reproduces from seed [54]; however, it is sometimes described as a facultative seeder [125] or resprouter [88,271,275]. It is generally a poor resprouter [54,67,269].
Pollination and Breeding System
Plants produce perfect flowers [225] and require cross-pollination [145]. They are pollinated by beetles, bees, flies, moths, butterflies, and other insects [137,145]. Observations of white brittlebush pollinators between Blythe and Needles, California, indicated that one species of soft-winged flower beetle, Tanaops abdominalis, occurred on white brittlebush 10 times more often than all other species combined [137].
Seed Production and Predation
Plants may flower and set seed in their first year. For example, white brittlebush seedlings planted at Rancho Santa Ana Botanical Gardens flowered and set seed within a year [91] and in coastal sage scrub, postfire resprouts flowered and set seed within the first year after fire [269].
White brittlebush flowers after at least 20 mm of cool-season rain followed by about 415 degree days (i.e., the sum of mean daily temperatures above 10 °C to equal 415). An additional 30 mm of precipitation after the trigger may be required to maintain flower production [39], or there may be minimum required precipitation from the previous spring or fall [175]. Flower production is higher during wet years. In Death Valley National Park, 0% to 100% of white brittlebush flowered each year over 37 years, with flowering greater in years with greater total winter-spring precipitation [86]. High temperatures may truncate the flowering period [39,175].
There is little quantified information available regarding white brittlebush seed production, although Padgett et al. (1999) stated that “a typical Encelia farinosa shrub produces tens of thousands of seeds” [190]. In southwestern Utah, white brittlebush plants produced between 8 and 77 flower heads/plant depending on the year [145]. In the Sonoran Desert, intraspecific competition for water lowered seed (i.e., achene) production. White brittlebush plants without neighbors were more likely to produce flowers and seeds compared with those with neighbors (79.8% versus 29.6%). White brittlebush plants without neighbors produced 87 seeds/flower head and 329 seeds/twig, while plants with neighbors produced 40 seeds/flower head and 98 seeds/twig [83]. Burning may reduce flowering and fruiting density [145]. For more information, see Plant Response to Fire.
Birds, rodents, and ants consume white brittlebush seeds [35,57,165,232,239], although some ant species may avoid them [46].
Seed Dispersal
Seeds lack pappi [94], are relatively light, and are primarily dispersed short distances by gravity and wind [54,165,269] but also likely by animals [165,232]. Seeds are generally concentrated under white brittlebush canopies [36]. Strong winds are likely needed to disperse seeds long distances [165]. Rodents eat seeds and may cache them, which could result in some seed dispersal [232]. Birds may scatter seeds as they pluck seeds from ripe seed heads [165].
Seed Banking
One study indicates that white brittlebush has at least a short-term persistent, soil-stored seed bank [36] but more information is needed on the longevity of soil-stored seeds. At Tumamoc Hill, 43% of the seeds from the previous year remained in the soil in April, 1 month before seed dispersed, despite 3 germination events during the previous growing season. This suggests white brittlebush seeds can remain viable in the soil seed bank at least 11 months after dispersal [36].
Two studies in the Sonoran Desert reported high densities of viable white brittlebush seeds in the soil seed bank [36,112]. In paloverde-mixed cacti desert scrub in central Arizona, total white brittlebush seed density was 5,895 seeds/m2 in September [112]. At Tumamoc Hill, total white brittlebush seed density was 313 seeds/m2 in April [36]. Neither study provided information on white brittlebush abundance in the plant communities.
Buffelgrass plants may help retain white brittlebush seeds in the soil seed bank. In Saguaro National Park, white brittlebush seeds were twice as dense in soil samples collected under white brittlebush canopies in buffelgrass-invaded patches (1,015 seeds/m2) as under white brittlebush canopies in uninvaded patches (427 seeds/m2), even though white brittlebush cover was six-fold lower in buffelgrass-invaded patches. Higher white brittlebush seed densities in soils of buffelgrass-invaded patches may be due to preinvasion seed retention and persistence or post invasion seed accumulation due to the physical vegetation structure of buffelgrass. White brittlebush seed viability was not tested [5].
Seed viability as high as 55% has been reported [190], but germination is often low (1%–5%) when white brittlebush seeds are sown in restoration projects or germinated under laboratory conditions [120,190]. A large portion of seeds (about 40%–65%) are empty or contain dead embryos [190].
Germination
Throughout its range, mass germination events of white brittlebush seeds are relatively common after large rainfall events [25,39,41,86,277]. At Tumamoc Hill, seeds germinated between 9 and 24 days after 8 of 17 rain events of at least 19 mm in winter or spring. No seeds germinated when temperatures fell below freezing in the 9 days after such events [36]. In Death Valley National Park, germination events occurred more often in El Niño years than non-El Niño years and depended on cumulative winter and spring precipitation. Germination events (25+ seedlings/480 m2) occurred in 18 of 37 years, and of these only 11 were large germination events (500+ seedlings/480 m2) [86].
While white brittlebush can germinate prolifically after cool-season or late summer rain events (see Seasonal Development), its germination in greenhouse and laboratory settings is often low (≤20%) [91,120,164,190]. However, one study found germination as high as 76%. Szarek et al. (1996) recorded peak germination capacity at 76% after 13 days of misting. Germination was also high after 1 day of misting for seeds collected from the Sonoran Desert when diurnal temperature variation was low and temperature moderate (53% with 15 °C/15 °C daytime/nighttime temperatures compared with 6% with 21 °C/9 °C daytime/nighttime temperatures). Germination increased by 10% to 20% when seeds from each treatment were misted for 2 days. No germination occurred at constant temperatures <5 °C or >25 °C. However, 42% of seeds germinated when they were maintained at 27 °C for 7 days followed by a germination period at 15 °C for 7 days [241]. A similar study found germination of seeds collected from the Mojave Desert was highest (14%) at 40 °C compared with cooler temperatures. Germination did not exceed 2% for temperatures between 5 and 30°C [120]. Mean germination time is about 7 days under greenhouse or laboratory conditions [164,241].
Variables that appear to affect germination in greenhouse and laboratory settings include moisture levels, storage times and temperatures, and seed treatments. For example, misting seeds to mimic conditions where seeds are soaked by heavy rains had higher germination capacities compared with unmisted seeds. Seeds that were misted reached maximum germination (76%) after 13 days [241]. White brittlebush seeds stored for 1 year had a greater germination capacity than seeds recently collected. However, germination decreased from 6% to 2% after 50 days of storage at 4 °C followed by a germination period at 15 °C, suggesting that cool conditions (i.e., overwintering) may not improve germination [241]. Another laboratory study found seeds stored for 6 months at cool temperatures (5–10 C°) had two to three times greater germination than seeds stored at room temperature suggesting that cool, dry conditions may help seeds retain viability in storage [190]. Application of gibberellic acid increased germination of viable seeds from 2% to 11% and from 6% to 17% for seeds from three populations [190], suggesting that a small portion of seeds may be dormant at maturity. No data were available on light requirements of white brittlebush seeds, although Padgett et al. (1999) stated that they do not seem to have “any specific light requirements” [190].
Seedling Establishment and Mortality
White brittlebush seedling establishment is typically low and is positively associated with precipitation [41,86]. Episodic mass germination events often occur after sufficient rainfall, but conditions for establishment are less common. In Death Valley National Park, seedlings established (i.e., survived to 1 year old) in 18 of 37 years, but in only 6 of those years did 20+ seedlings establish in the 480 m2 population monitoring plot. The number of seedlings that established increased with increasing cumulative spring and winter precipitation [86]. On Tumamoc Hill, 18.7% of white brittlebush seedlings established—an “unusually high percentage”—during a year when precipitation was just over 200% of average, including some rain in the typically hot and dry months of late spring [41]. Over the first 5 postfire years in sage scrub, seedling recruitment averaged 21,100 seedlings/ha on eight sites. Ten percent, 19%, 18%, 0%, and 53% of total seedlings recruited in postfire years 1 through 5, respectively. Precipitation was approximately 80% of average in the first, third, and fourth years, 175% of average in the second year, and 250% of average in the fifth year [125].
High numbers of white brittlebush seedlings may establish after fire [54,125,141,160], sometimes resulting in nearly monotypic stands in burned areas [61,236]. In sage scrub, seven white brittlebush seedlings established for each parent plant during the first year after fire [125]. For more information, see Plant Response to Fire.
In the Mojave Desert, one study found rodent herbivory reduced survival of white brittlebush seedlings in burned and unburned plots. Rodent herbivory can strongly suppress plant recruitment and influence plant composition and succession over time [216].
Plant Growth and Mortality
White brittlebush individuals are relatively short-lived, with an average lifespan of 16 years in the northern Sonoran Desert [37]; however, the maximum lifespan can be much longer. Early estimates indicated that the maximum lifespan was 32 years or less [99,220]. However, repeat photography in the Grand Canyon, Arizona, showed at least one individual survived for at least 72 years in an area protected from livestock grazing, more than twice previous estimates of maximum longevity. The authors suggest that their reports of longer lifespans for several species may be due to the exclusion of grazing [43]. In Death Valley National Park, the upper age limit was estimated at 66 years [86]. Median lifespan is apparently much shorter than maximum lifespan [32,37].
Mortality rates of white brittlebush plants are high, with highest mortality for young plants. Of the white brittlebush individuals that survived for at least 1 year during a 37-year study in Death Valley National Park, 50% died within 5 years, 90% died within 24 years, and 98% died within 37 years. Many of the 2% of individuals surviving the entire study were mature (i.e., flowering and fruiting) at the time the study began [86]. At Tumamoc Hill, 83% of 369 white brittlebush individuals died before they were 8 years old during a 72-year study; the oldest individual was 32 years old [99]. In a wet year at Tumamoc Hill, 18.1% of 816 seedlings survived 1 year after germination. Survival dropped to 0.3% during their second and third year, and only 0.1% of the initial cohort survived 4 years after germination [41]. Small plants are more likely to die than larger, more robust plants before they are 4 years old [32].
Mortality of established white brittlebush plants occurs as episodic events that are associated with large-scale climate shifts. Over a 37-year study in Death Valley National Park, six mortality events occurred. Mortality events correlated with a shift from the warm phase of the Pacific Decadal Oscillation to the cold phase, with mortality lagging 1 year behind the shift [86]. Smaller-scale weather events, particularly unusually frequent or hard freezes, may also cause mortality events [55].
High soil moisture availability may improve white brittlebush growth and survival, but study results vary. Over 37 years, spring canopy leaf cover varied from <5% in drought years to nearly 100% in years with the wettest springs [86]. In Saguaro National Park, 93% of 94 white brittlebush individuals planted and watered on roadside berms and decommissioned vehicle pullouts survived the first year, despite below-average precipitation. Plants were watered six times during the summer months and protected from herbivory with cages. Proximity to a road may have been an additional contributing factor to the high survival rate [9], as precipitation can accumulate and increase soil moisture next to roads [52]. In contrast, in southern Nevada, seeding and outplanting methods that included watering and protection from rodents were ineffective in establishing white brittlebush. All 40 outplants died within 3 years and no white brittlebush seedlings were observed in any seeded areas over 3 years. Precipitation was below average during the first 2 years of the study, which may have influenced seed and outplant response even though irrigation was provided for some treatment types [7]. Another study found no difference between white brittlebush survival within (86%) and outside (78%) of catchment basins that were comprised of two berms funneling water to the point of a “v”, encompassing 4 to 25 m2. White brittlebush was one of only three species that survived nearly as well outside of catchments as inside [81].
When accompanied with soil water, high soil temperatures during the growing season may aid white brittlebush growth. Optimal root development occurs at 25 °C, and warm temperatures following cool-season precipitation stimulates new root growth. In the northwestern Sonoran Desert, total root surface area and rooting depth was greater on south-facing slopes (291 cm2 and 13.4 cm) than north-facing slopes (225 cm2 and 10.6 cm). South-facing slopes had warmer soils, which favored new root growth when soil water became available during a wet cool-season period with 169 mm of rainfall. The area of new roots was 2.7 times greater on south- than north-facing slopes. Optimal root growth on warm south-facing slopes may promote larger plants; mean canopy leaf area was 57% higher for plants on south-facing slopes than on north-facing slopes [182].
Vegetative Reproduction and Regeneration
White brittlebush can resprout from the root crown after top-kill [271]. It is most often described as a poor resprouter [54,151,160,210,211,269,274], but some sources describe it as a moderate to strong resprouter [88,271,275]. Type and severity of disturbance may impact resprouting ability. White brittlebush apparently does not resprout after stem breakage, which causes mortality after flooding in washes and limits persistence of white brittlebush in active washes [85].
For more information regarding resprouting after fire see Immediate Fire Effects.
Successional Status
White brittlebush is often considered a pioneer species [161,210,223] but is also prevalent in later stages of succession [42,118]. In desert plains grassland communities, white brittlebush follows initial annual colonizers and is succeeded by perennial grasses [272].
While important in undisturbed communities, it is generally more prevalent in disturbed communities. In a review of vegetation reestablishment following a variety of disturbances in the Mojave and Sonoran deserts, Abella (2010) found that white brittlebush had approximately 50% higher occurrence and 60% higher relative abundance in reestablished disturbed vegetation than in undisturbed vegetation, on average [4]. White brittlebush may exhibit different postdisturbance responses based on weather patterns [248], predisturbance vegetation community [12,199], and other site characteristics.
White brittlebush may become more prevalent in Sonoran desert scrub communities after disturbance [166,178], including fire [61,199,210]. After fire, high numbers of white brittlebush seedlings establish [54,141,160,249], sometimes resulting in nearly monotypic stands in burned areas [61,236]. White brittlebush may become dominant in disturbed areas only when weather or site conditions are suitable for establishment [31,159,236]. Where their ranges overlap, white brittlebush often replaces creosotebush as the dominant shrub after severe or repeated fires [17,54,142,213,234].
White brittlebush may dominate communities in early or intermediate stages of succession after soil disturbance. Thirty-eight years after topsoil was cleared from a pipeline right-of-way, white brittlebush codominated with other early successional species, including burrobrush and brownplume wirelettuce. An adjacent 8-year-old pipeline right-of-way was similarly cleared, and topsoil was reapplied to some sections after construction. Creosotebush dominated plots with reapplied topsoil, and white brittlebush and white bursage dominated plots where topsoil was not reapplied. White brittlebush was not present on untreated control or restored plots. Treatments on restored plots included raking the soil surface, applying artificial desert varnish, and planting native species. White brittlebush was not planted [8]. In Grand Canyon National Park, white brittlebush was a significant component of vegetation cover and density on intermediate-aged debris flow terraces (32–285 years old; 2.4%–14.1% cover, 1,000–5,000 plants/ha) but did not contribute to vegetation cover or density on young (5–28 years old) or old (485–3,100 years old) terraces [42].
Immediate Fire Effects
White brittlebush is easily damaged by fire and is often killed or top-killed [54,151,210,274]. Seedlings are more likely to be killed by fire than mature plants [115]. Although mature white brittlebush plants are susceptible to charring and scorching, they are rarely entirely consumed by low-intensity fires. After relatively low-intensity fires in the western Sonoran Desert, 78% of mature white brittlebush plants were scorched, 20% were burned, and the remaining 2% were unburned. Scorched plants retained all or most of their branches and some dried leaves—only leaves and branches closest to the ground burned. Most leaves and small branches of burned plants were consumed by fire [54,160]. In Sonoran paloverde-cactus mixed desert scrub, from 8% to 15% of white brittlebush biomass was consumed during two June wildfires, and from 61% to 79% of photosynthetic tissue was killed [210] (see Plant Response to Fire). No information was available regarding the effects of heating or burning on white brittlebush seeds.
White brittlebush is often described as a poor resprouter after fire [54,151,160,210,211,269,274], but some sources describe it as a moderate to strong resprouter [88,271,275]. Its ability to resprout varies widely and is influenced by fire intensity [271], aspect [151] (see Plant Response to Fire: Central & Southern Californian Coastal Sage Scrub), and plant age [211], although research is limited. Higher fire intensities likely reduce its resprouting ability. The maximum rate of heat release above which root crowns fail to resprout is <170 kcal/sec/m2 [269]. In coastal sage scrub, white brittlebush showed “strong resprouting” after a “relatively cool” October fire, but only a single individual resprouted after a higher-intensity fire in July (maximum rate of heat release of 40 kcal/sec/m2) [271]. In the western Sonoran Desert, 0% (0 of 20) of completely burned individuals resprouted and 6.4% (5 of 78) of scorched individuals resprouted. Fire intensity was likely higher around completely burned individuals [54]. Five months after a June fire in Sonoran paloverde-mixed cactus desert scrub, white brittlebush mortality was higher where nonnative invasive perennial buffelgrass and crimson fountaingrass carried the fire compared to where winter annuals carried the fire; only 14% (3 of 21) of white brittlebush plants resprouted where the fire was carried by annual grasses and 0% resprouted where the fire was carried by perennial grasses. Invasive perennial grasses provided higher fuel loads than winter annuals, resulting in higher fire intensity [274]. On the other hand, 1 year after a May fire in the same vegetation type, 93% of white brittlebush resprouted, although details of fire intensity and fuel load were not provided [275]. Young, mature plants may be better able to resprout after fire than older, woodier plants [211].
Postfire Regeneration Strategy
- Small shrub, a sprouting root crown
- Ground residual colonizer (on site, initial community)
- Initial off-site colonizer (off site, initial community) [237]
Fire Adaptations
White brittlebush establishes from seeds from on- and off-site seed sources after fire [54,61,141]. Seeds are primarily dispersed short distances [54,165,269] by gravity, wind, and animals (see Seed Dispersal) and seedlings readily establish in postfire environments [12,54,141]. White brittlebush may also establish after fire from undamaged seeds in the soil seed bank, but there is little research on the effects of fire on white brittlebush seed banks (see Plant Response to Fire).
White brittlebush may resprout after fire [271,275]. Resprouting ability is highly variable and influenced by aspect [151], fire intensity [115,271], plant age [211], or other factors [127] (See Immediate Fire Effects).
Plant Response to Fire
White brittlebush is often more abundant and more prominent in recently burned than long unburned areas [110,142,160,210]. After fire, white brittlebush populations most often increase [12,24,54,61,141,153,199,222,234], but they may remain stable or decrease [31,115,125,151,269,275]. Regeneration is primarily from seeds [54], although resprouting plants contribute to postfire recovery to varying degrees, from comprising most postfire regeneration [275] to comprising very little [54,210,236]. When white brittlebush establishes in postfire communities, it may dominate in a stable state, especially in creosotebush scrub [234,240]. Alternatively, populations may decrease after fire if buffelgrass precludes it from establishing, if fire intervals are very short [12,115], or if freezes or other mortality events occur [55].
After fire, white brittlebush seedlings often establish in high numbers [54,141,160]—particularly in years with above average precipitation—sometimes resulting in nearly monotypic stands [61,236]. White brittlebush seedling establishment is influenced by several factors, including postfire weather [61,125,240], site characteristics (e.g., soil temperature and aspect) [12,151], fire characteristics (e.g., uniformity) [141], and its presence in the prefire plant community [199,234]. Seedling establishment and regeneration may be from on- or off-site seed sources but because seeds are primarily dispersed short distances (see Seed Dispersal), on-site seed sources—such as from the soil seed bank or from unburned or resprouting plants—may be relatively more important [269]; however, data are very limited. In some cases, such as in large, burned areas without on-site seed sources, white brittlebush may take many years to recover populations, only doing so after several generations of white brittlebush have spread inwards towards the middle of a burn [211,234,269].
White brittlebush may dominate repeatedly burned areas [235,236] although results vary [12]. For example, after two fires 5 years apart in Sonoran paloverde-mixed cacti desert scrub, white brittlebush cover and density increased from prefire cover and density (1% and 25 plants/ha) for at least 19 years after the second fire (15% and 1,365 plants/ha) [240]. In creosotebush scrub in the western Colorado Desert, white brittlebush comprised most perennial cover in once-burned and twice-burned plots, with 17.9% and 16.4% cover respectively, but did not occur in unburned stands. Once-burned plots burned 20 years before sampling, and twice-burned plots burned 20 and 3 years before sampling [235]. In contrast, in paloverde-mixed cacti desert scrub, white brittlebush cover and density were lower in plots that were burned 4 times in 30 years (0.5% and 175 plants/ha) than in unburned plots (0.9% and 253 plants/ha), although differences were not significant [12]. White brittlebush may be sparse or absent after repeated burns if the fire interval is so short that it prevents newly established white brittlebush plants from reaching reproductive maturity or if fire intensity is so high that it kills all plants and seeds in the area [115].
Limited data suggest that burning may reduce white brittlebush flower and fruit density. Individual plants produced similar numbers of flowers and fruits in plots located along burned edges (45 flowers/plant and 17 fruits/plant) and unburned plots (37 flowers/plant and 20 fruits/plant) 7 to 9 years after wildfire in the Mojave Desert in southwestern Utah. However, on a ground area basis, there were fewer flowers and fruits in burned edge (0.18 flowers/m2 and 0.09 fruits/m2) than unburned (6.3 flowers/m2 and 3.1 fruits/m2) plots. While not significant, cover and density of white brittlebush trended lower in burned interior plots than in the other plots, perhaps partially explaining why no flowers or fruits were produced in burned interior plots [145].
Mojave-Sonoran Bajada & Valley Desert Scrub
In recently burned creosotebush scrub, white brittlebush density and cover may be similar or lower than on unburned sites but may be substantially higher on sites that burned 12 to 28 years prior [160,185,234]. While longer term studies that quantify white brittlebush abundance are not available, observations suggest that white brittlebush may dominate an alternative stable vegetation community indefinitely [54,142,234,236]. Along a 2- to 28-year-old chronosequence at the northwestern extent of the Colorado Desert, white brittlebush dominated sites that burned 12 or more years prior. White brittlebush had higher cover and density in stands 12, 20, 24, and 28 years after fire compared to unburned reference stands (table 3). In the oldest stand, there was still hardly any creosotebush present (cover was <0.1%) [234]. On alluvial fans at the base of the San Jacinto Mountains, white brittlebush comprised 18% cover of 19% total vegetation cover 14 years after fire. No cover by creosotebush and white bursage and no creosotebush seedlings suggests that white brittlebush will remain dominant in that community unless conditions become more suitable for establishment of creosotebush and white bursage [160]. Repeated fires are likely to reinforce dominance by white brittlebush and facilitate monocultures because white brittlebush is likely to recover more quickly than associated shrubs, which are likely to be killed by subsequent fires [235,236].
| Time since fire (years) | Cover – burned | Cover – unburned | Density – burned | Density – unburned |
|---|---|---|---|---|
| 2 | 0.1 | 0 | 1.5 | 0.8 |
| 2 | <0.1 | 0.4 | 3.8 | 5.2 |
| 9 | 0 | 0 | 0 | 0 |
| 12 | 9.8* | 2.5 | 94.5* | 26.3 |
| 20 | 27.1* | 1.3 | 281* | 9.8 |
| 24 | 6.6 | 1.7 | 60.7* | 12.5 |
| 28 | 12.6* | 0.6 | 82.5* | 3.5 |
White brittlebush seedling establishment may be higher in burned than unburned creosotebush scrub communities for at least 5 years after fire [54], especially after repeated fires [236]. For example, in creosotebush-white bursage scrub in the western Sonoran Desert, mean white brittlebush seedling density in burned plots in the first postfire growing season (1,460 seedlings/ha) was more than 16 times that in unburned plots. In the second and third postfire growing seasons, white brittlebush seedling density in burned plots (7,010 seedlings/ha) was 1.2 times the seedling density in unburned plots. Heavy rains occurred in the third winter after fire, and white brittlebush seeds germinated in burned and unburned plots in high numbers. By three to five growing seasons after fire, white brittlebush seedling cover was <1% in unburned plots and nearly 3% in burned plots [54].
Sonoran Paloverde-Mixed Cacti Desert Scrub
Information on the fire ecology of white brittlebush in paloverde-mixed cacti desert scrub comes from studies in southern Arizona. These studies found white brittlebush density and cover generally decrease immediately after fire due to white brittlebush being killed or top-killed [61], then may increase up to 14 years since fire [199], 25 years since fire [24], or 29 years since fire [240]. It is possible that density and cover may continue to increase, but longer-term studies are lacking. For example, after a June prescribed fire, white brittlebush density decreased from 82 plants/ha before fire to 14 plants/ha 4 days after fire, then increased 762% over preburn levels to 707 plants/ha 9 months after fire, apparently due to high germination in burned areas [61]. In Saguaro National Park, mean white brittlebush density and cover increased relative to prefire levels for at least 29 years in once-burned plots and at least 19 years in twice-burned plots. Mean white brittlebush density and cover also increased in unburned plots, possibly due to weather conducive to seedling establishment or grazing cessation [240] (fig. 6). In a chronosequence study, white brittlebush cover was higher in 10-year-old (21%) and 14-year-old (29%) burned plots than in 26-year-old burned plots (0.5%) and unburned control plots (2.5%). The author notes that because the distribution of white brittlebush is patchy, cover values in chronosequence plots may not represent time-since-fire trends [199].
Figure 6—Grouped bar charts display A) mean canopy cover (%) and B) mean density (plants/ha) of white brittlebush with time since fire in 10 plots in Saguaro National Park. Plots were either burned once in 1989 (light green bars), burned twice in 1994 and 1999 (medium green bars), or unburned by recent fires (dark green bars). Plots were surveyed in 1976 (before fire), 2007, and 2018. Plots unburned by recent fires were unburned for at least 31 years when surveyed in 2007 and for at least 42 years when surveyed in 2018. Livestock grazing ceased in 1978. Charts created from data in Appendix S6 and S7 in Summers et al. (2021) [240].
White brittlebush may gain prominence in paloverde-mixed cacti desert scrub after fire. For example, 19 months after a June 1974 wildfire, burned plots were dominated by triangle bur ragweed, and white brittlebush provided the second greatest cover of any species, with 11.6% relative cover and 0.52% absolute cover. Unburned plots were dominated by redstem stork’s bill and red brome. White brittlebush provided the tenth greatest cover of any species in unburned plots, with 0.5% relative cover and 0.13% absolute cover [153]. On another site, white brittlebush made the greatest relative increase in density of any species 3 to 4 years after two wildfires [210].
Repeated fires may allow white brittlebush to maintain prominence in Sonoran paloverde-mixed cacti desert scrub for long periods. For example, in an area burned by five fires over 43 years (mean fire interval: 8.6 years), white brittlebush had the second-highest importance value, with 62% relative cover and 4% relative density. In the unburned control plot, it had the fifth-highest importance value, with 4% relative cover and 7% relative density [199].
White brittlebush’s response to fire in paloverde-mixed cacti desert scrub may be influenced by soil temperature [182], aspect [12], fire characteristics [141], and postfire weather [61,141,240]. A chronosequence of four sites representing 5- to 21-year-old fires, found that white brittlebush cover and density were higher in 17-year-old burned plots but were lower in 21-year-old burned plots, relative to their respective unburned control plots (table 4), a result attributed to soil temperature and aspect of the burned sites, and possibly fire characteristics. Hyperthermic soils (mean annual soil temperature: ≥22 °C) in 17-year-old burned plots may have favored postfire dominance by white brittlebush compared with thermic soils (mean annual soil temperature: 15–22 °C) in other burned plots. In the 21-year-old burned plots, dominant vegetation was similar to that in adjacent unburned plots possibly because the community was better able to recover due to a more mesic northeastern aspect. Alternatively, the researchers speculated that fire intensity may have been lower or the fire patchier in the 21-year-old burned plots [12] (table 4).
| Fire (years since fire) | Dominant vegetation – burned | Dominant vegetation – unburned | Cover – burned | Cover – unburned | Density – burned | Density– unburned |
|---|---|---|---|---|---|---|
| River Fire (5) | Triangle bur ragweed | Yellow paloverde | 0.5* | 0 | 198* | 53 |
| Vista Fire (7) | Whitethorn acacia | Yellow paloverde | 0.2 | 0.1 | 28 | 33 |
| Massacre Fire (17) | White brittlebush | Triangle bur ragweed | 9.6* | 0 | 2,252* | 0 |
| Siphon Fire (21) | Triangle bur ragweed | Triangle bur ragweed | 0.0* | 0.2 | 25* | 53 |
Below-average precipitation during the first few postfire growing seasons may contribute to low or reduced white brittlebush density in burned areas [61,141]. For example, white brittlebush density averaged 707 plants/ha 9 months after a June prescribed burn, consisting mostly of seedlings, but only 63 plants/ha 13 months after a May wildfire in an adjacent area burned the previous year. Cool-season precipitation the growing season following the May wildfire was 58% of the precipitation that fell the growing season following the June prescribed fire, which likely was less conducive for germination and establishment of white brittlebush in the area burned by wildfire [61]. Following a June prescribed fire in 1983, white brittlebush density increased from 12 to 29 months after fire (1,022 plants/ha to 2,872 plants/ha), then declined slightly 35 months after fire (2,265 plants/ha), a result attributed to a hot, dry growing season [141].
In paloverde-mixed cacti desert scrub, resprouting white brittlebush plants contribute to postfire recovery to varying degrees, from comprising most postfire regeneration [275] to comprising very little [210], but information is lacking regarding the specific factors contributing to such variation. For example, 1 year after the May 1993 Vista View Fire, total plant cover was 15% and white brittlebush density averaged 54 plants/ha on burned areas. Most (93%) white brittlebush plants were resprouts. In unburned control plots, total plant cover was 41% and white brittlebush density averaged 60 plants/ha. Details of plant damage and fire intensity were not reported [275]. In contrast, 4 years after a June wildfire, 27 individuals occurred in burned areas and none were resprouts. During the fire, 15% of white brittlebush biomass was consumed and 79% of photosynthetic tissue was killed. Three years after a second June wildfire in the same area, 133 individuals occurred in burned areas, only one of which was a resprout. During that fire, 8% of white brittlebush biomass was consumed and 61% of photosynthetic tissue of white brittlebush was killed [210].
Buffelgrass Communities
Large areas of paloverde-mixed cacti desert scrub have been converted to buffelgrass pastures for cattle grazing in Sonora, Mexico [96]. White brittlebush may be reduced by fire in these communities if fine fuels are sufficiently dense, providing higher fuel loads [115]. Invasive perennial grasses burn hotter than invasive annual grasses due to higher fuel loads [274].
Fires carried by buffelgrass and other invasive perennial grasses are more likely to kill white brittlebush than fires carried by winter annuals. During the June Bighorn Fire in Catalina State Park, Arizona, white brittlebush, paloverde, and saguaro were burned more severely where the fire was fueled by buffelgrass and crimson fountaingrass compared to areas where it was fueled by winter annuals. Five months after the fire, white brittlebush showed little evidence of recovery. In unburned plots, it had about 17% cover of live plants, while in burned plots it had <1% cover of dead plants and no cover of live plants. The authors anecdotally described fire behavior and noted that a fuel break kept the fire out of most of the invasive perennial grass fuels, but where buffelgrass and crimson fountaingrass burned, saguaro and white brittlebush did not resprout and paloverde resprouted weakly. The fire failed to spread in white brittlebush, saguaro, and paloverde communities without continuous fuel loads from invasive grasses [274]. On buffelgrass pastures in Sonora where fuel loads were relatively low (< 0.7 metric tons/ha), white brittlebush mortality was low and white brittlebush resprouted vigorously after fire. On pastures with higher fuel loads (>4.5 metric tons/ha), white brittlebush mortality was near 100% after fire [115]. For more information on this study, see the Research Project Summary by Smith [227].
Mojave Mid-Elevation Mixed Desert Scrub
In Mojave blackbrush desert scrub in Utah, white brittlebush prominence increased 6 years after four independent wildfires. Its density and cover were similar between burned (0.011 plants/m2 and 0.37%) and unburned (0.007 plants/m2 and 0.33%) plots, but its importance was greater in burned plots. It had the eleventh highest density of any species in unburned plots and the second highest density in burned plots, only exceeded by white bursage. Similarly, it had the ninth highest cover of any species in unburned plots, and the third highest cover in burned plots, exceeded only by white bursage and creosotebush. White brittlebush and Anderson wolfberry were the only species of 12 shrub species examined that increased after fire [110].
Central & Southern Californian Coastal Sage Scrub
Much of the evidence for white brittlebush reproducing primarily by seed after fire comes from coastal sage scrub [125,151,169]. One study covering 16 fires in southern California found that only a small proportion of white brittlebush resprouted after fire but that postfire seedling recruitment was high, on average. Following October and November wildfires, an average of 20% of white brittlebush plants resprouted and seven seedlings were produced for each prefire parent plant during the first postfire growing season. Regression analysis indicated that white brittlebush seedlings occurred at nearly the same proportion of the postfire seedling community the first year after fire as it did in the prefire community. However, because it resprouts less than associated species in coastal sage scrub, it was relatively less abundant than associated species after fire. Over the first 5 years after fire, seedling recruitment averaged 21,100 seedlings/ha on eight sites. Ten percent, 19%, 18%, 0%, and 53% of total seedlings recruited in postfire years 1 through 5, respectively. Precipitation was approximately 80% of average in the first, third, and fourth years, 175% of average in the second year, and 250% of average in the fifth year. Consecutive years of below-average precipitation may have prevented white brittlebush establishment in the fourth year [125].
After a fire in coastal sage scrub near Loma Linda, California, white brittlebush resprouting was greatest on north-facing slopes but seedling density was least, likely due to higher moisture conditions on north-facing slopes. Four months after a June wildfire, nearly 30% of white brittlebush resprouted on mesic north-facing slopes, while only 4% to 8% resprouted on relatively xeric east-, west-, and south-facing slopes. Eighteen months after the fire, 77% of regeneration on north-facing slopes was from resprouts, while only 4% to 8% was from resprouts on east-, west-, and south-facing slopes. Prefire white brittlebush density was 237 plants/ha on north-facing slopes, and 2,171 to 2,456 plants/ha on east-, west-, and south-facing slopes. Eighteen months after fire, white brittlebush density was <10 plants/ha on north-facing slopes and 1,924 to 4,695 plants/ha other slopes. Seedlings occurred in much higher numbers on east-, west-, and south-facing than on north-facing slopes [151]. East-, west-, and south-facing slopes may have higher densities of white brittlebush than north-facing slopes because warmer soil and air temperatures are favorable for germination [36] and root development [182] (see Seasonal Development), while mesic north-facing slopes potentially experience lower fire intensity, which is more favorable for resprouting [151,271]. A study in coastal sage scrub in San Bernardino County indicated that white brittlebush resprouting was greater after a low-intensity October fire than a high-intensity July fire [269,271]. For more information on this study, see Immediate Fire Effects.
Californian Xeric Chaparral Group
No published studies addressed white brittlebush’s response to fire in xeric chaparral.
Fuel Characteristics
The Native American Ethnobotany Database (2024) reported that the resinous branches of white brittlebush can be used to make “quick fires” [171], and another author indicated that it contains high resin and oil contents [152], suggesting that branches are highly flammable when dry. However, white brittlebush plants are rarely entirely consumed by low-intensity fires. Branches and leaves close to the ground are typically charred or scorched by fire carried by fine fuels, while green leaves typically remain on higher branches [160] (see Immediate Fire Effects).
White brittlebush individuals do not accumulate organic matter or support high densities of annual species at their base [167,207,234], which may limit fuel connectivity [234]. However, white brittlebush communities may be more able to carry fire even without a continuous layer of annuals than communities dominated by other long-lived perennials (e.g., creosotebush) [234]. In southern California, perennial vegetation in white brittlebush-creosotebush scrub 28 years after fire was denser and more evenly spaced than in unburned creosotebush scrub, indicating greater fuel continuity. Moisture content of white brittlebush may be very low relatively early in the dry season, which may also increase flammability relative to creosotebush [236], which may lessen the importance of herbaceous fuels necessary to carry fire [234].
In most years, desert scrub communities in which white brittlebush occurs are fuel-limited, because typically discontinuous vegetation and sparse understories limit ignition and fire spread. However, after 1 or more years with above-average precipitation, establishment of annual forbs and grasses can result in relatively thick and continuous patches of fine fuels that are sufficient to carry fire [50,89,90,102,158,209,274,279]. Nonnative invasive plants can add abundant and novel fuels to desert scrub ecosystems, increasing fuel biomass and continuity, and thus increasing the potential for uncharacteristically large and frequent fires [50,89,102,274,279]. Greater size and frequency of fires may create a feedback loop that results in an invasive grass/fire cycle [47,73,131,155,279].
Nonnative invasive plants that have impacted or have the potential to impact fuels in white brittlebush communities include the annual grasses compact brome, red brome [224], cheatgrass [278], Bermudagrass [59], and common Mediterranean grass [51], and the perennial grasses buffelgrass [116,156], crimson fountaingrass, and Lehmann lovegrass [254,274]. Annual invasive forbs such as redstem stork’s bill [49,113] and Sahara mustard [117] can also contribute to cover and continuity of fine surface fuels. Historically, native annuals did not occur in densities sufficient to carry fire as often as contemporary nonnatives, so fire intervals were longer [138].
Nonnative invasive perennial grasses, especially buffelgrass, crimson fountaingrass, and Lehmann lovegrass, may increase fuel loading relative to nonnative annuals or native species, resulting in greater fuel continuity and higher fire intensity [274]. Higher fuel loads by invasive perennial grasses increases white brittlebush mortality after fire [115,274] (see Buffelgrass Communities). Buffelgrass creates a persistent, year-to-year fire hazard that can burn in any month [89,156]. See the FEIS Species Review about buffelgrass for more details. In the desert regions of the southwestern United States and northern Mexico, nonnative invasive grasses have the potential to replace less fire-adapted native species and establish a grass-fire cycle [156]. Climate change is likely to further exacerbate invasion by nonnative, fire-adapted grasses and increase the length of fire seasons in the North American warm deserts [1] (see Management Under a Changing Climate).
Fire Regimes
White brittlebush is important in plant communities with vastly different historical mean fire intervals. Estimated mean historical fire intervals derived from LANDFIRE succession modeling range from <100 years in Southern California Dry-Mesic Chaparral and North American Warm Desert Riparian Systems to >1,000 years in Sonoran Paloverde-Mixed Cacti Desert Scrub [138] (see LANDFIRE Biophysical Settings section). In Sonoran Paloverde-Mixed Cacti Desert Scrub, fire is thought to have been rare to infrequent based on a lack of fuels to carry fire in most years and a lack of fire adaptations [12,251] and slow recovery rate [2,210] of dominant plants (e.g., saguaro). See Fire Regimes of Sonoran Desert Scrub Communities for more information. In Southern California Coastal Sage Scrub and Southern California Dry-Mesic Chaparral, historical mean fire intervals are modeled at 150 years and 51 years, respectively, based on relatively dense, shrub-dominated vegetation that is capable of carrying fire in most years [138].
Contemporary fire regimes in many ecosystems in which white brittlebush occurs may be outside their historical range of variability in some areas due to increases in human-caused ignitions and invasions of nonnative plants—especially grasses—that are driving shorter fire intervals [50,127,212,279]. Plant communities especially vulnerable to fire regime shifts include coastal sage scrub, dry-mesic chaparral, and Mojave and Sonoran desert scrub. For example, in coastal sage scrub and chaparral, nonnative annual grasses increase fine fuel loads and may burn annually, eliminating most native shrubs and resulting in type conversion to nonnative annual grasslands [62]. In Mojave and Sonoran desert scrub, increased human access and subsequent human-caused ignitions, and increases in receptive fine fuel loads and continuity caused by nonnative plant invasions (see Fuel Characteristics), are resulting in more frequent, larger, and more severe fires than occurred historically. Rapid recovery of herbaceous fuels—especially of nonnative grasses—may lead to repeated fires on a site. Repeated and/or large patches of high-severity fire can lead to vegetation type conversions that may persist indefinitely [132,279].
There is some debate regarding how often fire occurred historically in coastal sage scrub and chaparral communities [23,138]. Observations of modern fires indicate that very short (<10–25 years) fire intervals do not allow enough time for postfire shrub recovery, and nonnative invasive grasses may outcompete and replace native species, especially after multiple fires [123,128,138,277]. Very long fire intervals may result in conversion of coastal sage scrub and chapparal communities to other vegetation types, but more information is needed [138,260]. In general, fire intervals that are too short are a larger concern than fire intervals that may be too long [128]. Large wildfires in chaparral and coastal sage scrub are typically driven by extreme weather (i.e., Santa Ana winds) [138].
Climate change and invasive species are likely to continue to alter fire regimes in white brittlebush communities in the arid Southwest, which are likely to see more ignitions, larger fires, and continued spread of nonnative invasive grasses [1] (see Management Under a Changing Climate).
See these FEIS publications for information on historical fire regimes in plant communities in which white brittlebush is most common:
- Fire regimes of Sonoran desert scrub communities
- Fire regimes of creosotebush-white bursage desert scrub communities
- Fire regimes of Mojave mid-elevation mixed desert scrub communities
- Fire regimes of coastal sage scrub communities
- Fire regimes of California chaparral communities
For additional fire regime information, search FEIS for this species by entering the species name or acronym on the home page and selecting “Fire Regime” as the publication type.
Fire Management Considerations
Fire management considerations for white brittlebush vary depending on plant community. Fire is generally considered harmful in Mojave and Sonoran desert scrub communities where white brittlebush occurs. While white brittlebush cover and density often increase after fire, fire was historically infrequent in these communities (see Fire Regimes) and many dominant species are not resilient to frequent, large, and/or severe wildfires, especially in communities invaded by nonnative grasses and forbs [279]. Dominant species such as saguaro, paloverde, and creosotebush, may have high mortality rates and low establishment rates after fire [210,236]. If they are reduced or eliminated by fire, burned areas may become dominated by native “ruderal” species, such as white brittlebush [142,210,222], or by nonnative species for decades or longer [279]. Because postfire recovery of predisturbance vegetation may not occur for decades or even centuries [88,234], if at all, the primary fire management consideration is fire prevention, part of which requires fuels management in the form of reducing cover of nonnative invasive plants [279], and/or fuels that might sustain large-scale, high-intensity fires [170,244].
The National Park Service typically manages human-caused wildfires and desert wildfires with a full-suppression policy [170,244]. However, use of prescribed fire as a management tool is sometimes recommended in the Mojave and Sonoran Deserts for research, fuels reduction of nonnative invasive grasses, and cultural values [244,252,253]. At the upper-elevational extent of white brittlebush’s range in the Sonoran Desert, lightning-caused wildfires may be allowed to burn to achieve resource benefits, such as fuels reduction [244].
While fire has likely been an important process in coastal sage scrub and chaparral communities historically, with modeled estimated mean fire intervals of about 50 to 150 years [138], contemporary fuels and fire regimes have shifted in some areas due to increased ignitions and fine fuel loads (see Fire Regimes). Prescribed fire is generally not recommended in these communities except in strategic locations to reduce fuel loads to prevent fire from spreading across the wildland-urban interface. Additional fire management recommendations are to contain wildland fires within easily defended boundaries and to maintain a sustainable fire regime in chaparral (i.e., at least 25 years) [66,128]. Wildfires are typically suppressed in these communities [128,138]. However, full-suppression management policies appear to have little influence on total area burned or number of large fires in chaparral because large fires generally occur under extreme weather conditions [128]. Because chaparral and coastal sage scrub are vulnerable to type conversion to nonnative invasive grasslands, management considerations in these plant communities also include limiting spread of invasive grasses and restoring native vegetation [14,65].
Federal Status
None [256]
Other Status
According to NatureServe, white brittlebush is critically imperiled (S1) in Utah but has no status rank in the other three US states where it is native. It is ranked as globally secure (G5).
Information on conservation status of plant species and vegetation communities in the United States is available at NatureServe.
Importance to Wildlife and Livestock
High resin and oil contents make white brittlebush relatively unpalatable to wildlife and livestock [152]. However, it can be seasonally important in the diets of deer and cattle. In Sonora, Mexico, mule deer and cattle ate white brittlebush at variable rates within and among seasons, years, and locations. For example, it comprised from 0.1% to 7.3% of mule deer diets in the spring and from 0.2% to 10.7% in the fall. At one location, white brittlebush comprised 1.4% of fall forage in one year, and 10.7% of fall forage the following year at the same location. It comprised 0.5% of cattle diets in summer and 2.3% in fall. [11]. In southeastern California, white brittlebush provided up to 19% of desert mule deer diets in the summer, up to 15% in the spring, and less than 4% in the fall and winter [149]. Annual climate or phenological differences may influence the degree to which deer consume white brittlebush [11]. Bighorn sheep sometimes consume white brittlebush [35,150,215].
White brittlebush supports the specialist leaf beetle, the encelia leaf beetle. It consumes white brittlebush foliage [30,229,276], which may reduce white brittlebush fitness, particularly in combination with other stressors [191].
White brittlebush communities provide important habitat for two federally threatened species: California gnatcatchers [31] and Mohave desert tortoise (see Cover Value) [29].
Figure 7—A bighorn sheep extends his neck to eat white brittlebush flowers.
Palatability and Nutritional Value
Deer and bighorn sheep consume white brittlebush leaves and flowers [19] (fig. 7).
Cattle and domestic sheep readily eat flowers of white brittlebush, but rarely eat the leaves [76].
See the following studies for information on the nutritional value of white brittlebush leaves and flowers: [10,134,206,215].
Cover Value
Experts ranked white brittlebush as having moderate cover value for Sonoran desert tortoise [222]. It is also used extensively as cover by the Mohave desert tortoise [142,183,184]. In the western Sonoran Desert, Mohave desert tortoise preferred white brittlebush communities over other major vegetation types [28].
Value for Rehabilitation or Restoration of Disturbed Sites
White brittlebush is often used in seed mixes for revegetation efforts in southwestern Arizona [45,48,100], California [133], and Nevada [27,35], especially along roadsides [20]. Its range may be expanding in the United States as a result of its use in restoration projects [20,27,35,45,48,100,133] (see General Distribution).
White brittlebush is described as “hardy” and “easy to grow” from seeds or young transplants [35], although success of seeding and outplanting can be highly variable (e.g., [21,64]). For example, along California highways in the Mojave Desert, white brittlebush established well at only two of five sites, with an average of 3 plants/18 m2 on sites where it established. Seeds were both drilled and broadcast after disking and/or harrowing then mulched, fertilized, and irrigated [64]. Because seed viability may be very low (1%–5%) under horticultural conditions [120,189], Everett (2012) noted that brittlebush seeds should be sown “rather thickly” [92]. Survival of planted individuals may be high if soil moisture is available [9] (see Plant Growth and Mortality).
White brittlebush is a nurse plant for the large columnar cactus species saguaro and organpipe cactus [80,98,192,193,233].
White brittlebush is competitive with nonnative invasive grasses such as cheatgrass [264], red brome [6], and buffelgrass [56,115]. Other sources suggest white brittlebush may be displaced by buffelgrass if buffelgrass is provided temporary competitive release by such disturbances as freezes or fires that kill white brittlebush, and white brittlebush may be unable to reestablish [55].
Other Uses
People of the Papago Tribe, the Cahuilla Nation, and the Pima use white brittlebush for several purposes. An extraction of blossoms, leaves, and stems can be used to remedy toothaches, resins secreted by the plant can be heated and applied to the chest to relieve pains [26,100], and a poultice of the plant can be used as an analgesic [171]. White brittlebush resin can also be used as a chewing gum [60,121,171], a glue for arrow points [119], a waterproofing agent [60], and as varnish [171]. The resinous branches can be used to build “quick fires” [171].
Catholic priests use white brittlebush resin as an incense [100,121], inspiring the common name “incienso” [70,119].
White brittlebush is often planted for its attractive and hardy nature as a low water landscape plant [35,100,175].
Laboratory studies indicate that white brittlebush compounds have antibacterial properties [78].
Additional Management Considerations
While white brittlebush populations are generally secure throughout most of the species’ range (but see Other Status), some populations are threatened by buffelgrass invasion [5,55,187,258,259,274]. In Saguaro National Park near Tucson, Arizona, white brittlebush was an indicator of non-buffelgrass patches. White brittlebush occurred in all plots in all patch types but had six times more cover in patches without buffelgrass than in patches with buffelgrass. White brittlebush seeds occurred at higher densities in buffelgrass patches, possibly because seeds were retained by the buffelgrass structure, but did not germinate before or after invasion [5]. Where white brittlebush is reduced by an episodic mortality event, such as a significant freeze, buffelgrass may establish during the temporary release from competition and exclude white brittlebush seedlings from establishing thereafter [55]. In parts of Arizona and Sonora, buffelgrass has replaced native ruderal species, such as white brittlebush, along highways and in other disturbed areas [259]. In contrast, white brittlebush may establish or reestablish in buffelgrass pastures managed for livestock grazing, where it is generally seen as a nuisance species. Several studies describe integrated management methods that include prescribed fire, manual and mechanical removal, and/or herbicides to kill white brittlebush and other woody plants and maintain buffelgrass [96,114,115,152,227].
White brittlebush populations may increase and individuals may live longer where livestock grazing is excluded. White brittlebush maximum longevity in areas where grazing has historically been excluded was more than twice (72 years) previous recorded maximum longevity in other populations (32 years) [43] (see Plant Growth and Mortality). In 1976, 2 years before grazing ceased in Saguaro National Park, white brittlebush occurred with an average of 5% cover on ten vegetation monitoring plots. In 2018, white brittlebush occurred with an average of 22% cover. Some plots had burned in one or two fires, but white brittlebush cover and density increased to a similar degree in unburned plots [240].
White brittlebush may be showing signs of toxic effects from nitrogen saturation in coastal sage scrub [15], although soil nitrogen is likely only at its highest levels in the dry season, so it is unclear whether shrubs are dying from high nitrogen levels [16].
White brittlebush contains allelopathic compounds that may inhibit seedling growth and kill young plants [34,103]. Allelopathy may account for low densities of annuals around white brittlebush plants compared to similar shrubs, although such low densities may instead be due to its branching structure and lack of organic material around its base [22,144,167].
Management Under a Changing Climate
The deserts of the southwestern United States are particularly vulnerable to climate change and increasing climate variability because desert organisms already live near their limits for water and temperature stress. Many climate models predict that precipitation will decrease and shift in timing across most of the warm deserts of the American Southwest and Mexico [1,18,203]. Shifting precipitation patterns include increased variability in rainfall amount and timing [198], which may result in soil water content below plant thresholds more often [203]. Evidence of a warming and drying climate is already apparent in the Southwest [68,188], including in the Mojave [104], Sonoran [263], and Chihuahuan [107] deserts, as well as throughout California [77].
White brittlebush may be vulnerable to deficits of cool-season precipitation. For example, minimum cool-season precipitation requirements for germination [36] (see Germination) and flowering [39] (see Seed Production and Predation) of white brittlebush may be met less often and/or may be followed by hotter, drier periods, potentially resulting in decreased establishment [91] or truncated flowering periods [39,175]. White brittlebush and other drought deciduous shrubs may have increased water demands as their habitat warms [18], which may cause them to lose their leaves earlier in the season, decrease overall photosynthetic capacity, and reduce available soil water [86].
Under a warmer, drier climate, white brittlebush could flower earlier in the year due to heat sum requirements for flowering being met earlier in the season, similar to patterns observed in other desert shrubs due to the heat island effect [174]. Sonoran Desert shrubs, including white brittlebush, flowered 21 to 40 days earlier in the year in 2004 than in 1894, a result attributed to climate warming [38]. Phenology shifts could have far-reaching impacts if flowering times or other phenological events misalign with pollinator activity, precipitation timing, or other ecological or climatic factors, although plant responses to climate change and likely ecosystem impacts are not well understood [201].
Under a warmer, drier climate, white brittlebush could shift its range [104] into areas where its distribution was previously limited by low temperatures, such as into higher elevations or northern latitudes. For example, in Saguaro National Park, white brittlebush occurred only below 1,100 m in initial surveys in 1976, but after several fires over 40 years, white brittlebush expanded onto sites above 1,100 m [240] (see Plant Response to Fire). Climate models project fewer frost days in the desert southwest [18], which could allow white brittlebush to expand into areas that have historically been unsuitable habitat due to freezing thresholds [246,262]. It may be expanding its range in Arizona, California, and Nevada due to its inclusion in roadside seeding mixes [20,27,35,45,48,100,133], possibly exacerbated by concurrent expansion of suitable habitat.
In the western United States, wildfire activity is predicted to continue to increase [1,101,195], and white brittlebush is likely to become more prevalent throughout much of its range in response [12,17,24,54,61,141,142,153,199,213,222,234]. However, very frequent fire reduces white brittlebush populations if fire is so frequent that white brittlebush fails to regenerate [115]. Reduced spring precipitation may not meet precipitation thresholds for flowering and germination [39,41], which could reduce seed production and seedling establishment. Over 37 years in Death Valley National Park, mean annual temperatures increased by 1.2 °C and winter–spring precipitation showed a downward trend. Over the study period, white brittlebush exhibited a long-term decrease in shrub abundance, despite several substantial establishment periods. Ehleringer (2018) postulates that high density of white brittlebush on shallow soils will drive increasing competition for a decreasing water supply, exacerbating competitive interactions and limiting population sizes [86]. Increases in fire frequency may also promote invasive species establishment and spread and intensify wildfire-invasion feedback loops in aridlands [1], which may interfere with white brittlebush establishment and persistence.
| Biophysical Setting | Mean Fire Interval (years) | Fire Severity Percent (%) | ||||||
|---|---|---|---|---|---|---|---|---|
| Code | Fire Regime Group | Low | Mixed | Replacement | All | Low | Mixed | Replacement |
| Series 10820 - Mojave Mid-Elevation Mixed Desert Scrub | ||||||||
| 10820_12_15_16_17_23_24 | V-B | 815 | 815 | 0 | 0 | 100 | ||
| 10820_14 | V-A | 406 | 406 | 0 | 0 | 100 | ||
| 10820_4_13 | V-A | 399 | 399 | 0 | 0 | 100 | ||
| Series 10870 - Sonora-Mojave Creosotebush-White Bursage Desert Scrub | ||||||||
| 10870_4_6_12_13_14_15_17_24_25 | V-A | 329 | 329 | 0 | 0 | 100 | ||
| Series 10900 - Sonoran Granite Outcrop Desert Scrub | ||||||||
| 10900_14 | V-B | 513 | 513 | 0 | 0 | 100 | ||
| Series 10920 - Southern California Coastal Scrub | ||||||||
| 10920_4 | IV-B | 150 | 150 | 0 | 0 | 100 | ||
| Series 10960 - California Maritime Chaparral | ||||||||
| 10960_4 | IV-B | 124 | 124 | 0 | 0 | 100 | ||
| Series 11090 - Sonoran Paloverde-Mixed Cacti Desert Scrub | ||||||||
| 11090_14 | V-B | 1056 | 1056 | 0 | 0 | 100 | ||
| 11090_15_25 | V-B | 1049 | 1049 | 0 | 0 | 100 | ||
| 11090_4_13 | V-B | 1284 | 1284 | 0 | 0 | 100 | ||
| Series 11100 - Southern California Dry-Mesic Chaparral | ||||||||
| 11100_4 | IV-A | 51 | 51 | 0 | 0 | 100 | ||
| Series 11552 - North American Warm Desert Riparian Systems - Stringers | ||||||||
| 11552_13_14_15 | V-B | 735 | 735 | 0 | 0 | 100 | ||
| Summary | ||||||||
| Minimum | 51 | 51 | 0 | 0 | 100 | |||
| Maximum | 1284 | 1284 | 0 | 0 | 100 | |||
| Mean | 576 | 576 | 0 | 0 | 100 | |||
| Median | 460 | 460 | 0 | 0 | 100 | |||
| Percentage of fires in 3 fire severity classes, derived from LANDFIRE succession modeling. Replacement-severity fires cause >75% kill or top-kill of the upper canopy layer; mixed-severity fires cause 26%-75%; low-severity fires cause <26%. | ||||||||
| LANDFIRE. 2020. Biophysical settings models and descriptions, [Online]. Washington, DC: U.S. Department of Agriculture, Forest Service; U.S. Department of the Interior; U.S. Geological Survey; Arlington, VA: The Nature Conservancy, (Producers). Available: https://www.landfirereview.org/search.php [96496] | ||||||||
Table A1
| Life form | Common name | Scientific name |
|---|---|---|
| Cactus | buckhorn cholla | Cylindropuntia acanthocarpa |
| Cactus | coastal pricklypear | Opuntia littoralis |
| Cactus | organpipe cactus, organ pipe cactus | Stenocereus thurberi |
| Cactus | pachycereus | Pachycereus spp. |
| Cactus | saguaro | Carnegiea gigantea |
| Cactus | teddybear cholla, teddy-bear cholla | Cylindropuntia bigelovii |
| Forb | brownplume wirelettuce | Stephanomeria pauciflora |
| Forb | hairy desertsunflower | Geraea canescens |
| Forb | redstem stork's bill* | Erodium cicutarium |
| Forb | Sahara mustard* | Brassica tournefortii |
| Graminoid | Bermudagrass* | Cynodon dactylon |
| Graminoid | big galleta | Pleuraphis rigida |
| Graminoid | buffelgrass* | Pennisetum ciliare |
| Graminoid | cheatgrass* | Bromus tectorum |
| Graminoid | common Mediterranean grass* | Schismus barbatus |
| Graminoid | compact brome* | Bromus madritensis |
| Graminoid | crimson fountaingrass* | Pennisetum setaceum |
| Graminoid | Lehmann lovegrass* | Eragrostis lehmanniana |
| Graminoid | Mediterranean grass* | Schismus spp. |
| Graminoid | red brome* | Bromus rubens |
| Shrub | Anderson wolfberry | Lycium andersonii |
| Shrub | beloperone | Justicia californica |
| Shrub | black sage | Salvia mellifera |
| Shrub | blackbrush | Coleogyne ramosissima |
| Shrub | brittlebush, white brittlebush | Encelia farinosa |
| Shrub | burrobrush | Hymenoclea salsola |
| Shrub | bush arrowleaf | Pleurocoronis pluriseta |
| Shrub | button brittlebush | Encelia frutescens |
| Shrub | California brittlebush | Encelia californica |
| Shrub | California flannelbush | Fremontodendron californicum |
| Shrub | catclaw acacia | Senegalia greggii |
| Shrub | ceanothus | Ceanothus spp. |
| Shrub | Central Desert encelia | Encelia asperifolia |
| Shrub | chamise | Adenostoma fasciculatum |
| Shrub | coastal sagebrush, California sagebrush | Artemisia californica |
| Shrub | creosotebush | Larrea tridentata |
| Shrub | crucifixion thorn, crucifixion-thorn | Canotia holacantha |
| Shrub | desert ironwood, desert-ironwood | Olneya tesota |
| Shrub | desert lavender, desert-lavender | Hyptis emoryi |
| Shrub | desertholly, desert-holly | Atriplex hymenelytra |
| Shrub | Eastern Mojave buckwheat | Eriogonum fasciculatum |
| Shrub | jojoba | Simmondsia chinensis |
| Shrub | laurel sumac | Malosma laurina |
| Shrub | lemonade sumac | Rhus integrifolia |
| Shrub | manzanita | Arctostaphylos spp. |
| Shrub | Mohave rabbitbrush, Mojave rabbitbrush | Ericameria paniculata |
| Shrub | Mojave yucca | Yucca schidigera |
| Shrub | ocotillo | Fouquieria splendens |
| Shrub | Palmer encelia | Encelia palmeri |
| Shrub | Parish's goldeneye | Viguiera parishii |
| Shrub | Parry's beargrass, Parry's bear-grass | Nolina parryi |
| Shrub | physicnut | Jatropha cuneata |
| Shrub | plains encelia | Encelia halimifolia |
| Shrub | rough jointfir, rough joint-fir | Ephedra aspera |
| Shrub | sacahuista | Nolina microcarpa |
| Shrub | San Luis purple sage | Salvia leucophylla |
| Shrub | Schott’s pygmycedar, Schott's pygmy-cedar | Peucephyllum schottii |
| Shrub | sweetbush | Bebbia juncea |
| Shrub | triangle bur ragweed | Ambrosia deltoidea |
| Shrub | Vizcaino encelia | Encelia densifolia |
| Shrub | white bursage, burrobush | Ambrosia dumosa |
| Shrub | white sage | Salvia apiana |
| Shrub | yellow paloverde | Parkinsonia microphylla |
| Tree | blue paloverde | Parkinsonia florida |
| Tree | elephant tree | Bursera microphylla |
| Tree | Joshua tree | Yucca brevifolia |
| Tree | mesquite | Prosopis spp. |
| Tree | paloverde | Parkinsonia spp. |
| Tree | velvet mesquite | Prosopis velutina |
| Tree | whitethorn acacia | Vachellia constricta |
Table A2
| Life form | Common name | Scientific name |
|---|---|---|
| Bird | California gnatcatcher | Polioptila californica |
| Insect | ants | Formicidae |
| Insect | bees | Apoidea |
| Insect | beetles | Coleoptera |
| Insect | butterflies | Lepidoptera |
| Insect | encelia leaf beetle | Trirhabda geminata |
| Insect | flies | Diptera |
| Insect | moths | Lepidoptera |
| Mammal | bighorn sheep | Ovis canadensis |
| Mammal | cattle | Bos taurus |
| Mammal | desert mule deer | Odocoileus hemionus crooki |
| Mammal | domestic sheep | Ovis aries |
| Mammal | mule deer | Odocoileus hemionus |
| Mammal | rodents | Rodentia |
| Reptile | Mohave desert tortoise | Gopherus agassizii |
| Reptile | Sonoran desert tortoise | Gopherus morafkai |
1. Abatzoglou, John T.; Kolden, Crystal A. 2011. Climate change in western US deserts: Potential for increased wildfire and invasive annual grasses. Rangeland Ecology & Management. 64(5): 471-478. [92501]
2. Abella, S. R. 2009. Post-fire plant recovery in the Mojave and Sonoran Deserts of western North America. Journal of Arid Environments. 73(8): 699-707. [81859]
3. Abella, S. R.; Prengaman, K. A.; Embrey, T. M.; Schmid, S. M.; Newton, A. C.; Merkler, D. J. 2012. A hierarchical analysis of vegetation on a Mojave Desert landscape, USA. Journal of Arid Environments. 78: 135-143. [84158]
4. Abella, Scott R. 2010. Disturbance and plant succession in the Mojave and Sonoran Deserts of the American Southwest. International Journal of Environmental Research and Public Health. 7(4): 1248-1284. [95419]
5. Abella, Scott R.; Chiquoine, Lindsay P.; Backer, Dana M. 2012. Ecological characteristics of sites invaded by buffelgrass (Pennisetum ciliare). Invasive Plant Science and Management. 5(4): 443-453. [96003]
6. Abella, Scott R.; Craig, Donovan J.; Chiquoine, Lindsay P.; Prengaman, Kathryn A.; Schmid, Sarah M.; Embrey, Teague M. 2011. Relationships of native desert plants with red brome (Bromus rubens): Toward identifying invasion-reducing species. Invasive Plant Science and Management. 4(1): 115-124. [98829]
7. Abella, Scott R.; Craig, Donovan J.; Suazo, Alexis A. 2012. Outplanting but not seeding establishes native desert perennials. Native Plants. 13(2): 81-89. [87413]
8. Abella, Scott R.; Newton, Alice C.; Bangle, Dianne N. 2007. Plant succession in the eastern Mojave Desert: An example from Lake Mead National Recreation Area, southern Nevada. Crossosoma. 33(2): 45-55. [99092]
9. Abella, Scott R.; O'Brien, Kara L. 2015. Revegetating disturbance in National Parks: Reestablishing native plants in Saguaro National Park, Sonoran Desert. Natural Areas Journal. 35(1): 18-25. [97226]
10. Albert, Steven K.; Krausman, Paul R. 1993. Desert mule deer and forage resources in southwest Arizona. The Southwestern Naturalist. 38(3): 198-205. [22140]
11. Alcala-Galvan, Carlos Hugo; Krausman, Paul R. 2012. Diets of desert mule deer in altered habitats in the lower Sonoran Desert. California Fish and Game. 98(2): 81-103. [96437]
12. Alford, Eddie J.; Brock, John H. 2002. The effects of fire on Sonoran Desert plant communities. Final Report: RMRS-99164-RJVA. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station. 111 p. [Alford's Dissertation]. On file with: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Fire Sciences Laboratory, Missoula, MT. [47514]
13. Allen, Edith B. 2004. Restoration of Artemisia shrublands invaded by exotic annual Bromus: A comparison between southern California and the intermountain Region. In: Hild, Ann L.; Shaw, Nancy L.; Meyer, Susan E.; Booth, D. Terrance; McArthur, E. Durant, compilers. Seed and soil dynamics in shrubland ecosystems: Proceedings; 2002 August 12-16; Laramie, WY. RMRS-P-31. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: 9-17. [48994]
14. Allen, Edith B.; McDonald, Christopher; Hilbig, Bridget E. 2019. Long-term prospects for restoration of coastal sage scrub: Invasive species, nitrogen deposition, and novel ecosystems. In: Narog, Marcia, tech. coord. Proceedings of the chaparral restoration workshop, California; 2013 June 17-20; Arcadia, CA. Gen. Tech. Rep. PSW-GTR-265. Albany, CA: U.S. Department of Agriculture, Forest Service, Pacific Southwest Research Station: 1-18. [99323]
15. Allen, Edith B.; Padgett, Pamela E.; Bytnerowicz, Andrzej; Minnich, Richard. 1998. Nitrogen deposition effects on coastal sage vegetation of southern California. In: Bytnerowicz, Andrzej; Arbaugh, Michael J.; Schilling, Susan L., tech. coords. Proceedings of the international symposium on air pollution and climate change effects on forest ecosystems; 1996 February 5-9; Riverside, CA. Gen. Tech. Rep. PSW-GTR-166. Berkeley, CA: U.S. Department of Agriculture, Forest Service, Pacific Southwest Research Station: 131-139. [64971]
16. Allen, Edith B.; Sirulnik, Abby G.; Egerton-Warburton, Louise; Kee, Sheila N.; Bytnerowicz, Andrzej; Padgett, Pamela E.; Temple, Patrick J.; Fenn, Mark E.; Poth, Mark A.; Meixner, Thomas. 2005. Air pollution and vegetation change in southern California coastal sage scrub: A comparison with chaparral and coniferous forest. In: Kus, Barbara E.; Beyers, Jan L., tech. coords. Planning for biodiversity: Bringing research and management together: Proceedings of a symposium for the South Coast ecoregion; 2000 February 29 - March 2; Pomona, CA. Gen. Tech. Rep. PSW-GTR-195. Albany, CA: U.S. Department of Agriculture, Forest Service, Pacific Southwest Research Station: 79-95. [64505]
17. Allen, Edith B.; Steers, Robert J.; Dickens, Sara Jo. 2011. Impacts of fire and invasive species on desert soil ecology. Rangeland Ecology & Management. 64(5): 450-462. [84648]
18. Archer, Steven R.; Predick, Katharine I. 2008. Climate change and ecosystems of the southwestern United States. Rangelands. 30(3): 23-28. [96376]
19. Arizona Fish and Game Department. 1977. The Arizona white-tailed deer. Special Report No. 6. [Federal Aid in Wildlife Restoration Act: Project W-53-R]. Phoenix, AZ: Arizona Fish and Game Department. 108 p. [61802]
20. Baldwin, Bruce G.; Goldman, Douglas H.; Keil, David J.; Patterson, Robert; Rosatti, Thomas J.; Wilken, Dieter H., eds. 2012. The Jepson manual. Vascular plants of California. 2nd ed. Berkeley, CA: University of California Press. 1568 p. [86254]
21. Banerjee, Monisha J.; Gerhart, Vanda J.; Glenn, Edward P. 2006. Native plant regeneration on abandoned desert farmland: Effects of irrigation, soil preparation, and amendments on seedling establishment. Restoration Ecology. 14(3): 339-348. [64254]
22. Barbour, Michael G. 1973. Desert dogma reexamined: Root/shoot productivity and plant spacing. The American Midland Naturalist. 89(1): 41-57. [49000]
23. Barro, S. C.; Conard, S. G. 1991. Fire effects on California chaparral systems: An overview. Environmental International. 17(2-3): 135-149. [15760]
24. Barron, Kara Lynn. 2018. Fire and reseeding effects on Arizona upland plant community composition and a preliminary floristic inventory of Cave Creek Regional Park. Phoenix, AZ: Arizona State University. 75 p. Thesis. [96491]
25. Baskin, Carol C.; Baskin, Jerry M. 2001. Seeds: Ecology, biogeography, and evolution of dormancy and germination. San Diego, CA: Academic Press. 666 p. [60775]
26. Bean, Lowell John; Saubel, Katherine Siva. 1972. Telmalpakh: Cahuilla Indian knowledge and usage of plants. 1st ed. Banning, CA: Malki Museum Press. 225 p. [35898]
27. Beatley, Janice C. 1976. Vascular plants of the Nevada Test Site and central-southern Nevada: Ecologic and geographic distributions. [Washington, DC]: U.S. Energy Research and Development Administration, Office of Technical Information, Technical Information Center. 308 p. Available from U.S. Department of Commerce, National Technical Information Service, Springfield, VA. TID-26881/DAS. [63152]
28. Berry, Kristin H.; Yee, Julie L.; Lyren, Lisa M. 2020. Feral burros and other influences on desert tortoise presence in the western Sonoran Desert. Herpetologica. 76(4): 403-413. [97463]
29. Bertelsen, David C. 2018. Thirty-seven years on a mountain trail: Vascular flora and flowering phenology of the Finger Rock Canyon Watershed, Santa Catalina Mountains, Arizona. Desert Plants. 34(1-2): 1-247. [95896]
30. Bethke, James A.; Redak, Richard A. 1996. Seasonal occurrence of the herbivore Trirhabda geminata (Coleoptera: Chrysomelidae) on Encelia farinosa in southern California. Annals of the Entomological Society of America. 89(6): 843-848. [99107]
31. Beyers, Jan L.; Pena, Ginger C. 1995. Characteristics of coastal sage scrub in relation to fire history and use by California gnatcatchers. In: Weise, David R.; Martin, Robert E., technical coordinators. The Biswell symposium: Fire issues and solutions in urban interface and wildland ecosystems; 1995 February 15-17; Walnut Creek, CA. Gen. Tech. Rep. PSW-GTR-158. Albany, CA: U.S. Department of Agriculture, Forest Service, Pacific Southwest Research Station: 153-154. [26940]
32. Bitter, Nicholas Q.; Ehleringer, James R. 2021. Machine learning prediction of mortality in the common desert shrub Encelia farinosa. Ecological Informatics. 64: 1574-9541. [99122]
33. Bolling, John D.; Walker, Lawrence R. 2000. Plant and soil recovery along a series of abandoned desert roads. Journal of Arid Environments. 46(1): 1-24. [40875]
34. Bonner, James. 1950. The role of toxic substances in the interactions of higher plants. Botanical Review. 16(1): 51-65. [99328]
35. Boone, Jim. 2024. Brittlebush (Encelia farinosa), [Online]. In: Bird and Hike: Birding, hiking, and naturalizing around Las Vegas. Desert Wildlife Consultants, LLC, (Producer). Available: https://www.birdandhike.com/Veg/Species/Shrubs/Enceli_far/_Enc_far.htm. [99123]
36. Bowers, Janice E. 1994. Natural conditions for seedling emergence of three woody species in the northern Sonoran Desert. Madrono. 41(2): 73-84. [97230]
37. Bowers, Janice E. 2005. Effects of drought on shrub survival and longevity in the northern Sonoran Desert. Journal of the Torrey Botanical Society. 132(3): 421-431. [96659]
38. Bowers, Janice E. 2007. Has climatic warming altered spring flowering date of Sonoran Desert shrubs. The Southwestern Naturalist. 52(3): 347-355. [68657]
39. Bowers, Janice E.; Dimmitt, Mark A. 1994. Flowering phenology of six woody plants in the northern Sonoran Desert. Bulletin of the Torrey Botanical Club. 121(3): 215-229. [49034]
40. Bowers, Janice E.; McLaughlin, Steven P. 1987. Flora and vegetation of the Rincon Mountains, Pima County, Arizona. Desert Plants. 8(2): 50-94. [495]
41. Bowers, Janice E.; Turner, Raymond M.; Burgess, Tony L. 2004. Temporal and spatial patterns in emergence and early survival of perennial plants in the Sonoran Desert. Plant Ecology. 172(1): 107-119. [96658]
42. Bowers, Janice E.; Webb, Robert H.; Pierson, Elizabeth A. 1997. Succession of desert plants on debris flow terraces, Grand Canyon, Arizona, U.S.A. Journal of Arid Environments. 36(1): 67-86. [27546]
43. Bowers, Janice E.; Webb, Robert H.; Rondeau, Renee J. 1995. Longevity, recruitment and mortality of desert plants in Grand Canyon, Arizona, USA. Journal of Vegetation Science. 6(4): 551-564. [42371]
44. Bowers, Michael A. 1988. Plant associations on a Sonoran Desert bajada: Geographical correlates and evolutionary source pools. Vegetatio. 74(2-3): 107-112. [4408]
45. Brady, E. LeRoy. 1991. Use of native plants for roadside revegetation. In: Rangeland Technology Equipment Council, 1991 annual report. 9222-2808-MTDC. Washington, DC: U.S. Department of Agriculture, Forest Service, Technology and Development Program: 15-16. [17081]
46. Briggs, C. M.; Redak, R. A. 2016. Seed selection by the harvester ant Pogonomyrmex rugosus (Hymenoptera: Formicidae) in coastal sage scrub: Interactions with invasive plant species. Environmental Entomology. 45(4): 983-990. [97150]
47. Brooks, M. L.; Matchett, J. R. 2006. Spatial and temporal patterns of wildfires in the Mojave Desert, 1980-2004. Journal of Arid Environments. 67(Supplement): 148-164. [65283]
48. Brooks, Margaret A. 1995. Evaluating roadside revegetation in central Arizona. In: Roundy, Bruce A.; McArthur, E. Durant; Haley, Jennifer S.; Mann, David K., compilers. Proceedings: Wildland shrub and arid land restoration symposium; 1993 October 19-21; Las Vegas, NV. Gen. Tech. Rep. INT-GTR-315. Ogden, UT: U.S. Department of Agriculture, Forest Service, Intermountain Research Station: 68-73. [24827]
49. Brooks, Mathew L.; Chambers, Jeanne C. 2011. Resistance to invasion and resilience to fire in desert shrublands of North America. Rangeland Ecology & Management. 64(5): 431-438. [84701]
50. Brooks, Matthew L.; Minnich, Richard A. 2006. Southeastern deserts bioregion. In: Sugihara, Neil G.; van Wagtendonk, Jan W.; Shaffer, Kevin E.; Fites-Kaufman, Joann; Thode, Andrea E., eds. Fire in California's ecosystems. Berkeley, CA: University of California Press: 391-414. [65559]
51. Brooks, Matthew L.; Minnich, Richard A.; Matchett, John R. 2018. Southeastern deserts bioregion. In: van Wagtendonk, Jan W.; Sugihara, Neil G.; Stephens, Scott L.; Thode, Andrea E.; Shaffer, Kevin E.; Fites-Kaufman, Jo Ann, eds. Fire in California's ecosystems. 2nd ed. Oakland, CA: University of California Press: 353-378. [93914]
52. Brooks, Matthew; Lair, Bridget. 2005. Ecological effects of vehicular routes in a desert ecosystem. Report prepared for the United States Geological Survey, Recoverability and Vulnerabilty of Desert Ecosystems Program. U.S. Dept. of Interior, U.S. Geological Survey. 24 p. [99102]
53. Brown, David E., ed. 1982. Biotic communities of the American Southwest--United States and Mexico. Desert Plants. Special issue: 4(1-4): 1-342. [62041]
54. Brown, David E.; Minnich, Richard A. 1986. Fire and changes in creosote bush scrub of the western Sonoran Desert, California. The American Midland Naturalist. 116(2): 411-422. [537]
55. Burgess, Tony L.; Bowers, Janice E.; Turner, Raymond M. 1991. Exotic plants at the Desert Laboratory, Tucson, Arizona. Madrono. 38(2): 96-114. [15362]
56. Burquez, Alberto; Martinez-Yrzar, Angelina; Miller, Mark; Rojas, Karla; de los Angeles Quintana, Mara; Yetman, David. 1998. Mexican grasslands and the changing aridlands of Mexico: An overview and a case study in northwestern Mexico. In: Tellman, Barbara; Finch, Deborah M.; Edminster, Carl; Hamre, Robert, eds. The future of arid grasslands: Identifying issues, seeking solutions; 1996 October 9-13; Tucson, AZ. Proceedings RMRS-P-3. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: 21-32. [29256]
57. California Native Plant Society. 2020. Brittlebush: Encelia farinosa, [Online]. In: Calscape. Sacramento, CA: California Native Plant Society (Producer). Available: https://calscape.org/Encelia-farinosa-(Brittlebush). [2024, September 2]. [99124]
58. Cannon, William Austin. 1911. The root habits of desert plants. Washington, DC: The Carnegie Institution of Washington. 96 p. [5003]
59. Carey, Jennifer H. 1995. Cynodon dactylon. In: Fire Effects Information System, [Online]. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Fire Sciences Laboratory (Producer). Available: www.fs.usda.gov/database/feis/plants/graminoid/cyndac/all.html [2024, January 11]. [98834]
60. Castetter, Edward F. 1935. Ethnobiological studies in the American Southwest. I. Uncultivated native plants used as sources of food. Biological Series No. 4: Volume 1. Albuquerque, NM: University of New Mexico. 62 p. [35938]
61. Cave, George H.; Patten, Duncan T. 1984. Short-term vegetation responses to fire in the upper Sonoran Desert. Journal of Range Management. 37(6): 491-496. [610]
62. Cione, Nancy K.; Padgett, Pamela E.; Allen, Edith B. 2002. Restoration of a native shrubland impacted by exotic grasses, frequent fire, and nitrogen deposition in southern California. Restoration Ecology. 10(2): 376-384. [43206]
63. Clark, Curtis. 1998. Phylogeny and adaptation in the Encelia alliance (Asteraceae: Helliantheae). Aliso. 17(2): 89-98. [99125]
64. Clary, Raimond F., Jr.; Slayback, Robert D. 1983. Plant materials and establishment techniques for revegetation of California desert highways. Transportation Research Record. 969: 24-26. [99100]
65. Cleland, Elsa E.; Funk, Jennifer L.; Allen, Edith B. 2016. Coastal sage scrub. In: Mooney, Harold; Zavaleta, Erica, eds. Ecosystems of California. 1st ed. Berkeley, CA: University of California Press: 429-448. [99322]
66. Conard, Susan G.; Weise, David R. 1998. Management of fire regime, fuels, and fire effects in southern California chaparral: Lessons from the past and thoughts for the future. In: Pruden, Teresa L.; Brennan, Leonard A., eds. Fire in ecosystem management: Shifting the paradigm from suppression to prescription: Proceedings, Tall Timbers fire ecology conference; 1996 May 7-10; Boise, ID. No. 20. Tallahassee, FL: Tall Timbers Research Station: 342-350. [35654]
67. Conrad, C. Eugene. 1987. Common shrubs of chaparral and associated ecosystems of southern California. Gen. Tech. Rep. PSW-99. Berkeley, CA: U.S. Department of Agriculture, Forest Service, Pacific Southwest Forest and Range Experiment Station. 86 p. [4209]
68. Cook, Benjamin I.; Mankin, Justin S.; Anchukaitis, Kevin J. 2018. Climate change and drought: From past to future. Current Climate Change Reports. 4(2): 164-179. [99098]
69. Cooper, William Skinner. 1922. The broad-sclerophyll vegetation of California: An ecological study of the chaparral and its related communities. Publ. No. 319. Washington, DC: The Carnegie Institution of Washington. 145 p. [6716]
70. Cronquist, Arthur; Holmgren, Arthur H.; Holmgren, Noel H.; Reveal, James L.; Holmgren, Patricia K. 1994. Intermountain flora: Vascular plants of the Intermountain West, U.S.A. Vol. 5: Asterales. New York: The New York Botanical Garden. 496 p. [28653]
71. Crosswhite, Frank S.; Crosswhite, Carol D. 1984. A classification of life forms of the Sonoran Desert, with emphasis on the seed plants and their survival strategies. Desert Plants. 5(4): 131-161. [45807]
72. Cunningham, G. L.; Strain, B. R. 1969. An ecological significance of seasonal leaf variability in a desert shrub. Ecology. 50(3): 400-408. [4598]
73. D'Antonio, Carla M.; Vitousek, Peter M. 1992. Biological invasions by exotic grasses, the grass/fire cycle, and global change. Annual Review of Ecology and Systematics. 23: 63-87. [20148]
74. Daniel, Thomas F.; Butterwick, Mary L. 1992. Flora of the South Mountains of south-central Arizona. Desert Plants. 10(3): 99-119. [19896]
75. Davis, Frank W.; Stine, Peter A.; Stoms, David M. 1994. Distribution and conservation status of coastal sage scrub in southwestern California. Journal of Vegetation Science. 5(5): 743-756. [25944]
76. Dayton, William A. 1931. Important western browse plants. Misc. Publ. No. 101. Washington, DC: U.S. Department of Agriculture. 214 p. [768]
77. Diffenbaugh, Noah S.; Swain, Daniel L.; Touma, Danielle. 2015. Anthropogenic warming has increased drought risk in California. PNAS. 112(13): 3931-3936. [99093]
78. Downum, Kelsey R.; Villegas, Sergio; Rodriguez, Eloy; Keil, David J. 1989. Plant photosensitizers: A survey of their occurrence in arid and semiarid plants from North America. Journal of Chemical Ecology. 15(1): 345-355. [7658]
79. Drennan, Philippa M.; Nobel, Park S. 1997. Frequencies of major C3, C4, and CAM perennials on different slopes in the northwestern Sonoran Desert. Flora. 192(3): 297-304. [99329]
80. Drezner, T. D. 2006. Plant facilitation in extreme environments: The non-random distribution of saguaro cacti (Carnegiea gigantea) under their nurse associates and the relationship to nurse architecture. Journal of Arid Environments. 65(1): 46-61. [61810]
81. Edwards, Fred; Bainbridge, David; Zink, Thomas; Allen, M. 2000. Rainfall catchments improve survival of container transplants at Mojave Desert site. Ecological Restoration. 18(2): 100-103. [99101]
82. Ehleringer, J. R.; Bjorkman, O. 1978. Pubescence and leaf spectral characteristics in a desert shrub, Encelia farinosa. Oecologia. 36(2): 151-162. [99126]
83. Ehleringer, James R. 1984. Intraspecific competitive effects on water relations, growth and reproduction in Encelia farinosa. Oecologia. 63(2): 153-158. [99127]
84. Ehleringer, James R. 1988. Comparative ecophysiology of Encelia farinosa and Encelia frutescens: I. energy balance considerations. Oecologia. 76(4): 553-561. [99130]
85. Ehleringer, James R.; Clark, Curtis. 1988. Evolution and adaptation in Encelia (Asteraceae). In: Gottlieb, L. D.; Jain, S. K. (eds). Plant Evolutionary Biology. Dordrecht, Netherlands: Springer: 221-248. [99129]
86. Ehleringer, James R.; Sandquist, Darren R. 2018. A tale of ENSO, PDO, and increasing aridity impacts on drought-deciduous shrubs in the Death Valley region. Oecologia. 187(4): 879-895. [96992]
87. Ehlerlinger, James R.; Cook, Craig S. 1984. Photosynthesis in Encelia farinosa gray in response to decreasing leaf water potential. Plant Physiology. 75(3): 688-93. [99128]
88. Esque, Todd C.; Schwalbe, Cecil R. 2002. Alien annual grasses and their relationships to fire and biotic change in Sonoran Desert scrub. In: Tellman, Barbara, ed. Invasive exotic species in the Sonoran region. Arizona-Sonora Desert Museum Studies in Natural History. Tucson, AZ: University of Arizona Press; Arizona-Sonora Desert Museum: 165-194. [48660]
89. Esque, Todd C.; Schwalbe, Cecil R.; Haines, Dustin F.; Halvorson, William L. 2004. Saguaros under siege: Invasive species and fire. Desert Plants. 20(1): 49-55. [91075]
90. Esque, Todd C.; Webb, Robert H.; Wallace, Cynthia S. A.; van Riper, Charles, III; McCreedy, Chris; Smythe, Lindsay. 2013. Desert fires fueled by native annual forbs: Effects of fire on communities of plants and birds in the lower Sonoran Desert of Arizona. The Southwestern Naturalist. 58(2): 223-233. [87384]
91. Everett, Percy C. 1957. A summary of the culture of California plants at the Rancho Santa Ana Botanic Garden, 1927-1950. Claremont, CA: The Rancho Santa Ana Botanic Garden. 223 p. [7191]
92. Everett, Percy C. 2012. A second summary of the culture of California plants at the Rancho Santa Ana Botanic Garden, 1950-1970. Version 1.1. Claremont, CA: The Rancho Santa Ana Botanic Garden. 541 p. [94252]
93. Federal Geographic Data Committee, Vegetation Subcommittee. 2021. U.S. National Vegetation Classification, [Online]. Vers. 2.03. Washington, DC: Federal Geographic Data Committee, (Producer). Available: http://usnvc.org/ [2024, Jan 20]. [90140]
94. Flora of North America Editorial Committee, eds. 2024. Flora of North America north of Mexico, [Online]. Flora of North America Association (Producer). Available: http://www.efloras.org/flora_page.aspx?flora_id=1. [36990]
95. Forseth, I. N.; Ehleringer, J. R.; Werk, K. S.; Cook, C. S. 1984. Field water relations of Sonoran Desert annuals. Ecology. 65(5): 1436-1444. [77996]
96. Franklin, Kim A.; Lyons, Kelly; Nagler, Pamela L.; Lampkin, Derrick; Glenn, Edward P.; Molina-Freaner, Francisco; Markow, Therese; Huete, Alfredo R. 2006. Buffelgrass (Pennisetum ciliare) land conversion and productivity in the plains of Sonora, Mexico. Biological Conservation. 127(1): 62-71. [55821]
97. Gallaher, T. J.; Brock, K.; Imada C. T.; Imada, K.; Walvoord, N. 2020. Encelia farinosa, [Online]. In: Plants of Hawaii. Honolulu, Hawaii: Bishop Museum (Producer). Available: http://www.plantsofhawaii.org. Accessed: [2024, April 25]. [99131]
98. Godinez-Alvarez, Hector; Valverde, Teresa; Ortega-Baes, Pablo. 2003. Demographic trends in the Cactaceae. The Botanical Review. 69(2): 173-203. [47043]
99. Goldberg, Deborah E.; Turner, Raymond M. 1986. Vegetation change and plant demography in permanent plots in the Sonoran Desert. Ecology. 67(3): 695-712. [4410]
100. Gormally, Josh. 2011. Plants found along the effluent dominated stretch of the Middle Santa Cruz River. Desert Plants. 27(1): 19-31. [83580]
101. Goss, Michael; Swain, Daniel L.; Abatzoglou, John T.; Sarhadi, Ali; Kolden, Crystal A.; Williams, A. Park; Diffenbaugh, Noah S. 2020. Climate change is increasing the likelihood of extreme autumn wildfire conditions across California. Environmental Research Letters. 15(9): 094016. [94595]
102. Gray, Miranda E.; Dickson, Brett G.; Zachmann, Luke J. 2014. Modelling and mapping dynamic variability in large fire probability in the lower Sonoran Desert of south-western Arizona. International Journal of Wildland Fire. 23(8): 1108-1118. [88832]
103. Gray, Reed; Bonner, James. 1948. An inhibitor of plant growth from the leaves of Encelia farinosa. American Journal of Botany. 35(1): 52-57. [99327]
104. Guida, Ross J.; Abella, Scott R.; Smith, William J.; Stephen, Haroon; Roberts, Chris L. 2014. Climate change and desert vegetation distribution: Assessing thirty years of change in southern Nevada's Mojave Desert. Professional Geographer. 66(2): 311-322. [98840]
105. Hassler, Michael. 2024. Encelia farinosa. In: World Plants (1994 - 2024): Synonymic checklist and distribution of the world flora [Online]. Version 19.2. Bruchsal, Germany: World Plants (Producer). Available: www.worldplants.de. [2024, April 17]. [99132]
106. Herbel, Carlton H. 1994. SRM 508: Creosotebush-tarbush. In: Shiflet, Thomas N., ed. Rangeland cover types of the United States. Denver, CO: Society for Range Management: 67-68. [67048]
107. Hoerling, Martin P.; Dettinger, Michael; Wolter, Klaus; Lukas, Jeff; Eischeid, Jon; Nemani, Rama; Liebmann, Brant; Kunkel, Kenneth E. 2013. Present weather and climate: Evolving conditions. In: Garfin, Gregg; Jardine, Angela; Merideth, Robert; Black, Mary; LeRoy, Sarah, eds. Assessment of climate change in the southwest United States. National Climate Assessment 2013 Report. Washington, DC: Island Press: 74-96. [97414]
108. Holland, Robert F. 1986. Preliminary descriptions of the terrestrial natural communities of California. Sacramento, CA: California Department of Fish and Game. 156 p. [12756]
109. Holzman, Barbara. 1994. SRM 211: Creosote bush scrub. In: Shiflet, Thomas N., ed. Rangeland cover types of the United States. Denver, CO: Society for Range Management: 20-21. [66670]
110. Horn, Kevin J.; Wilkinson, Joseph; White, Steven; St. Clair, Samuel B. 2015. Desert wildfire impacts on plant community function. Plant Ecology. 216(12): 1623-1634. [90139]
111. Horton, Jerome S. 1960. Vegetation types of the San Bernardino Mountains. Tech. Pap. No. 44. Berkeley, CA: U.S. Department of Agriculture, Forest Service, Pacific Southwest Forest and Range Experiment Station. 29 p. [10687]
112. Hosna, Rachel Kathryn. 2020. Legacies of fire and presence of woody vegetation on soil seed banks of four North American desert systems. Las Cruces, NM: New Mexico State University. 57 p. Thesis. [98744]
113. Howard, Janet L. 1992. Erodium cicutarium. In: Fire Effects Information System, [Online]. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Fire Sciences Laboratory (Producer). Available: www.fs.usda.gov/database/feis/plants/forb/erocic/all.html [2024, January 11]. [98842]
114. Ibarra, Fernando A.; Martin, Martha H.; Torres, L. Ricardo; Silva, Martin F.; Morton, Howard L.; Cox, Jerry R. 1986. The brittlebush problem and potential control measures in buffelgrass pastures in Sonora, Mexico. In: Proceedings, Western Society of Weed Science; 1986 March 18-20; San Diego, CA. Vol. 39. Westminster, CO: Western Society of Weed Science: 57-66. [21947]
115. Ibarra-F, Fernando; Martin-R, M.; Cox, J. R.; Miranda-Z, H. 1996. The effect of prescribed burning to control brush species on buffelgrass pastures in Sonora, Mexico. In: Ffolliott, Peter F.; DeBano, Leonard F.; Baker, Malchus B., Jr.; Gottfried, Gerald J.; Solis-Garza, Gilberto; Edminster, Carleton B.; Neary, Daniel G.; Allen, Larry S.; Hamre, R. H., tech. coords. Effects of fire on Madrean Province ecosystems: A symposium proceedings; 1996 March 11-15; Tucson, AZ. Gen. Tech. Rep. RM-GTR-289. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station: 195-204. [28078]
116. Innes, Robin J. 2022. Pennisetum ciliare, buffelgrass. In: Fire Effects Information System, [Online]. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Missoula Fire Sciences Laboratory (Producer). Available: www.fs.usda.gov/database/feis/plants/graminoid/pencil/all.html. [98237]
117. Innes, Robin J. 2023. Brassica tournefortii, Sahara mustard. In: Fire Effects Information System, [Online]. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Missoula Fire Sciences Laboratory (Producer). Available: www.fs.usda.gov/database/feis/plants/forb/bratou/all.html. [98218]
118. Karpiscak, Martin M. 1980. Secondary succession of abandoned field vegetation in southern Arizona. Tucson, AZ: University of Arizona. 219 p. Dissertation. [53608]
119. Kartesz, John Thomas. 1988. A flora of Nevada. Reno, NV: University of Nevada. 1729 p. Dissertation. [42426]
120. Kay, Burgess L.; Graves, Walter L.; Young, James A. 1988. Long-term storage of desert shrub seed. Mojave Revegetation Notes. No. 23: 1-12. [98845]
121. Kearney, Thomas H.; Peebles, Robert H.; Howell, John Thomas; McClintock, Elizabeth. 1960. Arizona flora. 2nd ed. Berkeley, CA: University of California Press. 1085 p. [6563]
122. Keeler-Wolf, Todd. 2007. Mojave Desert scrub vegetation. In: Barbour, Michael G.; Keeler-Wolf, Todd; Schoenherr, Allan A., eds. Terrestrial vegetation of California. Berkeley, CA: University of California Press: 609-625. [82716]
123. Keeley, Jon E. 1995. Future of California floristics and systematics: Wildfire threats to the California flora. Madrono. 42(2): 175-179. [50362]
124. Keeley, Jon E.; Davis, Frank W. 2007. Chaparral. In: Barbour, Michael G.; Keeler-Wolf, Todd; Schoenherr, Allan A., eds. Terrestrial vegetation of California. Berkeley, CA: University of California Press: 339-366. [82706]
125. Keeley, Jon E.; Fotheringham, C. J.; Baer-Keeley, Melanie. 2006. Demographic patterns of postfire regeneration in mediterranean-climate shrublands of California. Ecological Monographs. 76(2): 235-255. [63206]
126. Keeley, Jon E.; Fotheringham, C. J.; Rundel, Philip W. 2012. Postfire chaparral regeneration under mediterranean and non-mediterranean climates. Madrono. 59(3): 109-127. [86241]
127. Keeley, Jon E.; Keeley, Sterling C. 1988. Chaparral. In: Barbour, Michael G.; Billings, William Dwight, eds. North American terrestrial vegetation. New York: Cambridge University Press: 165-207. [19545]
128. Keeley, Jon; Witter, Martha; Taylor, Robert. 2004. Challenges of managing fires along an urban-wildland interface ─ lessons from the Santa Monica Mountains, Los Angeles, California. In: The 3rd International Wildland Fire Conference; 2003 October 3-6; Sydney, Australia. Ulaanbaatar, Mongolia: Global Fire Modeling Center, Fire Ecology Research Group: 1-10. [99324]
129. Kemp, Paul R. 1989. Seed banks and vegetation processes in deserts. In: Leck, Mary Allessio; Parker, V. Thomas; Simpson, Robert L., eds. Ecology of soil seed banks. San Diego, CA: Academic Press: 257-281. [60865]
130. Kirkpatrick, J. B.; Hutchinson, C. F. 1977. The community composition of Californian coastal sage scrub. Vegetatio. 35(1): 21-33. [5612]
131. Klinger, Rob; Brooks, Matt. 2017. Alternative pathways to landscape transformation: Invasive grasses, burn severity and fire frequency in arid ecosystems. Journal of Ecology. 105(6): 1521-1533. [92303]
132. Klinger, Rob; Underwood, Emma C.; McKinley, Randy; Brooks, Matthew L. 2021. Contrasting geographic patterns of ignition probability and burn severity in the Mojave Desert. Frontiers in Ecology and Evolution. 9: 593167. [95796]
133. Koehler, Catherine E.; Montalvo, Arlee M. 2004. Encelia farinosa. In: Francis, John K., ed. Wildland shrubs of the United States and its territories: Thamnic descriptions. Volume 1. Gen. Tech. Rep. IITF-GTR-26. San Juan, PR: U.S. Department of Agriculture, Forest Service, International Institute of Tropical Forestry; Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: 302-306. [52153]
134. Krausman, Paul R.; Ordway, Leonard L.; Whiting, Frank M.; Brown, William H. 1990. Nutritional composition of desert mule deer forage in the Picacho Mountains, Arizona. Desert Plants. 10(1): 32-34. [7259]
135. Kuchler, A. W. 1964. Creosote bush-bur-sage (Larrea-Franseria). In: Kuchler, A. W. Manual to accompany the map of potential vegetation of the conterminous United States. Special Publication No. 36. New York: American Geographical Society: 43. [67338]
136. Kyhos, Donald W. 1967. Natural hybridization between Encelia and Geraea (Compositae) and some related experimental investigations. Madrono. 19(2): 33-43. [99133]
137. Kyhos, Donald W. 1971. Evidence of different adaptations of flower color variants of Encelia farinosa (Compositae). Madrono. 21(2): 49-61. [99134]
138. LANDFIRE. 2020. Biophysical settings models and descriptions, [Online]. Washington, DC: U.S. Department of Agriculture, Forest Service; U.S. Department of the Interior; U.S. Geological Survey; Arlington, VA: The Nature Conservancy, (Producers). Available: http://www.landfirereview.org/search.php [2024, February 2]. [96496]
139. Leitner, Lawrence A. 1987. Plant communities of a large arroyo at Punta Cirio, Sonora. The Southwestern Naturalist. 32(1): 21-28. [1439]
140. Levin, Geoffrey A. 1988. How plants survive in the desert. Environment Southwest. Summer: 20-25. [9239]
141. Loftin, Samuel Robert. 1987. Postfire dynamics of a Sonoran Desert ecosystem. Tempe, AZ: Arizona State University. 97 p. Thesis. [12296]
142. Lovich, Jeffrey E.; Ennen, Joshua R.; Madrak, Sheila V.; Loughran, Caleb L.; Meyer, Katherin P.; Arundel, Terence R.; Bjurlin, Curtis D. 2011. Long-term post-fire effects on spatial ecology and reproductive output of female Agassiz's desert tortoises (Gopherus agassizii) at a wind energy facility near Palm Springs, California, USA. Fire Ecology. 7(3): 75-87. [85188]
143. Lowe, Charles H. 1964. Arizona's natural environment: Landscapes and habitats. Tucson, AZ: The University of Arizona Press. 136 p. [20736]
144. Ludwig, J. A.; Cunningham, G. L.; Whitson, P. D. 1988. Distribution of annual plants in North American deserts. Journal of Arid Environments. 15(3): 221-227. [6656]
145. Lybbert; Andrew H.; Taylor, Justin; DeFranco, Alyssa; St. Clair, Samuel B. 2017. Reproductive success of wind, generalist, and specialist pollinated plant species following wildfire in desert landscapes. International Journal of Wildland Fire. 26(12): 1030-1039. [93080]
146. MacMahon, James A. 1988. Warm deserts. In: Barbour, Michael G.; Billings, William Dwight, eds. North American terrestrial vegetation. New York: Cambridge University Press: 231-264. [19547]
147. MacMahon, James A.; Wagner, Frederic H. 1985. The Mojave, Sonoran and Chihuahuan Deserts of North America. In: Evenari, Michael; Noy-Meir, Imanuel; Goodall, David W. eds. 12A: Hot deserts and arid shrublands. Ecosystems of the world. Amsterdam: Elsevier: 105-201. [95987]
148. Marks, John Brady. 1950. Vegetation and soil relations in the lower Colorado Desert. Ecology. 31(2): 176-193. [44004]
149. Marshal, Jason P.; Bleich, Vernon C.; Andrew, Nancy G.; Krausman, Paul R. 2004. Seasonal forage use by desert mule deer in southeastern California. The Southwestern Naturalist. 49(4): 501-505. [51969]
150. Martin, Alexander C.; Zim, Herbert S.; Nelson, Arnold L. 1951. American wildlife and plants. New York: McGraw-Hill. 500 p. [4021]
151. Martin, Bradford D. 1984. Influence of slope aspect on postfire reproduction of Encelia farinosa (Asteraceae). Madrono. 31(3): 187-189. [4936]
152. Martin-Rivera, Martha; Ibarra-Flores, Fernando; Guthery, Fred S.; Kublesky, William P., Jr.; Camou-Luders, Gustavo; Fimbres-Preciado, Jesus; Johnson-Gordon, Donald. 2001. Habitat improvement for wildlife in north-central Sonora, Mexico. In: McArthur, E. Durant; Fairbanks, Daniel J., comps. Shrubland ecosystem genetics and biodiversity: Proceedings. 2000 June 13-15; Provo, UT. Proc. RMRS-P-21. Ogden, UT: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: 356-360. [96423]
153. Mazurek, Ellen Jill. 1981. Effects of fire on small mammals and vegetation in the Upper Sonoran Desert. Tempe, AZ: Arizona State University. 88 p. Thesis. [55763]
154. Mcauliffe, Joseph R.; Van Devender, Thomas R. 1998. A 22,000-year record of vegetation change in the north-central Sonoran Desert. Palaeogeography, Palaeoclimatology, Palaeoecology. 141(3): 253-275. [97250]
155. McDonald, C. J.; McPherson, G. R. 2011. Fire behavior characteristics of buffelgrass-fueled fires and native plant community composition in invaded patches. Journal of Arid Environments. 75(11): 1147-1154. [83896]
156. McDonald, Christopher J.; McPherson, Guy R. 2013. Creating hotter fires in the Sonoran Desert: Buffelgrass produces copious fuels and high fire temperatures. Fire Ecology. 9(2): 26-39. [87751]
157. McGinnies, William G. 1983. Flowering periods for common desert plants: Southwestern Arizona. Tucson, AZ: University of Arizona, College of Agriculture, Office of Arid Lands. 1 p. [96404]
158. McLaughlin, Steven P.; Bowers, Janice E. 1982. Effects of wildfire on a Sonoran Desert plant community. Ecology. 63(1): 246-248. [1619]
159. Minnich, Richard A. 1983. Fire mosaics in southern California and northern Baja California. Science. 219(4590): 1287-1294. [4631]
160. Minnich, Richard A. 1995. Postfire succession in desertscrub communities of southern California. In: Fletcher-Jones, Anne, ed. The Desert Tortoise Council: Proceedings of 1994 symposium; 1994. San Bernardino, CA: Desert Tortoise Council: 93-112. [97152]
161. Minnich, Richard A. 1996. Fire ecology of exotic grasses in the California desert. In: Lovich, Jeff; Randall, John; Kelly, Mike, eds. Proceedings, California Exotic Pest Plant Council: Symposium '95; 1995 October 6-8; Pacific Grove, CA. Berkeley, CA: California Exotic Pest Plant Council: 61. [44031]
162. Minnich, Richard A.; Dezzani, Raymond J. 1998. Historical decline of coastal sage scrub in the Riverside-Perris Plain, California. Western Birds. 29(4): 366-391. [97183]
163. Mirov, N. T.; Kraebel, C. J. 1937. Collecting and propagating the seeds of California wild plants. Res. Note No. 18. Berkeley, CA: U.S. Department of Agriculture, Forest Service, California Forest and Range Experiment Station. 27 p. [9787]
164. Mirov, N. T.; Kraebel, Charles J.;. 1939. Collecting and handling seeds of wild plants. Civilian Conservation Corps: Forestry Publication No. 5. Washington DC: U.S. Government Printing Office. 52 p. [99109]
165. Montalvo, A. M.; Koehler, C. E.; Beyers, J. L. 2010. Plant profile for Encelia farinosa, [Online]. In: Native plant recommendations for southern California ecoregions. Riverside, CA: Riverside-Corona Resource Conservation District and U.S. Department of Agriculture, Forest Service, Pacific Southwest Research Station (Producer). Available: https://www.rcrcd.org/plant-profiles. [99135]
166. Morton, Howard L. 1994. SRM 507: Palo verde - cactus. In: Shiflet, Thomas N., ed. Rangeland cover types of the United States. Denver, CO: Society for Range Management: 66-67. [67047]
167. Muller, Cornelius H. 1953. The association of desert annuals with shrubs. American Journal of Botany. 40(2): 1-52. [4650]
168. Munz, Philip A. 1974. A flora of southern California. Berkeley, CA: University of California Press. 1086 p. [4924]
169. Myers, Marlyce A.; Ellestrand, Norman C. 1986. Post-fire succession at an inland (Riversidian) site of coastal sage scrub: Variation in community response. In: DeVries, Johannes J., ed. Proceedings of the chaparral ecosystems research conference; 1985 May 16-17; Santa Barbara, CA. Report No. 2. Davis, CA: University of California, California Water Resources Center: 129-132. [4833]
170. National Park Service. 2021. Resource stewardship strategy summary: Joshua Tree National Park, California. Twentynine Palms, CA: U.S. Department of the Interior, National Park Service, Joshua Tree National Park. 115 p. [99334]
171. Native American Ethnobotany Database. 2024. Encelia farinosa [Online]. Available: http://naeb.brit.org/about. [99153]
172. NatureServe. 2009. International ecological classification standard: Terrestrial ecological classifications. In: NatureServe Central Databases. Arlington, VA: NatureServe. 1172 p. Available online: http://downloads.natureserve.org/get_data/data_sets/veg_data/nsDescriptions.pdf [2024, July 11]. [94380]
173. NatureServe. 2024. NatureServe Explorer, [Online]. Arlington, VA: NatureServe (Producer). Available: https://explorer.natureserve.org. [94379]
174. Neil, Kaesha. 2008. Effects of urbanization on flowering phenology in the metropolitan phoenix region of USA: Findings from herbarium records. Tempe, AZ: Arizona State University. 117 p. Dissertation. [98852]
175. Neil, Kaesha; Wu, Jianguo; Bang, Christofer; Faeth, Stanley. 2014. Urbanization affects plant flowering phenology and pollinator community: Effects of water availability and land cover. Ecological Processes. 3(1): 17. [99136]
176. Nichol, A. A. 1952. [Revised]. The natural vegetation of Arizona. Tech. Bull. 68. Tucson, AZ: University of Arizona, Agricultural Experiment Station: 189-230. [Revisions by W. S. Phillips]. [3928]
177. Niering, W. A.; Whittaker, R. H.; Lowe, C. H. 1963. The saguaro: A population in relation to environment. Science. 142(3588): 15-23. [5093]
178. Niering, William A.; Lowe, Charles H. 1984. Vegetation of the Santa Catalina Mountains: Community types and dynamics. Vegetatio. 58(1): 3-28. [12037]
179. Nilsen, Erik T.; Sharifi, M. R.; Rundel, P. W. 1987. Leaf dynamics in an evergreen and a deciduous species with even-aged leaf cohorts, from different environments. The American Midland Naturalist. 118(1): 46-55. [63901]
180. Nilsen, Erik Tallak; Sharifi, M. Rasoul; Rundel, Philip W. 1984. Comparative water relations of phreatophytes in the Sonoran Desert of California. Ecology. 65(3): 767-778. [49012]
181. Nobel, Park S. 1997. Root distribution and seasonal production in the northwestern Sonoran Desert for a C3 subshrub, a C4 bunchgrass, and a CAM leaf succulent. American Journal of Botany. 84(7): 949-955. [99137]
182. Nobel, Park S.; Linton, Matthew J. 1997. Frequencies, microclimate and root properties for three codominant perennials in the northwestern Sonoran Desert on north- vs. south-facing slopes. Annals of Botany. 80(6): 731-739. [99138]
183. Nussear, K. E.; Tuberville, T. D. 2014. Habitat Characteristics of North American Tortoises. In: Rostal, D. C.; McCoy, E. D.; Mushinsky, H. R., eds. Biology and Conservation of North American Tortoises. Baltimore, MD: Johns Hopkins University Press: 77-84. [99103]
184. Nussear, Kenneth E. 2004. Mechanistic investigation of the distributional limits of the desert tortoise Gopherus agassizii. Reno, NV: University of Nevada. 24 p. Dissertation. [99105]
185. O'Leary, John F.; Minnich, Richard A. 1981. Postfire recovery of creosote bush scrub vegetation in the western Colorado Desert. Madrono. 28(2): 61-66. [3973]
186. Odening, Walter R.; Strain, B. R.; Oechel, W. C. 1974. The effect of decreasing water potential on net CO2 exchange of intact desert shrubs. Ecology. 55(5): 1086-1095. [17775]
187. Olsson, Aaryn D.; Betancourt, Julio; McClaran, Mitchel P.; Marsh, Stuart E. 2012. Sonoran Desert ecosystem transformation by a C4 grass without the grass/fire cycle. Diversity and Distributions. 18(1): 10-21. [95472]
188. Overpeck, Jonathan T.; Udall, Bradley. 2020. Climate change and the aridification of North America. PNAS. 117(22): 11856-11858. [99094]
189. Padgett, Pamela E.; Allen, Edith B. 1999. Differential responses to nitrogen fertilization in native shrubs and exotic annuals common to mediterranean coastal sage scrub of California. Plant Ecology. 144(1): 93-101. [63333]
190. Padgett, Pamela E.; Vazquez, Lucia; Allen, Edith B. 1999. Seed viability and germination behavior of the desert shrub Encelia farinosa Torrey and A. Gray (Compositae). Madrono. 46(3): 126-133. [36367]
191. Paine, T. D.; Redak, R. A.; Trumble, J. T. 1993. Impact of acidic deposition on Encelia farinosa gray (Compositae: Asteraceae) and feeding preferences of Trirhabda geminata horn (Coleoptera: Chrysomelidae). Journal of Chemical Ecology. 19(1): 97-105. [99106]
192. Parker, Kathleen C. 1988. Environmental relationships and vegetation associates of columnar cacti in the northern Sonoran Desert. Vegetatio. 78(3): 125-140. [6953]
193. Parker, Kathleen C. 1989. Nurse plant relationships of columnar cacti in Arizona. Physical Geography. 10(4): 322-335. [95474]
194. Parker, Kathleen C. 1991. Topography, substrate, and vegetation patterns in the northern Sonoran Desert. Journal of Biogeography. 18(2): 151-163. [14979]
195. Pechony, O.; Shindell, D. T. 2010. Driving forces of global wildfires over the past millennium and the forthcoming century. Proceedings of the National Academy of Sciences. 107(45): 19167-19170 [+ suppl.]. [90856]
196. Pendleton, Rosemary L.; Pendleton, Burton K.; Harper, Kimball T. 1989. Breeding systems of woody plant species in Utah. In: Wallace, Arthur; McArthur, E. Durant; Haferkamp, Marshall R., comps. Proceedings--symposium on shrub ecophysiology and biotechnology; 1987 June 30 - July 2; Logan, UT. Gen. Tech. Rep. INT-256. Ogden, UT: U.S. Department of Agriculture, Forest Service, Intermountain Research Station: 5-22. [5918]
197. Peterson, Eric B. 2008. International vegetation classification alliances and associations occurring in Nevada with proposed additions. Carson City, NV: Nevada Natural Heritage Program. 347 p. [77864]
198. Petrie, M. D.; Collins, S. L.; Gutzler, D. S.; Moore, D. M. 2014. Regional trends and local variability in monsoon precipitation in the northern Chihuahuan Desert, USA. Journal of Arid Environments. 103: 63-70. [98820]
199. Phillips, Barbara G. 1997. History of fire and fire impacts at Tonto National Monument, Arizona. Technical Report No. 59. Tucson, AZ: University of Arizona. 79 p. [95476]
200. Phillips, Donald L.; MacMahon, James A. 1978. Gradient analysis of a Sonoran Desert Bajada. The Southwestern Association of Naturalist. 23(4): 669-679. [97361]
201. Piao, Shilong; Liu, Qiang; Chen, Anping; Janssens, Ivan A.; Fu, Yongshuo; Dai, Junhu; Liu, Lingli; Lian, Xu; Shen, Miaogen; Zhu, Xiaolin. 2019. Plant phenology and global climate change: Current progresses and challenges. Global Change Biology. 25(6): 1922-1940. [98871]
202. Pockman, William T.; Sperry, John S. 2000. Vulnerability to xylem cavitation and the distribution of Sonoran Desert vegetation. American Journal of Botany. 87(9): 1287-1299. [35960]
203. Polley, Wayne H.; Briske, David D.; Morgan, Jack A.; Wolter, Klaus; Bailey, Derek W.; Brown, Joel R. 2013. Climate change and North American rangelands: Trends, projections, and implications. Rangeland Ecology & Management. 66(5): 512-528. [98950]
204. POWO. 2024. Plants of the world online, [Online]. Facilitated by the Royal Botanic Gardens, Kew, (Producer). Available: http://www.plantsoftheworldonline.org/. [97319]
205. Raunkiaer, C. 1934. The life forms of plants and statistical plant geography. Oxford, England: Clarendon Press. 632 p. [2843]
206. Rautenstrauch, Kurt R.; Krausman, Paul R.; Whiting, Frank M.; Brown, William H. 1988. Nutritional quality of desert mule deer forage in King Valley, Arizona. Desert Plants. 8(4): 172-174. [2768]
207. Richardson, Bland Z.; Monsen, Stephen B.; Bowers, Diane M. 1986. Interseeding selected shrubs and herbs on mine disturbances in southeastern Idaho. In: McArthur, E. Durant; Welch, Bruce L., compilers. Proceedings--symposium on the biology of Artemisia and Chrysothamnus; 1984 July 9-13; Provo, UT. Gen. Tech. Rep. INT-200. Ogden, UT: U.S. Department of Agriculture, Forest Service, Intermountain Research Station: 134-139. [1973]
208. Rodgers, Jane E.; Miller, Carol. 2008. Encelia Adans.: Encelia. In: Bonner, Franklin T.; Karrfalt, Robert P., eds. Woody plant seed manual. Agric. Handbook No. 727. Washington, DC: U.S. Department of Agriculture, Forest Service: 488-489. [79215]
209. Rogers, G. F.; Vint, M. K. 1987. Winter precipitation and fire in the Sonoran Desert. Journal of Arid Environments. 13(1): 47-52. [5547]
210. Rogers, Garry F.; Steele, Jeff. 1980. Sonoran Desert fire ecology. In: Stokes, Marvin A.; Dieterich, John H., technical coordinators. Proceedings of the fire history workshop; 1980 October 20-24; Tucson, AZ. Gen. Tech. Rep. RM-81. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station: 15-19. [16036]
211. Rundel, Philip W. 2007. Sage scrub. In: Barbour, Michael G.; Keeler-Wolf, Todd; Schoenherr, Allan A., eds. Terrestrial vegetation of California. Berkeley, CA: University of California Press: 208-228. [82701]
212. Schmid, Mary K.; Rogers, Garry F. 1988. Trends in fire occurrence in the Arizona upland subdivision of the Sonoran Desert, 1955 to 1983. The Southwestern Naturalist. 33(4): 437-444. [6103]
213. Schoenherr, Allan A.; Burk, Jack H. 2007. Colorado desert vegetation. In: Barbour, Michael G.; Keeler-Wolf, Todd; Schoenherr, Allan A., eds. Terrestrial vegetation of California. Berkeley, CA: University of California Press: 657-682. [82717]
214. Schuster, William S. F.; Sandquist, Darren R.; Phillips, Susan L.; Ehleringer, James R. 1992. Comparisons of carbon isotope discrimination in populations of aridland plant species differing in lifespan. Oecologia. 91(3): 332-337. [99139]
215. Seegmiller, Rick F.; Krausman, Paul R.; Brown, William H.; Whiting, Frank M. 1990. Nutritional composition of desert bighorn sheep forage in the Harquahala Mountains, Arizona. Desert Plants. 10(2): 87-90. [11943]
216. Sharp Bowman, Tiffanny R.; McMillan, Brock R.; St. Clair, Samuel, B. 2017. Rodent herbivory differentially affects mortality rates of 14 native plant species with contrasting life history and growth form traits. Oecologia. 185(3): 465-473. [94839]
217. Shiflet, Thomas N., ed. 1994. Rangeland cover types of the United States. Denver, CO: Society for Range Management. 152 p. [23362]
218. Shreve, Edith B. 1923. Seasonal changes in the water relations of desert plants. Ecology. 4(3): 266-292. [99140]
219. Shreve, Edith B. 1924. Factors governing seasonal changes in transpiration of Encelia farinosa. Botanical Gazette. 77(4): 432-439. [99141]
220. Shreve, Forrest; Hinckley, Arthur L. 1937. Thirty years of change in desert vegetation. Ecology. 18(4): 463-478. [4574]
221. Shreve, Forrest; Wiggins, Ira L. 1964. Vegetation and flora of the Sonoran Desert. Stanford, CA: Stanford University Press. Vol. 1-2. 1740 p. [21016]
222. Shryock, Daniel F.; Esque, Todd C.; Chen, Felicia C. 2015. A 30-year chronosequence of burned areas in Arizona - effects of wildfires on vegetation in Sonoran Desert tortoise (Gopherus morafkai) habitats. Open-File Report 2015-1060. Reston, VA: U.S. Geological Survey. 61 p. [97280]
223. Shryock, Daniel F.; Esque, Todd C.; Chen, Felicia C. 2015. Topography and climate are more important drivers of long-term, post-fire vegetation assembly than time-since-fire in the Sonoran Desert, US. Journal of Vegetation Science. 26(6): 1134-1147. [92517]
224. Simonin, Devin A. 2001. Bromus madritensis, Bromus rubens. In: Fire Effects Information System, [Online]. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Fire Sciences Laboratory (Producer). Available: www.fs.usda.gov /database/feis/plants/graminoid/brospp/all.html [2024, January 12]. [98867]
225. Simpson, Beryl B. 1977. Breeding systems of dominant perennial plants of two disjunct warm desert ecosystems. Oecologia. 27(3): 203-226. [96409]
226. Singhal, Sonal; Roddy, Adam B.; DiVittorio, Christopher; Sanchez-Amaya, Ary; Henriquez, Claudia L.; Brodersen, Craig R.; Fehlberg, Shannon; Zapata, Felipe. 2021. Diversification, disparification and hybridization in the desert shrubs Encelia. The New Phytologist. 230(3): 1228-1241. [99142]
227. Smith, Jane Kapler, compiler. 2008. Research Project Summary: Effects of June fires in shrubland converted to buffelgrass pasture in Sonora, Mexico. In: Fire Effects Information System, [Online]. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Fire Sciences Laboratory (Producer). Available: https://www.fs.usda.gov/database/feis/research_project_summaries/Ibarra-F96/all.html. [70145]
228. Smith, Stanley D.; Monson, Russell K.; Anderson, Jay E. 1997. Drought-deciduous shrubs. In: Physiological ecology of North American desert plants. Adaptations of desert organisms. Heidelberg, Germany: Springer-Verlag Berlin: 109-123. [99152]
229. Sorensen, Mary A.; Bethke, James A.; Redak, Richard A. 2010. Potential host plants of Trirhabda geminata (Coleoptera: Chrysomelidae): Impacts on survival, development, and feeding. Environmental entomology. 39(1): 159-163. [99108]
230. Spalding, Volney M. 1909. Distribution and movements of desert plants. Publication No. 113. Washington, D.C.: Carnegie Institution of Washington. 243 p. [95491]
231. St. John, Harold. 1973. List and summary of the flowering plants in the Hawaiian Islands. Hong Kong: Cathay Press Limited. 519 p. [25354]
232. Stamp, Nancy E.; Ohmart, Robert D. 1978. Resource utilization by desert rodents in the lower Sonoran Desert. Ecology. 59(4): 700-707. [49025]
233. Steenbergh, Warren F.; Lowe, Charles H. 1969. Critical factors during the first years of the saguaro (Cereus giganteus) at Saguaro National Monument, Arizona. Ecology. 50(5): 825-834. [19692]
234. Steers, Robert J.; Allen, Edith B. 2011. Fire effects on perennial vegetation in the western Colorado Desert, USA. Fire Ecology. 7(3): 59-74. [85189]
235. Steers, Robert J.; Allen, Edith B. 2012. Impact of recurrent fire on annual plants: A case study from the western edge of the Colorado Desert. Madrono. 59(1): 14-24. [86112]
236. Steers, Robert Jeremy. 2008. Invasive plants, fire succession, and restoration of creosote bush scrub in southern California. Riverside, CA: University of California Riverside. 197 p. Dissertation. [97153]
237. Stickney, Peter F. 1989. Seral origin of species comprising secondary plant succession in northern Rocky Mountain forests. FEIS workshop: Postfire regeneration. Unpublished draft on file at: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Fire Sciences Laboratory, Missoula, MT. 10 p. [20090]
238. Stromberg, Juliet C.; Krischan, Thomas M. 1983. Vegetation structure at Punta Cirio, Sonora, Mexico. The Southwestern Naturalist. 28(2): 211-214. [4335]
239. Suazo, A. A.; Craig, D. J.; Vanier, C. H.; Abella, S. R. 2013. Seed removal patterns in burned and unburned desert habitats: Implications for ecological restoration. Journal of Arid Environments. 88: 165-174. [87778]
240. Summers, Ryan; Hovland, Matthew; Gibson, Yvette; Mata-Gonzalez, Ricardo. 2021. Climate-mediated shifts toward bunchgrass dominance 40 years after grazing cessation in Saguaro National Park, Arizona, USA. Applied Vegetation Science. 24(4): e12623. [97294]
241. Szarek, Stan R.; Cole, Shelley; Ortiz, Gabriel; Kaufman, Stephanie. 1996. Factors influencing the germination of the desert shrub brittlebush (Encelia farinosa, Asteraceae). American Journal of Botany. 83(supplement): 88 Abstract. [99320]
242. Thomas, Jacob; Yusuf, Mohammed; Sivadasan, M.; El-Sheikh, Mohamed A. 2015. Taxonomy and distribution of two newly recorded genera in Saudi Arabia. Kuwait Journal of Science. 42(3): 158-169. [99099]
243. Thomas, Kathryn A.; Keeler-Wolf, Todd; Franklin, Janet; Stine, Peter. 2004. Mojave Desert Ecosystem Program: Central Mojave vegetation database. Final Report. Sacramento, CA: U.S. Geological Survey, Western Ecological Research Center & Southwest Biological Science Center. 265 p. [98862]
244. Thornburg, John. 2016. Saguaro National Park: Fire management plan. Tucson, AZ: National Park Service, Saguaro National Park. 110 p. [99333]
245. Thorne, Robert F.; Prigge, Barry A.; Henrickson, James. 1981. A flora of the higher ranges and the Kelso Dunes of the eastern Mojave Desert in California. Aliso. 10(1): 71-186. [3767]
246. Turnage, William V.; Hinckley, Arthur L. 1938. Freezing weather in relation to plant distribution in the Sonoran Desert. Ecological Monographs. 8(2): 530-550. [3789]
247. Turner, Raymond M. 1982. Mohave desertscrub. In: Brown, David E., ed. Biotic communities of the American Southwest--United States and Mexico. Desert Plants. 4(1-4): 157-168. [2374]
248. Turner, Raymond M. 1990. Long-term vegetation change at a fully protected Sonoran Desert site. Ecology. 7(2): 464-477. [10866]
249. Turner, Raymond M.; Bowers, Janice E.; Burgess, Tony L. 2005. Sonoran Desert plants: An ecological atlas. Tucson, AZ: The University of Arizona Press. 504 p. [94363]
250. Turner, Raymond M.; Brown, David E. 1982. Sonoran desertscrub. Desert Plants. 4(1-4): 181-221. [2375]
251. Turner, Raymond M.; Webb, Robert H.; Esque, Todd C.; Rogers, Garry F. 2010. Repeat photography and low-elevation fire responses in the southwestern United States. In: Webb, Robert H.; Boyer, Diane E.; Turner, Raymond M. Repeat photography: Methods and applications in the natural sciences. 1st ed. Washington, DC: Island Press: 224-235. [95737]
252. U.S. Department of the Interior, National Park Service, Death Valley National Park. 2007. Fire management plan: Death Valley National Park. Lone Pine, CA, Beatty, NV: U.S. Department of the Interior, National Park Service, Death Valley National Park. 110 p. [99326]
253. U.S. Department of the Interior, National Park Service, Mojave National Preserve. 2004. Mojave National Preserve: Fire management plan. Essex, CA: Mojave National Preserve. 84 p. [99321]
254. Uchytil, Ronald J. 1992. Eragrostis lehmanniana. In: Fire Effects Information System, [Online]. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Fire Sciences Laboratory (Producer). Available: https://www.fs.usda.gov/database/feis/plants/graminoid/eraleh/all.html [2024, January 11]. [99143]
255. USDA, NRCS. 2024. The PLANTS database, [Database]. Greensboro, NC: U.S. National Plant Data Team, (Producer). Available: https://plants.usda.gov/. [34262]
256. USDI, Fish and Wildlife Service. 2022. Endangered Species Program, [Online]. U.S. Department of the Interior, Fish and Wildlife Service (Producer). Available: https://www.fws.gov/endangered/. [86564]
257. Van Buren, Renee; Cooper, Janet G.; Shultz, Leila M.; Harper, Kimball T. 2011. Woody plants of Utah. A field guide with identification keys to native and naturalized trees, shrubs, cacti, and vines. Logan, UT: Utah State University Press. 513 p. [93924]
258. Van Devender, Thomas R.; Dimmitt, Mark A. 2006. Final report on conservation of Arizona upland Sonoran Desert habitat. Status and threats of buffelgrass (Pennisetum ciliare) in Arizona and Sonora. Project# 2004-0013-003. Tucson, AZ: Arizona-Sonora Desert Museum. 28 p. Available online: https://www.desertmuseum.org/programs/buffel_survey_report.pdf. 28 p. [2022, October 3]. [96072]
259. Van Devender, Thomas R.; Felger, Richard S.; Burquez, Alberto. 1997. Exotic plants in the Sonoran Desert region, Arizona and Sonora. In: Kelly, M.; Wagner, E.; Warner, P., eds. Proceedings, California Exotic Pest Plant Council symposium; 1997 October 2-4; Concord, CA. Volume 3. Berkeley, CA: California Exotic Pest Plant Council: 10-15. [44103]
260. Vogl, Richard J. 1982. Chaparral succession. In: Conrad, C. Eugene; Oechel, Walter C., technical coordinators. Proceedings of the symposium on dynamics and management of Mediterranean-type ecosystems; 1981 June 22-26; San Diego, CA. Gen. Tech. Rep. PSW-58. Berkeley, CA: U.S. Department of Agriculture, Forest Service, Pacific Southwest Forest and Range Experiment Station: 81-85. [6011]
261. Wagner, Warren L.; Herbst, Derral R.; Sohmer, S. H., eds. 1999. Manual of the flowering plants of Hawai'i. [Revised edition]. Volume 1. Bishop Museum Special Publication 97. Honolulu, HI: University of Hawai'i Press; Bishop Museum Press. 988 p. [70167]
262. Webb, Robert H.; Bowers, Janice E. 1993. Changes in frost frequency and desert vegetation assemblages in Grand Canyon, Arizona. In: Redmond, K. T.; Tharp, V. L., eds. Ninth annual Pacific climate (PACLIM) workshop; 21-24 April 1992; Pacific Grove, CA. Tech. Rep. 34. San Francisco, CA: Interagency Ecological Studies Program for the Sacramento-San Joaquin Estuary: 71-82. [99144]
263. Weiss, Jeremy L.; Overpeck, Jonathan T. 2005. Is the Sonoran Desert losing its cool? Global Change Biology. 11(12): 2065-2077. [95504]
264. Wells, Gail. 2012. Cheating cheatgrass: New research to combat a wily invasive weed. Fire Science Digest. No. 13. Boise, ID: National Interagency Fire Center, Joint Fire Science Program. 11 p. [85747]
265. Welsh, Stanley L.; Atwood, N. Duane; Goodrich, Sherel; Higgins, Larry C., eds. 2016. A Utah flora. 5th ed. Provo, UT: Brigham Young University. 987 p. [94185]
266. Went, F. W. 1948. Ecology of desert plants. I. Observations on germination in the Joshua Tree National Monument, California. Ecology. 29(3): 242-253. [12915]
267. Went, F. W.; Westergaard, M. 1949. Ecology of desert plants. III. Development of plants in the Death Valley National Monument, California. Ecology. 30(1): 26-38. [11102]
268. Westman, W. E. 1983. Xeric Mediterranean-type shrubland associations of Alta and Baja California and the community/continuum debate. Vegetatio. 52 (1): 3-19. [12000]
269. Westman, W. E.; O'Leary, J. F.; Malanson, G. P. 1981. The effects of fire intensity, aspect and substrate on post-fire growth of Californian coastal sage scrub. In: Margaris, N. S.; Mooney, H. A., eds. Components of productivity of Mediterranean climate regions--basic and applied aspects. The Hague, The Netherlands: Dr. W. Junk Publishers: 151-179. [13593]
270. Westman, Walter E. 1981. Factors influencing the distribution of species of Californian coastal sage scrub. Ecology. 62(2): 439-455. [11032]
271. Westman, Walter E.; O'Leary, John F. 1986. Measure of resilience: The response of coastal sage scrub to fire. Vegetatio. 65(3): 179-189. [63361]
272. Whitfield, Charles J.; Beutner, Edward L. 1938. Natural vegetation in the desert plains grassland. Ecology. 19(1): 26-37. [5251]
273. Whittaker, R. H.; Niering, W. A. 1975. Vegetation of the Santa Catalina Mountains, Arizona. V. Biomass, production, and diversity along the elevation gradient. Ecology. 56(4): 771-790. [35729]
274. Wilder, Benjamin T.; Jarnevich, Catherine S.; Baldwin, Elizabeth; Black, Joseph S.; Franklin, Kim A.; Grissom, Perry; Hovanes, Katherine A.; Olsson, Aaryn; Maulsa, Jim; Kibria, Abu S. M. G.; Li, Yue M.; Lien, Aaron M.; Ponce, Alejandro; Rowe, Julia A.; Soto, Jose R.; Stahl, Maya R.; Young, Nicholas E.; Betancourt, Julio L. 2021. Grassification and fast-evolving fire connectivity and risk in the Sonoran Desert, United States. Frontiers in Ecology and Evolution. 9: e955561 [+Supplement]. [96620]
275. Wilson, R. C.; Narog, M. G.; Koonce, A. L.; Corcoran, B. M. 1994. Postfire regeneration in Arizona's giant saguaro shrub community. In: DeBano, Leonard F.; Gottfried, Gerald J.; Hamre, Robert H.; Edminster, Carleton B.; Ffolliott, Peter F.; Ortega-Rubio, Alfredo. Biodiversity and management of the Madrean Archipelago: The Sky Islands of southwestern United States and northwestern Mexico. RM-GTR-264. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station: 669 p. [95503]
276. Wisdom, C. S. 1985. Use of chemical variation and predation as plant defenses by Encelia farinosa against a specialist herbivore. Journal of Chemical Ecology. 11(11): 1553-1564. [99104]
277. Zedler, Paul H. 1981. Vegetation change in chaparral and desert communities in San Diego County, California. In: West, D. C.; Shugart, H. H.; Botkin, D. B., eds. Forest succession: Concepts and application. New York: Springer-Verlag: 406-430. [4241]
278. Zouhar, Kris. 2003. Bromus tectorum. In: Fire Effects Information System, [Online]. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Fire Sciences Laboratory (Producer). Available: www.fs.usda.gov/database/feis/plants/graminoid/brotec/all.html. [92768]
279. Zouhar, Kristin. 2023. Fire regimes of Sonoran desert scrub communities. In: Fire Effects Information System, [Online]. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Missoula Fire Sciences Laboratory (Producer). Available: www.fs.usda.gov/database/feis/fire_regimes/Sonoran_desert_scrub/all.html. [98816]