Saltcedar
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Introduction and Spread
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Impacts
Water use
In arid western states, competing interest for surface and groundwater is controversial (Morrison 1996) and within some river systems there is risk that the water needs of growing human populations may not be met (Waggoner and Schefter 1990). Evapotransporation (ET) or water use rates of saltcedar are among the highest of any phreatophyte evaluated in southwestern North America (Brotherson et al. 1984; Van Hylckama 1974; White et al 2000), including native riparian trees (Busch and Smith 1995; Neill 1983). Its consumption of water can desiccate springs, drain pools, and even dry up perennial streams (Johnson 1986). Saltcedar in heavily infested areas of the Southwest is estimated to consume twice as much water per year as the major cities of southern California (Friederici 1995; Johnson 1986).
Although total water usage is considerably higher in saltcedar stands than in areas with native woody riparian species, transpiration rates of individual plants have been shown to be similar to those of several herbaceous plants and co-occurring phreatophytes, including willows, cottonwoods, and mesquite (Anderson 1982; Sala et al. 1996). The difference between total stands and native riparian habitat was attributed to greater leaf area index in saltcedar communities compared to any other riparian populations (Sala et al. 1996).
Saltcedar is a facultative phreatophyte, drawing moisture from the saturated zones above the water table and from less saturated soils in areas with deeper water tables (Everitt 1980). Thus, the longer a community has been invaded by saltcedar, the greater the capacity to lower the water table (Brotherson et al. 1984). Robinson (1965) cited studies that indicate saltcedar has the potential to consume 4 acre-ft of groundwater annually. Estimates of water use by saltcedar in New Mexico, Texas, and Arizona were variable, reflecting variations in weather and environment, as well as difficulties in estimating ET rates (Weeks et al. 1987; White et al. 2003). Johns (1989) summarized data from various studies attempting to determine water usage by saltcedar (Culler et al. 1982; Gatewood et al. 1950; Gay and Hartman 1982; Van Hylckama 1974) and concluded that saltcedar water use ranged from 3 to 7.5 acre-ft (0.9-2.3 m) depending on depth to groundwater.
On the Middle Rio Grande at the Bosque del Apache National Wildlife Refuge, ET rates from a dense stand of saltcedar was reportedly 4.35 ft (1.35 m) per year and 3.91 ft (1.19 mm) during the growing season (April 5 to November 21) (Bawazir 2000; King and Bawazir 2000). By comparison, a sparse stand of cottonwood used 2.97 ft (0.90 m) per year and 2.62 ft (0.80 m) during the same growing season. In another study along the Rio Grande reported by Dahm et al. (2002), a dense stand of saltcedar (111-122 cm per year) and a mature cottonwood stand with an extensive understory of saltcedar and Russian olive (123 cm per year) had the highest rates of annual evapotranspiration (ET). A mature cottonwood stand with a closed canopy had intermediate rates of ET (98 cm per year) and a less dense saltcedar stand had the lowest rates of ET (74-75 cm per year).
Water consumption for saltcedar and for replacement vegetation was estimated following root plowing in the Pecos River floodplain between Acme and Artesia, NM. From 1980 to 1982, water use by saltcedar was about 11.8 inches (0.3 m) greater than replacement vegetation (Weeks et al. 1987). In theory, this reduction in water use should have resulted in an increased base flow of the Pecos River of 10,000 to 20,000 acre-ft per year. However, clearing 21,500 acre (8,700 ha) of saltcedar by root plowing could not be used to verify base-flow gains. Hart (2002) estimates that saltcedar control by aerial spraying on the Pecos River for three years salvaged slightly more than 36,000 acre-ft of water.
Evapotranspiration measurements from a dense stand of saltcedar along the lower Colorado River in Arizona (using the Bowen ratio measurement system) ranged from 0.08 inches (2 mm) per day in the spring and fall up to 0.5 inches (13 mm) per day in midsummer (Gay 1986). Total ET for the growing season (Mar 23 to Nov 11) was estimated to be 66 inches (1,680 mm) per year.
Peak ET rates for saltcedar [0.43 in (11 mm) per day] were compared to winter wheat [0.31 in (8 mm) per day], cotton [0.35 in (9 mm) per day], and alfalfa [0.35 in (9 mm) per day] during the growing season in Arizona (Gay and Hartman 1982). Water use by saltcedar began with bud burst in early April [0.11 in (2.9 mm) per day], and reached a peak of about 0.43 inches (11 mm) per day near the summer solstice in late June. Phreatophytes (mainly saltcedar and mesquite) were removed from three reaches of the Gila River floodplain in Arizona in 1969 and resulted in an average combined reduction in ETof 19 inches (483 mm) per year (Culler et al. 1982).
Economic feasibility studies to estimate costs and benefits of saltcedar control have been conducted for some western U.S. waterways (Great Western Research 1989). Horton and Campbell (1974) estimated water savings (difference between saltcedar use and native vegetation use) resulting from saltcedar control on the Colorado River were as high as 3 acre-ft per year. It has been estimated that 568,000 acre-ft per year of water are lost along the Colorado and that water lost by saltcedar from the Bonneville Unit of the Central Utah Water Project had an estimated cost of $27 million annually (Brotherson and Field 1987).
Fire frequency
Saltcedar is a fire-adapted species with more efficient fire recovery mechanisms than nearly all native riparian species (Anderson et al. 1977b; Busch and Smith 1993). Following fire, saltcedar is better able to utilize volatilized nutrients, increase soil concentrations of mineral elements, increase soil pH, and reduce available moisture than native woody riparian species (Busch and Smith 1993). This adaptation has likely been a significant factor promoting its rapid colonization of watercourses (Busch and Smith 1992, 1993; Wiesenborn 1996).
In native riparian plant communities dominated by cottonwood, willows, or mesquite, wildfires are infrequent (Busch and Smith 1993). In contrast, intervals between fires are considerably shorter in saltcedar-infested areas. It has been hypothesized that saltcedar, like other plant species that readily resprout, might have developed adaptive characteristics that enhance flammability of communities where they grow (Busch and Smith 1992). This can lead to replacement of non-adapted fire communities dominated by cottonwood and willow (Busch and Smith 1992; Kerpez and Smith 1987). In support of this, Anderson et al. (1977b) demonstrated that 21 of 25 saltcedar stands along the lower Colorado River burned within a 15-year period. Fires burned 35% of saltcedar-dominated vegetation on the lower Colorado River floodplain between 1981 and 1992, compared to only 2% of communities of honey mesquite (Prosopis glandulosa Torr.) or screwbean mesquite (Prosopis pubescens Benth.) during the same time period (Busch 1995). Increased incidence of fire in saltcedar stands has been attributed to substantial accumulation of leaf litter, as well as dead and senesced woody material (Busch 1995; Busch and Smith 1993; Kerpez and Smith 1987). Fuel buildup by saltcedar promotes a fire every 10 to 20 years in North American desert riparian settings (Lovich et al. 1994).
Hydrology
Robinson (1965) suggested that dense saltcedar stands could increase in areas inundated by floods. The extensive root system of saltcedar is more stable and resistant to erosion than most native riparian trees and shrubs. When stream channels are stabilized, they become more immobile and inflexible (Graf 1978), which progressively restricts channel width by increasing sediment deposition. Narrowing of the water channel increases the rate of water flow and the potential and severity of subsequent floods (Egan et al. 1993; Frasier and Johnsen 1991; Friederici 1995; Kerpez and Smith 1987). A saltcedar-infested area on the Gila River in Arizona had a 30% increase in water flow velocity and 13% increase in water depth (Great Western Research 1989). As the river recedes, saltcedar establishes itself further into the channel. This process continues until stream flow is severely reduced. Saltcedar infestations increased on the Brazos River in northcentral Texas beginning in 1941. At this time, the mean width of the river channel along a 75-mi (121-km) stretch was 508 ft (155 m). By 1979, the mean width had been reduced to 217 ft (66 m). This narrowing in the channel width increased the incidence of flooding, as well as the area inundated by floodwaters (Blackburn et al. 1982). However, Everitt (1980) noted that while vegetation can promote local sediment deposition, the concept that vegetation over large areas can increase regional deposition of sediment is unfounded.
Sedimentation
Saltcedar infestations are also implicated for an increase in sedimentation buildup along waterways that results in flooding. Dense thickets reduce channel widths by forming a partial barrier to flood flow (Kerpez and Smith 1987), thereby decreasing the velocity at which river water flows (Sudbrock 1993, Brotherson and Field 1987, Frasier and Johnson 1991). This reduction in river flow increases sediment deposition along river channels and in holding facilities (Kerpez and Smith 1987, Sudbrock 1993). Saltcedar is thought to be at least partially responsible for the buildup of sediment along the Rio Grande. A survey conducted in 1957 of Elephant Butte Reservoir found that over 500 million cubic meters of storage capacity were lost to sedimentation (Hay 1972). Just a few km down river, Caballo Reservoir lost just over 2 million cubic meters. Over time, the buildup of sediment allows saltcedar seed to germinate along sandbars where the river has receded (Brotherson and Field 1987), adding to the infestation and further reducing channel widths (Frasier and Johnson 1991). The width of the Brazos River in Texas was reported to have been reduced by 71% over 40 years (Blackburn et al. 1982). This, in turn, leads to the additional reduction in streamflow which causes floodwater to disperse to areas that normally do not flood (Kerpez and Smith 1987).
Soil attributes
It appears likely that saltcedar increases soil salinity. Numerous salts and minerals, both macro- and micronutrients, are excreted by saltcedar glands (Berry 1970; Bosabalidis and Thomson 1984; Dreesen and Wangen 1981; Kleinkopf and Wallace 1974; Storey and Thomson 1994; Thomson et al. 1969). Leaves and stems contain concentrations of soluble salts in the range of 5 to 15% (Hem 1967). These salts are absorbed by the roots from deeper soil layers, transported though the plant, and concentrated in the leaves. The salts are eventually deposited on the soil surfaceunder the plant (Kerpez and Smith 1987) when deciduous leaves drop or following rainfall events. Consequently, salts are redistributed over time from deep within the soil profile to become concentrated on the soil surface of floodplains. Excessive surface deposits of salt can inhibit germination of other species (Egan et al. 1993), restricting competition for space and water with other under- or overstory vegetation (Brotherson and Field 1987).
Livestock
The literature is void of research investigating the importance of saltcedar in the diets of cattle, sheep and horses but these animals have been observed to occasionally browse saltcedar foliage and remove seedling plants. When given a preference, livestock will select other native herbage and shrubs before grazing saltcedar. Goats have been used as a biological tool for saltcedar control and are known to graze the plant heavily in special situations (Kris Havstad, USDA-ARS Jornada Experimental Range, personal communication).
Wildlife
Saltcedar provides limited food and shelter necessary for wildlife survival (Shrader 1977). Although some wildlife species successfully survive in saltcedar-dominated areas, most species are negatively affected by displacement of native riparian plant species and other habitat changes resulting from encroachment of saltcedar.
Some obligate riparian bird species can successfully utilize saltcedar (Ellis 1995). However, most of these continue to show preference for more diverse, native plant communities (Shrader 1977). Bird species include various doves, Gambel’s quail (Lophortyx gambelii), other granivores (seed feeders), or other types of ground feeding birds. Because saltcedar seeds are too small to be eaten by most animals (Neill 1983), bird populations generally forage in nearby agricultural fields. Doves, particularly white-winged doves (Zenaida asiatica) and mourning doves (Z. macroura), are among bird species that utilize saltcedar (Anderson et al. 1977a). Although saltcedar provides nesting sites for doves, their populations are usually higher in native plant communities, especially those dominated by mesquite (Shrader 1977), that provide more food for doves (Kerpez and Smith 1987). The endangered southwestern subspecies of the willow flycatcher (Empidonax trailii extimus) will also nest in saltcedar when willows are displaced (DeLoach et al. 1996). Other species, such as Gambel’s quail, prefer honey mesquite and cottonwood communities, but will utilize saltcedar for shelter, but not nesting (Shrader 1977). Saltcedar invasion may be responsible for shifts in riparian bird populations along the middle Rio Grande (Yong and Finch 1997).
Although some riparian bird species have continued to breed in saltcedar-dominated habitats, breeding densities of these populations have declined. On the Colorado River, migratory and resident spring and summer breeding species show a declining trend in saltcedar use from west to east. These birds are largely restricted to cottonwood-willow communities (Hunter et al. 1988). The Bell vireo (Vireo bellii) on the lower Colorado River is nearly excluded as a breeding species, and summer tanagers (Piranga rubra) and yellow-billed cuckoos (Coccyzus americanus) are also in serious danger of elimination from the lower Colorado River due to saltcedar infestations.
The majority of birds, particularly riparian species, are more directly dependent on native plant communities (Ellis 1995; Hunter et al. 1988). Waterfowl, frugivores (fruit and seed feeders), and insectivores, usually abundant in native riparian vegetation, almost completely avoid saltcedar (Brotherson and Field 1987; Kerpez and Smith 1987; Shrader 1977). Many frugivores feed on the fruit of desert mistletoe (Phoradendron californicum Nutt.). This parasitic plant typically grows on native woody species (Cohan et al. 1978) and is uncommon on saltcedar (Haigh 1996).
When comparing total bird density and species diversity, saltcedar stands consistently had lower values than communities dominated by cottonwood, willow, and mesquite (Anderson et al. 1977a; Cohan et al. 1978; Engel-Wilson and Ohmart 1978; Hunter et al. 1988; Kerpez and Smith 1987). In one study, cottonwood-willow communities were the most valuable to bird populations, followed by honey mesquite, desert wash, saltcedar, orchard, and arrowweed (Pluchea sericea [Nutt.] Cov.) (Anderson et al. 1977a). Along the Colorado River, native riparian areas sustained a density of 154 birds per 99 acre(40 ha), whereas the saltcedar-dominated areas had an average of four birds per 99 acre(40 ha) (Johnson 1986). Along the Pecos River in New Mexico, more birds were observed in 96 acre(39 ha) of cottonwood, willow, and mesquite communities than in 48,400 acre(19,600 ha) of saltcedar (Engel-Wilson and Ohmart 1978).
Restoration projects can have a dramatic effect on breeding bird populations in saltcedar stands. Anderson et al. (1977a) noted that the addition of one or more native tree species, even in small numbers, greatly enhanced overall attractiveness of an area to breeding pairs.
Bird species used both saltcedar and cottonwood habitats along the Rio Grande in central New Mexico, with three species using only saltcedar and six species using only cottonwood (Ellis 1995). Assuming the prediction by Howe and Knopf (1991) that saltcedar may completely replace cottonwood habitat along the middle Rio Grande in New Mexico over the next century, the richness of riparian bird species in that area is expected to decline.
Brown and Johnson (1989) argued that, while saltcedar habitat along the lower Colorado River was much less valuable for breeding birds than native riparian habitat, the reverse was true along the Colorado River in Grand Canyon National Park. Hunter et al. (1988) proposed that bird nests in saltcedar along the lower Colorado River experienced higher heat loads than nests in multi-layered cottonwood forests that afford more shade. Brown and Trosset (1988) noted that saltcedar stands in Grand Canyon National Park developed after construction of the Glen Canyon Dam. Since comparable vegetation was not present along the river prior to construction of the dam, saltcedar represented a new habitat type for that area. Regional saltcedar management strategies should consider this and other bird species.
With the exception of woodrats (Neotoma spp.) and the desert cottontail (Sylvilagus audubonii), other native mammals are not known to widely feed on mature saltcedar. In some instances, young sapling growth is utilized by rodents, rabbits and other mammals. When saltcedar was cleared from a 49-acre(20-ha) area along the lower Colorado River and replaced with native vegetation, the diversity of both birds and rodents increased significantly (Anderson and Ohmart 1985). In a study by Engel-Wilson and Ohmart (1978), mammals such as porcupine (Erethizon dorsatum) and beaver (Castor canadensis) had a high affinity for the cottonwood-willow habitats, but occurred in very low densities in saltcedar-dominated communities.
Insects
Willows and cottonwoods support a greater abundance of insect life than does saltcedar (Bailey et al. 2001; Neill 1983; Knutson et al. 2003). The principal insect species that thrives on saltcedar in the Southwest are cicadas and the European honeybee (Apis mellifera) (Horton and Campbell 1974). Saltcedar provides an early source of pollen for over-wintering bees (Kerpez and Smith 1987) and can be a beneficial species for honey production (Shrader 1977). Insect reproduction fluctuates dramatically on saltcedar compared with native riparian habitat (Carothers et al. 1976). Anderson (1994) studied the Apache cicada (Diceroprocta apache Davis) in a native riparian community and a saltcedar stand along the lower Colorado River. He found that although cicadas were abundant in both communities, the insects emerged later in native cottonwood and willow-dominated communities when migrating and nesting birds were present. This change in temporal availability of this key food resource may help explain the low population of breeding birds in saltcedar communities.
Faster decomposition of saltcedar litter was associated with a two-fold decrease in macro-invertebrate richness and a four-fold decrease in overall macro-invertebrate abundance relative to native cottonwood communities (Bailey et al. 2001). This study demonstrates that invasion by saltcedar affects leaf litter quality, which in turn affects stream macro-invertebrates. Thus, impacts on primary consumers and food web structure could affect higher trophic levels.
Rare, sensitive, or threatened species
The Federally Endangered southwestern willow flycatcher (Empidonax trailii extimus) is known to nest in saltcedar-dominated areas (USFWS 1993). This subspecies of the willow flycatcher is widely distributed in scattered remnant populations across much of the area where saltcedar is invasive. Although it prefers to feed and breed in riparian woodlands dominated by native plants including willows (Salix spp.), arroweed (Pluchea spp.), and Baccharis species, there has been concern that it might be further threatened if a biological control agent eliminated saltcedar over wide areas of the Southwest. Most scientists point out that even a highly successful biological control agent will not likely eliminate saltcedar and that where it is reduced, native plants favored by breeding and feeding birds are expected to establish (Lovich and de Gouvenain 1998).
Plant biodiversity/richness
In some sites, 70 to 80% of the total vegetative cover can consist of saltcedar. Such infestations lead to dramatic reductions in native indigenous woody and herbaceous plant composition and abundance (Engel-Wilson and Ohmart 1978; Hughes 1993; Lovich et al. 1994; Weeks et al. 1987).
Fire frequency, as well as other factors in saltcedar communities, has dramatically reduced native plant populations and hindered reestablishment of riparian species. When present in saltcedar-dominated communities, cottonwood populations often consist solely of mature or nearly mature trees. Because these communities lack a broad age distribution typical of vigorous reproducing populations, they are less able to spread and suppress saltcedar stands through shading (Shrader 1977). Along the Rio Grande, cottonwood regeneration has not occurred for 30 to 35 years, partially because human interventions such as dams and other structures have suppressed natural flooding events necessary for seedling establishment. Consequently, riparian woodlands are aging and are expected to dramatically decline within the next 50 years (Howe and Knopf 1991).
Positive impacts
Saltcedar habitats have been found to provide shelter and roosting for some avian species, the most important of which are white-tail and mourning doves (Anderson et al. 1977a). However, saltcedar areas have low forage value, which forces doves to use nearby fields for food supply (Kerpez and Smith 1987). Saltcedar also provides an early source of pollen for overwintering bees (Kerpez and Smith 1987) and can be a beneficial species for honey production (Shrader 1977).
Saltcedar was introduced partially to stabilize soils and mitigate stream bank erosion. In 1925, the New Mexico Extension News reported that saltcedar would be planted at the head of arroyos near Silver City to slow flow of water and reduce erosion. This perceived benefit resulted in increased flooding in many areas of the West (Graf 1978; Egan et al. 1993; Frasier and Johnsen 1991; Friederici 1995; Kerpez and Smith 1987).
Economics
A comprehensive economic analysis estimated ecosystem services lost to saltcedar invasion in the western United States (Zavaleta 2000). Values lost from saltcedar included irrigation and municipal water, flood control, hydropower, wildlife habitat, and river recreation. Dove hunting and sedimentation were considered benefits provided by saltcedar in the analysis. The 55-year value lost to ecosystem services in the western United States was $7.331 billion to $16.062 billion.
Zavaleta (2000) also estimated costs (in 1998 dollars) of a regionwide saltcedar eradication program including evaluation, control, revegetation, and monitoring. A cost of $3,006 per acre ($7,428 per ha) for a 20-year comprehensive program provided a net benefit of $3,312 to $6,975 per acre ($8,184 and $17,235 per ha). The cost:benefit ratio for the eradication program ranged from 2:10 to 3:32, with a total net benefit of $3.843 billion to $11.225 billion (assuming 0% discount rate).
Economic feasibility studies to estimate costs and benefits of saltcedar control have been conducted for some western U.S. waterways (Great Western Research 1989). Horton and Campbell (1974) estimated water savings (difference between saltcedar use and native vegetation use) resulting from saltcedar control on the Colorado River were as high as 3 acre-ft per year. It has been estimated that 568,000 acre-ft per year of water are lost along the Colorado and that water lost by saltcedar from the Bonneville Unit of the Central Utah Water Project had an estimated cost of $27 million annually (Brotherson and Field 1987).
Saltcedar clearing is accomplished using a combination of herbicide, burning, and mechanical control techniques (McDaniel and Taylor 2003). Complete land restoration costs range from $1,852 to $3,200 per acre ($750 to $1,300 per ha) on extensive floodplain areas on the Bosque del Apache NWR along the Rio Grande (Taylor and McDaniel 1998a). For saltcedar control alone, a combination of herbicide and burning cost from $114 to $225 per acre ($46 to $91 per ha) with expected control around 92% (McDaniel and Taylor 2003). Root plowing plus raking costs range from $300 to $700 per acre ($121 to $283 per ha) with plant control exceeding 97%. Saltcedar infestations intermixed with remnant stands of desirable trees, shrubs, or herbaceous cover are cleared with cut-stump treatments on a contractual basis averaging between $2,000 to $2,500 per acre ($809 to 1012 per ha) (Taylor and McDaniel 2003).
