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of wildlife habitat as a goal of agricultural programs. For example, the Wildlife Incentives Program was the first agricultural program developed exclusively for the creation and protection of wildlife habitat. The Conservation Reserve Program was originally presented in the 1985 Farm Act and reauthorized, in 1996, the setting aside or converting of as much as 36.4 million acres of environmentally sensitive farmland, through 2001. The Wildlife Habitat Incentives Program of 1996 set aside $200 million for restoration programs in the Everglades Agricultural Area provision (Day, 1996). All of this legislation, explicitly linked to agriculture, has benefited wildlife.

Negative Effects of Agriculture on Wildlife

The greatest negative effect of agriculture on wildlife is the conversion of natural vegetation to an agroecosystem. Habitat loss directly reduces biodiversity. At least 71 species and subspecies of vertebrates1 and at least 217 species of plants2 have gone extinct in North America (north of Mexico) since the arrival of Europeans. Over 95% of our original virgin forests are now gone from the lower 48 states (Postel and Ryan, 1991). The decline in large mammalian predators as a result of habitat loss to agriculture has resulted in an increase in the population of deer, which have now become an agricultural pest in many regions (Day, 1996). These losses of species and habitats are the result of the gamut of human activities, of which agricultural activities form a major part.

Given that a certain amount of land will be dedicated to providing for human food and nutrition, the next most significant effect is fragmentation. Habitat fragmentation is defined as the subdivision of continuous habitat over time; the most important large-scale cause of habitat fragmentation is expansion and intensification of human land use (Burgess and Sharpe, 1981; Harris, 1984). Fragmentation is considered to be an important cause of local extinction (Wilcox and Murphy, 1986). Fragmentation results in a loss of original habitat, a reduction in the sizes of the patches of remaining habitat, and an increase in the degree of interpatch distances, all of which increase the rate of local extinction (Harris, 1984; Wilcove et al., 1986). Even certain aspects of the edge effect, positive for some species, are detrimental to others. Edges can serve as potential ecological traps for breeding birds by concentrating nests in a small area where the risk of predation is high (Rudnicky and Hunter, 1993). There is also an apparently high rate of nest parasitism of breeding birds by cowbirds in edge habitats (Brittingham and Temple, 1983).

Activities other than deforestation associated with agriculture also pose threats to wildlife. The use of agricultural chemicals expanded after the end of World War II and their impact on avian populations became a national issue after the publication of Rachel Carson's Silent Spring in 1962. Several recent review volumes have addressed the impacts of agricultural chemicals on wildlife and discussed the attempts to mitigate these impacts (Kendall and Lacher, 1994; Colborn et al., 1996).

1 See The Nature Conservancy, 1992. Extinct Vertebrate Species in North America, unpublished draft list, March 4, 1992. The Nature Conservancy, Arlington, VA.

2 See Russell, C. and Morse, L., 1992. Extinct and Possibly Extinct Plant Species of the United States and Canada, unpublished report, review draft, 13 March 1992, The Nature Conservancy, Arlington, VA.

Although persistent pesticides, like DDT, which was an issue in Carson's time, are rarely used, other more acutely toxic compounds now pose a threat of mortality to wildlife. Others, so-called endocrine disrupters, may cause long-term reproductive damage (Colborn et al., 1996). Some of the most significant victims of pesticides have been nontarget species of insects. The loss of these potential pollinators will have far-reaching effects, even on many agricultural crops (Buchmann and Nabhan, 1996). States like California, which was once the largest U.S. user of agricultural toxicants, now have ambitious programs to reduce chemical use (Anderson, 1995). This includes compounds recently suspected of acting as endocrine-disrupting chemicals (Fry, 1995).

Noss and Cooperrider (1994) present a series of summary tables in chapter 3 of their book on the impacts of a variety of land-use practices. Concern for increased rates of local extinction and the concomitant loss of biodiversity as a result of agricultural development is a growing issue in tropical regions as well as temperate zones (Holloway, 1991).


In the U.S., Texas is one of seven states that grow rice (Texas Rice Task Force, 1993). The Texas rice crop is grown in the gulf prairies and marshes of the upper Texas coast (Gould, 1975). The native tall grass prairies historically extended inland from extensive coastal marshes for approximately 20 to 150 km. The prairies were characterized by nearly level to gently sloping topography interspersed with small, rain-filled depressions. Prior to the 1900s, the prairies of the upper Texas coast were grazed by herds of bison (Bison bison) and wild horses (Robertson and Slack, 1995). As the land was settled, bison and wild horses were replaced with free-ranging cattle and later with agricultural crops (Craigmiles, 1975; Stutzenbaker and Weller, 1989). Rice was first introduced to the coastal prairies in the mid-1800s. By 1954, a peak of 254,000 ha of rice were harvested on the gulf prairies (Hobaugh et al., 1989). Currently, about 110,000 ha of rice are harvested in Texas producing an aggregate addition to the Texas economy of almost $1 billion (Texas Rice Task Force, 1993). In addition, the economy of the rice growing region of the state is enhanced by significant expenditures for recreational hunting.

Rice fields are prepared for planting in late winter with actual planting occurring in March or April. Fields are flooded shortly thereafter until immediately prior to harvest in August. These flooded fields provided large expanses of wetlands for some resident birds to use. In portions of the Texas rice belt a second crop ("ratoon crop") results from resprouting from the initial planting and is harvested in October. Harvested fields contain waste grain and are left to stand fallow for up to 2 years. During the subsequent seasons, the fallow fields are grazed by cattle. Therefore, a typical, 3-year, rice-pasture rotation system involves three fields; during early winter rice is harvested in one field, another field is plowed in preparation for planting rice the next spring, and the third field is being grazed (Hobaugh et al., 1989).

At a landscape scale, the tall grass prairies of the upper Texas coast were a relatively homogeneous matrix of tall prairie grasses with small, scattered, natural depressions (Figure 1a). At smaller scales within the matrix, the landscape was heterogeneous, with grasses, forbs, and scattered natural wetlands with associated aquatic vegetation. Because of the intensive rice-cropping system, the resulting landscape is a reversed image of the native tall grass prairie environment (Figure 1b) — a heterogeneous mosaic at the landscape scale, with homogenous field-sized stands of vegetation, prepared fields, or pastures.

The rice-cropping system in Texas lies adjacent to a heavily industrialized region with over 30% of the U.S. petroleum industry and more than 50% of the U.S. chemical production occurring in this region (Robertson and Slack, 1995). In addition, the Houston-Galveston metropolitan area is the fourth largest metropolitan area in the U.S. As a result of these economic pressures, the area of wetland habitats has declined dramatically in Texas, and especially in coastal regions of the state (Anderson, 1996; Moulton et al., 1997). Current estimates of losses show that >35% (84,000 ha) of Texas' coastal marshes have been destroyed since the 1950s (Anderson, 1996).

The net effect of the landscape mosaic produced by the Texas rice-cropping system has been a dramatic change in use by migratory birds since the advent of rice agriculture. Lesser snow geese (Chen caerulescen), greater white-fronted geese (Anser albifrons), and Canada geese (Branta canadensis) only began to use the prairie after rice agriculture became established on the upper Texas coast (Hobaugh et al., 1989). Waterfowl were commonly associated with the small natural depressions in the native prairies (Mcllhenny, 1932). However, it wasn't until the 1940s and 1950s with mechanization of rice farming, extensive irrigation, and the 3-year rice rotation system that geese and waterfowl began to exploit the system fully (Hobaugh et al., 1989; Robertson and Slack, 1995). Waterfowl and geese are attracted to the mosaic of habitats because of the availability of waste grain after harvest in the fall and the extensive areas of standing and impounded water associated with roost ponds. Gawlik (1994) documented as much as 87 kg/ha in stubble fields immediately after harvest. Waste rice is an important source of food for wintering ducks, geese, and numerous granivorous passerines (Terry, 1996). Similarly, the importance of waste rice to wintering waterfowl has been documented for the Central Valley of California (Alisauskas et al., 1988; Brouder and Hill, 1995; Gawlik, 1994). In addition to rice grains, green vegetation emerging during the winter in harvested rice fields and in fields that had been prepared for the rice crop the following season, becomes an important source of food for geese (Hobaugh, 1985; Gawlik, 1994). Well over 2 million waterfowl and geese winter on the upper Texas coast with the bulk of these birds found using freshwater wetlands associated with rice agriculture (Haskins, 1996). The extensive use of rice-cultivated land by wintering lesser snow geese has been identified as a significant component of the observed high population growth rates. These high population densities have resulted in significant alteration to Arctic coastal salt marsh plant communities (Abraham and Jefferies, 1997).

Figure 1 Schematic representation of the Texas coastal tall grass prairie (A) before the advent of rice cultivation (mid-1980s) and (B) after rice production. Rice production has fragmented the larger landscape into patches of homogeneous stands of vegetation, prepared fields, or pastures but has greatly reduced landscape heterogeneity at small spatial scales.

Figure 1 Schematic representation of the Texas coastal tall grass prairie (A) before the advent of rice cultivation (mid-1980s) and (B) after rice production. Rice production has fragmented the larger landscape into patches of homogeneous stands of vegetation, prepared fields, or pastures but has greatly reduced landscape heterogeneity at small spatial scales.

McFarlane (1994) and Terry (1996) have documented the use of the rice system by more than 70 species of birds during an annual cycle. Most species were associated with wetland habitats such as roost ponds, flooded rice fields, and natural depressions. Sheridan et al. (1989) have documented at least 22 species of colonial-nesting waterbirds nesting in 42 colonies located on the upper Texas coast including rice lands. In addition, over 35 species of migratory shorebirds were documented for the upper Texas coast, with 16 species found by Terry (1996). Migratory shore birds take advantage of the wetland habitats associated with rice agriculture, as well as moist, open fields prepared for next year's rice crop.

Migratory Birds, Agroecosystems, and Agricultural Chemicals

Swainson's hawks (Buteo swainsoni) are long-distance migrants whose habits in North America have been well documented (England et al., 1997). Breeding habitat in North America consists of open grassland and shrub steppe semiarid ecosystems from Mexico to the prairie provinces of Canada. Birds nest in trees adjacent to large fields, often utilizing agricultural grassland habitat for locating food and other daily activities (Bloom, 1980). Hawks hunt on the ground or in midair, opportunistically eating insects, small mammals, reptiles, and birds (Bednarz, 1986). Adults primarily feed on mammals and birds to supply the nutritional requirements to growing nestlings. Fledglings, aggregates of nonbreeding hawks, and aggregates of premigratory mixed-age hawks forage primarily for insects such as grasshoppers and dragonflies (Woffinden, 1986; Johnson et al., 1987).

Breeding in western North America during the boreal summer, Swainson's hawks migrate to the nonbreeding grounds in South America with the advancing austral summer, maintaining both climatic and habitat similarities (Figure 2). The journey of up to 10,000 km in each direction takes less than 2 months, and, once settled in southern South America, hawks generally reside in the agricultural grasslands of the Argentine pampas (White et al., 1989; Woodbridge et al., 1995; Goldstein, 1997). This habitat is similar to agricultural prairies found throughout their North American range.

Utilizing agricultural areas west and north of the capital city of Buenos Aires during the nonbreeding season, Swainson's hawks were found in the Argentine provinces of La Pampa, Cordoba, Buenos Aires, Santa Fe, and San Luis (Goldstein, 1997). Hawks were encountered sunbathing and foraging in freshly tilled fields or in fields whose crop height was less than 40 cm. Crops used included alfalfa, corn, sorghum, soybean, and sunflower. Flocks of hawks followed insect outbreaks, traveling over a small region as stages in crop growth changed throughout the season. Other insectivorous species, such as the Chimango caracara (Milvago chimango), burrowing owl (Athene cunicularia), Franklin's gull (Larus pipixcan), and southern lapwing (Vanellus chilensis) followed these insect outbreaks as well. Aplomado falcons (Falco femoralis) also feed on agricultural sites in the pampas, predating insectivorous songbirds.

Increasing monoculture and more intensively managing alfalfa in these regions have also resulted in heavy reliance on agrochemicals for crop protection from insect pests. Subsequent hot and dry conditions of the pampas during the austral summers of 1994-95 and 1995-96 led to severe grasshopper outbreaks, exacerbating the problem of reliance on chemical controls. Typically, the inexpensive organophos-phate insecticide monocrotophos (MCP) was used for grasshopper controls. During this time, when insect outbreaks and chemical controls were at their maximum, the largest flocks of Swainson's hawks, up to 12,000 birds, were seen.

Agrochemical controls during the austral summers from 1994 through 1996 led to 19 documented hawk mortality incidents, accounting for approximately 6000 dead Swainson's hawks over two seasons (Goldstein et al., 1996; Goldstein, 1997). Hawks died in fields while foraging for grasshoppers, in roosts after returning from foraging bouts, and along the trajectory from fields to roosting trees. The agrochemical MCP

Figure 2 Swainson's hawk (Buteo swannsoni breeding range, migratory route, and nonbreed-ing range, with a list of the common names for the Swainson's hawk used across the Americas. (Courtesy of M. Fuller, unpublished data.)

was determined responsible for mortality in birds from all 6 sites sampled and in 17 of 19 sites overall, based on forensic analysis and farmer testimony (Goldstein, 1997). The mortality incidents were highly publicized in the scientific and lay news media, resulting in the establishment of an international working group whose function it was to resolve potential future conflicts between agricultural production and wildlife habitat use prior to the 1996-97 austral summer season. University scientists, agrochemical representatives, conservation activists, and government personnel from Argentina, the U.S., and Canada joined together to designate an MCP-free zone in the area of previous Swainson's hawk mortality.

During 1996, use of MCP in alfalfa or as a grasshopper control agent was made illegal in Argentina. An ecotoxicology program was initiated, with field and laboratory training for government agents, students, and veterinarians living in the pampas. Grassroots campaigns described grasshopper-eating hawks as allies to farmers during the time when they were required to transition from MCP to another chemical. The OP dimethoate and the synthetic pyrethroid cypermethrin were most frequently chosen as chemical alternatives. With the successful removal of MCP from this zone, hawk mortality was completely eliminated.

Agricultural Practices in Coffee Agroecosystems

Coffee (Coffea arabica) originated in Africa and was introduced to Latin America in the early 18th century by the Dutch. Nearly one third of the world coffee now comes from Latin America where it is the leading agricultural commodity for many countries and the leading source of foreign exchange. In all, 44% of the permanent cropland is now coffee, including 750,000 ha in Central America (Perfecto et al., 1996).

Coffee is a shade-tolerant species and was traditionally grown under the canopy of taller trees, often native species. Coffee in the traditional system was allowed to grow fairly tall (3 to 5 m) under a 60 to 90% cover of shade. Plants were grown at a relatively low density (1000 to 2000/ha), took 4 to 6 years to first harvest, and had a life span of over 30 years. Soil erosion was low and there was little need for agrochemical use (Perfecto et al., 1996). Several factors influenced the shift to a more-intensified approach to cultivation, called sun coffee. First, the spread of coffee leaf rust to Latin America caused phytopathologists to reason that the problem would be minimized if coffee were grown in the sun as, therefore, less moisture would accumulate on the leaves. This led to the development of more densely planted, high-yield varieties that would produce up to four times the kilograms per hectare of traditional plantations. Sun coffee is kept shorter (2 to 3 m) and planted at densities of 3000 to 10,000/ha. Time to first harvest is shorter (3 to 4 years), but plantation life span is less (12 to 15 years). In addition, there is a greater input of agricultural chemicals and a higher likelihood of erosion. The high cost of inputs, however, made sun coffee nearly 50% more expensive than shade coffee (Perfecto et al., 1996). This does not include the environmental cost of sun coffee production. Nevertheless, sun coffee has spread throughout the region and now is the most common practice in Colombia (60% of all production; Perfecto et al., 1996).

Concern over neotropical migratory birds (NTMBs) has refocused attention on shade coffee. Wunderle and Waide (1993) surveyed overwintering neotropical migrants in the Bahamas and the Greater Antilles and observed that shade coffee plantations provided habitat for species normally restricted to forests. Russell Green-berg of the Smithsonian Migratory Bird Center and colleagues have conducted several studies on the levels of biodiversity, including neotropical migrants, that are supported in sun vs. shade plantations. As a generalization, shade coffee supports more biodiversity than sun coffee; however, there is a great deal of variation among types of shade coffee that merits examination.

Greenberg et al. (1997) did a comparison of bird species composition in fragments of forest, matorral (second-growth shrub land), and three types of coffee plantations: sun coffee, shade coffee with a Gliricidia sepium overstory, and shade coffee with a canopy of several species of the genus Inga. Both Gliricidia and Inga and nitrogen-fixing legumes are commonly planted in shade coffee plantations in Latin America. Forest remnants had the highest richness, followed by the matorral and Inga sites. Sun coffee and Gliricidia had the lowest species richness, this in spite of the fact that the Gliricidia shade plantations were at a lower elevation with a potentially larger pool of species to draw from. This suggests that shade coffee alone can support somewhat higher richness than sun coffee, but the canopy species is important. When the canopy is composed of a single species, the gain in richness approaches early second growth, but is still less than forest fragments and far less than undisturbed forest. Greenberg recommends that plantations should be as structurally diverse as possible and a mixed agroforestry system could provide this kind of habitat.

There are data to suggest that shade coffee provides habitat for other species as well. Estrada and colleagues (1993; 1994) found that bat species richness in shade coffee plantations in Veracruz, Mexico was higher than in adjacent agricultural fields but lower than in forest; the same was true for terrestrial mammals. Biodiversity of other groups of vertebrates and invertebrates is higher as well (Perfecto et al., 1996).

Several conservation organizations are now promoting shade coffee and mixed agroforestry systems of production as more sustainable and environmentally benign. The Rainforest Alliance has launched an ECO-OK certification program for coffee to support more sustainable practices (Wille, 1994). Conservation International has a similar Sustainable Coffee Initiative with sites in Mexico, Guatemala, Colombia, and Peru which promotes shade coffee and organic techniques (Greenberg, 1996).

Trees as Row Crops: Plantation Forestry and Wildlife

Currently in the U.S., approximately 490 million acres of land are used for commercial timber production (American Forest and Paper Association, AF&PA, 1996). The majority is held by private individuals (59%), with the remainder held by the forestry industry, national forests, and other public agencies (Powell et al., 1992). Timber is important economically; however, the impacts of timber production on wildlife vary and are often extensive. The process of clear-cutting and reforestation creates an artificial cycle of disturbance, resulting in truncated succession, a loss of species richness, and a loss in structural and functional diversity. Native stock is often replaced by nonnative or genetically "improved" species which results in a loss of genetic purity or genotypes of native stock. Pesticides and herbicides are also used with unknown, secondary consequences on the native flora and fauna (Noss and Cooperrider, 1994).

One of the most obvious and frequently cited impacts of timber production on biodiversity is an increase in habitat fragmentation. Habitat fragmentation simply refers to the subdividing of a continuous habitat over time (Pickett and Thompson, 1978; Foster, 1980) and processes that affect an intact forest may be exaggerated when the forest community is fragmented (Noss, 1983). Fragmentation of forests also increases the proportion of edge to interior habitat as the size of the forest decreases (Ranney et al., 1981). Numerous studies on NTMBs have cited increased competition, nest predation, and cowbird parasitism associated with fragmentation and increased edge habitat (Brittingham and Temple, 1983; Small and Hunter, 1988; Yahner and Scott, 1988; Wilcove and Robinson, 1990; Rudnicky and Hunter, 1993).

Another major impact of timber production on biodiversity is the loss of the habitat itself. Of the virgin forests in the U.S., 85% had been destroyed by 1980, with losses of approximately 95 to 98% in the lower 48 states (Postel and Ryan, 1991). Species dependent on these forest habitats, such as the endangered red-cockaded woodpecker (Picoides borealis) suffer when these habitats are destroyed or altered (Yahner, 1995). The red-cockaded woodpecker has almost been eradicated from its former range, in part, due to the destruction of its primary habitat, longleaf pine forests. Longleaf pine forests have been reduced by approximately 98% and are the most endangered forest type (Noss, 1989). This reduction is due to logging, fire suppression, and replacement with faster growing, more economical loblolly pine (Dickson et al., 1995). Other species that have experienced declines, or have been extirpated at least partially because of logging practices, include the spotted owl (Strix occidentals), the red wolf (Canis niger), the wood duck (Aix sponsa), the ivory-billed woodpecker (Campephilus principalis), and the passenger pigeon (Ectopistes migratorius) (Bellrose, 1976; Paradiso and Nowak, 1982; Robinson and Bolen, 1984; Gill, 1990; Block et al., 1995).

Although the majority of forest management practices appear to have a detrimental impact on biodiversity, some practices do benefit some species. For examples, small clear-cuts (<10 ha) appear to benefit some songbirds and mammals which require dense, brushy, vegetation for cover and food (Scott and Yahner, 1989; Hughes and Fahey, 1991; Yahner, 1993). Edge does benefit some game animals, such as the white-tailed deer (Odocoileus virginianus) (Yoakum and Dasmann, 1971), and many game managers emphasize the creation and maintenance of edge despite its potential negative impacts on other species.

Timber management practices have recently come under fire from various organizations and this increased concern over the impacts of timber management has led in part to the implementation of the Sustainable Forestry Initiative (SFI) by the AF&PA. Several priorities outlined by SFI directly emphasize biodiversity (AF&PA, 1996). As advances, innovations, and technologies are developed and implemented, these negative effects on biodiversity may be ameliorated in the future.


There is little question that, over time, agriculture has resulted in a loss of biodiversity. The losses have been particularly serious in areas where little natural habitat was left among agroecosystems, such as in the central plains of the U.S. The prairie fauna once included grizzly bears, wolves, mountain lions, elk, deer, and millions of bison, all of which are now extinct throughout most of their former grassland ranges. The eastern deciduous forest of the U.S. was nearly completely deforested at the turn of the century, although remnant patches of forest remained on hilltops and in valleys. As agriculture shifted westward, many species were able to recolonize, regenerating forests and reducing the long-term severity of the agricultural impact.

If lands are properly managed in a mosaic, losses of biodiversity need not be permanent. Studies indicate that extensive areas of tropical forest, once thought to be virgin, were at one time under intensive cultivation. Pristine forests in the Darien gap in Panama were apparently subjected to over 4000 years of human disturbance, and are probably no older than 350 years (Bush and Colinvaux, 1994). Diversity is extremely high with no indication that this history of agricultural activity depleted the biodiversity of the region. These forests never were subjected to extensive deforestation, however, and this likely prevented large-scale local extinctions. Indeed, Mellink (1991) observed that the presence of isolated farms in the San Luis Potosi plateau of Mexico actually increased regional bird species richness.

The implementation of policies recommended under the 1996 Farm Act will help to preserve patches of natural habitat and will facilitate the protection of residual populations of native species. If patches are kept small and isolated, however, they may not protect viable populations over long periods of time. Areas of native vegetation should be as large as possible and interconnected via corridors of habitat to maximize their effect. The extensive nature of modern agriculture is much different than the practices of the precolonial people of the Darien region in Panama, and the long-term consequences on wildlife populations under current management practices will be negative.

Change in Community Structure

Agroecosystems simplify the environment. Generally, they contain fewer species than the native flora, and they contain less diversity in foliage structure than native ecosystems. They also contain far fewer species of invertebrates and vertebrates than natural ecosystems. Thus, the communities of organisms associated directly with agroecosystems represent but a tiny subset of the total biodiversity of the region, and the community structure is also simplified. These changes affect not only species diversity, but also the functional diversity of agroecosystems. When agroecosystems are extensive and remaining habitat is fragmented, local extinctions that result from fragmentation will also reduce the species and functional diversity of the region as a whole. All of these effects have been observed and documented (McNeely et al., 1990; Day, 1996).

Changes in community structure can result in changes in ecosystem function. A common consequence of human activities is the extirpation of large predators. The elimination of predator populations can allow herbivore populations to increase, thus increasing the impact of herbivores on plant community structure and function. This can lead to large-scale changes in community structure and function (Rasmussen, 1941; Paine, 1966). Changes in species composition can also affect community processes in unexpected ways by altering the functional diversity of communities (Tilman et al., 1997; Hooper and Vitousek, 1997). Thus, the effects of altered species diversity can extend far beyond the boundaries of agroecosystems and can result in major shifts in ecosystem function as well as composition.

Recommendations for the Mitigation of Impacts

The joint production of crops and wildlife is a relatively new concept (Howitt, 1995). Paoletti et al. (1992) present a table of choices among farming systems that either reduce or enhance biodiversity in the regional agroecosystem. Indeed, given the preponderance of lands under some form of human management, we must begin to assign a larger role to preserving biodiversity in agroecosystems (Pimental et al., 1992). In the U.S., the realization that most biodiversity exists on private lands has led to the development of new initiatives to make conservation more attractive to private landowners. California is a habitat mosaic of federal and private lands and an extensive region of intensive agricultural development. California also is the most biologically diverse state with the largest number of federally listed candidates or endangered species in the country (Scott et al., 1995). Two thirds of these listed species are on private lands. A federally sponsored program called Habitat Conservation Plans (HCP) provides a mechanism where landowners agree to an overall plan to protect an endangered species and its habitat in exchange for a permit to alter some portion of the habitat in the planning area (Scott et al., 1995). HCPs can vary in their geographic scope from a single parcel or landowner to larger areas and multiple landowners. This mechanism was instituted first in California over a controversy concerning several threatened species of butterflies. The success of this experience led to the 1982 amendment to the Endangered Species Act, allowing HCPs. Although the majority of these HCPs are in California, they will be implemented nationwide, with approximately 40 plans approved and at least another 150 in progress (Beatley, 1995).

There are numerous efforts to reduce the use and impact of agricultural chemicals on agroecosystems (Kendall and Lacher, 1994). The U.S. Environmental Protection Agency has supported the creation of dialogue groups to assist in the resolution of conflicts over agricultural chemical use and the protection of wildlife (Avian Effects Dialogue Group, 1994) and has developed a new paradigm for the assessment of environmental risk (U.S. Environmental Protection Agency, 1992). California has become a world leader in the development of methods to reduce pesticide use (Anderson, 1995). The sustainable agriculture movement also emphasizes the use of native biota for pest control (Miller and Rossman, 1995).

The impact of agricultural activities on tropical wildlife is of growing concern. Economic factors strongly influence agricultural practices in the tropics (McNeely and Norgaard, 1992). Frequently, pressures for colonization result from economic hardship in urban areas, and the lack of technical expertise of colonists results in a sequence of poor land-use practices from deforestation through inappropriate agriculture to abandoned pasture land. This land-use succession has been referred to as "nutrient mining" (Southgate and Clark, 1993). Southgate and Clark (1993) make the point that farmers and ranchers in countries where crop and livestock yields have improved seldom encroach on natural habitats. In countries with poor yields and increasing populations, new areas are continually being cleared. Programs that increase yields effectively buy time for the implementation of population control. However, most donor and foreign aid agencies are currently reducing support of agricultural development programs. Protectionism in developed countries also inhibits the transfer of technologies to developing countries, inhibiting their competitiveness. Some environmental groups unwittingly contribute to this process by ignoring or downplaying agricultural development and emphasizing conservation efforts only, which can exacerbate rather than mitigate pressures on protected areas. The dynamics of land use on a mosaic landscape are complex and interrelated; it is not possible to view wildlife conservation and agricultural development as independent processes. Agricultural development projects funded by international aid agencies must address conservation and the mitigation of environmental impacts, and conservation projects must look more closely at the interrelationship between agricultural productivity and deforestation.

Public awareness of the impacts of agroecosystems on biodiversity in general, and wildlife in particular, has led to the development of new laws and regulations in countries throughout the world. There is a new appreciation of the concept of the management of the whole landscape for multiple uses, and a better understanding of the interrelations among landscape units. Agriculturists and conservationists alike are coming to an agreement that the old way of doing things is no longer the best way, considering the diverse expectations of a more-educated, globally connected populace.

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