Forest Management and Southern Pine Beetle Outbreaks: A Historical ...

For. Sci. 62(2):166 –180 http://dx.doi.org/10.5849/forsci.15-071

REVIEW ARTICLE

forest management

Forest Management and Southern Pine Beetle Outbreaks: A Historical Perspective Stephen R. Clarke, John J. Riggins, and Frederick M. Stephen The southern pine beetle (Dendroctonus frontalis) (SPB) is an eruptive pest of pine forests in the southeastern United States. Numerous studies have been conducted on the relationships among SPB population dynamics, climatic factors, natural enemies, and competitors, but the influence of changes in forest management through time on SPB activity has received little attention. Forest management dictates the configuration and condition of hosts available for SPB populations, whereas suppression has an impact on population levels of SPB and their associates and also affects the area and distribution of the susceptible host type remaining after treatment. In contrast to the frequent and widespread SPB outbreaks in the last half of the 20th century, recent SPB activity in the Southeast has been localized and short-lived. Reports from the mid-1800s through the mid-1900s indicate that outbreaks then were also less common. In this review, we examine how changes in forest management practices have played a significant role in the history of SPB outbreaks. Keywords: southern pine beetle, forest management, population dynamics, suppression

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he southern pine beetle (Dendroctonus frontalis) (SPB) has been documented as a major mortality agent in pine forests in the southeastern United States for more than 120 years (Hopkins 1892, Thatcher et al. 1980, Coulson and Klepzig 2011). Loblolly pine (Pinus taeda), shortleaf pine (Pinus echinata), pitch pine (Pinus rigida), and Virginia pine (Pinus virginiana) are the primary hosts in the southeastern United States, but all southern yellow pine species including slash pine (Pinus elliottii) and longleaf pine (Pinus palustris) may be killed. Marginal hosts such as spruce are occasionally attacked and killed (Murphy 1917). There are up to eight overlapping generations per year (Hain et al. 2011). The range of SPB also includes Arizona, New Mexico, Mexico, and Central America (Clarke and Nowak 2009) and has recently expanded north into New Jersey (2009), New York (2014), and Connecticut (2015).1 The SPB is among the most aggressive pine bark beetles, capable of rapidly killing vast numbers of trees (Paine et al. 1997). Infestations (spots) are initiated when dispersing females locate a susceptible host, often one with weakened defenses such as a lightningstruck pine. Females release the aggregation pheromone frontalin, which when combined with host odors attracts conspecific males and females (Renwick and Vite´ 1969). Arriving males release several pheromones including verbenone and endo-brevicomin, the latter synergizing the attraction of frontalin at low concentrations (Sulli-

van 2011). At higher concentrations, verbenone and endobrevicomin help shift attacks to adjacent pines (Sullivan 2011). If the SPB population density is sufficient, SPB pheromone production can result in mass-attack behavior that overwhelms the defenses of healthy host trees, resulting in rapid tree death and expanding infestations (Gara and Coster 1968). If suitable host trees are present, immigrating, reemerging, and emerging adults can combine to create one or more expanding spot heads on the periphery of the infestation (Cameron and Billings 1988, Billings 2011a). Infestations can grow rapidly, adding ⬎1 new tree/day (Ayres et al. 2011). Large, unsuppressed infestations may reach thousands of hectares in size in less than 1 year (Clarke and Billings 2003, Billings 2011a). Historically, SPB populations were classified into two groups: epidemic (outbreak) or endemic (low density). Outbreaks have been defined as occurring when one or more infestations were detected for every 1,000 acres of susceptible pine host type (Price et al. 1998). Endemic periods included all other population levels. From 1960 through the late 1990s, the Gulf Coastal Plain experienced outbreaks every 6 –9 years, each lasting 2– 4 years (Birt 2011). The regularity of SPB outbreaks during this period caused many foresters and entomologists to believe that this pattern was the “norm” for SPB populations. However, in the late 1990s severe SPB population declines occurred in Arkansas, Louisiana, and Texas, and no active,

Manuscript received June 13, 2015; accepted December 18, 2015; published online February 4, 2016. Affiliations: Stephen R. Clarke ([email protected]), USDA Forest Service, Forest Health Protection, Lufkin, TX. John J. Riggins ([email protected]), Mississippi State University. Frederick M. Stephen ([email protected]), University of Arkansas. Acknowledgments: We thank Ronald F. Billings (Texas A&M Forest Service) and Jim Guldin (USDA Forest Service, Southern Forest Research Station) for their informative comments on an earlier draft of this article. We also thank the anonymous reviewers for their constructive suggestions. This review benefited greatly from discussions on the history of southern forest entomology at the East Texas Forest Entomology Seminar. 166

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expanding SPB infestations have been detected in these states since 1999 (Clarke 2012). The current absence of SPB-caused tree mortality in the western Gulf Coastal Plain raises many questions, including whether their absence is an aberration in the historical context of regional SPB dynamics or symptomatic of normal but poorly understood population fluctuations. There has been little consistency in attempts to explain historical SPB population oscillations, as variability occurs across its range and through time (King 1972). Factors linked to regional SPB population fluctuations can generally be divided into the following broad categories: abiotic factors (weather); biotic factors including predation (Reeve et al. 1998, Reeve 2011), parasitism (Berisford 2011), intra- and interspecific competition (Stephen 2011), and mutualists and phoronts (Hofstetter et al. 2006); and host suitability and availability (Lorio 1986). Weather and climatic variables set the parameters under which SPB can survive and define the rate at which they develop (Wagner et al. 1979). Extreme cold winters and very hot summers can cause significant brood mortality (McClelland and ` et al. 2007, Friedenberg et al. 2008), whereas Hain 1979, Trân rainfall and unfavorable temperatures may inhibit flight (Moser and Dell 1979, Geer et al. 1981, Thompson and Moser 1986). The availability of lightning-struck pines as epicenters for infestation development or as refuges during periods of low numbers or adverse conditions affects SPB population viability (Coulson et al. 1983, Hain et al. 2011). Predictive models based on abiotic factors have yielded mixed results. In two studies using SPB population density in East Texas, Kalkstein (1976) correlated outbreaks with climatic conditions, whereas Turchin et al. (1991) found that climatic variables were not significantly linked to population density changes. Models describing regional SPB outbreaks have included climatic variables, but the main predictor for current year activity was the proximity of active infestations the previous year (Gumpertz et al. 2000, Duehl et al. 2010). Duehl et al. (2010) even illustrated that average climatic conditions were more likely to spawn outbreaks than extreme conditions. Kroll and Reeves (1978) developed a model that used mean February temperature and rainfall in the previous spring, summer, and fall to predict the number of summer SPB infestations. However, their model fails to predict the absence of infestations experienced in the past 15 years west of the Mississippi River. These models were developed from data from 1960 through the mid-1990s, a period of frequent outbreaks, and appear to have little utility in forecasting long-term regional trends in SPB activity. Incorporating predation as a predictive variable, Billings (1988a) used the ratio of SPB and its major predator, Thanasimus dubius (Coleoptera: Cleridae), as the basis for a survey system for SPB population levels across the South. Turchin et al. (1999) found a density-dependent relationship between SPB and T. dubius, theorizing that their interaction can be a factor in SPB population collapses. Competition with Ips spp. and Monochamus titillator for host resources has also been recognized as an important factor in SPB outbreak dynamics (Coulson et al. 1976, Clarke and Billings 2003, Stephen 2011). The amount and species of fungi carried by SPB may be a factor in SPB population trends (Bridges 1983). The role of these SPB associates in long-term population trends is complicated by other factors affecting their numbers. For example, T. dubius preys on a variety of beetles, including Ips spp. (Mizell and Nebeker 1982), whereas Ips population numbers are often tied to drought or storm events (Connor and Wilkinson 1983).

The density and condition of pines can affect SPB population levels, but host availability has received relatively little attention in SPB population fluctuations. Unsuppressed outbreaks may kill 40 – 60% of the susceptible host type within the outbreak area (Clarke and Billings 2003, Maingi and Luhn 2005), as even healthy pines are killed once infestations become large and expand rapidly. Large acreages affected by SPB may revert to mixed pine-hardwood stands and remain at low hazard for considerable periods (Coleman et al. 2008). Although forest stand conditions have not generally been included in modeling of SPB outbreaks, they are the major factors used in the development of numerous SPB hazard rating systems (Mason et al. 1985) and infestation growth models (Stephen and Lih 1985). Pine basal area, total basal area, pine species, and mean annual growth increment are variables influential in predicting spot growth potential at the stand level (Mason et al. 1985). These stand characteristics are the result of forest management, or the lack thereof. The science of forestry in the southeast United States has changed dramatically over the past 150 years, yet the importance of prevailing forest management practices as a driver of regional SPB population dynamics has not received enough attention. Silvicultural systems are an important influence on the development and spread of SPB infestations (Guldin 2011), and Bennett (1965, 1968) and Hedden (1978) correlated prevailing silvicultural practices with recent SPB outbreaks. This relationship can be expanded to cover the recorded trends of SPB activity. The objective of this article are to describe historical patterns of forest management and to discuss how they are linked with SPB outbreak frequency.

Historical Patterns of Forest Management and SPB Populations Clarke et al. (2000) and Schowalter et al. (1981) provided an overview of how forest management practices could affect SPB population patterns in the West Gulf Coastal Plain (WGCP). In the following sections we describe the historical patterns of forest management and SPB activity in the southeast United States. Unfortunately, historical information on SPB population levels is plagued by several problems. Reports of SPB activity are sketchy and inconsistencies abound. It is unclear whether a lack of data for a particular time frame indicates that SPB populations were low or absent or whether pine mortality due to SPB was not reported or attributed to other causes. Hopkins (1903) noted that observers may have mistaken secondary bark beetles as the cause of mortality when the trees were actually killed by SPB. The definition of an outbreak or epidemic is also unclear. Balch (1928) indicated that epidemics may be widespread or local and stated that an epidemic occurs whenever SPB is capable of killing healthy pines. Texas historically reported SPB activity as the number of infestations per 1,000 acres, whereas Louisiana reported numbers of trees killed per 1,000 acres (Kalkstein 1976). Thatcher (1960), King (1972), and Price et al. (1992) provide lists of SPB outbreaks from the early 1900s through 1960, yet their outbreak years do not always match. For example, Price et al. (1992) list a southwide SPB infestation from 1911 to 1924, but Thatcher (1960) does not. In the following sections we attempt to reconcile the outbreaks of SPB before 1960 and to correlate SPB population patterns with the forest management practices in vogue during the different timespans. Forest Science • April 2016

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Civil War to World War I (WWI) Historically, shortleaf and longleaf pine were the primary species used for timber, whereas loblolly pine was a less abundant but important source (Fernow 1896). Before the arrival of European settlers, pure longleaf pine forests and savannas covered 37.2 million ha of the coastal plain region from Virginia westward to East Texas, with pine-hardwood and slash pine forests covering another 5.26 million ha (Frost 1993). Longleaf pine occupied the drier soils on ridges, plateaus, and south-facing slopes (Bonnicksen 2000, p. 227–236). Large-scale surface fires initiated by lightning burned through the longleaf pine forests every 1–3 years, helping to thin weaker trees and maintain this species in open, pure stands (Bonnicksen 2000, Frost 2006). Native Americans also contributed to the fire frequency, burning some pine forests once or twice a year (Frost 2006). Loblolly pine was restricted to moister sites and wet bottomlands, and stands were patchy compared with those of longleaf pine (Bonnicksen 2000, Stanturf et al. 2002). Shortleaf pine was the dominant pine species north of the longleaf pine region, generally occurring in mixed pine-hardwood stands (Fernow 1896). Slash pine had a smaller range and was limited to east of the Mississippi River (Lohrey and Kossuth 1990). Timber production in the South accelerated after the Civil War as supplies in the Northeast became depleted. There was little in the way of actual forest management. Cut-out and get-out was the primary strategy as migratory lumbermen worked their way through the region (Walker 1991, p. 101–102, Billings 2014). Forest conservation and renewal were not a high priority, only the number of board-feet that could be harvested (Clark 1984, p. 13). Timber theft was also rampant. Forests near navigable waterways were the first to be cut, but the expansion of the railroads soon allowed access into large forested areas (Williams 1989, p. 244 –261, Walker 1991, p. 89). Between 1870 and 1920 most of the virgin timber in the South was logged (Frost 1993), and approximately 5.1 million of the more than 49 million ha of the original pre-European pine forests remained by 1927 (Schultz 1997, p. 1– 6). Although clearcutting was the main harvest method, occasionally some seed trees were left or the stands were high graded, with the remaining stems achieving merchantability in 12–15 years (Walker 1991, p. 131). Six to eight seed trees per acre were left after harvest in some cutover longleaf pine stands (Duzan 1980). Shortleaf pine stands were often cut to diameter or stump limits (Mattoon 1915, Reynolds 1980). The smaller trees remaining on the land often were used for fence rails, ballast, or rail ties (Reynolds 1980, Clark 1984, p. 6). The residual pines sometimes provided for a second-growth forest, but hardwoods became more dominant in many sites as they were not removed (Reynolds 1980). The naval stores industry tapped the remaining longleaf pines for resin, with North Carolina the hub of the activity in the late 1800s (Butler 1998, p. 177). The process used often left the pines weakened and subject to destruction by fire or insects. As longleaf pines disappeared from North Carolina, the naval store industry shifted southward to Georgia and Florida in the early 1900s. New techniques resulted in far fewer instances of tree death after resin extraction (Butler 1998). Before 1920, southern pine plantations were scarce (Frost 1993), and less than 202.3 ha were successfully reforested (Wakeley 1954). The few seed trees left in longleaf pine stands were usually not sufficient to restock the cutover areas, often due to feral hogs (Frost 1993). Longleaf pine seed production was also irregular, with good seed crops occurring every 5–7 years (Earley 2004). Much of the 168


cleared forestland was used for agriculture or homesteads. Agricultural practices of the period resulted in significant soil erosion (Locke et al. 2010). As soils were eventually depleted and the boll weevil decimated cotton crops, vast acreages of agricultural land were abandoned throughout the South between the end of the Civil War and the start of WWI. Natural pine regeneration, primarily from the loblolly pine in bottomlands, began to encroach in these old fields and cutover areas which had converted to grassy savannas. Changes in fire management practices contributed to stand conversion. Native Americans burned the forests, and these surface fires benefited fire-adapted species such as longleaf pine. The forest fragmentation from the establishment of crops and homesteads by European settlers prevented landscape-scale low-intensity fires that had served to preserve pure longleaf pine forests, allowing hardwoods and loblolly pine to become established in the understory (Frost 2006). The low-intensity surface fires were replaced by uncontrolled wildfires, as fires set to burn brush or to clear duff in pine stands used for naval stores often escaped (Clark 1984, p. 9, 21). The resultant devastating forest fires could destroy regeneration, including seedlings and saplings of longleaf pine (Garren 1943). Wildfire sometimes benefited shortleaf pine, as burning resulted in sprouting and the development of coppice shortleaf pine stands (Mattoon 1915). The threat of wildfires led to the initiation of the fire suppression movement (Fowler and Konopik 2007), which allowed pine species not adapted to fire such as loblolly pine to replace longleaf pine. The absence of active forest management contributed to the development of SPB outbreaks. Extensive pine mortality in North Carolina in 1797 may have been caused by SPB (Walker 1991, p. 59), and Balch (1928) suggests that an SPB outbreak occurred in South Carolina around 1800. Price et al. (1998) described other potential SPB activity during this time period. The SPB was officially described in 1868 (Zimmermann 1868), and the identity and impact of SPB were initially clarified by A.D. Hopkins, the first head of the Division of Forest Insect Investigations in the US Department of Agriculture (USDA). Hopkins stated that the first known SPB outbreak was in 1842 and also reported that SPB killed a large amount of longleaf pine timber in Texas from 1882 to 1885 (Hopkins 1903, 1921). An extensive outbreak from 1890 to 1893 in Virginia, West Virginia, Pennsylvania, and North Carolina affected more than 75,000 square miles, killing pines of all sizes, even nonforest ornamentals (Hopkins 1903, 1911, USDA Forest Service 1947). Widespread infestations occurred Southwide between 1891 and 1911, with high tree mortality and economic losses of 2 million dollars from 1908 to 1911 (Hopkins 1921, Balch 1928, USDA Forest Service 1947). Chittenden (1904) documented the return of SPB to Georgia in 1903, although Newell (1904) erroneously stated that this activity was the furthest south that SPB has been reported. Craighead (1925) listed outbreaks in North Carolina from 1902 to 1905 and in Virginia and Tennessee from 1913 to 1916. However, only 2,000 trees were affected in the latter outbreak. No large-scale mortality was reported from 1912 to 1921. Based on these patterns, Hopkins (1911, 1921) stated that outbreaks occur at relatively long intervals. Balch (1928) claimed that outbreaks had occurred at 10-year intervals since 1890, although it appears that he probably meant every 10 years somewhere in the Southeast and not necessarily at the same location. Outbreaks were characterized by 1–3 years of intense activity; then became exceedingly rare in the intervening periods (Craighead 1925, St. George

and Beal 1929). The status of SPB populations between these outbreaks is unknown. The dominance of longleaf pine stands, which generally have fewer impacts from SPB than shortleaf and loblolly pine stands (Burns and Honkala 1990, Friedenberg et al. 2007), plus the lack of regeneration after clearcutting probably influenced the low frequency of outbreaks. Infestations may have been primarily confined to loblolly pines in low-lying areas. As the longleaf pine forests in the uplands became overmature with slow radial growth, the trees became susceptible to SPB (Belanger 1980), and infestations may have moved into these areas resulting in the development of large outbreaks. Suppression actions were minimal (USDA Forest Service 1947), and most outbreaks were allowed to run their course, although cutting and burning infested trees for fuel contributed to the decline of outbreaks in the South Atlantic and Gulf States in 1910 –1911 (USDA 1913, p. 149). Hopkins (1899) attributed the decline of an outbreak in West Virginia to cold, wet conditions coupled with an increase in fungi infecting the beetles. Craighead (1925) linked the onset of outbreaks with periods of rainfall deficiency and high precipitation during attack and brood development with population declines. Host depletion may also have played a role in halting infestation growth and outbreak collapse.

WWI to WWII The lumber industry declined in the South once much of the mature sawtimber had been harvested. Around 46% of all timber cut in the United States came from the South in 1909 (Williams 1989, p. 272), but production subsequently diminished and remained low through the Great Depression in the 1930s. The large demand for cotton after WWI quickly dropped when prices fell, leading many southern farmers to turn to pine production for an income (Elliott and Mobley 1938, p. 49 –50). Many of these second-growth forests were in poor shape due to logging, fire, and overgrazing. The annual growth was slow at around 2.1 m3/ha. Half of southern US forests were in farm woodlands, much of it loblolly pine that had encroached into abandoned fields. The establishment of peckerwood (small-scale) sawmills led to the depletion of many of these second-growth forests before they reached maturity (Clark 1984). The demand for rail ties also contributed to clearing of the remaining mixed forest stands (Anonymous 1906, Yarnell 1998). To improve the poor forest conditions, increased forest management was initiated during this era. Forestry was promoted on the state level as southern states began to form forestry commissions or departments (Clark 1984, p. 49). Nationally, forest reserves that had been established by the Forest Reserve Act of 1891 and managed by the Division of Forestry in the Department of the Interior were transferred by President Theodore Roosevelt to the Department of Agriculture. The Division of Forestry became the USDA Forest Service, with Gifford Pinchot named as the first Chief. Pinchot and the USDA Forest Service began promoting active forest management, and during the Great Depression, many landowners sold their forested property to the federal government for the National Forest System, increasing the acreage under professional management (Walker 1991, p. 145). The 1924 Clarke-McNary Act authorized federal cost-sharing support for state tree seed and nursery programs as well as increased educational and technical assistance to landowners (Zimmerman 1976, Billings 2014). Soon thousands of landowners and timber companies were receiving technical advice from state foresters supported by state and federal funds (Steen 1976). The USDA Southern Forest Experiment Station was established in 1921

and began investigating methods of forest conservation and the restoration of cutover lands. In the late 1920s and 1930s the cut-out and get-out philosophy evolved to selective management, and forestry began to emerge as a science as American foresters returned home after training in Europe, introducing scientific principles to forest management (Elliott and Mobley 1938, p. 3). With the use of European forestry principles, attempts were made to manage the remaining mature forests by selection cutting, but this silvicultural practice generally failed because of the high grading of the stand during the first cut (Smith 1972). Longleaf pine was still considered the most valuable pine because of its straight, tall growth and its importance for the production of naval stores. Despite its value, longleaf stands across the South continued to decline. For example, by 1935 only 2.9% of virgin longleaf pine stands were uncut, and 43% of previous longleaf pine type in Texas and Louisiana that had been previously cutover remained in clearcuts (Bridges and Orzell 1989). Only around 6.1 million ha of longleaf pine remained southwide by 1946 (Frost 1993). Feral hogs and the exclusion of fire continued to affect the ability of longleaf pine to regenerate (Frost 1993), and natural loblolly pine regeneration continued to encroach in cutover areas. Slash pine was also valued for its fast growth, resin production, and use for construction timber (Elliott and Mobley 1938, p. 288 –289). Some managed, mature forests were maintained in longleaf or slash pine for production of naval stores until the extraction procedures left the trees weakened and susceptible to fire and insect damage (Butler 1998). The SPB outbreaks before WWI documented above also reduced the acreage of pure, mature pine stands. Sparse pine regeneration has been documented in the large areas affected by SPB (Coleman et al. 2008), and hardwoods frequently became dominant after SPB removed the pine overstory (St. George and Beal 1929). The mixed composition of the forests after infestations reduced the potential for frequent SPB outbreaks (Balch 1928, Hoffman and Anderson 1945). In addition, the few living pines left within areas of near 100% mortality (escape trees) have been shown to be more resistant to SPB attack (Strom et al. 2002). Before intensive forest management, these trees may have provided a seed source for establishment of new pine forests in the affected area, perhaps reducing the impacts or delaying the occurrence of the next outbreak. The lack of forest management, which was still in its infancy, the widespread host removal from previous outbreaks, and other factors discussed above probably contributed to a sharp decrease in SPB activity between 1920 and 1950. Only one extensive southwide outbreak was reported, occurring between 1922 and 1923 (Howard 1924, Craighead 1925, St. George and Beal 1929). A small outbreak during this time frame developed in overmature longleaf and slash pines in Florida that had been severely damaged by turpentining (Wyman 1924). Thereafter, most reported outbreaks were localized and short-lived. Scattered infestations were noted at various sites across the South from 1924 to 1948 (St. George and Beal 1929, St. George 1930, Hoffman and Anderson 1945, Hetrick 1949), and the USDA Forest Service (1947) listed localized outbreaks in 1925–1927, 1931–1932, 1937–1938, and 1945–1946. Marlatt (1933) reported that SPB activity in 1933 in the northern part of its range was at its highest level in 40 years, but the report from the following year did not include any information that the outbreak had continued (Strong 1934). The Texas Forest Service recorded SPB activity in 1919 and 1926 (Billings 1988b), but it is doubtful that these were large-scale outbreaks given their short duration. Forest Science • April 2016

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Similarly, timber companies in Texas reported high levels of SPB activity in 1938 –1939 and 1944 –1947 (Billings 1988b), but these flairs also were confined to small areas. An isolated outbreak occurred in southern Pennsylvania in 1932–1933 (Knull 1934). Suppression of active SPB infestations was rare during this period. Most foresters were hesitant to attempt direct controls because they believed outbreaks were a result of long dry periods and that heavy rains would lead to population collapses (Elliott and Mobley 1938, p. 334). Others considered tree loss during periods of endemic population densities as beneficial (Gisborne 1950). Salvage of infested trees occurred only on a small percentage of infestations (St. George and Beal 1929). Despite the lack of suppression of most SPB infestations, it was during this time that the groundwork for modern SPB integrated pest management strategies was laid. One of the first major operational direct control actions occurred in 1939 on an infestation of over 100 ha in a remnant stand of virgin timber in Texas (Walker 1991, p. 191–192). United States and Texas Forest Service entomologists recommended cutting a half mile swath of pines around the infested area (Billings 1988b). Another group of infestations in Texas in 1944 –1945 was controlled by removal of the brood trees and the felling of a quarter mile buffer around each spot (Billings 1988b). Other direct suppression methods were tested. Craighead and St. George (1938) experimented with tree injections of chemicals for bark beetle control, but considered it impractical for SPB because of the rapid permeation of blue-stain fungi into the sapwood and the necessity of treating infested pines soon after attack. St. George and Craighead (1929) suggested that if infested pines could not be removed, then the trees should be felled and debarked, and the bark and tops burned. They also recommended felling and laying the trees in a north-south direction to allow solar radiation to kill the brood. Chapman (1942) provided similar suppression advice for infestations in Arkansas and Louisiana. It is not known whether these tactics were ever applied operationally. Another factor potentially affecting SPB populations was outbreaks of Ips bark beetles. St. George (1925) documented widespread tree mortality in Texas, Louisiana, Mississippi, and Alabama due to Ips spp. The widespread infestations and tree mortality in 1931–1932 that were often attributed to SPB were actually due to Ips bark beetles (Cary 1932a, 1932b). The tree loss due to these Ips outbreaks contributed to reduced availability of host type for SPB. In addition, the extraction of turpentine from longleaf and slash pine promoted attacks by the black turpentine beetle, D. terebrans, and Ips bark beetles, increasing populations of SPB competitors (Outland 2004, p. 102–106). Although SPB activity was low, changes in forestry that would affect future population levels were taking place. A 5-fold increase in the need for pulpwood was predicted in the 1930s in the South after the discovery that southern pines could produce acceptable pulp and paper (Reed 1995). Only about 1,214 ha of pine plantations existed in the South in 1926 (South and Buckner 2004). Paper and pulp companies used pine regeneration on former agricultural lands and constructed many mills to produce their product during the 1930s (Williams 1989, p. 287). The Great Southern Lumber Company planted approximately 2,833 ha of pines near Bogalusa, LA, between 1920 and 1926 (Wakeley 1954). In response to the great depression, President Franklin Roosevelt with his New Deal established the Emergency Conservation Work program in 1933, which evolved into the Civilian Conservation Corps in 1937 (Clark 1984, p. 74). These agencies planted more than 0.6 million ha in pines 170


across the South during the 1930s and early 1940s. Most of the pine plantations were established with loblolly or slash pine (Frost 2006, Zhang et al. 2010). The recommended tree spacings for loblolly and shortleaf pine were dense: 5 ft ⫻ 5 ft (4,304 stems/ha) or 6 ⫻ 6 (2,990 stems/ha) for pulp and timber and 3 ⫻ 3 (11,960 stems/ha) or 4 ⫻ 4 (6,726 stems/ha) for erosion control (Mobley and Hoskins 1956, p. 360). Slash and longleaf pine were planted at 6 ⫻ 6 to 6 ⫻ 10 (1,794 stems/ha) for timber and 10 ⫻ 10 (1,077 stems/ha) or 12 ⫻ 12 (746 stems/ha) for naval store production. Unmanaged woodlands were also converted to pine plantations (Frost 1993). Pulpwood production increased from almost zero in 1920 to 14 million cords in 1953 (Mobley and Hoskins 1956, p. 27). These dense pine plantations would provide the high-hazard habitat for SPB outbreaks in the last half of the 20th century. Wildfire suppression also increased dramatically during this period and firefighting became an emphasis of the Forest Service and Civilian Conservation Corps. Fire had been a tool for land clearing, but its use decreased after forests were converted to agriculture. Wildfires burned an average of 15.4 million ha annually from 1931 to 1935, but increased fire prevention and suppression efforts led to sharp reductions in acreage burned thereafter (Southern Forest Resources Analysis Committee 1969).

WWII to Late 1900s Despite the initiation of planting efforts in the 1930s, much of the former forested land in the South remained understocked in the 1950s. Only 2 million acres were in planted pine forests, and the remainder of the forestlands was naturally regenerated stands with little management. Wakeley (1954) estimated that 5.3 million ha in the South were poorly or nonstocked at the time of WWII. The USDA Forest Service began promoting treeplanting as a method to turn these ha into productive forests (USDA Forest Service 1958a). The Soil Bank Program, established in 1956, contributed in the conversion of farmland into forests (Dangerfield et al. 1998). Between 1925 and 1980, approximately 11.7 million ha were planted or reseeded in the southern states (Williams 1989, p. 479). The acreage of planted pine reached 12.1 million ha by 1997 (Smith et al. 2001). A major portion of the acreage planted in the South was on land retired from crop use (Williams 1989, p. 472). Wetlands were also drained and used for tree production (Walker 1991). Recommendations were to plant the species that showed the best performance rather than the species that formerly occupied the site (Wakeley 1954). Therefore, loblolly pine was the tree species of choice in many areas across the South, and slash pine plantations often replaced longleaf pine in the eastern half of the Gulf Coastal Plain (Bridges and Orzell 1989). The loss of longleaf pine acreage continued, with 76 and 82% decreases in Louisiana and Texas, respectively, between 1955 and 1985 (Outcalt 1996). Only 1.3 million ha of longleaf pine remained southwide by 1994 (Frost 1993). Stocking levels in plantations were usually very dense, with a recommended density of 6 ⫻ 6 ft or 8 ⫻ 8 ft (Wakeley 1954). Foresters were cautioned not to use tree spacing of 10 ⫻ 10 ft or wider where poor survival might be expected (Walmer et al. 1975). The push to promote regeneration was accompanied by an increased emphasis on forest management. The end of the war and the baby boom that followed caused an increased demand for lumber for housing. Active management of forests became more common, with forestry practices in the South on both private and federal lands geared toward the production of fast-growing pines in pure stands that could be harvested for pulp or sawtimber. In the 1940s, the

government began providing small landowners with incentives to manage their forests (Walker 1991, p. 214). The trees they grew could then be purchased by industry. Even-aged management was the primary silvicultural technique, and clearcutting was the harvest method of choice (Baker et al. 1996). Site preparation methods became widely used, including shearing, disking, chopping, and burning (Fox et al. 2004). The objectives of site preparation were to eliminate competition and reduce logging slash (Fox et al. 2007). Fertilization of pine plantations eventually became commonplace, with more than 4.4 million ha treated between 1969 and 2000 (Fox et al. 2004). Fire exclusion was a standard practice, and the area affected by wildfire decreased from 38 million ha/year in the early 1930s to 2 million/year in the mid-1960s (Southern Forest Resources Analysis Committee 1969). The area burned annually fell to only 1 million ha by 1999 (South and Buckner 2004). The reduction in burning in the forests resulted in increased hardwood encroachment in pine stands over time. Forest managers turned to chemical control to reduce competition. The herbicide 2,4,5-T was applied extensively to release young pine plantations because it was inexpensive and pines were resistant (Lowery and Gjerstad 1991). These practices resulted in a dramatic increase in softwood timber production in the South from 85 to 150 million m3 between 1952 and 1992 (Wear 1996). However, not all forests were productive, and by the 1970s a majority of the 41% of timberlands owned by nonindustrial landowners and the 30% owned by farmers were poorly managed. Larger forested tracts became more common, with more acreage owned by state and federal governments, industry, and large private landowners (Clark 1984). The forest industry increased its share of southern forests from 17 to 23% between 1952 and 1992 (Wear 1996). Urban encroachment also contributed to the decrease in small privately owned forests. These changes resulted in an overall reduction in forested land between the 1960s and the early 1980s (Clark 1984, Wear 1996), but a higher percentage of forests under active management. Forest management intensified dramatically in the 1980s and 1990s, particularly on industry and other private forestlands (Siry 2002). Environmentalism also came into vogue in the 1960s and 1970s (Walker 1991, p. 203). Recreation, preservation, and wildlife were now viewed as equally if not more important than timber when determining how forests on federal lands should be managed, and the management emphasis of dominant use evolved to multiple use and then to environmentally sensitive, multiple use (Yaffee 1999). Environmental groups pushed for wilderness and old-growth, and lawsuits were initiated to halt clearcutting and other harvest methods to preserve forestlands. The Wilderness Act in 1964 led to the consideration and designation of roadless areas as wilderness. Despite improved forest management after WWII, the South provided vast areas of susceptible host type for SPB during the last half of the 20th century as a result of overstocking, high grading, extreme fire suppression, and other practices used before WWII. Forests were dense and loblolly pine was a much larger component than in presettlement times (Perkins and Matlack 2002, Van Lear et al. 2004). Many pest eruptions during the last half of the 1900s were apparently promoted by forest and fire management practices (Coulson and Stephen 2006), and southern forests during this time period experienced widespread, frequent, and often economically and ecologically important SPB outbreaks (Figure 1). An outbreak began in Texas in 1949 and lasted through 1951 (Lee 1954, Thatcher 1960). Mississippi and Alabama had outbreaks starting in 1952, subsiding in 1956 (USDA Forest Service 1954, 1956). In

Figure 1. Volume of trees killed by the SPB, Dendroctonus frontalis, in the southeastern United States, 1960 –2014. Data are from “A History of Southern Pine Beetle Outbreaks in the Southeastern United States through 2004” (www.srs.fs.usda. gov/econ/data/spb/) and the SPB Data Portal (svinetfc12. fs.fed.us/SPB_DataPortal/).

1957 Louisiana had localized SPB activity evident for the first time in more than 40 years (USDA Forest Service 1957), although the extent of infestations may not have reached the outbreak level. The first reported SPB activity in Arkansas since the early 1900s occurred in 1969 (USDA Forest Service 1969). In the eastern half of its range, a widespread SPB outbreak began in 1952–1953 in northwest South Carolina, northeast Georgia, east Tennessee, and west North Carolina, not subsiding until winter 1957 (USDA Forest Service 1958b). Ford (1951) reported a 10-county outbreak in eastern North Carolina in 1950 –1951 and described an area in Nash County that had extensive timber losses 5 years earlier. A severe outbreak occurred in central Virginia from 1953 to 1956, and 6 years later in 1962 another outbreak affected 6 counties in the Piedmont region (Knox and Schroeder 1963). Once outbreaks began in areas of the Gulf Coastal Plain, they seemed to reoccur every 6 –10 years over the next 30 – 40 years. SPB activity transitioned from periodic to chronic (Williamson and Vite´ 1971). The Texas Forest Service began tracking SPB infestations in East Texas in 1959 and can document five major outbreaks between 1960 and 1994, with the most destructive from 1984 to 1986 (Billings 1995). Hain and McClelland (1979) reported frequent eruptions in SPB activity in North Carolina between 1960 and 1977, with a severe outbreak occurring from 1973 to 1976. Price et al. (1998) provided data on SPB beginning in 1960, although data before 1973 from some states are missing. The number of infestations is not included, just the volume of timber affected; thus, volume is used to represent the frequency and intensity of SPB activity in Figure 1. Based on the volume lost, states (with the exception of Florida) bordering the Gulf of Mexico or the Atlantic Ocean had four outbreaks between 1973 and 1996. It is likely that these states had at least one to two additional outbreaks in the 25 years before this period. The USDA Forest Service began tracking infestation levels on National Forests in 1985 using the Southern Pine Beetle Information System (SPBIS) database (Peacher 2011), and the peaks in SPB infestation were similar to those presented in Price et al. (1998). The length and severity of these outbreaks varied considerably, but most lasted at least 2 years. Forest Science • April 2016

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The large increase in pine acreage and the conversion from longleaf pine to loblolly pine were major factors in the frequent, repeated eruptive behavior. The dense loblolly pine monocultures planted extensively across the South starting in the 1930s resulted in forests with increased susceptibility to SPB infestation. Timber stand improvement removed hardwoods producing green-leaf volatiles that can inhibit SPB’s response to its pheromones (Dickens et al. 1992, Sullivan et al. 2007). Once one pine was successfully colonized, the infestation could spread rapidly in the pure, crowded pine stands. Initially infestations and outbreaks appeared concentrated in mature stands of pines. These were areas that had been planted or had seeded into pine in cutover areas and abandoned fields. These forests often were unmanaged and grew into dense, mature stands of susceptible hosts by the mid-1900s. By the 1980s, many of the old, high-hazard natural stands had been harvested or killed by SPB, and subsequently pine plantations began experiencing SPB activity (Cameron and Billings 1988). Many of these plantations had not been thinned and were becoming highly susceptible to SPB, with even higher hazard in dense pine stands without hardwood removal (Kushmaul et al. 1979, Hedden and Reed 1980, Mason et al. 1985), as increased competition reduced tree vigor and radial growth (Ku et al. 1980). Longleaf pine stands had been affected by SPB before WWI; however, they had very little mortality after WWII compared to that for shortleaf, Virginia, and loblolly pine stands (Burns and Honkala 1990, Clarke et al. 2000, Friedenberg et al. 2007). One study indicated that the susceptibility of individual longleaf pines to SPB is similar to that of loblolly pines (Martinson et al. 2007); conversely, Hodges et al. (1979) found that longleaf pines were more resistant due to high flow and yield of oleoresin. Friedenberg et al. (2007) suggested the SPB have behavioral mechanisms that cause them to avoid longleaf pine. Although the susceptibility of pine species to SPB may vary, stand factors also may be involved in the reduced impacts observed. The remaining longleaf pine stands are generally located on sites suitable for this species; thus, the trees may be more vigorous than loblolly pines on poor sites. The spacing in longleaf pine stands is also wider than that for loblolly pine, and the host tree maps prepared by the Forest Health Technology Enterprise Team illustrate a lower basal area for longleaf pine stands (most ⬍13.8 m2/ha) southwide than for loblolly pine stands.2 The proximity of adjacent pines affects the rate of spread (Johnson and Coster 1978, Hedden and Billings 1979), so colonization of a lightning-struck or weakened longleaf pine by SPB would be less likely to produce an expanding infestation than a similar situation in a loblolly pine stand. Initially pest detection and suppression programs in the South were weak, and a push began to treat insect outbreaks similar to fire (Kowal 1955). As state and federal forest insect programs expanded in the last half of the 20th century, applications of SPB suppression treatments became widespread, and outbreaks were no longer allowed to run their course in accessible areas in most forests. In the early days of SPB suppression (1949 –1970), trees involved in expanding infestations were felled and either removed or sprayed with BHC in diesel oil (Billings 2011b). The insecticide applications also killed the natural enemies and competitors of SPB (Williamson and Vite´ 1971). The lack of effectiveness in reducing areawide SPB populations, coupled with detrimental effects on SPB associates and environmental concerns, led to the diminished use of insecticides after 1970. Cut-and-remove became the recommended suppression option, as the SPB in the trees were removed from the site and the 172


landowner received a financial return on the treated trees (Swain and Remion 1981, Billings 2011a). However, all affected trees were usually removed, including trees vacated by SPB that still contained predators and competitors, which take longer to emerge than SPB. Although removal of infested trees to mills has not been demonstrated to result in infestations along the transport route, it could augment and help maintain low-level populations of SPB outside of the outbreak area, as 100 SPB can emerge from 1 ft2 of infested bark (Bennett 1965). Cut-and-leave was used wherever cut-and-remove was not practical (Swain and Remion 1981). The felling of the infested trees plus a buffer of uninfested pines usually stopped infestation expansion (Clarke and Billings 2003), but SPB brood survival in the felled trees is similar to that in untreated standing pines (Hodges and Thatcher 1976). Despite limited impact on SPB brood survival, cut-and-leave treatments in summer do not result in an increase in new infestations in the surrounding pine forests (Billings and Pase 1979, Fitzgerald et al. 1994). Leaving the felled trees in place, as well as the pines vacated by SPB, supplied ample breeding material for SPB competitors (Connor and Wilkinson 1983, Stephen 2011) and also helped maintain clerid populations. Host depletion was no longer a major factor limiting the onset of new outbreaks except in inaccessible areas. Although SPB suppression was effective in minimizing tree and economic losses, the large acreages of susceptible host type not killed by SPB during outbreaks due to suppression measures remained available for future infestation if timely harvests or other preventative measures were not used (Clarke and Billings 2003). Thus, the swift suppression measures contributed to the prevalence of high-hazard stands across the South and probably helped fuel the more frequent outbreaks observed between 1960 and the mid-1990s. Although still eruptive, primary bark beetles (Hain et al. 2011), SPB populations no longer experienced extreme boom-crash cycles with extended periods of low activity. Sufficient numbers of beetles were available during the crash periods to allow a rapid return to outbreak status. Although SPB suppression may have influenced the increased periodicity of outbreaks, it also significantly reduced the total tree loss over time based on comparisons between areas with suppression and no suppression (Clarke and Billings 2003, Maingi and Luhn 2005). In addition, the areas affected by infestations and their treatment generally remained as pine stands due to reseeding from the protected pines around the edge of the openings and the previous removal of hardwood competition. The establishment of federal Wildernesses across the South in the 1980s also affected SPB populations. The Four Notch area in Texas was under consideration for wilderness status when an SPB outbreak started in 1983 (Billings and Varner 1986). Initially the infestations were allowed to run their course due to pressure from environmental activists, but the spots soon merged into large expanding infestations that threatened private land. Once the decision to suppress infestations was finally made, the spots were difficult to control. Based on this experience, suppression actions were applied swiftly in the newly established Texas Wildernesses in 1985 and 1986. These entries into Wilderness were fought by environmental activist groups and ultimately led to the adoption of the Final Environmental Impact Statement for the Suppression of the SPB (USDA Forest Service 1987). This document limited suppression in Wilderness, and large acreages of pine mortality were the result (Clarke and Billings 2003). Kisatchie Hills Wilderness in Louisiana and the Sipsey Wilderness in Alabama also had major pine loss due to SPB, as did the Big Thicket National Preserve in Texas in the mid-1970s.

The effects of the large unsuppressed infestations in Wildernesses and Preserves on SPB population numbers in the surrounding areas are not known, but it is likely that large numbers dispersed into nearby forests, perhaps increasing the intensity and/or extent of outbreaks.

Late 1900s to Early 2000s Although the intensive forest management initiated in the 1980s continued into the 2000s (Siry 2002), the focus of management in the southeastern United States began to change substantially. The debate over the environmental impacts of widespread timber harvests in southern forests that began in the 1970s resulted in increased management of forested areas for recreation, wildlife, and wilderness rather than just wood products (Walker 1991, p. 227). This environmental awareness led to a decrease in intensive pine plantation management and clearcutting on Federal forestland in the 1980s (Baker et al. 1996). Large-scale clearcutting was replaced by smaller clearcuts (⬍30 ha) or thinning. A major effort was undertaken to return longleaf to its original distribution on suitable sites (Brockway et al. 2005). Although species conversion has a long way to go, the acreage planted in longleaf pine continues to increase. In addition, the SPB Prevention Program was initiated in 2003 (Nowak et al. 2008). This program has provided funds to private, state, and federal forest managers for hazard reduction activities such as thinning and longleaf conversion, and more than 0.4 million ha have been treated to date. Commercial thinning programs are predicated on having outlets and markets for the wood removed, and an additional loss of sawmills and other wood-processing plants southwide could threaten the continuation and expansion of thinning operations. Another major forest management emphasis over the past 30 years was to return fire to pine ecosystems. Many federally managed forests across the South are now being burned on a 3- to 5-year rotation, and state forestry commissions recommend a similar burn frequency on privately owned pine forests. These fires reduce competition for the pines and favor the regeneration of fire-associated ecosystems such as longleaf pine-bluestem and shortleaf-bluestem, and thus are an important component in the restoration of these pine types (Guldin 2008). The establishment of seed orchards to produce genetically improved pines for replanting efforts began in the 1950s, and by the 2000s most seedlings planted were from second-generation orchards (McKeand and Svensson 1997, McKeand et al. 2003). Although genetically improved strains were selected for increased growth characteristics and not based on susceptibility to SPB, their use could still affect SPB populations. Fast-growing, vigorous pines may be able to repel potential colonizers by entombing the beetles in resin (Paine and Stephen 1988), and poor radial growth has been associated with increased stand hazard to SPB (Belanger 1980). However, improved genetic stock may quickly reach a susceptible size, accelerating the development of high-hazard stands requiring early and/or more frequent thinning for SPB prevention. The gains in seedling survival rates have led to a decrease in planting density for southern yellow pines, with a recommended maximum spacing of 8 ⫻ 10 ft (Nowak et al. 2008). Amateis and Burkhart (2012) determined that optimal spacing for loblolly pine pulpwood production was 8 ⫻ 8 ft, with wider spacing better for sawtimber growth. The reduction in intraspecific competition should also improve tree vigor. Although high-quality pine seedlings are now available, the acreage in pine plantations has decreased, and landownership (and associated man-

agement goals and practices) has also changed dramatically. In 1992, 60% of pine plantations were on industry land versus 6.9% on other private land (Wear 1996). Natural pine stands in the South were still dominated by loblolly pine (South and Buckner 2003), and loblolly pine was still the primary pine species planted on private lands (McKeand et al. 2003). Federal land managers shifted away from planting loblolly pine, and openings were either allowed to reseed naturally or regenerated with shortleaf or longleaf pine where appropriate. In the period from 2012 to 2015, the Southern Region of the USDA Forest Service regenerated 9,376 ha in longleaf pine, 3,193 ha in shortleaf pine, and only 79 ha in loblolly pine (George Weick, USDA Forest Service, Facts database, pers. comm., Jan. 13, 2016). The recovery plan for the red-cockaded woodpecker (RCW), Picoides borealis, drives forest management on many southeastern National Forests. The RCW is an endangered species that uses living pines (preferably longleaf) for nesting and breeding (Conner et al. 2001). Court-ordered management and the RCW Recovery Plan have spurred much of the longleaf conversion and prescribed burning described above. RCW clusters and recruitment stands are maintained at low pine basal areas (usually ⬍13 m2/ha) and at extended rotation ages (120 years for loblolly and 200 years for longleaf). Although longer stand rotations are now common on federal forestlands, shorter rotations have become the norm elsewhere. Removals steadily increased over the past 50 years on private lands, compared with a recent decline on public forests (USDA Forest Service 2001). Most forestland in the South once owned by timber companies has been sold to investment groups such as real estate investment trusts (REITs) and timberland investment and management organizations (TIMOs) (Stanturf et al. 2003). These groups manage their forests for quick turnaround, usually between 10 and 15 years (Fox et al. 2007). To increase profit margins, REITs and TIMOs may only thin when economically justified. However, the stands are usually harvested before or soon after becoming a high hazard for SPB infestation. The frequency and severity of SPB outbreaks has subsided (Figure 1). In the Western Gulf Coastal Plain, SPB activity has been at extremely low levels since the late 1990s. A 3-year major SPB outbreak in Texas collapsed in 1994, followed by localized infestations in 1997. No infestations have been detected since then. No infestations have been reported in Arkansas since 1997, and Louisiana has recorded only one infestation west of the Mississippi River since 1998. Mississippi experienced a low-level outbreak from 2000 to 2004, with losses reaching only a fraction of those sustained in previous outbreaks in the 1970s and 1980s. A 1-year outbreak occurred on the Homochitto National Forest in southwest Mississippi in 2012, whereas another isolated outbreak from 2013 to 2014 was confined to the Trace Unit of Tombigbee National Forest in northeastern Mississippi. In the eastern United States, a widespread outbreak occurred in North Carolina, South Carolina, Tennessee, Kentucky, Virginia, Georgia, and Florida from 1999 to 2003 (Nowak et al. 2008). Much of the tree mortality in the northern part of the range occurred in mountainous areas where accessibility was limited for both forest management and infestation suppression. An outbreak in Georgia affected 3,440 ha in 2007 (USDA Forest Service 2009) but collapsed the next year. Smaller 1-year outbreaks also developed in South Carolina and Alabama in 2005 (USDA Forest Service 2006). All other scattered SPB activity across the South has remained below outbreak levels. These short-term outbreaks are a divergence from the 2- to 3-year outbreaks typical in the last half of the 20th century (Hain et al. 2011). In general, the scattered activity Forest Science • April 2016

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has remained at non-outbreak levels in the eastern portion of the SPB range in the United States since 2002. The reduced SPB activity over the last 10 to 12 years throughout most of the South can be partially tied to the changes in forest management described above and the intensive management practices implemented in the 1980s (Siry 2002). On private lands, increased thinning has reduced the susceptibility of formerly highhazard stands, and short rotation periods result in the harvest of many pine stands before they become high hazard. The SPB prevention program has proven effective, as an evaluation of a recent outbreak in Mississippi found that almost every expanding SPB infestation occurred in unthinned stands (Nowak et al. 2015). Although thinning and the conversion from loblolly pine to longleaf pine has reduced SPB hazard throughout the South, many ha still await treatment. The reintroduction of fire to the ecosystem may also affect forest risk from SPB attack. Prescribed burning may temporarily increase host susceptibility and should be avoided in young pine plantations in areas experiencing SPB outbreaks (Cameron and Billings 1988). However, the reduction of competition from fire can increase tree vigor and could improve stand resistance in the long term (Boyle et al. 2004), and fewer infestations develop in recently burned stands than in unburned stands (Nowak et al. 2015). RCW management also has mixed effects on SPB susceptibility. RCW cavity trees have been preferentially attacked by SPB (Conner and Rudolph 1995), and the longer rotation ages of stands preserves old and overmature pines that often have increased susceptibility (Belanger 1980, Hicks 1980). Although individual RCW cavity trees may be preferentially colonized by SPB, the lower pine basal area in RCW-managed stands reduces the SPB hazard and inhibits the development of expanding infestations (Nebeker et al. 1995). The swift suppression of infestations in previous outbreaks served to create small patches throughout the pine forests. Forest ownership patterns also changed, as the number of private forest holdings less than 100 acres increased, whereas larger private parcels decreased in number (Birch 1997). This patchiness, combined with the initiation of the SPB Prevention Program, created a more diverse and fragmented landscape for SPB. Forest fragmentation reduces impacts from SPB (Cairns et al. 2008, Costanza et al. 2012) and was probably a major factor in the reduction in the frequency and magnitude of outbreaks in the South in the past 15 years. Southern pine beetle control strategies have also changed in response to modern timber industry trends. The loss of timber markets and mills led to an increase in cut-and-leave treatments instead of the cut-and-remove suppression technique favored in the 1960s–1990s. During the outbreaks on the Kisatchie National Forest in the mid-1980s and the National Forests in Texas in the early 1990s, almost 3/4 of all infestations requiring suppression were treated by cut-and-remove (Nettleton et al. 1988, Clarke and Billings 2003). In contrast, all infestation treatments on the Oconee National Forest in Georgia in 2007 used cut-and-leave (SPBIS database). Cut-and-remove was used on only 11 of 213 infestations treated during the 2012 outbreak on the Homochitto National Forest (Meeker 2013). Even though SPB are not removed from the forest in cut-and-leave treatments, this method still has some advantages over cut-and-remove. On National Forestlands, cut-and-leave treatments do not require a timber sale contract, so treatments can be applied on average 2 weeks faster than cut-and-remove treatments. Delays in treatment implementation result in increases in infested trees and higher SPB population numbers. SPB predators 174


and competitors are left at the site in cut-and-leave treatments, plus the southern Ips and Monochamus species readily attack the recently felled pines, which can accelerate their population increases and intensify interspecific competition (Connor and Wilkinson 1983, Stephen 2011). In addition, infested materials from cut-and-remove treatments were routinely transported long distances, often from outbreak areas to sawmills in areas with low-level SPB populations, potentially helping augment and maintain local populations. Rising fuel costs and the increased use of cut-and-leave have greatly reduced the movement of infested logs, and most trees from cut-and-remove treatments are processed locally. As a comparison, large SPB outbreaks and widespread activity have not decreased in frequency in Central America and southern Mexico since the 1960s and continue to plague their pine stands (Bennett 1965, Billings et al. 2004). Historically, very little forest management has been implemented in the region. Suppression applications, primarily cut-and-leave, have increased over the past 30 years, but delays in application usually result in spots expanding to several ha before treatment. The large forested areas killed will not reach high-hazard status for many years, but SPB infestations continue to develop in the remaining pine stands, causing frequent eruptions in activity. SPB infestations have also recently developed in the northeastern United States, with mortality occurring in New Jersey, New York, and Connecticut. This northern movement is probably driven by climatic factors combined with a lack of management of the pine forests. The recent trend of SPB outbreak frequency and duration in the South can be expected for the near future if current management practices and fragmentation patterns continue (Costanza et al. 2012). The potential for extensive and damaging outbreaks still exists, particularly on federal lands. The 2013 National Insect and Disease Risk Map3 indicated large areas of high hazard for SPB infestation on National Forest and National Park Service ownerships. Funding and staffing shortfalls, coupled with a lengthy environmental analysis process, have not allowed thinning to keep up with tree growth. Longer forest rotations and the number of ha set aside for limited or no thinning and pest suppression also have contributed to increased hazard. Although longleaf pine restoration efforts are continuing, the number of longleaf pine seedlings planted is only a fraction of the number of loblolly pine seedlings deployed (McKeand et al. 2003). The closure of mills and the lack of outlets for the stems removed during thinnings could slow the pace of forest management. These and other possible changes in forest management could ultimately result in new patterns of SPB outbreak frequency and intensity in coming years.

Summary Forest management practices shape the structure and condition of hosts available for SPB populations, regulating the initiation, directionality, and rapidity of expanding infestations, both in current times and in the following decades. Control tactics affect the population levels of SPB and their associates and govern the amount and location of susceptible host type remaining after treatments. Although prompt SPB suppression ensures that susceptible host type remains after an outbreak, it provides forest managers with a window of opportunity to implement prevention measures before the onset of another outbreak. Forest management and infestation suppression have been intrinsically linked with historic SPB activity (Figure 2). In the 1800s,

Figure 2. Relationship of forest management and SPB suppression practices with SPB outbreak frequency, longevity, and intensity, 1865–2014.

unmanaged contiguous forests allowed the development of expanding, coalescing SPB infestations affecting large areas. These eruptions were infrequent, as the infested areas often converted to hardwood, and it took time for any pine regeneration to emerge and stands to become susceptible again. The widespread removal of pine stands and their conversion to agriculture in the late 1800s through the early 1900s significantly reduced the amount of pine acreage, resulting in infrequent and localized SPB activity throughout the first half of the 1900s. The widespread establishment of pine plantations (monocultures), the introduction of forest science to forest management, and an increase in SPB suppression contributed to the frequent, cyclic SPB behavior in the coastal plain during the latter half of the century, although the latter two also helped reduce the overall acreage affected by outbreaks. Intensive forest management practices initiated in the 1980s, including increased emphases on species restoration, tree genetic improvement, and SPB integrated pest management, including prevention and prompt suppression, have played a role in the decrease in SPB activity across the South observed thus far in the 2000s. Though forest management practices are important in regulating SPB population dynamics, climatic conditions and population levels of the natural enemies and competitors of SPB are equally and sometimes more important. However, the manipulation of forest stand structure plus applications of infestation suppression tactics are the factors that can be used by forest managers to reduce SPB impacts and influence the severity and extent of population fluctuations. The historical patterns of forest management and the associated SPB population levels should provide foresters and forest entomologists with insight on future SPB trends and the appropriate silvicultural methods to reduce impacts. We hope this article will serve as the foundation for future investigation into the long-term

impacts of southwide forest management practices on the periodicity, intensity, and expanse of SPB outbreaks. Endnotes 1. For more information, see Connecticut Department of Energy and Environmental Protection (DEEP), www.ct.gov/deep/cwp/view.asp?a⫽2697&q⫽5,63,452. 2. For more information, see foresthealth.fs.usda.gov/portal. 3. For more information, see foresthealth.fs.usda.gov/nidrm/.

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