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The following document is also available for download.
Clickable Table of Contents I. Presettlement Vegetation and Natural Disturbance
II. Vegetation and Land Use After European Settlement
Transition hardwoods, also called transition hardwoods-white pine-hemlock (Westveld et al. 1956) or white pine-hemlock-hardwood (Lull 1968), contains species characteristic of zones to the north (northern hardwoods) and to the south (central hardwoods). Here, we consider Westveld et al.’s (1956) central hardwoods-white pine-hemlock and transition hardwoods-white pine-hemlock together, matching Lull’s (1968) map, on the basis that Westveld et al. (1956) described the boundary between the two zones as difficult to distinguish and that, unlike the Westveld et al.’s (1956) central hardwoods zone, both zones contained white pine as a dominant species. The transition hardwoods vegetation zone forms a band extending from the coast of Maine and New Hampshire through Massachusetts and northern Connecticut into southern New York and central Pennsylvania (Figure 1). Transition hardwood forests also extend north along major waterways that include the Connecticut and Hudson Rivers, the Saint Lawrence Seaway, and Lake Champlain. Transition hardwood forests are characterized by hardwoods such as beech (Fagus grandifolia), birches (Betula lenta, B. alleghaniensis, B. populifolia), maples (Acer saccharum, A. rubrum), oaks (Quercus rubra, Q. velutina, Q. montana, Q. alba), hickories (Carya ovata, C. tomentosa, C. glabra), ash (Fraxinus americana), and formerly, chestnut (Castanea dentata). White pine (Pinus strobus) and hemlock (Tsuga canadensis) are the characteristic conifers of the region. The transition hardwoods zone occurs in some of the most populated regions of the northeast, both historically and currently. As such, it contains some of the most heavily disturbed and modified forests in the Northeast. In the decades following the French and Indian Wars, European settlement brought extensive forest clearing and agriculture, which peaked in the latter half of the 19 th century (Foster 1999). By the mid-late 1800s, some regions of New England had experienced clearing in excess of 60%, with all virgin timber cut by 1850 (Cogbill et al. 2002). As a result of the opening of the West by the railroads and the growth of cities during the Industrial Revolution, thousands of farms were abandoned, resulting in reforestation of much of the region. Former agricultural fields and pastures became dominated by even-aged white pine forest, which later fueled a second logging boom beginning around the turn of the 20 th century as the pines reached a merchantable size (Hall et al. 2002). Forests then became dominated by another relatively even-aged cohort of young hardwoods that established as advanced regeneration beneath the pines. Original and subsequent cutting (including much that must be considered as high-grading) generally left these forests in a depleted and deteriorated condition (Westveld et al. 1956). Although our understanding of silviculture in this type is now greatly improved, the region faces new threats, principally from forest fragmentation and parcelization in a rapidly urbanizing environment. The challenges for forest and biodiversity conservation in this region are significant. An understanding of how historical land use has modified species composition and ecological processes over the last 400 years, or an understanding of each area’s “land-use legacy,” is critical to designing appropriate management strategies (Hall et al. 2002, Foster et al. 2003).
I. Presettlement Vegetation and Natural Disturbance
There are few satisfactory methods for studying pre-European vegetation and disturbance, but the use of multiple methods paints a more accurate picture of forests of the distant past than the use of one single method. Lorimer and White (2003) list a number of these techniques, including 1) pollen and charcoal from lake sediment cores, 2) land surveys that record witness trees and written descriptions and accounts, and 3) reconstructions from modern old-growth stands. Whitney (1994) provides a thorough review of the techniques for studying presettlement vegetation. In addition to these methods, there is 4) historical information available on the frequency and intensity of major disturbances since European settlement. This historical information, combined with observations on species responses to these disturbances, allows inference of pre-European disturbance regime, species composition, and stand structure. We should also consider a common misconception about the records of surveyors and other early travelers: that they reflect a snapshot of the unspoiled forest landscape, devoid of the influence of humans in general or, more specifically, of humans of European ancestry. As Callicott (1991) writes,
This is particularly true of much of the transition hardwood zone, where Native American nations had well-developed agricultural societies. Much of the early ecological analysis of surveys, land records, and even remnant old-growth stands (as described above) took place in an era when the anthropological work of Kroeber (1939) represented the state-of-the-art, and suggested a low population density for this period. Later work by Dobyns (1966), however, suggested that a wave of European diseases traveled ahead of actual European colonists, and that these settlers might have encountered a native population reduced by as much as a factor of 10. The exact magnitude of the population reduction remains hotly contested in the anthropological literature, but its existence is hardly in doubt. A particularly poignant example begins at what may be taken as the historic (though not geographic) heart of the land-use story in transition hardwoods: the Mayflower colonists landed at Plymouth not in unbroken wilderness, but at the site of an abandoned agricultural settlement recently devastated by pestilence, perhaps initiated through contact with European fishermen or explorers. Samuel de Champlain had mapped the settlement and its cabins as inhabited not 15 years earlier (Russell 1976).
As extensive human land use has obscured, or homogenized, the natural distribution of forest species across the northeastern landscape (Fuller et al. 1998, Hall et al. 2002), palynology can be used to roughly delineate and describe presettlement forests. Within the transition hardwoods zone, pollen in lake sediment cores has shown that in presettlement times, species composition varied between uplands and lowlands within the region (Fuller et al. 1998). In Massachusetts, the central uplands before European settlement had a higher representation of pollen from shade-tolerant, mature forest species, including hemlock, sugar maple, yellow birch, and beech (Fuller et al. 1998), consistent with a description of higher elevations within the transition hardwoods zone provided by Westveld et al. (1956). In the Berkshire-Taconic plateau of northwestern Connecticut and western Massachusetts, hemlock and chestnut were found to be dominant (Deevey 1943). In other upland sites in central Massachusetts, oak and chestnut were both important species, with relative pollen abundance of more than 40% each (Foster et al. 2002a). Oak dominated and chestnut was minimal up to 500 years before European settlement, then chestnut increased dramatically at the expense of oak, probably due to the onset of more mesic conditions (Foster et al. 2002a). However, 100-200 years prior to European settlement (circa 1450), beech, sugar maple, and hemlock all began to decline while oaks increased (Fuller et al. 1998). The timing of this vegetation shift corresponded to the beginning of a climatic period known as the Little Ice Age (Fuller et al. 1998). During this time New England and Europe experienced extreme winters and cool summers. A decline in precipitation may have been responsible for the vegetational change, as cold-tolerant species such as spruce and fir did not increase, but oaks, which are generally well adapted for drier conditions, increased. An alternate hypothesis, that an increase in fire frequency might have favored sprouting oaks over other species, was thought not likely to be responsible for the shift, as charcoal was more abundant in lowlands than in the uplands where the increase in oak was most pronounced (Fuller et al. 1998). Pre-European lowlands and major river valleys in Massachusetts were dominated by oak, chestnut, and hickory, with some pine (white and pitch) and low abundance of hemlock, beech, and sugar maple (Davis 1958, Fuller et al. 1998). Greater amounts of charcoal were found in sediment cores taken from lakes at lower elevations than lakes in the uplands. The presence of higher amounts of charcoal is consistent with the dominance of oak in lowlands, as periodic fire is thought to be necessary for the existence and long-term maintenance of oak forests. However, climate and substrate are also important (Abrams 1992, Fuller et al. 1998, Foster et al. 2002). Fuller et al. (1998) and Parshall and Foster (2002) found a gradient in fire frequency from the coastal and southern lowlands to the north and west, with highest amounts of charcoal in pine barrens areas, intermediate amounts in oak, pine, and hardwoods (transition hardwoods) forests, and lowest amounts of charcoal in northern hardwoods forests at higher elevations. They hypothesized that differences in forest composition between uplands and lowlands (Fuller et al. 1998) and across forest types (Parshall and Foster 2002) probably resulted from interactions among climate, substrate, fire, and Native American populations, the latter highest in the valleys where fire was most frequent.
Surveys and descriptions of forests in the transition hardwoods zone written before or at the time of colonization are the basis of a number of studies describing presettlement vegetation. Witness trees, recorded as boundary markers between towns and properties by early surveyors, provide the only direct quantitative evidence available on species composition before the onset of European land use practices. However, we should note that witness trees are not an unbiased sample of the vegetation encountered by surveyors, and that although surveyors were generally capable observers of the landscape, their observations dealt with their own concerns and not those of the early 21 st century forestry and ecology community. Cogbill et al. (2002) compiled town proprietor surveys of 389 towns in New England. Their data analysis clearly indicated differences in vegetation along latitudinal and elevational lines. Northern states ( Vermont, New Hampshire, Maine, upstate New York; mean 43-45 degrees N latitude) generally were dominated by beech with a mix of hemlock, birches, spruces, and maples, and were representative of the northern hardwoods and spruce-fir vegetation zones. Southern states ( Massachusetts, Connecticut, Rhode Island, downstate New York; mean < 43 degrees north latitude) were dominated by oaks with a mix of hickories and chestnut, all species that were characteristic of the central hardwoods zone. The transition between them was restricted to a 1-2 town-wide tension zone between northern hardwoods and central hardwoods forests. This tension zone, called ‘central pine’ by Cogbill et al. (2002), was generally located in western and northern Massachusetts, southern New Hampshire, and southwestern Maine. It contained more pine and hemlock witness trees than any other area, but was compositionally more similar to the oak (central hardwoods) zone than the northern hardwoods. The boundary of the central pine tension zone with northern hardwoods was quite distinct and followed the northern boundary of Westveld et al.’s (1956) transition hardwoods, also used in this study. Cogbill et al. (2002) theorized that because the central pine tension zone crossed a range of variation in altitude, latitude, climate, bedrock type, and past disturbance regimes, the distinct northern ecotone must be related to a different factor: i.e., the northern limit of frequent fire in New England. Conversely, the southern boundary of central pine was rather diffuse, similar to the situation reported by Westveld et al. (1956). The narrowness of the central pine zone in Cogbill’s (2002) study may be at odds with maps of modern transition hardwoods vegetation zone in New England, including the one presented in this study. The confusion seems to be caused by the location of the southern boundary of transition hardwoods, usually depicted as the southern limit of conifer codominance in the otherwise oak-dominated forests of southern New England. Nineteenth century writings describing early forests, such as that of an 1885 gazetteer’s description of the Pisgah forest in southwestern New Hampshire, often reported dense stands of fine white pine along river valleys and hemlock on the uplands (Cline and Spurr 1942). Cogbill et al. (2002) were quick to point out that the idea that pine and hemlock were widespread dominants throughout New England was a misconception, and that in actuality these two species were major dominants only in a narrow band between northern and central hardwoods. Additionally, chestnut was not as widespread as some accounts would suggest, having less than 10% of witness trees in most locations. Nineteenth century white pine forests, having established following agricultural abandonment (Baldwin 1940, Westveld et al. 1956, Hall et al. 2002) or some other large-scale disturbance (e.g., fire; Cline and Spurr 1942, Spurr 1956a), were often to the south of its historical range of dominance [i.e., the tension zone described by Cogbill et al. (2002)]. Some authors suggested that the distributions of forest species before European settlement were more closely tied to climate and natural disturbance, while postsettlement distributions depended heavily on human influence (Fuller et al. 1998, Hall et al. 2002). Outside of New England, studies of witness trees in the transition hardwoods zone have shown elevational differences akin to those reported by the palynological studies of Fuller et al. (1998) in New England. Abrams and Ruffner (1995) conducted a study of witness tree distributions in the transition hardwoods regions of north central Pennsylvania. They found that within the region, the Allegheny Mountains uplands were dominated by white oak, chestnut, and red maple, with pitch pine on the plateau tops, while lowland stream valleys had a higher incidence of hemlock, white pine, and birch. The higher incidence of white pine in lowlands is similar to a description of Massachusetts by Fuller et al. (1998). However, unlike the transition hardwoods forests in New England, higher elevations in the Allegheny Mountains completely lacked shade-tolerant beech as witness trees. Within the Allegheny Mountains, the western Allegheny Front was dominated by more oak (white oak and chestnut oak) and less hemlock than the eastern Ridge and Valley Province, which had more white pine and less oak. Elevational and geographic differences in species composition in the presettlement transition hardwoods of north central Pennsylvania were strongly influenced by landform (Abrams and Ruffner 1995). Upper elevations and the western Allegheny Front area were more xeric, favoring oak, chestnut, and fire- and drought-tolerant pitch pine. Lower elevations such as stream valleys and protected coves, especially in the Ridge and Valley Province, had a higher proportion of mesic species such as white pine, hemlock mixed in with the oaks. The abundance of red maple across all landforms in the Allegheny Mountains contradicted the once popular idea that it was only present in lowland swamps in presettlement times (Abrams and Ruffner 1995). Red maple was more abundant in the Allegheny Mountains than anywhere else in the transition hardwoods zone prior to European settlement. The pre-European boundary between the transition hardwoods and the northern hardwoods was probably set by the northern limit of fire. Fire was most common in pine barrens, central hardwoods, and transition hardwoods (decreasing order of importance; Cogbill et al. 2002, Parshall and Foster 2002). The return interval of fire in northern hardwoods, however, can reach thousands of years (Fahey and Reiners 1981). Early accounts of fire exist for the transition hardwoods zone, as explorers reported park-like woods and tall grass in forest understories in coastal Massachusetts and Maine, as well as in river valleys in Massachusetts and New York [e.g., Morton (1634) as summarized by Day (1953) and Russell (1983)]. Charcoal has been found in lake sediments throughout the transition hardwoods zone (e.g., Fuller et al. 1998, Parshall and Foster 2002), with more charcoal found in lowlands and river valleys where Native American populations were highest (Lorimer and White 2003). Native Americans were credited with starting fires in presettlement accounts (e.g., Morton 1634), but the extent of this practice is unknown (Parshall and Foster 2002, Lorimer and White 2003). Most fires were thought to occur in the early spring or late fall when there were no leaves on deciduous trees, allowing the litter and understory plants to dry out (Lorimer and White 2003). For a discussion of presettlement fire in the Northeast and the role of Native Americans, see Part 1.
3) Reconstruction of Vegetation and Disturbance from Contemporary Stands
In addition to the interpretation of sediment cores and historical records, a third way to study the composition and dynamics of presettlement forests involves the study of contemporary old-growth stands (Lorimer and White 2003). Like the first two methods of inferring past forest composition, sampling old growth has its limitations. Cogbill et al. (2002) pointed out that <1% of forests in the northeastern United States is true old growth. These areas probably remained uncut because they had unusual histories or extreme landscape settings that made them inaccessible, and as such they may not represent the common presettlement landscape (Cogbill et al. 2002). In combination with other methods, the study of old-growth remnants remains an important tool for studying presettlement forests (Lorimer and White 2003). However, given the relatively frequent nature of stand-replacing disturbance in this zone (as discussed below), these relict areas should not be taken as representative of broad areas of the landscape. Old-growth stands, whether defined structurally (i.e. old large trees, an uneven-aged structure and relatively shade-tolerant species composition) or demographically (e.g., Oliver and Larson 1996), are extremely rare within the transition hardwoods zone because of widespread clearing for agriculture and clearcutting that occurred from the time of settlement through the end of the 19 th century. The few stands that survived exploitation have been subject to natural catastrophes in the 20 th century (e.g., the 1938 hurricane which affected most of southern New England), and to increasing anthropogenic stresses such as the modification of natural herbivore populations. The following studies are examples of some old-growth stands. Both original species composition and the effects of natural disturbance on forest composition and dynamics will be described.
Remnant Forest Composition and Disturbance – New England
The Pisgah forest in southwestern New Hampshire lies within the range of transition hardwoods and along the tension zone in New England (Cogbill et al. 2002), and was an example of an old-growth remnant until 1938 when much of it was destroyed by the “great hurricane”. Studies of the Pisgah forest prior to the 1938 hurricane combined with post-hurricane reconstruction of the stand have provided a clear picture of the structure and disturbance history of an undisturbed transition hardwoods forest. The Pisgah tract was an historic anomaly in the region because some stands within it were continuously forested throughout the post-settlement period. This was due in part to its inaccessibility and also to one family’s long-term management strategy of protecting it as a reserve source of firewood and timber (Cline and Spurr 1942). However, because of Pisgah’s status as being relatively free of human disturbance, it may model the type of forest found in the transition hardwoods zone of southern New England before widespread modification by Europeans. Early studies of the Pisgah area described the forest as composed of broad-leaved trees and hemlock 80-100 feet high, interspersed with white pines towering at 150 feet (Spurr 1956b). Before the great hurricane of 1938 (which felled practically all the mature trees in Cheshire County, NH), the composition of stand remnants at Pisgah generally agreed with early accounts, with 60% of trees being pine or hemlock, and with maples, beech, birches, and red oak making up the remainder (Cline and Spurr 1942). Past disturbances were thought to be responsible for the presence of white pine, paper birch, red oak, red maple, and chestnut, all of which are less shade-tolerant than hemlock and beech. At the time of Cline and Spurr’s study, even in its ‘virgin’ state, 80% of the forest showed evidence of having been affected by natural disturbances – fire, wind, ice, insects, or pathogens – at some point in the past. Fires occurred in the original Pisgah forest, as evidenced by written history, fire scars on trees, and the presence of soil charcoal (Cline and Spurr 1942). Fires were more destructive before the 1940s, as there was no available fire control and no artificial firebreaks such as roads. Flammable litter that accumulated after wind disturbances could fuel blazes touched off by lightning. Even before the 1938 hurricane, Cline and Spurr (1942) reported that wind was the most common cause of death of Pisgah’s older trees, especially on thin soils or exposed sites, and evidence of fire and wind damage (extent unspecified) were often found together. Cline and Spurr (1942) observed that many emergent white pines had been struck by lightning both at Pisgah and at the nearby Harvard Forest in northern Massachusetts, and Henry and Swan (1974) showed through forensic analysis of debris and charcoal on the forest floor that the entire stand was destroyed in 1665 by a forest fire, probably fuel-enhanced by dead trees killed in a hurricane in 1635. Native Americans may have also started fires to thin the forest understory or to drive game (see Part 1), and Cline and Spurr (1942) cite an 1886 account that a local tribe used frequent fire before 1720 to keep fields open for cultivation in the Pisgah area. Additionally, European settlers routinely burned the trees they cut in order to clear frontier land for agriculture, probably leading to further conflagrations in adjacent wildlands. Patchy disturbance by wind and fire created small canopy gaps, coarse woody debris, and tip-up mounds and pits. Small gaps of a few trees allowed recruitment in the understory, mostly of hemlock, red oak, red maple, paper birch, and black birch, and old (>50 yrs) closed gaps looked much like the surrounding forest (>60% pine and hemlock; Cline and Spurr 1942). In contrast, severe disturbances of wind and fire that destroyed the canopy led to the formation of even-aged stands of white pine with half as much hemlock and less red oak and red maple as regenerated in partial disturbances (Cline and Spurr 1942). Following the 1665 catastrophic fire, white pine had the most regenerating stems, followed by white oak and hemlock, spruce, red maple, and poplar (forensic examination of a single plot; Henry and Swan 1974). The fire produced a layer of charcoal on the forest floor, and all white pine in the plot germinated within 13 years of the fire. In contrast, 80% of the hemlock, more shade-tolerant than pine, germinated within 37 yr of the fire. In Pisgah, and in presettlement transition hardwoods in general, the presence of even-aged cohorts of white pine signified not a climax condition, but evidence of former large-scale disturbance (Cline and Spurr 1942).
Remnant Forest Composition and Disturbance – Pennsylvania and New York
In old-growth fragments of transition hardwoods in Pennsylvania (Tionesta Scenic and Natural Areas, Woodbourne Forest) and New York (Niles Huyck Preserve), climax species were similar to those found in New England, and included beech, hemlock, sugar maple, and white ash (Runkle 1982). Sites with large scale disturbances were avoided in Runkles’s (1982) study, but in northwest Pennsylvania, large scale disturbances occur often enough to generate stands of white pine, such as those at Heart’s Content and Cook Forest. As in New England, windstorms were the most common large scale natural disturbance in Pennsylvania and New York transition hardwoods. Other widespread disturbances included the chestnut blight, beech bark disease and hurricanes (in New York). Small-scale disturbance predominated in Runkle’s (1982) highly selected New York and Pennsylvania old growth, as small canopy openings covered a mean of 4.3% of the land area. New gaps formed in 1% of the forest area per year (Runkle 1982, Lorimer 1989) as individual or groups of adjacent trees died. Gap size ranged from less than 25 m 2 to 200-375 m 2 (Runkle 1981, Lorimer 1989), with 1% of the forest area closing by sapling height growth (not lateral spread) per year. Lateral closure of gaps was less likely to occur in gaps greater than 280 m 2 in area (Lorimer 1989). Runkle (1982) documented a natural canopy tree (gap formation) rotation time of ~100 years, but varying from ~50 to ~200 years. The most common cause of tree death in old growth was not reported in these studies. (Very few studies actually report etiology of canopy openings.) Regeneration in gaps depended on gap area. Larger gaps up to 2000 m 2 yielded higher woody species diversity, basal area, and number of stems (Runkle 1982), and faster-growing stems. The species dominant in the surrounding matrix of old growth were found in all gap sizes (Runkle 1982), and their regeneration was adequate to perpetuate the canopy composition of the original stand (possibly with some reciprocal replacement of sugar maple and beech; Runkle 1981). The response of other species to gap size varied depending on their life history. Natural small-scale disturbance regimes favored shade tolerant species but allowed for the persistence of opportunists at low densities across the landscape, primarily in large gaps (Runkle 1982). While Runkle’s (1982) study provides a compelling picture of old-growth remnants within the transition hardwood zone, the highly selected nature of those remnants suggests they provide an unrepresentative picture of what pre-European forests may have resembled. Indeed, given the prevalence of hurricanes and other large disturbances over the region, other kinds of forest structures and dynamics seem particularly probable. It is to these large disturbances that we turn next.
4) Observations on Stand Replacing Disturbances Stand replacing disturbances, caused by either wildfire or tropical cyclonic storms, were important in the transition hardwoods region. The role of fire was evident in many of the studies reviewed above. Here, we focus on hurricanes. Along the eastern seaboard (primarily New England and New York), hurricanes are a major stand-replacing disturbance event that impacts ecosystem structure, function, and dynamics of coastal forests (Boose et al. 2001). Hurricanes, like fire, form a strong regional gradient in frequency and intensity, strongest near the coast and declining to the northwest. Boose et al. (2001) analyzed 67 hurricanes in New England since 1620, and found a return interval for F2 hurricanes (Fujita Scale) ranging from 85 to 150 years in central and transition hardwoods ( Connecticut, Rhode Island, Massachusetts, and southern New Hampshire). Damage from hurricanes can range from loss of leaves and branches (F0) to scattered blowdowns and small gaps (F1) to extensive blowdowns and large gaps (F2). Destruction at individual sites depends not only on hurricane intensity, but on the topography and prior history of the stand, which could make it more susceptible to damage (Boose et al. 2001). White pine is more susceptible to hurricane damage than hardwoods; however, increasing age (>50 yr) makes all forests more vulnerable. Exposed south-facing slopes are particularly vulnerable to hurricanes (Foster 1988, Carlton and Bazzaz 1998a). The most significant hurricane to hit the transition hardwoods region in the last 100 years was the 1938 hurricane. Lorimer and White (2003) estimated that within the old-field white pine region (which includes the transition hardwoods zone) the hurricane caused more than 75% tree mortality on 40% of the landscape and more than 50% mortality on 60% of the landscape. Only 25% of the landscape had lower than 25% mortality, while 15% of the landscape had no damage. In New England, 3 billion board feet of timber were blown down on 600,000 acres of forest, including half of all of the white pine in the region (Spurr 1956a). Regeneration following the hurricane was primarily from advanced growth beneath the white pine in all stands that were at least middle aged ( Baldwin 1940). These species included mixed hardwoods such as red oak, red maple, white ash, and sugar maple, except in cool moist areas where hemlock and balsam fir were predominant ( Baldwin 1940). One third of all regeneration was from the establishment of light-seeded early successional species such as gray birch, pin cherry, and black cherry, but these were short lived species and the final composition was predicted to mirror the advanced regeneration (Spurr 1956). In general, hardwoods that normally followed logging of white pine forests were the ones that benefited from the hurricane (Spurr 1956a). In the Pisgah old-growth forest in southern New Hampshire, the 1938 hurricane had a devastating effect, destroying all stems larger than 17 cm in diameter on Henry and Swan’s (1974) study site. It is likely that the forest as a whole suffered in a similar manner. The forest had been weakened by three lesser windstorms in the previous 40 years, allowing extensive damage to occur in 1938 (Henry and Swan 1974). Following the hurricane, regeneration was enhanced by the advanced growth of beech, paper birch, and hemlock that had established in former canopy gaps. New arrivals to the site included red maple, black birch, sugar maple, and striped maple. All mature white pine were destroyed and they did not regenerate (Henry and Swan 1974, Foster 1988). Understory hemlock capitalized on the disturbance to increase in height and density, whereas red maple and black birch established in pits and on mounds, respectively, that were created by windthrow (Henry and Swan 1974). In other areas at Pisgah [outside of Henry and Swan’s (1974) study area], paper birch had more of a role in regeneration on high slopes, associated with red maple, black cherry, and black birch (Spurr 1956a). On low slopes, hemlock and beech were released from advance growth. Where no salvage logging occurred, hurricane accelerated succession rather than setting it back; salvage logging injured and killed the advanced growth, as well as scarified the soil, leading to eventual dominance of early successional species (Spurr 1956a). After the hurricane, 1.5 billion board feet, or 40% of damaged timber, was salvaged throughout the region (Foster et al. 1997). Experimental hurricane blowdowns simulating the 1938 hurricane in the Harvard Forest in north central Massachusetts have led to further insights about the altered environment caused by hurricane damage and how it can affect ecosystem processes. Carlton and Bazzaz (1998a) reported that immediate environmental changes after mechanically pulling down 70% of the trees included uprooting and the formation of mounds and pits, which are persistent microsites for the regeneration of particular species, especially birches (Carlton and Bazzaz 1998b). Most forests in the eastern U.S. have between 14-50% forest floor area in mounds and pits, though a single catastrophic storm usually causes less than 10% of the forest floor to be converted to mound and pit microtopography (Carlton and Bazzaz 1998b). Mounds represent extreme soil conditions: dry, with low organic matter and nutrients (Beatty 1984). Pits collect moisture, organic matter, and nutrients over time. Downed logs and tree crowns form other microsites for seedling establishment as they rot. Further, the amount and spectrum of light to hit the forest floor also increases following a hurricane, which in turn warms the soil and alters decomposition processes over the course of years. Light levels and nitrification rates were much more heterogeneous in blowdown sites than undisturbed forests ( Carlton and Bazzaz 1998a). Immediately following the simulated hurricane disturbance at Harvard Forest, seed rain was abundant and heterogeneous because of the location of residual canopy trees ( Carlton and Bazzaz 1998b). Seeds were concentrated in pits and tended to slide off vertical surfaces. Birch seedlings composed 90% of all new seedlings. Most seedlings germinated on scarified horizontal surfaces rather than mounds or pits, but mounds became more favorable the second year after disturbance. Seedling mortality was caused by extrinsic factors in winter or factors related to resource limitation and fitness during the growing season, small seedlings being most susceptible ( Carlton and Bazzaz 1998b). The bare soil of mounds promoted regeneration of small-seeded species such as black birch on moister sites. However, trees establishing on mounds were subject to mortality from frosts in autumn, frost heaving or snowshoe hare browsing in winter, and drought in summer. Pits accumulated thick leaf litter and water, burying or drowning seedlings ( Carlton and Bazzaz 1998a). Other seedlings died of shading or drought, but in general, conditions in the blowdown area favored early-successional species. Three years after the simulated hurricane disturbance, midsummer light levels at 40 locations were still three times higher than in the undisturbed forest understory ( Carlton and Bazzaz 1998a). Soil resource levels, nitrogen cycling, and moisture in the blowdown area were not significantly different from the undisturbed forest, though organic matter was only slightly higher in undisturbed forest (Foster et al. 1997, Carlton and Bazzaz 1998a). The hurricane produced structural reorganization in the forest without altering overall ecosystem processes, and productivity returned to 71% of pre-disturbance levels within four years (Foster et al. 1997). In central and transition hardwoods forests further from the coast, such as those in Pennsylvania, windstorms such as tornadoes and downbursts may play an important role in disturbing large patches (though tornado return interval at any given point is 10,000 – 20,000 years; Lorimer and White 2003).
II. Vegetation and Land Use After European Settlement
The transition hardwood zone was heavily settled by Native American groups, in comparison to the northern hardwood zone (Russell 1980). However, dramatic changes were initiated beginning with the first European contacts. Once Europeans colonized North America , they had tremendous influence on the character, composition, and amount of forest vegetation. The earliest colonies in what was to become the northeastern United States were established in the early 1600s in coastal New York , Massachusetts , and New Hampshire , and spread progressively inland. Population size and associated land use intensity in the Northeast peaked around 1850 but was followed by a mass exodus to the cities and to the more fertile farmland of the Midwest . European land use has had tremendous impact on transition hardwoods, and those impacts can be summarized in five major categories: 1) agriculture and forest clearing followed by abandonment and revegetation, 2) forest harvesting, often high grading, 3) an initial increase in fire frequency followed by effective fire exclusion, 4) modification of natural animal populations, and 5) introduction of nonnative pests and diseases.
1) Agriculture and Abandonment
Hall et al. (2002) documented changes in land use in Massachusetts over 400 years, illustrating patterns that applied to most of New England and the northeast. Using colonial land surveys, 19 th century maps showing forested lands, and agricultural censuses, Hall et al. (2002) documented changes in the composition, structure, and distribution of forests at the statewide level. Agricultural land use in Massachusetts peaked between 1830 and 1885 when 50-60% of the land was in pasture, hay, or agricultural fields, and sheep and cattle exceeded 650,000 (Cogbill et al. 2002, Hall et al. 2002). In 1830, forest blocks were small, averaging 119-204 ha. At the peak of agricultural development, forests were relegated to poor lands such as mountains, swamps, and dry sand plains. Remaining forests were relentlessly harvested for timber and fuel during the agricultural period (Cogbill et al. 2002). After 1850, many northeastern farms were abandoned with the discovery of productive agricultural lands in the West, the development of railroads, and the lure of the industrial revolution in the major cities (Hall et al. 2002). Farms on rocky uplands and coastal sand plains were abandoned first, and relatively fertile river valleys last. Forests increased statewide as agricultural land decreased; only 7% of Massachusetts is in agriculture today (Foster 1999, Hall et al. 2002). The Harvard Forest in northcentral Massachusetts is a typical example of a forest originating after clearing and farm abandonment in the transition hardwoods region. Spurr (1956b) reviewed the period of settlement (1733-1907), noting that almost all of the virgin forest in the area was cut by the end of the 1700s. During the settlement period, 9% of the land was continuously forested but heavily utilized for timber and firewood, whereas the remaining 91% was cleared for agricultural use, although not necessarily all at once. The primary use of cleared land was upland pasture (75%), and the remaining ground was plowed for cultivation (16%). Abandonment allowed the establishment of even-aged stands, composed primarily of white pine (Spurr 1956b). Baldwin (1940) noted that many white pine stands throughout the region resulted from preferential grazing of hardwoods by cattle in former pastures. As the stands developed, the overstory was almost completely composed of white pine, some reaching over a meter in diameter and exceeding 35 m in height ( Baldwin 1951). Hemlock-hardwood understories eventually developed beneath the pines, with hardwoods including beech, red oak, yellow birch, red maple, sugar maple, and others. White pine did not successfully regenerate beneath its own canopy ( Baldwin 1951). Stands of this type were common from New England to western Pennsylvania in the transition hardwoods region (Baldwin 1951, Howard and Lee 2002) until logged or destroyed by 1938 hurricane. Deforestation and reforestation of the northeastern United States had effects that reverberated through centuries after the original land uses had ceased. Modern vegetation patterns on the northeastern landscape still reflect 19 th century land use practices (Foster et al. 2003). The age structure of modern forests are often unimodal as a result of old-field abandonment or a clearing event. However, most of these forests are stratified mixed-species stands, and have complex diameter distributions and vertical structure. Many forests were cleared or cut repeatedly for firewood, producing even-age stands and multistemmed trees that remain today (Hall et al. 2002). The redevelopment of forest cover in the transition hardwood zone has been dramatic. For example, in 1885, only 30% of the forest lands in Massachusetts had trees greater than 12.2 m tall. By 1971, the percentage had grown to 77% as trees grew taller and sawtimber increased (Hall et al. 2002). However, even with this increase, eastern forests currently lack structural elements such as large trees, large coarse woody debris, tip-up mounds, pits, and large standing dead trees (Foster et al. 2003). Although much of the Northeast has reforested, significant differences have been documented in understory flora of primary and secondary, post-agricultural woods. Primary forests, in this context, refer not necessarily to old growth, but to areas that remained in forests throughout the historic period. These areas were often heavily used for firewood and may even have been clearcut. Secondary forests are areas that experienced a transition in land use, including clearage for farmland or pasture. Unlike windstorms and historic timber harvest, which had relatively minor effects on ground-level vegetation, disturbance by plowing and grazing eliminated native vegetation in large areas, and many forest understory plants have not been able to recolonize secondary forest stands (Whitney and Foster 1988, Donohue et al. 2000, Bellemare et al. 2002). Poor colonizing ability may limit the spread of species associated with continuously forested patches into formerly deforested areas, even though agricultural use ended more than a century ago and both areas are now forested (Whitney and Foster 1988, Bellemare et al. 2002). Bellemare et al. (2002) report that species without morphological adaptations for seed dispersal are rare in most post-agricultural forests, reducing understory diversity far below presettlement levels. Environmental differences between forests, although present, had less influence over current vegetation than previous land use. Bedrock outcrops, where present, have sometimes served as refugia throughout the agricultural period for species that are poor dispersers, enhancing recovery in nearby areas (Bellemare et al. 2002). For canopy tree species, agricultural land use obscured natural regional patterns of abundance that were historically controlled by climate and natural disturbance, homogenizing the forest composition on the broad scale (Fuller et al. 1998). In analyzing pollen from pre- and post-settlement periods over much of northern Massachusetts, Fuller et al. (1998) documented a sharp decline on uplands of long-lived, slow growing, shade tolerant species (hemlock, beech, sugar maple, and yellow birch). Post-settlement, they were replaced by species more tolerant of disturbance (white pine and chestnut initially, then oak, red maple, and birch during reforestation). Likewise, in north central Pennsylvania, hemlock and beech have been replaced in dominance by red and sugar maple, red oak, black cherry, and black birch (Abrams and Ruffner 1995). On lowland sites in New England transition hardwoods, sediment core pollen showed that overall forest composition did not change markedly, though land cover was reduced (Fuller et al. 1998). Disturbance-tolerant species, which were originally more abundant on lowlands due to fire in presettlement times, were maintained by logging and fire caused by 18 th and 19 th century European land use practices. Regionally, oak, birch, and red maple became more abundant, and are still high. Chestnut declined because of the blight and beech bark disease is currently attacking beech. Oak may now be declining in the lowlands because of fire suppression. In Pennsylvania lowlands, pine, white oak, and hickory have been reduced, while chestnut oak, red oak, black cherry, and black birch have increased since settlement (Abrams and Ruffner 1995). Elimination of chestnut facilitated an increase of oak on ridge sites. South of the Allegheny plateau, increases in red maple, black cherry, and other mesic species may be a result of fire exclusion in the 20 th century (Abrams and Ruffner 1995). Similar vegetation changes have been reported for forests of the Allegheny Plateau in northwest Pennsylvania, the Catskills of New York, and northern Vermont (Abrams and Ruffner 1995).
2) Farm Woodlot Use, Forest Harvesting, and High Grading
During the period after European settlement, most primary forests underwent significant change as a result of natural and anthropogenic disturbance. Oliver and Stephens (1977) reconstructed one such stand, a woodlot at the Harvard Forest that had remained in continuous forest cover since at least 1500 A.D. The stand had been disturbed by several major hurricanes (including two before 1800 and two after). The stand had been clearcut in 1803, and partially cut at least 9 times between 1803 and 1952. Despite this history of extensive use, the stand was still characterized by pit-and-mound topography and included trees up to 58 cm DBH. Regeneration was episodic; initiation of new cohorts was confined to major disturbance events, with minor events producing few or no stems recruiting to the canopy. The pattern of change in both primary and secondary forests was substantially modified by the wave of forest harvesting that followed some 50 years after agricultural abandonment. Around the turn of the 20 th century, after establishment and growth of white pine stands on old agricultural land, a new round of timber harvest began in the northeast. These pine forests were either clearcut or ‘selectively’ logged. Westveld et al. (1956) argued that both practices left the forests of the region in a degraded state, although the passage of time has left some reason to moderate their conclusion. Certainly, the widespread clearcutting reduced regional stocking levels. Selective logging often left behind pines that were too small; these pines in turn provided structural complexity, and often early economic return, to the stands regrowing around them. Pines that were left due to gross defect, all too common among old-field pine, presented further challenges. However, even these pines were often utilized in areas with a strong box or bucket industry. The logged forests were succeeded by even-aged hardwoods that had established beneath the pine overstory (Spurr 1956a). The conversion to hardwoods was hastened in 1938 by the hurricane in southern New England, and in the Harvard Forest hardwoods increased 44-56% and hemlock-hardwoods by 5-12% between 1907 and 1946, while white pine was reduced to 2-26% (Spurr 1956a). Red oak and red maple were prominent in most hardwood associations, oak on drier sites and maple on more mesic sites. In the western part of the range of transition hardwoods, such as in northern Pennsylvania where agriculture may not have altered such a high percent of the landscape as in New England, logging also transformed the forested area. Between 1797 and 1880, 'selective' cutting (high-grading) had removed nearly all of the high quality timber (especially white pine) from accessible sites (Abrams and Ruffner 1995). The period between 1880 and 1930 became known as the “clearcut era,” when much of the remaining timber was stripped from the landscape. Hemlock was utilized for tanning, and low quality wood fueled a large charcoal-iron industry that began as early as 1774 and continued through the early 1900s (Abrams and Ruffner 1995). As a result of indiscriminate and careless logging practices, recurring slash fires were common, especially in areas dominated by conifers, eliminating reproduction of fire-intolerant late-successional species like beech and hemlock. White pine and its seed source were eliminated from much of the area, and as a result the tree species composition has shifted toward red oak and red maple in comparison to pre-European forests. Currently red maple is increasing and oak is decreasing, possibly as a result of widespread fire suppression. These changes were consistent in the Allegheny Mountains across the western Front to the Ridge and Valley Province in the east (Abrams and Ruffner 1995). As elsewhere, chestnut was eliminated by the blight and is no longer represented in forest canopies in Pennsylvania.
3) Modification of Fire Frequency
The incidence of fire in the transition hardwood zone before European settlement was higher than in the northern hardwood zone to the north, where catastrophic fire frequencies may have been less than once per millenium (Fahey and Reiners 1981). Lorimer and White (2003) review much of the literature on fire frequency in the transition hardwood zone (this zone is treated as part of their “eastern oak forest” region). Although there was considerable regional variability, fire – and especially frequent but light understory fire – appears to have been an important element in the transition hardwood forest (Lorimer and White 2003). Explorers, surveyors, and settlers often characterized the forests of coastal Massachusetts as open or parklike, and cited native use of fire as a primary reason (Whitney 1994). The degree to which such conditions were characteristic of the interior is unclear, as disease and social disruption may well have modified native fire use patterns in the several decades it took for settlers to penetrate far beyond the coast (Russell 1983). When Europeans arrived in North America, cleared forests, and built settlements, they temporarily elevated fire frequency above pre-European levels. Palynological studies comparing presettlement to postsettlement vegetation in New England found an increase in the amount of charcoal deposits in lake sediments at times that corresponded to the influx of grass and herbaceous pollen, lower organic content, and higher sedimentation associated with European land clearance (Fuller et al. 1998, Parshall and Foster 2002). Increased fire frequency may have resulted from the deliberate use of fire in land clearing and an increase in accidental ignitions as the European population grew. However, Lorimer and White (2003) review a series of studies suggesting that in many areas fire frequencies did not change after European settlement, but remained relatively constant. What is clear is that for much of the history of the transition hardwood forest, fire was a frequent feature. Wildland fire undoubtedly decreased as the proportion of forest in the landscape declined in the late 1700s and early 1800s. However, during the wave of old-field pine reforestation after agricultural abandonment, fire returned. The regional incidence of fire increased with logging, reaching a maximum around 1900, before fire control began (Abrams and Ruffner 1995). Since the end of World War II, forests that had historically burned were protected from fire as the result of better firefighting techniques, increased access into burn-prone areas, improved equipment, and increased efficiency (Fahey and Reiners 1981). Subsequently major periods of fire in the 20 th century occurred only when drought conditions favored them (Fahey and Reiners 1981). The overall suppression of fires in the transition hardwoods, especially in Pennsylvania, has led to a decrease in oak and hickory and an increase in red maple. Abrams and Ruffner (1995) and Lorimer and White (2003) predict a future decline of the oak component if fire is not reintroduced.
4) Alteration of Animal Communities
As the previous sections have indicated, farm abandonment and timber harvests have had a substantial influence on the forest habitats of this region. Animal populations responded to the “successional wave” that resulted after the abandonment of farmlands in a relatively short period (circa 1850-1920). At present, mid-successional stands dominate the region and some species are becoming more abundant as tree sizes increase with succession. Pileated woodpeckers (Dryocopus pileatus), the largest woodpecker in the region, has apparently responded to the increasing availability of large diameter trees that are used for nesting and foraging (DeGraaf and Yamasaki 2001). However, it is important to note that few species are dependent on the mid-successional stands that now dominate the region. Of the more than 260 vertebrates associated with forest habitats in the northeastern United States (DeGraaf et al. 1992, Scanlon 1992), the majority utilize food or cover found within several seral stages. As a result, wildlife communities are hampered by the lack of structural complexity and habitat elements associated with the extremes of forest succession – regenerating habitats and old-growth stands. Populations dependent on regenerating forests in particular are experiencing conspicuous declines (e.g., Litvaitis 1993, Wagner et al. 2003). This group includes several species of birds (Dettmers 2003), mammals (Litvaitis 2001), reptiles (Kjoss and Litvaitis 2001), and a variety of butterflies and moths (Wagner et al. 2003). Although populations of these species are probably declining from unprecedented levels of abundance that resulted with the upsurge of early-successional habitats in the first half of the 20 th century (Litvaitis 1993, Foster et al. 2002b), several species are in jeopardy of regional extinctions (Litvaitis et al. 1999). In pre-colonial landscapes, small, disturbance-generated openings were likely important habitats for these species. However, the matrix that now comprises many landscapes within the transition hardwoods zone is very different from historic conditions and includes agricultural fields, suburbs, industrial parks, and extensive road networks. Small patches in these altered environments may no longer function as suitable habitat. The current plight of New England cottontails (Sylvilagus transitionalis) may illustrates this point. New England cottontails occupy a variety of habitats (e.g., old field, shrub-dominated wetland, and regenerating forest) that are characterized by dense understory vegetation (Barbour and Litvaitis 1993). Such habitats are ephemeral, and historically, were the result of some biotic or abiotic disturbance. In pre-Columbian landscapes, spatial extent of these disturbances likely varied with larger disturbance-generated patches of habitat more common along the Atlantic coast and small disturbances characteristic of inland forests (Lorimer and White 2003). In both landscapes, cottontail populations probably shifted in time and space. Present-day populations of cottontails also occupy a variety of disturbance-generated habitat, but those occupying small patches of habitat are now highly susceptible to extirpation (Barbour and Litvaitis 1993, Litvaitis and Villafuerte 1996). On small patches of habitat (<3 ha), rabbits encounter food shortages (Villafuerte et al. 1997). In response to these shortages, cottontails forage away from escape cover and are killed by predators at approximately twice the rate as cottontails on large patches where per capita food resources are more abundant (Barbour and Litvaitis 1993, Villafuerte et al. 1997). Although a change in foraging behavior by cottontails occupying small patches was probably a response among historic populations, contemporary populations of rabbits must now contend with the prevalence of generalist predators, especially coyotes (Canis latrans) and foxes (Vulpes vulpes). These carnivores are capable of exploiting a variety of habitats, and their populations and others [including raccoons (Procyon lotor)] have increased in response to converting secondary forests to other land uses and the removal of larger predators [particularly wolves (Canis lupus) and cougars (Felis concolor)] that preyed upon them (Oehler and Litvaitis 1996). Thus, as cottontails and populations of other prey species have declined in response to the loss of early-successional habitats, their predators have increased to levels that are probably greater than ever before. As a result, predation is the major proximate factor limiting populations of cottontails (Barbour and Litvaitis 1993, Smith and Litvaitis 2000). Survival rates of cottontails in small patches are now so low that these patches essentially function as demographic sinks (Barbour and Litvaitis 1993, Brown and Litvaitis 1995, Villafuerte et al. 1997). The pattern of local extinction and subsequent recolonization that likely characterized populations of New England cottontails in pre-Columbian landscapes is no longer viable. Few individuals disperse from small patches of habitat (Barbour and Litvaitis 1993) and those that do encounter habitats with limited cover where they are vulnerable to intense predation (Brown and Litvaitis 1995). Present-day populations of New England cottontails, therefore, are dependent on large patches of habitat close to each other to assure long-term survival (Litvaitis and Villafuerte 1996). Other species affiliated with early-successional habitats are eliminated locally if only small patches of habitat are available. These include golden-winged warblers (Vermivora chrysoptera) that do not nest in patches <10 ha (Confer and Knapp, 1981) and large-bodied snakes [e.g., black racers (Coluber constrictor )] that seem limited to large patches (>10 ha) of regenerating forests in human-dominate landscapes (Kjoss and Litvaitis 2001). Providing such habitats presents an obvious challenge (see below). In addition to increases among generalist predators, populations of white-tailed deer (Odocoileus virginianus) also have increased in response to landscape modifications and the elimination of large predators. In some areas, the density of deer is currently greater than before the region was first settled by Europeans (McCabe and McCabe 1984, Foster et al. 2002). Elevated populations of deer are having a variety direct, indirect, and community-wide effects (reviews by Côté et al. 2004, Rooney and Waller 2003). Because deer forage selectively, they are capable of altering the abundance, species richness, and competitive interactions of forest plants (e.g., Alverson et al. 1988, Tilghman 1989). The effects of browsing by dense populations of deer have been widely recognized as affecting forest tree composition (e.g., Leopold et al. 1947, Webb et al. 1956). In the transitional hardwoods zone, Whitney (1985) chronicled 50 years of change at an old-growth stand in Pennsylvania. During that period, the local deer population increased dramatically with a pulse of early-successional habitat that followed farm abandonment and clearcutting between 1890 and 1920. Overbrowsing of seedlings and saplings by deer severely altered the diameter-class distribution of the stand, creating a age/size gap for several species. Although much of the research on the effects of deer foraging has focused on trees, forests herbs represent most of the plant diversity in temperate forests, and therefore, warrant attention. Herbs also may be more vulnerable to herbivory because they never grow tall enough to escape browsing (Rooney and Waller 2003). Studies with natural exclosures (Rooney and Waller 2001) and deer-free islands (Balgooyen and Waller 1995) have shown that foraging by deer can reduce the abundance of some herbs. Among Trillium, browsing can reduce growth and reproduction (Rooney and Waller 2001). Augustine et al. (1998) indicated that herbs that naturally occur at low densities may be especially vulnerable to local extirapations. Indirect effects of deer browsing also can be substantial. In central Massachusetts, elevated deer populations (where large predators are absent and hunting prohibited for several decades) forage intensively on blackberry (Rhus allegheniensis) that colonized forest openings. Blackberry seedlings normally promote the establishment of tree seedlings (Horsley and Marquis 1983); however, when they are removed by foraging deer, openings are taken over by a competitor, hay-scented fern (Dennstaedia pinctilobula). This fern, a species that deer avoid, then becomes very abundant and inhibits the establishment of tree seedlings (George and Bazzaz 1999) and also excludes small herbs (Rooney and Dress 1997). Stands of hay-scented fern may persist for decades, resulting in a savannah-like community that Stromayer and Warren (1997) considered an “alternate stable state”. The effects of abundant deer populations may also affect the composition of animal communities. De Calesta (1994) used enclosure experiments to reveal that deer can reduce vertical complexity of forest understories resulting in a substantial decline in the abundance and diversity of shrub-nesting songbirds. In summary, the animal communities associated with the transitional hardwoods zone have been substantially altered by historic and current land uses. Forest communities now contain an abundance of habitat generalists that are capable of affecting biotic interactions and modifying local species pools. In substantially modified portions of this region, the loss of large predators and fragmentation of remaining habitats will present substantial challenges to efforts to maintain biotic diversity.
5) Introduced Species and Diseases
Since European settlement, the structure and composition of transition hardwoods has been altered not only by human disturbance, but also by various blights and introduced species. These invasive organisms have reduced populations of, or effectively eliminated, some native species. Chestnut blight is a vivid example of the effects of an invasive disease on eastern forests, while the hemlock wooly adelgid presents a current serious concern. Glossy buckthorn is an example of an introduced non-native species that has the potential to modify succession at the level of the forest understory.
Chestnut Blight
Chestnut blight (Chryphonectria parasitica), a fungus imported from China, completely eliminated chestnut from its historical place in the canopy of virgin stands and secondary forests in America. Though evidence shows that chestnut pollen was absent or at low levels in New England from the last glaciation until about 2500 years ago, it became abundant at that time (Paillet 2002). Written records (1904-1914) indicate chestnut comprised 50% of the timber volume on well-drained slopes on non-calcareous bedrock in southern New England and Appalachians. Chestnuts were most populous on north and east slopes, and oaks and hickories were its associates. The northern maximum of the old range of chestnut corresponds well to the northern border of transition hardwoods (Paillet 2002). Today chestnut is not reproducing sexually, as sprouts are killed before they reach maturity. Historically chestnut had great economic importance; they possessed straight, rot-resistant wood and a fast growth rate. Their stumps sprouted when cut, and they produced edible nuts. Formerly, chestnut established after a disturbance that opened the canopy, allowing advanced regeneration of saplings. For example, in the Pisgah forest of New Hampshire, all chestnut was even-aged (180-190 years) at the time of the blight (about 1915 in New England; Cline and Spurr 1942). Chestnuts had established after a single northeasterly storm followed by fire. In general, the abundance of chestnut across the Northeast was related to natural disturbance, then human disturbance, and finally the blight. The distribution of seedlings that survived the blight (only to be killed back and sprout repeatedly) is related to old-field abandonment patterns in the late 1800s. As a result of chestnut’s demise, hemlock, beech, and sugar maple are replacing them in New England (Cline and Spurr 1942). In Pennsylvania, chestnut blight facilitated an increase in chestnut oak and red oak on ridgetops (Abrams and Ruffner 1995). The species still exists everywhere in its range as an understory shrub (shade-tolerant when young) however (Paillet 2002).
Hemlock Wooly Adelgid
The hemlock wooly adelgid (Adelges tsugae) was another import from Asia, an aphid-like insect native to Japan. It arrived on the east coast of the United States in 1950, where it had no natural predators. By 1985 it had made its way to New England, migrating north at approximately 30 km per year (Orwig et al. 2002). Its host, hemlock, is an important component of old growth forests in the transition hardwoods region and provides vital wildlife habitat. It is shade-tolerant and slow growing. The impending loss of hemlock threatens to homogenize species composition and structure, as its foliage provides habitat and shelter in shaded understories. Orwig et al. (2002) showed that damage to hemlock resulting in mortality and loss of vigor is related to the duration of infestation. Sites infested for a longer duration of time had more foliage loss and higher mortality. Mortality appears to be weakly related to aspect, where drier western slopes tend to hasten an infested hemlock’s demise, but overstory composition (mixed vs. hemlock stands), slope, and elevation make no difference. Reaction to the adelgid in the transition hardwoods zone has been to salvage and preemptively log many hemlock stands, which is accelerating hemlock’s disappearance from the landscape, especially in the south (Kizlinski et al. 2002, Orwig et al. 2002). Since the mid-1980s, 4290 ha of hemlock stands were preemptively logged in Connecticut and Massachusetts (Orwig et al. 2002). Orwig et al. (2002) predict an increasing rate of both infestation and preemptive salvage. Harvesting affects regeneration dynamics, future forest composition and structure, and wildlife habitat, producing more rapid and pronounced microenvironmental and vegetation changes than chronic infestation damage (Kizlinski et al. 2002).The loss of hemlock from much of the transition hardwood zone has profound implications, as hemlock has unique structural characteristics for bird habitat (Tingley at al. 2002). Black-throated green warbler, blackburnian warbler, hermit thrush, and Acadian flycatcher, all associated with intact stands, suffered declines where hemlock was eliminated. Other species of more open woods increased, including eastern wood pee-wee, brown-headed cowbird, tufted titmouse, white-breasted nuthatch, red-eyed vireo, hooded warbler, and several woodpeckers (Tingley et al. 2002). Forest composition changes as hemlock declines: black birch, brambles, sedges, and hayscented fern increase in logged areas, and black birch is an important component of the regeneration even when salvage logging does not occur (Kizlinski et al. 2002). Black birch, maple, and oak establishment in the understory is abundant in stands with high hemlock mortality (Orwig et al. 2002). These developments are broadly consistent with the general pattern of stand initiation in the transition hardwood zone. The decomposition of forest floor was shown to be faster in damaged areas than undamaged areas, releasing nutrients into the soil. Where salvage or preemptive logging has occurred, decomposition rates are higher. Kizlinski et al. (2002) recommend that silviculture that allows for reestablishment of tree seedlings prior to harvesting (i.e. by allowing the adelgid to gradually break up the hemlock canopy) would lessen ecological impacts of hemlock removal. An approach to salvage (or presalvage) harvest that more deliberately mimics the shelterwood system might have similar effects, and would reduce losses to decay.
Invasive Species A wide array of invasive species are of concern throughout the transition hardwood zone. Many invade the forest understory, even in stands with little recent disturbance. An online list and atlas of invasive species in New England is maintained by the University of Connecticut (http://invasives.eeb.uconn.edu/ipane/). Many of these species successfully invade forests, including closed canopy forests, and can interfere with or potentially displace regeneration of native trees and plants. Many are also tied to land use history, because they are often found at or near the sites of abandoned farms, or are spread along roads and power lines. Glossy buckthorn (Rhamnus frangula) exemplifies this emerging challenge. Glossy buckthorn, a shrub of European origin, has naturalized throughout much of the northeastern United States. First observed in 1898, glossy buckthorn increased dramatically since the 1970s and 80s (Frappier et al. 2003a). Frappier et al. (2003a) studied the effects of glossy buckthorn on ground-level vegetation in four transition hardwood forests in southern New Hampshire that were of old-field origin and dominated by white pine. They found that buckthorn density was inversely related to woody seedling density, herb cover and species richness. Buckthorn density explained more variation in tree seedling density than environmental variables such as canopy openness, pH, or soil texture. Frappier et al.’s (2003a) suggest that high abundance of this invasive species can cause reduced native seedling density and alter ground level plant abundances, negatively affecting recruitment and richness. Resource limitation by shading was suggested as the most likely mechanism for this effect. In a separate study, Frappier et al. (2003b) found that the spread rate of glossy buckthorn was 6.7 meters per year, and over 35 years it had dominated an appreciable area at one site in New Hampshire. A lag in the spread of buckthorn during the first 20 years of establishment would have made it difficult to predict the present effects of this species early in its invasion. Glossy buckthorn now forms dense thickets at this site, and will present a challenge to future regeneration of the stand. Because glossy buckthorn can regenerate from buried seed, simple control through herbicide may prove difficult.
The land-use pattern in the modern-day transition hardwood zone is heavily influenced by, but not identical to, the pattern created during European colonization and settlement. That pattern was, in turn, influenced by the pattern of settlement by native Americans. The early European settlers tended to locate their villages at the sites of native villages that had been abandoned or depopulated by disease (Russell 1976); these were often well-situated, and the availability of cleared land for agriculture was a major attraction. Early European roads between major villages often followed tracks used by native Americans; these corridors became highways, and are now areas of major development pressure within the transition hardwood zone. In other ways, the early European economy of the region makes itself felt: some modern observers have complained (not without justification) that the traffic pattern of metropolitan Boston was designed by 17 th century cattle. In the forested setting, abandoned homesites persist, stone walls are a constant reminder of past land use, dates of farm abandonment influence forest availability for harvest, and operational details such as the layout of skid trails and landings are often driven by the spatial arrangement of long-disused agricultural lanes. One persistent feature inherited from the early settlement pattern in this region is the prevalence of small, nonindustrial forest ownership in parcels from tens to perhaps a few hundred hectares. Many early villages in the region followed an ancient European pattern, with small fenced crofts surrounding village homes and the bulk of the town held as common, undivided forest and pasture. However, over time, much of the common land was divided and assigned to individuals, and scattered outlying farms also became common. The footprint of both types of ownership can be seen in the New England landscape today (Russell 1976). In many rural towns, it is only within recent decades that populations have recovered to, or perhaps surpassed, their peak before agricultural abandonment in the mid 1800s, while in others, it now far exceeds that peak; but much of the pattern of residential development is driven by roads and parcel boundaries established centuries before. Strikingly, the minimum lot sizes adopted by some towns to limit the number (but often inadvertently increase the sprawl) of new residences mirror the tract sizes of early village crofts (perhaps a few hectares). These new “crofts” are not usually completely cleared, but remain in forest, and instead of a self-sufficient flow of food and fuel to residents enmeshed in a local system of limited markets, they are expected to provide a variety of benefits – many aesthetic and psychological – to the citizens of a highly mobile civilization linked to global markets. However, continuing reductions in parcel size have raised serious concerns about the sustainability of forestry in much of the region (Kittredge et al. 1996). The converse of this dominance of small, nonindustrial forest owners is the relative lack of large industrial, or even large nonindustrial, forest holdings. There are, however, notable exceptions arising from a wave of land consolidations in the late 19 th and early 20 th centuries. Notable among these are the watershed protection holdings for the water supplies of New York, Boston, and several smaller cities and towns (e.g., Barten et al. 1998). In several instances, well-heeled families assembled large ownerships from abandoned farms or cutover lands, often paying what the sellers considered at the time to be exorbitant rates. Some of the major research forests in the area, including the Great Mountain Forest in northwestern Connecticut (Winer 1956), the Yale-Myers Forest in northeastern Connecticut (Meyer and Plusnin 1945) and the Yale-Toumey Forest in southwestern New Hampshire (Toumey and Hawley 1916, Toumey 1932), the Caroline A. Fox Research Forest of the New Hampshire Division of Forests and Lands (Baldwin 1947), and the Harvard Forest in Massachusetts (Fisher 1931) were assembled in this fashion. The opening of the 21 st century may herald a similar wave of consolidation, as local and regional land conservancies consolidate remaining undeveloped lands. Despite the highly parcelized and diverse land tenure in the region, much of the transition hardwood zone remains in a “mid-successional wave” with serious consequences for the maintenance of biodiversity. In part, this reflects both the wave of agricultural abandonment in the late 19 th century and the resynchronizing effect of the 1938 hurricane. In part, it also reflects (possibly misguided) forest harvesting practices that persist in much of the nonindustrial forest land of the transition hardwood zone, as will be discussed below. Fragmentation (as distinct from parcelization) is also a serious concern in the region, though roads, residential clearings, and the outward sprawl of commercial development along main travel corridors are primary drivers (Sundquist and Stevens 1999), with forest harvesting playing little role.
Recently, several prominent authors (Franklin et al. 2002, Seymour et al. 2002) have called for a shifting silvicultural paradigm to reflect “presettlement” or “natural” patterns of disturbance, arguing that such an approach would be more likely to conserve both ecosystem functions and biodiversity. While a definition of “natural” as devoid of human influence (e.g. Hunter 1996) is clearly artificial and ethnocentric when applied to the transition hardwood zone, with its long history of habitation and use by diverse human societies, it is well worth asking what a forestry paradigm based on pre-European disturbance regimes might mean. Certainly it does not mean the small-gap paradigm suggested for “the northeast” by Seymour et al. (2002); that paper defined the Northeast as the northern hardwood and Acadian mixed-conifer zones, two zones with strikingly different pre-European disturbance regimes. The short discussion of gap-phase dynamics in northeastern forests by Lorimer (1989) also seems to refer to areas north of the transition hardwoods. Rather, as the review by Lorimer and White (2003) indicates, the transition hardwood zone before European contact almost certainly reflected a regime of frequent large-scale stand-replacing disturbances, such as hurricanes, coupled with the common occurrence of understory – and perhaps the occasional catastrophic – fire. Can fire be reintroduced successfully into the transition hardwood landscape? Certainly, its previously common occurrence (Wade et al. 2000, Parshall and Foster 2002, Lorimer and White 2003), and potential impact on the relative dominance of oak, pine, and shade-tolerant species (Lorimer 1984; Lorimer et al. 1994; Abrams 1992, 2001) suggest that it would be important. However, reintroduction of fire may not even be biologically straightforward. For example, in one study of prescribed fire in a transition hardwood stand with a mountain laurel (Kalmia latifolia) dominated understory (Ducey et al. 1996, Moser et al. 1996), a single understory burn merely intensified the dominance of laurel; a fire hot enough to cause some canopy mortality was required for resprouting oak to reach a competitive position. The relatively low intensity of modern prescribed fires, in comparison to fires of the pre-European period, has been cited for the failure to obtain desired oak regeneration (McGee et al. 1995). Franklin et al. (2003) suggested that the relatively large size of otherwise fire-intolerant but shade-tolerant hardwoods meant that simple reintroduction of fire would be inadequate to obtain their control. From a social standpoint, the infrastructure necessary to manage a coherent and safe prescribed fire effort is probably beyond the capability (and interest) of most nonindustrial private forest owners in the region. Although efforts at the formation of landowner cooperatives have been promising for other management activities (Stevens et al. 1999, Klosowski et al. 2001), it is not clear that such cooperatives developed for management of prescribed fire would find adequate support. Finally, the air quality impacts of widespread reintroduction of prescribed fire would likely prove problematic. Much of the transition hardwood zone is downwind of large urban areas within the region, and significant coal-fired power stations in the upper Midwest (CCRC 1998). The addition of particulates and other pollutants from large-scale use of prescribed fire would undoubtedly create legal and policy challenges. As Oliver and Larson (1996), Boose et al. (2002), and Lorimer and White (2003) indicate, given the high frequency of hurricanes in this zone, the landscape before European settlement would nearly always have been dominated by the most recent post-hurricane cohort. However, the composition of the landscape would have fluctuated dramatically in the proportion of stand initiation and stem exclusion stands, an issue to which we will return later. The silviculture of single-cohort stands has long been a topic of research in the region. The regeneration of major species is reasonably well understood (e.g. Hibbs 1982a, 1982b; Kittredge and Ashton 1990, 1995; Abrams 1992; Ashton 1992; Ashton and Larson 1996). In particular, oaks are strongly dependent on either stump sprouts, or well-established advance regeneration, for their dominance in the next cohort (Smith and Ashton 1993). Furthermore, it is often desirable to protect the stems of young eastern white pine from damage by the white pine weevil (Pissodes strobii); this can be accomplished by leaving partial shade until the pines have reached a predetermined height (Peirson 1922, Maughan 1930). Both factors argue strongly for the use of some form of shelterwood system for single-cohort silviculture (Hannah 1988, Seymour 1995). The general development of single-cohort mixed-species stands in the transition hardwood zone is described by Oliver (1978). Broadly speaking, such stands undergo a prolonged period of dynamic stratification, with initial dominance by paper, black and yellow birches and red maple. Eventually oaks and, where they are present, hickories, penetrate and then overtop the stratum of initial dominants. Pines may exist as scattered dominants or superdominants throughout development, while a variable lower stratum of shade-tolerants such as beech, hemlock, and sugar maple is a characteristic feature. Quantitative guidance for intermediate treatments, such as thinnings, is not well developed for this region; for example, many foresters rely on the midwestern oak density management guidance of Gingrich (1967) in writing prescriptions for this region. However, given the extreme dependence of dollar value on individual tree grade (even for sawlog-size material of the same species, prices may vary by over an order of magnitude depending on presence or absence of defect), individual crop-tree approaches to thinning are eminently suitable, and individual-tree level information (as opposed to stand-level averages) may be most useful (Kittredge 1988, Ward 2002). Given the land-use history of the transition hardwood zone, and the “mid-successional wave” phenomenon, many stands are lacking in structural legacies such as large live trees, large snags, and large downed dead trees. Notable exceptions occur where wolf trees (either former lone trees amid pasture, or as rows along old roads) are present in abundance. It is, of course, impossible to develop large snags or downed logs without first growing large live trees. One approach to developing such legacies within single-cohort stands is to employ some form of green-tree retention (e.g., Miller and Kochenderfer 1998). In a shelterwood context, this falls under the traditional rubric of “standards;” if a sufficient number of standards are retained long into the next rotation, the stand may be considered truly a two-cohort stand. Other forms of green tree retention, either dispersed or aggregated, may be suitable in this zone. However, there is virtually no quantitative guidance for foresters either for selection of retention trees, or for predicting their impact on the next cohort. This presents particular difficulties given the importance of grade to the value of hardwoods, and the propensity of oaks (as with other genera in the Fagaceae), along with sugar maple, to develop crooks following partial shade or suppression (Sorenson and Wilson 1964, Trimble 1968, Oliver and Larson 1996). Given the pre-European disturbance regime, and the relative shade-intolerance of the dominant species in the transition hardwood forest, a single-tree selection scheme appears to be an inappropriate silvicultural system in this type. However, given that the frequency and intensity of pre-European hurricane and fire damage varied over the transition hardwood region (Boose et al. 2001, Parshall and Foster 2002), with some areas escaping stand-scale disturbance for perhaps more than a century or two, the rather limited use of uneven-aged methods could be justified. Such even-aged management would be most applicable on small landholdings where there is a desire for a continuing flow of products (or dollars). Even more important for many small woodlot owners is the aesthetic impact of recent harvests (Kelty et al. 2003). These factors suggest some form of group selection or group shelterwood, including a variety of irregular shelterwood techniques, might be useful. Group selection was studied as early as Cline 1924 in this forest type. Kelty et al. (2003) review some of the challenges of developing a group selection system in an existing transition hardwood single-cohort structure. They point out that relatively large openings (>0.12 ha) are required to maintain tree species diversity, and that elimination of trees in lower strata is also required. Perhaps the most critical issue facing transition hardwood silviculture is the continuation of wide-spread high-grading. For example, much of the timber harvest on nonindustrial private forests in Massachusetts continues to be the selective removal of oaks and pines (Kittredge et al. 2003). It is possible that much of this unsustainable practice occurs out of ignorance, despite the well-established knowledge base for successful oak and pine regeneration. Some may also be perpetrated by foresters misapplying diameter-distribution based regulation (e.g. the BDq method, Guldin et al. 1996) in a selection system, under the mistaken belief that a reverse-J diameter distribution and a complex species mix implies an uneven-aged stand. Rather than producing a desired uneven-aged stand, such an approach can merely simplify the stand structure and hasten the dominance of the shade-tolerants (such as beech, red maple, and hemlock) in lower canopy strata. Application of diameter-distribution regulation in a single-cohort structure violates the assumption that diameter and age are correlated, and can lead to undesirable outcomes (Kelty et al. 2003).
3) Maintaining a Diverse Landscape
In some respects, attempting to reconstruct pre-colonial forests conditions in a region that has been substantially modified by development and dense road networks misses the obvious point that several centuries of human activity have had some irreversible effects. The hurdles presented by parcelization and existing land uses further complicate any efforts to restore forests to a more pristine state. Even if we were able to adequately describe natural disturbance regimes, we could not reconstruct pre-colonial forest communities given the density of humans and the alterations they have imposed much of this region. Recognition of these limitations has prompted some to recommend that providing a range of forest habitats capable of supporting viable populations of native species is a more pragmatic approach than attempting to simulate natural disturbances (e.g., Litvaitis et al. 1999, Confer and Pascoe 2003). Such an approach may in fact deviate from historic conditions in highly-modified landscapes (Litvaitis 2003). Perhaps these conditions present the greatest challenges.
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