Tree Physiology 28, 537–549
© 2008 Heron Publishing—Victoria, Canada
Changes in composition, structure and aboveground biomass over
seventy-six years (1930–2006) in the Black Rock Forest, Hudson
Highlands, southeastern New York State
W. S. F. SCHUSTER,1,2 K. L. GRIFFIN,3 H. ROTH,4 M. H. TURNBULL,5 D. WHITEHEAD6 and
D. T. TISSUE7
1
Black Rock Forest Consortium, 129 Continental Road, Cornwall, NY 12518, USA
2
Corresponding author ([email protected])
3
Lamont-Doherty Earth Observatory, Columbia University, 61 Route 9W, 6 Biology, Palisades, NY 10964, USA
4
Department of Environmental Science, Barnard College, 3009 Broadway, 404 Altschul Hall, New York, NY 10027, USA
5
School of Biological Sciences, University of Canterbury, Private Bag 4800, Christchurch, New Zealand
6
Landcare Research, P.O. Box 40, Lincoln 7640, New Zealand
7
Department of Biology, Texas Tech University, Lubbock, TX 79409-3131, USA
Received March 1, 2007; accepted June 9, 2007; published online February 1, 2008
Summary We sought to quantify changes in tree species
composition, forest structure and aboveground forest biomass
(AGB) over 76 years (1930–2006) in the deciduous Black
Rock Forest in southeastern New York, USA. We used data
from periodic forest inventories, published floras and a set of
eight long-term plots, along with species-specific allometric
equations to estimate AGB and carbon content. Between the
early 1930s and 2000, three species were extirpated from the
forest (American elm (Ulmus americana L.), paper birch
(Betula papyrifera Marsh.) and black spruce (Picea mariana
(nigra) (Mill.) BSP)) and seven species invaded the forest
(non-natives tree-of-heaven (Ailanthus altissima (Mill.) Swingle) and white poplar (Populus alba L.) and native, generally
southerly distributed, southern catalpa (Catalpa bignonioides
Walt.), cockspur hawthorn (Crataegus crus-galli L.), red mulberry (Morus rubra L.), eastern cottonwood (Populus deltoides
Bartr.) and slippery elm (Ulmus rubra Muhl.)). The forest canopy was dominated by red oak and chestnut oak, but the
understory tree community changed substantially from mixed
oak–maple to red maple–black birch. Density decreased from
an average of 1500 to 735 trees ha –1, whereas basal area doubled from less than 15 m2 ha –1 to almost 30 m2 ha –1 by 2000.
Forest-wide mean AGB from inventory data increased from
about 71 Mg ha –1 in 1930 to about 145 Mg ha –1 in 1985, and
mean AGB on the long-term plots increased from 75 Mg ha –1
in 1936 to 218 Mg ha –1 in 1998. Over 76 years, red oak
(Quercus rubra L.) canopy trees stored carbon at about twice
the rate of similar-sized canopy trees of other species. However, there has been a significant loss of live tree biomass as a
result of canopy tree mortality since 1999. Important constraints on long-term biomass increment have included insect
outbreaks and droughts.
Keywords: basal area, canopy, carbon, density, environmental
change, forest inventory, long-term, mortality, oak, Quercus,
red maple.
Introduction
Understanding the dynamics of the global carbon (C) cycle
and its underlying drivers is critical to an explanation of the behavior of our planet’s climate system (IPCC 2001). Forests
store and cycle most of the earth’s terrestrial biomass and thus
play a dominant role in the global C cycle (Dixon et al. 1994,
Landsberg and Gower 1997). Measuring forest carbon and its
temporal fluxes has become a matter of global interest (Brown
2002). Temperate deciduous forests in the northern hemisphere comprise some of the world’s most substantial C sinks
(Pacala et al. 2001, Myneni et al. 2001), which partially offset
anthropogenic increases in atmospheric CO2 concentration
and the associated climate consequences. In particular, forest
ecosystems in the USA appear to function as globally important sinks (Birdsey and Heath 1995, Goodale et al. 2002), but
many of the spatial and temporal details, and details about the
controlling factors, remain elusive (Liu et al. 2006).
Long-term and large-scale studies are critical for quantifying landscape-scale contributions to C sequestration and responses of forests to environmental change (Körner 2003).
Critical carbon releases, important in determining long-term
sequestration and fluxes, occur rarely and often rapidly, and
are thus usually missed by short-term studies. Improving our
understanding of long-term C budgets in large, topographically heterogeneous regions, including the influences of rare
events, remains a major goal (Chapin et al. 2006). Factors important to long-term C flux such as disturbance history, suc-
538
SCHUSTER, GRIFFIN, ROTH, TURNBULL, WHITEHEAD AND TISSUE
cessional age and community composition, interact in complex fashion with other biotic and abiotic factors to regulate
long-term forest C storage (Ryan et al. 1997, Caspersen et al.
2000). Long-term studies of forests as they age are important
to quantify changes in sequestration rates and determine if mature C-sink forests today will remain sinks in the face of environmental change (Hyvönen et al. 2007). Although ecosystem
theory states that net C uptake will decrease to zero as ecosystems approach maturity (Odum 1969, 1971), some studies
have documented a continuing C-sink potential in older forests
(e.g., Schulze et al. 2000, Carey et al. 2001, Zhou et al. 2006).
Therefore, assessing the contributions of, and interactions
among, land use history, forest age and ecophysiological responses to changing environmental parameters is a requirement for predicting future CO2 uptake and C dynamics (Albani
et al. 2006).
In October 2006 the Canopy Processes group of IUFRO (International Union of Forest Research Organizations) held a
workshop entitled “Regional Forest Responses to Environmental Change” at the Bartlett Experimental Forest, Harvard
Forest, and Black Rock Forest, all located in the northeastern
USA. Recent studies have indicated that the eastern USA
functions as a long-term regional net carbon sink (Albani et al.
2006 and references therein). Stated goals of the workshop
were “to assess the state of knowledge of regional forest responses to global change” and “to introduce researchers and
students to the breadth of research and the regional characteristics of forests in the northeastern USA.”
Long-term records from inventories and study plots in the
Black Rock Forest, a 1530-ha oak-dominated forest research
station in southeastern New York State, enable analyses of
how forests of this particular region have changed over the past
several decades. Forests of the surrounding Highlands Physiographic Province have generally been aggrading (i.e., accumulating biomass) following recovery from repeated clearcutting
and widespread conversion to agricultural use during the 19th
and early 20th centuries (Tryon 1930). We used forest inventory records beginning in 1930 and repeated measurements
between 1936 and 2006 on undisturbed plots, in conjunction
with species-specific allometric equations, to document how
live, aboveground forest biomass (about 50% C; Birdsey
1992) and key parameters of forest composition and structure
have changed in this forest over the intervening period. The results suggest proximate causes for some of the changes and
provide a foundation for further studies to determine which
patterns represent regional responses to environmental
change.
Methods
Site description
The Black Rock Forest (BRF) is a 1550-ha oak-dominated
forest preserve in southeastern New York State (41°24′ N,
74°01′ W; Figure 1). It is located within the Highlands Physiographic Province, an approximately 8000 km2 uplands characterized by deciduous forest and underlain by metamorphic
rocks of the Reading Prong formation (Fenneman 1938,
Braun 1967, Schuberth 1968). Through land acquisition and
donations the forest increased in size from 1258 ha in 1928, to
1416 ha at the time of the second forest-wide inventory in
1985, to 1550 ha in 2006 (Figure 1). The topography is rocky
Figure 1. Map of the Black Rock
Forest (BRF) showing inventory
plots (1985: stars; 1985 and 2000:
circled stars) and long-term plots
(numbered and unnumbered
squares) with inset map showing
location within the Highlands
Physiographic Province (shaded).
TREE PHYSIOLOGY VOLUME 28, 2008
BIOMASS CHANGES IN THE BLACK ROCK FOREST
with steep slopes and elevations ranging from 110 to 450 m
above sea level. Mean annual precipitation is 1.2 m and air
temperature is strongly seasonal, with monthly averages ranging from –2.7 °C in January to 23.4 °C in July (Ross 1958,
Turnbull et al. 2001). The soils are mostly medium textured
loams, with bedrock or glacial till parent material at depths
ranging from 0.25 to 1 m (Olsson 1981). Soil reaction is
acidic, availability of nutrients is low, and site index ranges
from poor to good (Lorimer 1981; examples in Table 1).
European settlement in the area began around 1700 and the
land was repeatedly logged, with a small proportion of the forest completely converted to agriculture and livestock pasture
and then abandoned by 1900 (Raup 1938). Frequent clearcuts
and fires resulted in a preponderance of hardwood sprout regeneration (Tryon 1939). The BRF became established as a research forest in 1928 (Tryon 1930) and was a unit of the Harvard University Forest system from 1949 to 1989. Since 1989,
the BRF has been operated as a field station and nature preserve by the Black Rock Forest Consortium, a group of academic institutions from the surrounding region (Mahar 2000,
Buzzetto-More 2006).
Tree species composition
Changes in forest-wide tree species composition were explored by comparing historical records from four sources. The
first was a 1938 list of the vascular plants of the BRF based on
botanical inventories and vegetation transects made in 1936
and 1937 (Raup 1938). The second was a 1949 list of common
tree species in the forest from an operations report of the forest
research station (Tryon and Finn 1949). The third was a list of
tree species encountered during a 1985 forest inventory based
on 218 sample plots around the forest (Figure 1, Friday and
Friday 1985). The fourth was a 2003 complete flora of the vascular plants of the BRF based on the results of field surveys
completed during 1990–1993 and 1996–2000. Voucher specimens from this survey were filed at the Brooklyn Botanic Garden (BKL), the New York State Museum Herbarium (NYS)
and the BRF. In this paper, we define a tree as a woody perennial plant generally with a single stem and capable of reaching
a height of more than 5 m. Neither the second nor third data
539
sources purported to include all tree species in the forest, but
they were consulted to provide information on the timing and
details of composition changes.
Long-term plot structural measurements and biomass
estimation
Sixteen long-term plots ranging in size from 0.04 to 0.1 ha
were established in the forest between 1931 and 1936 to monitor forest growth (Tryon 1939; Figure 1). Eight plots have remained undisturbed, and on these plots all trees greater than
2.54 cm dbh were assigned to a crown class (dominant,
codominant (canopy trees); intermediate, suppressed (understory trees)) and were measured for diameter at the same
marked location on the trunk about every five years through
1993. Since 1994, measurements of dbh and crown class have
been performed each year. In total, measurements were made
on 1224 trees of 22 species. Most measurements were made in
July or later months, when the majority of diameter growth
had already occurred (Karnig and Stout 1969), and therefore
represent status at the end of that year’s growing season. In
four of the years, measurements were made between April and
early May, and are here taken to represent status at the end of
the previous year’s growing season.
The eight remaining long-term plots are located at intermediate forest elevations (Plots 1–8 in Figure 1). Table 1 lists
stand age for these plots determined from forest records and
confirmed by examination of increment cores (Lorimer 1981,
D’Arrigo et al. 2001), mean height of the canopy trees, slope
and soil characteristics, and site index (calculated from the
heights of dominant and codominant oaks in 1998; Schnur
1937). Based on these data, half of the plots are on sites of
good quality, with the other half on sites ranging from fair to
poor quality, typical for many forests in the Highlands
(Lorimer 1981). The plots were originally established in pairs,
one within an area that was experimentally thinned and the
other in a nearby area left undisturbed as a control. Thinning
operations removed diseased and dying trees, as well as those
species considered to be less economically desirable (e.g.,
gray birch (Betula papyrifera), bigtooth aspen (Populus
grandidentata). However, among the remaining thinned plots
Table 1. Site conditions and stand data for eight long-term plots examined in this study.
Plot
Stand age
in 2000
Slope (%)
Aspect
Soil pH1
Height in 19982
(Mean ± SE)
Site index3
(1998 est.)
1
2
3
4
5
6
7
8
115
115
115
115
95
95
90
90
11
9
11
8
2
2
11
9
NW
NW
NW
NW
W
NE
NE
NE
3.65
3.70
3.85
4.00
4.55
4.25
3.90
3.85
24.2 ± 0.7 a
24.7 ± 0.8 a
18.2 ± 1.0 b
15.8 ± 0.9 b
23.8 ± 1.4 ac
24.6 ± 0.8 ac
20.6 ± 0.4 d
22.2 ± 0.8 cd
61 (average)
61 (average)
43 (poor)
39 (poor)
61 (average)
62 (average)
53 (fair)
57 (fair-average)
1
Mean of top and subsoil composites from eight subsamples per plot.
Mean heights followed by same letter do not differ significantly at α = 0.05, LSD test.
3
For mixed oaks (Schnur 1937).
2
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540
SCHUSTER, GRIFFIN, ROTH, TURNBULL, WHITEHEAD AND TISSUE
(i.e., Plots 2, 4, 6, and 8) only Plots 2 and 6 were significantly
reduced in density and biomass compared with their paired
control plots (BRF data).
Density, basal area, and aboveground biomass (see below)
were calculated for each plot for each measurement date. All
plots were measured in 1936 and every year since 1994, but in
the interim different pairs of plots were often not measured in
the same year, and thus we used linear interpolation to estimate parameters between measurement periods. Values from
all eight plots were averaged, because among-plot differences
were not of focal interest, and 95% confidence intervals were
calculated.
We used previously derived allometric regression equations
(Brenneman et al. 1978) to estimate live aboveground tree biomass (AGB) from dbh measurements for the most common
species in this study: red oak (Quercus rubra), chestnut oak
(Quercus prinus), white oak (Quercus alba), red maple
(Acer rubrum), sugar maple (Acer saccharum), yellow birch
(Betula alleghaniensis), black birch (Betula lenta), pignut and
shagbark hickories (Carya spp.), white ash (Fraxinus
americana), basswood (Tilia americana), black cherry (Prunus serotina) and eastern hemlock (Tsuga canadensis). Complete harvest, drying, and weighing of 11 red oaks and 11
chestnut oaks (BRF data) indicated that these equations were
the most accurate for the Black Rock Forest’s dominant species among the many equations available (e.g., Ter-Mikaelian
and Korzukhin 1997). For the remaining less common tree
species we used the general New-York-State-derived equations of Monteith (1979) for either hardwoods or softwoods.
Live biomasses in tree stumps, roots and understory vegetation
are omitted from these formulae. We estimated AGB for each
plot by summing the individual tree AGB estimates and dividing by plot area. These biomass data were not transformed because they were normally distributed and only mildly
heteroscedastic (Sokal and Rohlf 1981).
To compare the biomass increment rates of canopy trees of
different species, we selected data for all canopy trees of the
six most common species with dbh in 1936 between 10 and
20 cm that survived until 2006, representing in total 103 trees.
Aboveground biomass values from allometric equations were
averaged by species and year over these trees.
squares that were subsequently disturbed). We estimated
aboveground biomass for each tree with allometric equations
and then estimated plot aboveground biomass per hectare
(AGB) by summing for all trees on each plot and dividing by
the area. We then regressed estimated 1930 AGB for each plot
on mean wood volume in cords (1 cord is about 2.26 m3 ) per
hectare for the stand in which each plot was located. The resulting regression relationship, Ba = 2.169V, where Ba is aboveground biomass in Mg ha –1 and V is wood volume in cords per
hectare, had an r 2 of 0.68 (P < 0.002). Forest-wide 1930 AGB
was estimated as the stand-area-weighted mean of these values.
The 1985 inventory subdivided the forest into 71 stands
based on species composition, canopy height and canopy
cover (Friday and Friday 1985). Three sample points (occasionally more) were located at regular intervals along the long
axis of each stand, resulting in a total of 218 sample plots (star
symbols in Figure 1). All trees greater than 5 cm dbh in these
sample plots were tallied with a 10-factor basal area prism and
measured for dbh and crown class, totaling 2078 trees of 37
species. The AGB was estimated for each tree and plot as
above, plot values were averaged for each stand, and forest-wide 1985 AGB was estimated as the area-weighted mean
for each of the 71 stands. The 1930 and 1985 forest-wide AGB
estimates are not directly comparable because of different inventory methods and areas added to the forest between inventories. A total of 56 experimental thinnings and timber harvests were carried out in the interim, impacting about
one-third of the forest, and 10 hectares of conifers were
planted in the 1940s on formerly cleared areas omitted from
the first inventory (Tryon and Finn 1949, Harrington and
Karnig 1975).
In 2000 (or in a few cases 1999 or 2001), a total of 45 of
these 218 inventory plots were resampled (circled stars in Figure 1), by the same methods as described above. These plots
were selected to quantify changes (1) within the Cascade
Brook watershed in the southeastern part of the forest and (2)
within and around mature stands of eastern hemlock in the
northern portion of the forest. Simple plot means were calculated for 1985 and 2000 to quantify AGB and its change over
the interval across this subset of plots.
Results
Forest inventories and biomass estimates
Forest-wide inventories were completed in 1930 and 1985. In
the 1930 inventory, the forest was subdivided into 150 stands
based on differences in species composition and tree density.
Standard cruise methods for the time, examining all trees of
cordwood size (i.e., those greater than 10 cm dbh; Tryon
1930), were used to determine stand area, stand age (based on
historical records and tree ring counts), density by species and
mean wood volume for each stand.
We developed a regression equation to estimate aboveground biomass for the 150 stands inventoried in 1930 from
inventory volume estimates in combination with recorded diameter measurements of trees on 12 original unthinned plots
(Plots 1, 3, 5, 7, in Figure 1 and eight others indicated by
Tree species composition
Between the 1930s and 1990s, three tree species were extirpated from the BRF and eleven tree species were added for a
net increase of eight species (Table 2). American elm (Ulmus
americana L.) was common on lower slopes and near streams
but was completely eradicated after 1949 by the introduced
fungal Dutch elm disease (Ophiostoma ulmi (Buism.) Nannf.),
which was spread through northeastern North America primarily by the native elm bark beetle (Hylurgopinus rufipes
Eich.; Gibbs 1981). Black spruce (Picea mariana) was present
in higher-elevation wetlands in the 1930s but failed to survive
or reproduce and was not reported in forest records after 1949.
Paper birch (Betula papyrifera Marsh) was rare in the BRF in
TREE PHYSIOLOGY VOLUME 28, 2008
BIOMASS CHANGES IN THE BLACK ROCK FOREST
541
Table 2. Tree species in the Black Rock Forest 1938–2003.
Scientific name
Common name
Acer pensylvanicum L.
Acer rubrum L.
Acer saccharinum L.
Acer saccharum Marsh.
Acer spicatum Lam.
Ailanthus altissima (Mill.) Swingle
Alnus incana (L.) Moench
Alnus serrulata (rugosa) (Dryand) Willd.
Amelanchier canadensis (L.) Medik.
Betula alleghaniensis Britt. (B. lutea Michx.f.)
Betula lenta L.
Betula papyrifera Marsh.
Betula populifolia Marsh.
Carpinus caroliniana Walt.
Carya cordiformis (Wang.) Koch
Carya glabra (Mill.) Sweet.
Carya ovata (Mill.) Koch
Castanea dentata (Marsh.) Borkh.
Catalpa bignonioides Walt.
Cornus florida L.
Crataegus crus-galli L.
Crataegus macrosperma Ashe
Fagus grandifolia Ehrh.
Fraxinus americana L.
Fraxinus nigra Marsh.
Fraxinus pennsylvanica Marsh.
Ilex verticillata (L.) A. Gray
Juglans cinerea L.
Juglans nigra L.
Juniperus virginiana L.
Larix decidua Mill.
Larix laricina (DuRoi) Koch
Liriodendron tulipifera L.
Malus pumila Mill.
Morus rubra L.
Nyssa sylvatica Marsh.
Ostrya virginiana (Mill.) Koch
Picea abies (L.) Karst.
Picea glauca (Moench) Voss
Picea mariana (nigra) (Mill.) B.S.P.
Pinus banksiana Lamb.
Pinus resinosa Soland.
Pinus rigida Mill.
Pinus strobus L.
Platanus occidentalis L.
Populus alba L.
Populus deltoides Bartr.
Populus grandidentata Michx.
Populus tremuloides Michx.
Prunus pennsylvanica L. f.
Prunus serotina Ehrh.
Prunus virginiana L.
Pyrus communis L.
Quercus alba L.
Quercus bicolor Willd.
Quercus coccinea Muenchh.
Quercus ilicifolia Wang.
1
Striped maple
Red maple
Silver maple
Sugar maple
Mountain maple
Tree of heaven
Speckled alder
Smooth alder
Shadbush
Yellow birch
Black birch
Paper birch
Gray birch
Ironwood
Bitternut hickory
Pignut hickory
Shagbark hickory
American chestnut
Southern catalpa
Dogwood
Cockspur hawthorn
Hawthorn
Beech
White ash
Black ash
Green ash
Winterberry
Butternut
Black walnut
Eastern redcedar
European larch
Tamarack
Tulip poplar
Common apple
Red mulberry
Black gum
Hop-hornbeam
Norway spruce
White spruce
Black spruce
Jack pine
Red pine
Pitch pine
White pine
Sycamore
White poplar
Eastern cottonwood
Bigtooth aspen
Quaking aspen
Pin cherry
Black cherry
Choke-cherry
Cultivated pear
White oak
Swamp white oak
Scarlet oak
Scrub oak
Year of report1
1938
1949
1985
2003
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
Notes
Found on land, added after 1985
First reported in 1980s
One tree remaining in 2003
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
Raup 1938, Tryon and Finn 1949, Friday and Friday 1985, Barringer and Clemants 2003.
TREE PHYSIOLOGY ONLINE at http://heronpublishing.com
First reported in 1990s
First reported in 1990s
Only cultivated trees
First reported in 1990s
Only cultivated trees
Not reported after 1949
Only cultivated trees
First reported in 1990s
First reported in 1985
Continued overleaf
542
SCHUSTER, GRIFFIN, ROTH, TURNBULL, WHITEHEAD AND TISSUE
Table 2 (Cont'd). Tree species in the Black Rock Forest 1938–2003.
Scientific name
Quercus montana (prinus) Willd.
Quercus rubra L. (Q. borealis Michx. f.)
Quercus velutina Lam.
Rhamnus cathartica L.
Robinia pseudo-acacia L.
Sassafras albidum (officinale) (Nutt.) Nees
Tilia americana L.
Tsuga canadensis (L.) Carr.
Ulmus americana L.
Ulmus rubra Muhl.
Common name
Chestnut oak
Red oak
Black oak
European buckthorn
Black locust
Sassafras
Basswood
Eastern hemlock
American elm
Slippery elm
Total
1
Year of report1
1938
1949
1985
2003
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
×
40
64
56
40
Notes
Not reported after 1949
First reported in 1980s
Raup 1938, Tryon and Finn 1949, Friday and Friday 1985, Barringer and Clemants 2003.
the 1930s (Raup 1938). It was not reported for many years and
was not listed in the 2003 published flora, although a single individual in poor health was subsequently located (K. Barringer, pers. comm.).
Of the eleven species added to the forest tree species list
subsequent to the 1938 published flora, three (black walnut
(Juglans nigra), Norway spruce (Picea abies), and jack pine
(Pinus banksiana)) were planted. One species, silver maple
(Acer saccharinum), was on land added to the forest after 1985
and has not been found elsewhere in the forest naturally. The
remaining seven species are all thought to have invaded the
forest naturally. None of these seven is common in the forest
and their exact dates of arrival are unknown. Two of these,
tree-of-heaven (Ailanthus altissima), and white poplar
(Populus alba), are species not native to North America and
are known for their spreading, invasive nature (Alien Plant Invaders of Natural Areas, Plant Conservation Alliance, National Park Service, http://www.nps. gov/plants/alien/). The
other five new species, southern catalpa (Catalpa
bignonioides), cockspur hawthorn (Crataegus crus-galli), red
mulberry (Morus rubra), eastern cottonwood (Populus
deltoides), and slippery elm (Ulmus rubra), have range distributions with centers substantially to the south of the BRF (Little 1971, 1977).
On the eight long-term plots, red oak and chestnut oak trees
together comprised 64% of the canopy trees in 1936 and 70%
of the canopy trees in 2006 (Figure 2). Chestnut oak was the
most common species in the canopy in 1936 but was replaced
by red oak as the most common canopy tree by 2006. Red maple decreased from 10% of canopy trees in 1936 to 4% in
2006. Only four of 43 red maple canopy trees survived the period and only two of 102 understory red maples in 1936 had
become canopy trees by 2006. Sugar maple increased from 5%
of canopy trees in 1936 to 9% in 2006. Black birch and yellow
birch both increased slightly from 4 to 7% of canopy trees.
Basswood and shade-intolerant species such as gray birch
were eliminated from the canopy by 2006.
Changes were greater in the understory than in the canopy
of the long-term plots (Figure 2). Oak trees decreased from
37% of the understory trees in 1936 to 13% in 2006. Red maple increased from 27% of understory trees in 1936 to 41% in
2006, whereas sugar maple decreased from 13 to 9%. Black
birch increased from fifth most common understory tree in
1936 to second most common in 2006 and yellow birch also
increased. Basswood and shade-intolerant species such as gray
birch were eliminated from the understory by 2006, as also occurred in the canopy. Black gum (Nyssa sylvatica) tripled from
2% of the understory in 1936 to 6% by 2006, whereas a group
of other shade-tolerant understory trees including shadbush
(Amelanchier canadensis) and striped maple (Acer pennsylvanicum) increased to total 9% of the understory by 2006.
Density and basal area
On the eight long-term plots, mean tree density decreased
from about 1500 trees per hectare in 1936 to 735 tree per hectare in 2006 (Figure 3). Density steadily declined to about
950 trees per hectare by the mid 1960s and remained near that
value until the early 1980s when it began declining again. In
the 1930s, there was substantial among-plot variation in density because the youngest plots had a great many trees,
whereas the two other plots had been thinned. This variance
decreased steadily until 1978 and then increased as some plots
experienced new recruitment whereas others continued to thin.
Since 1988, variance in density has continued to decrease
along with absolute density.
Mean basal area on the long-term plots doubled from less
than 15 m2 ha –1 in 1936 to 29 m2 ha –1 in 1998 (Figure 3). Basal
area increment was low between 1946 and 1954, between
1961 and 1971, and between 1979 and 1986, and was negative
from the end of the 1981 growing season through the end of
1984. After peaking in 1998, mean basal area decreased by
about 10% to a low of 26 m2 ha –1 in 2005, rebounding slightly
in 2006. Mean maximum annual basal area increment was
0.52 m2 ha –1 year –1, whereas the mean basal area increment
over the entire 70-year period was 0.17 m2 ha –1 year –1.
Aboveground biomass
Forest-wide AGB based on Black Rock Forest inventory data
TREE PHYSIOLOGY VOLUME 28, 2008
BIOMASS CHANGES IN THE BLACK ROCK FOREST
543
Figure 2. (A) Canopy and (B) understory
tree species composition on eight
long-term plots in the Black Rock Forest
in 1936 (open bars) and 2006 (filled bars).
averaged 71.4 ± 3.4 Mg ha –1 in 1930, ranging from little more
than zero in young stands to as much as 161 Mg ha –1 (Figure 4). Stand age at the time ranged from 20 to 80 years and
wood volume ranged from 0 to 168 m3 ha –1 (Tryon 1930). In
Figure 3. Mean (A) tree density and (B) basal area in eight long-term
plots in the Black Rock Forest from 1936 to 2006, with 95% confidence intervals.
1985, the second forest-wide inventory documented mean
AGB of 144.9 ± 5.1 Mg ha –1. Although not directly comparable because of different methods and sample areas, these data
indicate a long-term AGB accumulation rate of roughly
1.3 Mg ha –1 year –1 over the intervening 55 years.
Aboveground biomass on the long-term plots increased
from 75 Mg ha –1 in 1936 to 183 Mg ha –1 in 1985, for a mean
Figure 4. Forest-wide aboveground live tree biomass (AGB) as estimated by 1930 and 1985 inventories (䉲), mean AGB from a subset of
45 inventory plots in 1985 and 2000 (䊏) and mean AGB on a series of
eight long-term plots in the Black Rock Forest from 1936 to 2006 (䊉),
with 95% confidence intervals.
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544
SCHUSTER, GRIFFIN, ROTH, TURNBULL, WHITEHEAD AND TISSUE
Figure 5. Mean estimated aboveground biomass per tree for all trees
of major canopy tree species between 10 and 20 cm diameter at breast
height in 1936 that persisted on eight long-term plots in the Black
Rock Forest through 2006.
AGB accumulation rate of 2.2 Mg ha –1 year –1. The long-term
plots had higher mean AGB in 1985 compared with the forest-wide mean from the 1985 inventory (183.4 versus
144.9 Mg ha –1 ), although the 95% confidence intervals overlapped (Figure 4). The 45 plots from the 1985 inventory that
were measured in 2000 had higher AGB than the forest-wide
mean because these plots included more than a dozen mature
stands of eastern hemlock and other stands growing on deeper
soils derived from thick deposits of glacial till (Denny 1938).
Between 1985 and 2000, mean AGB on these 45 plots increased from 167 to 195 Mg ha –1 (1.8 Mg ha –1 year –1 ),
whereas AGB on the eight long-term plots increased from 183
to 215 Mg ha –1 (2.1 Mg ha –1 year –1 ).
Four periods of generally lower biomass increment on the
long-term plots (1946–1954, 1961–1971, 1979–1986 and
1999–2005) correspond with periods of decreased basal area
increment, because both were calculated from dbh measurements (Figure 4). The most recent decline in AGB, which occurred from the end of the 1998 growing season through the
end of the 2005 growing season, represented a loss of 9.1% in
total live aboveground tree biomass.
Among canopy trees with diameters between 10 and 20 cm
in the 1930s that survived to 2006, red oaks increased in
aboveground biomass at more than twice the mean rate for
trees of all other species (Figure 5). Red oak’s proportion of
total aboveground biomass, averaged across all plots, increased from 27 to 45% over this period. Chestnut oak growth
rates were less than those for red oak, but were still greater
than for other species. Yellow birch and black birch and red
maple and sugar maple canopy trees alive for the entire period
all averaged about 400 kg dry aboveground biomass in 2006.
Discussion
Changes in tree species composition and structure
A notable feature of the changes in the tree species list for the
BRF between the 1930s and the late 1990s is the impact of introduced species: one of the three species extirpated from the
forest, American elm, was eliminated by an introduced fungus, and two of the seven tree species that moved into the forest are introduced plants widely considered to be invasive
(tree-of-heaven and white poplar). Barringer and Clemants
(2003) reported that 20% of the modern flora of the BRF are
introduced species, a proportion that has increased over time.
From an examination of floristic records, Robinson et al.
(1994) documented that the proportion of introduced species
in the flora of Staten Island, New York increased from 19% to
more than 33% between 1879 and 1991. A second notable feature of the change in tree species list is that the pattern consistent with predictions about the way tree species ranges will
change as a result of climate warming (Iverson et al. 1999,
2007). Two of the three tree species extirpated from the forest
(black spruce and paper birch) are northern species with southern range margins near the BRF, and the five native tree species that have moved into the forest (southern catalpa, cockspur hawthorn, red mulberry, eastern cottonwood, and slippery
elm) all have ranges predominantly to the south of the BRF
(Little 1971, 1977). Nearby meteorological stations recorded
a mean air temperature increase of 1.0 °C over the 20th century, mainly through fall and winter warming and increased
summer minimum daily temperatures (Warrach et al. 2006).
Thus aside from plantings of Norway spruce and black walnut,
the spread of introduced species and climate warming appear
responsible for all of the changes in forest species richness
over six decades.
The eight remaining long-term plots are generally typical of
the forest in composition and aboveground biomass (Figure 4)
and provide more detail on compositional and structural
changes since 1936. The aggrading character of the BRF is
clearly indicated by the doubling of basal area on the
long-term plots from less than 15 m2 ha –1 in the 1930s to
nearly 30 m2 ha –1 by 2000. Mean basal area of another upland
oak forest in south-central New York doubled to 32 m2 ha –1 in
a shorter period from 1935 to 1985 (Fain et al. 1994). Mortality was substantial in the BRF during much of this period as a
result of natural thinning and was exacerbated after periods of
drought (Tryon and Finn 1949, Karnig and Lyford 1968). During the most recent period of mortality (1999–2005), basal
area of the long-term plots was reduced by an average of 10%.
Mortality has been greatest on the initially densest plots and all
plots are now converging on a density of about 700 stems
(> 2.5 cm dbh) per hectare.
The forest canopy of these plots underwent relatively minor
compositional change between 1936 and 2006, with red oak
and chestnut oak remaining dominant, followed by maples and
birches. Canopy red oak and chestnut oak in the BRF have
high photosynthetic capacities, water-use efficiencies and
photosynthetic nitrogen-use efficiencies on both dry and wet
sites that likely contribute to their long-term persistence
(Turnbull et al. 2001, 2002). In contrast, the forest understory
changed dramatically between 1936 and 2006, with all oak
species decreasing, red maple increasing, and black birch becoming the second most common tree in the understory.
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BIOMASS CHANGES IN THE BLACK ROCK FOREST
Larger compositional changes in the understory compared
with the canopy have been documented between 1965 and
2004 in a nearby old-growth stand of eastern hemlock (Weckel
et al. 2006). These changes portend possible major future
turnover in the composition of regional forest canopies.
The lack of successful oak reproduction in the understory is
a widespread phenomenon (e.g., Abrams 1992, Lorimer 1994,
Drury and Runkle 2006) due in part to the relatively poor
physiological performance of Quercus species in shade. Oaks
have a host of adaptations to drought (waxy xeromorphic
leaves, low water potential for stomatal closure, high
photosynthetic rates in dry conditions) and disturbances such
as fire (thick bark, deep roots, resprouting ability) that adapt
them to periodic disturbance (Abrams 1992, Johnson 1993,
Turnbull et al. 2001, 2002). In the absence of fires because of
the practice of fire suppression, oak seedlings and saplings are
often out-competed by thin-barked, fire-sensitive tree species
(Chapman et al. 2006, Drury and Runkle 2006). Although
acorns and oak seedlings are common around BRF, but even
where disturbances have occurred, oak seedlings have rarely
grown into young trees because of deer browsing(BRF data,
also see Horsley et al. 2003).
Young red maple trees have become established more abundantly than any other species in BRF. Nagel et al. (2002)
showed that red maples in the BRF have lower energy and resource requirements for leaf construction and maintenance
than co-occurring oaks, which may promote establishment
and persistence in low light. Lorimer (1984) proposed that maples may become canopy trees in eastern deciduous forests
formerly dominated by oaks. However, on the BRF long-term
plots red maple has largely failed to occupy or persist well in
the canopy. The ecophysiological properties of real maple
place it at a disadvantage on dry sites compared with red oak
and chestnut oak (Turnbull et al. 2001, 2002). Tree-ring studies of red maples in the Harvard Forest have revealed reduced
growth increments since 1992 (Pederson 2005). In the BRF,
the more shade-tolerant, gap-facultative (sensu Orwig and
Abrams 1994) sugar maple has made the understory–canopy
transition more often than red maple. Black birch and yellow
birch have both increased in the understory as well as in the
canopy.
The reason for the loss of moderately shade-tolerant basswood from these plots is unclear. The species exhibited steady
decline and was nearly eliminated from the long-term plots by
1970. Basswood was also lost from a nearby old-growth eastern hemlock forest between 1965 and 2004 (Weckel et al.
2006) but has maintained its importance in some other
long-term plots (Fain et al. 1994, Woods 2000).
Changes in forest biomass
In general, the aboveground biomass patterns on the BRF
long-term plots are consistent with forest-wide estimates from
periodic forest inventories. These results document that aboveground biomass increased by a factor of nearly three over the
past 76 years. Despite fluctuations in productivity, the BRF apparently functioned as a C sink for most of the 20th century, although the net change also includes unmeasured changes in
545
belowground carbon pools. This conclusion may extend to
many forested mountainous areas in the eastern United States
(Liu et al. 2006).
Most forest stands in the BRF have current live aboveground biomass between 150 and 250 Mg ha –1, typical for mature eastern North American deciduous forests. Jenkins et al.
(2001) analyzed the USDA Forest Service Forest Inventory
and Analysis (FIA) database for the mid-Atlantic region, and
reported a mean AGB of 199 Mg ha –1 for mature (above 20 m2
ha –1 basal area), closed-canopy oak–hickory forests that had
not experienced recent losses as a result of logging, fire, disease or insects. The AGB varies substantially across the BRF
because of differences in age, growth rates, disturbance and
management practices, and some stands have AGB values
greater than 300 Mg ha –1, which is within the range reported
for old-growth eastern hardwood forests (220 to 330 Mg ha –1;
Jenkins et al. 2001). For comparison, the hardwood forest at
the Harvard Forest eddy-covariance tower site had a total AGB
of 200 Mg ha –1 in 2000 (Barford et al. 2001) and older hemlock–white pine stands in the Harvard Forest had a mean total
AGB of 320 Mg ha –1 (J. Hadley, pers. comm.). The experimental forest at Hubbard Brook, New Hampshire, located 340
km northeast of the BRF, had a mean AGB of 162 Mg ha –1 in
low-elevation northern hardwood stands in the early 1970s and
197.8 Mg ha –1 in one control watershed in 1992 (Likens et al.
1994).
Given that deciduous tree biomass is composed of about
50% C (0.498 g C g –1 wood; Birdsey 1992), estimated aboveground C in live trees in the BRF averaged 35.7 Mg C ha –1 in
1930, 72.5 Mg C ha –1 in 1985 and about 95 Mg C ha –1 in 2000,
with some stands containing more than 150 Mg C ha –1.
Belowground C stores, however, generally dominate ecosystem C (Dixon et al. 1994). Because stump and coarse root biomass averages about 20% of total tree biomass for deciduous
trees in the region (Jenkins et al. 2001), estimated mean total
live tree carbon in the BRF increased from 45 Mg C ha –1 in
1930 to 91 Mg C ha –1 in 1985 to 119 Mg C ha –1 in 2000. Adding an average of 5% for C in coarse woody debris from surveys on the long-term plots (data not shown) and an estimate
of soil organic C of 50 Mg C ha –1 for this region (USGCRP
2000) yields a rough average of 175 Mg C ha –1 for total ecosystem C in 2000, though the actual relationship between
above- and belowground C in this ecosystem is not known. For
comparison, Birdsey and Heath (1995) estimated that total
ecosystem C across all northeastern USA timberlands averaged 217.5 Mg C ha –1.
Based on the known age of the stands where the BRF
long-term plots are located, annual AGB increments after
clearcutting up to the time of plot establishment in the 1930s
averaged 2.1 Mg ha –1 year –1. The period of highest long-term
plot AGB increment was from after plot establishment in the
1930s to the early to mid 1960s, with a mean rate of 2.9 Mg
ha –1 year –1 (range 2.1–4.2 Mg ha –1 year –1 ). Increases in biomass were primarily a result of the growth of surviving trees,
because recruitment was minimal during this period.
Eddy-covariance studies in the Harvard Forest recorded a
mean net ecosystem uptake of 2.0 Mg C ha –1 year –1 from the
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546
SCHUSTER, GRIFFIN, ROTH, TURNBULL, WHITEHEAD AND TISSUE
atmosphere since the early 1990s, with aboveground woody
growth accounting for just over half of the C uptake (Barford
et al. 2001). This indicates roughly similar annual aboveground carbon sequestration in the BRF and the Harvard Forest during the 1990s. However, since the end of the 1998 growing season, mean annual AGB increment in the BRF has been
–2.1 Mg ha –1 year –1.
The temporal pattern of biomass change on the long-term
plots is characterized by a generally steady increase interrupted by four periods of slower or even negative biomass increment (1946–1954, 1961–1971, 1979–1986 and 1999–
2005). Although inconclusive, the timing of this pattern suggests some of the most important factors controlling long-term
biomass and C dynamics. The first and third periods of lower
biomass increment on the long-term plots were almost certainly caused by insect outbreaks. In the late 1940s–early
1950s, the golden oak scale (Asterolecanium variolosum
(Ratzeburg)) attacked oak trees, particularly chestnut oak
(BRF records). It may have been responsible for the mortality
of chestnut oak trees between 1946 and 1948 and generally
slower oak growth from 1945 to 1954. The apparent length of
this period of reduced growth in the long-term plots, like those
in the 1960s and 1980s, is in part an artifact due to infrequent
measurement. The BRF chestnut oak tree ring chronologies
pinpoint a period of low growth between 1947 and 1949
(D’Arrigo et al. 2001). In the early 1980s, outbreaks of the introduced gypsy moth (Lymantria dispar L.) caterpillar caused
significant growth reductions and direct mortality of chestnut
oak and other tree species, resulting in a slight decline in AGB
on the long-term plots. Nearly every tree in the forest was defoliated in 1981 and the impact was so large that 1981 has become a marker year for BRF tree-ring studies (D’Arrigo et al.
2001). In the Harvard Forest, extensive gypsy moth defoliation
in 1981 resulted in smaller tree ring widths of red oak and red
maple trees than during any previous drought (Pederson
2005). Although absent in long-term plots, reductions in eastern hemlock growth and up to 50% hemlock mortality have
occurred in BRF stands since the introduced hemlock wooly
adelgid (Adelges tsugae Annand) was first noted in the forest
in 1992 (Kimple and Schuster 2002). Before the first records
used in this study, the decimation of American chestnut
(Castanea dentata (Marsh.) Borkh.) by the chestnut blight
(Cryphonectria parasitica (Murrill) Barr) between 1915 and
1918 caused a substantial loss of tree biomass in the BRF, but
was followed by increased growth of the remaining trees
(Stout 1956).
The second and fourth periods of reduced biomass increment coincided with severe droughts. Environmental factors
such as precipitation and temperature substantially control annual C flux in some ecosystems (e.g., Goulden et al. 1996,
Braswell et al. 1997, White et al. 1999). Our results indicate
that the regional drought lasting from 1962 to 1966 resulted in
slow aboveground tree growth on the long-term plots, a pattern
manifest in annual ring widths of eastern hemlock and chestnut oak (D’Arrigo et al. 2001). The concomitant high mortality, especially of scarlet oak and chestnut oak (Karnig and
Lyford 1968), had a large impact on stand biomass, and three
of the eight long-term plots lost AGB during this period.
The most recent period of AGB reduction on the long-term
plots, from 1999 to 2005, can be characterized in detail because of the initiation of annual measurements in 1994.
Growth was poor overall in the years 1997, 1999 and 2004,
and 1999–2001 and 2003–2005 were periods of high mortality for both canopy and understory trees of chestnut oak, red
oak, and red maple. Increased mortality simultaneously occurred on other dry and upper-slope sites in the BRF and surrounding Highlands. Although a detailed analysis is beyond
the scope of this paper, the dry years of 1995, 1999 and
2001–2002 (BRF data) may well be underlying factors.
Previous land use, i.e., extensive forest clearing, is a major
reason for the overall pattern of carbon sequestration and biomass uptake in the BRF. As the forest stands have aged, there is
some evidence of decreasing annual forest carbon storage (i.e.,
lower mean biomass increment after the 1960s) but some
stands more than 100 years old have exhibited undiminished
AGB increments. Extrapolations indicate that ABG increments during the initial 25–45 years after clearcutting of the
stands around the long-term plots were similar to long-term
mean rates of about 2.2 Mg ha –1 year –1. The most rapid accumulation of AGB on most plots occurred between the early
1930s and the early 1960s, when stand ages increased from
25–45 years to 55–75 years. The period of lowest overall
net biomass gain then ensued from the early 1960s to the
mid-1980s, but this may have been due more to the external biotic and abiotic factors than to increasing stand maturity. Human impact on forest biomass via tree removal may be a factor
in the relatively low forest-wide AGB increment of 1.3 Mg
ha –1 year –1 between the forest inventories of 1930 and 1985.
Thinning initially lowers AGB but has been shown to increase
increment growth in the remaining trees for a decade or more
(Karnig and Stout 1969). However, several timber harvests between 1930 and 1985 extracted substantial tree biomass (BRF
data).
Species composition has significantly influenced long-term
AGB trends and C storage in the BRF. Red oak canopy trees on
the long-term plots in the 1930s gained biomass throughout
the period to 2006 at more than twice the rate of red maple,
sugar maple, black birch and yellow birch trees of similar initial size. American chestnut was a dominant in the forest but
was eliminated by 1918 (Stout 1956), with red oak being the
major beneficiary. In some other areas, tree-ring analyses have
shown increased red oak growth following chestnut demise
(Pederson 2005). Increasing red oak growth rates around
much of New England have been documented after 1950, especially in older populations (Pederson 2005).
We do not believe the large biomass increases for red oak
and chestnut oak trees are artifacts of the allometric equations
because the match with BRF data on size–biomass relationships was the primary reason for our selection of these equations, and the Brenneman et al. (1978) equations were developed in locations with similar site conditions and species composition. The equations also project less red oak aboveground
TREE PHYSIOLOGY VOLUME 28, 2008
BIOMASS CHANGES IN THE BLACK ROCK FOREST
biomass at given diameters compared with most of the other
species. These patterns of greater biomass gain by red oak and
chestnut oak compared with other species are consistent with
studies of the photosynthetic and respiratory responses of
these dominant BRF tree species to environmental variation
(Turnbull et al. 2001, 2002). Lorimer (1981), Abrams (1998)
and others have previously documented widespread expansion
of red maple, which now dominates the understory and midcanopy of many eastern deciduous forests. However, although
common in the BRF, red maple has contributed little to
long-term biomass and C accumulation. If the current oak-dominated canopy in the BRF and the surrounding Highlands
Region is eventually replaced by the understory dominants red
maple and black birch, the change will likely be accompanied
by a new release of C.
Our results show that mortality has been a controlling factor
in long-term aboveground biomass patterns in the BRF, episodically having a greater influence than interannual variation
in wood biomass increment. Understory trees are always at
high risk of mortality, but on the long-term plots some canopy
red oak and chestnut oak trees with reasonable growth rates
have also died since 1999, significantly reducing AGB.
Jenkins and Pallardy (1995) reported that oak mortality was
greater for trees that grew quickly for many years prior to a
drought compared with those that grew more slowly. The observed reductions in live biomass have substantially augmented the BRF detrital carbon pool. Carbon in coarse woody
debris on the long-term plots was inventoried after the 2002
growing season and ranged from 5 to 33 Mg ha –1, averaging
about 5% of the total aboveground carbon on the plots.
Other environmental factors with potential for substantial
effects on forest carbon and stand dynamics include increasing
atmospheric CO2 concentration, acid precipitation, anthropogenic nitrogen deposition and increasing temperatures (Medlyn et al. 2000, McKenzie et al. 2001, Aber et al. 2001, Driscoll
et al. 2001, Hyvönen et al. 2007). For example, the BRF currently receives high annual input of N from the atmosphere,
with a mean wet deposition rate of 6–7 kg N ha –1 year –1 (National Acid Deposition Program 2002). Acid deposition may
contribute to sugar maple dieback and reductions in forest productivity through depletion of base cations in soils (Likens et
al. 1996, Long et al. 1997, Driscoll et al. 2001). But the effects
of this and the other factors listed above can be difficult to determine. Further study of these factors may reveal their importance, singly and in combination, in forest C dynamics.
Implications and additional research
The results of this study are consistent with other evidence indicating that North American temperate forests have functioned as significant C sinks for many decades. Black Rock
Forest and other forests in the Highlands Province are still capable of sequestering carbon, especially in areas with canopies
dominated by red oak and chestnut oak. But recent changes in
stand structure, disturbance regimes, and influxes of new tree
species combined with recent canopy mortality indicate the
potential for substantial forest change in the future. Increased
547
red maple and black birch in the understory, and tree-ofheaven (an introduced species with extremely rapid growth
and abundant seed production (Knapp and Canham 2000)) in
gaps, could result in a future forest that would likely store less
carbon and provide a reduced set of ecosystem services
compared with the current oak forest.
Our understanding of multiple environmental stress interactions at the forest level remains limited despite its global importance. Long-term monitoring studies of oak forests such as
BRF are rare (Chapman et al. 2006) but such databases are
critical to develop baselines for validating models and predicting future scenarios (Aber et al. 2001). Long-term studies can
provide a basis for understanding the influences of multiple
controlling factors, but they need to be large in scope and
duration. Studies to quantify belowground changes in roots
and soil organic matter are needed to more accurately track
ecosystem-level biomass and carbon dynamics.
Acknowledgments
We acknowledge Hal Tryon and Hugh Raup for the original groundwork that enabled this study. We thank Ben Stout and Jack Karnig for
continuing the forest measurements between 1949 and 1992 and
Kathleen and James Friday for conducting the 1985 inventory. John
Brady, Aaron Kimple, Matthew Munson, John Canella, Michael
White, Kristina Kipping, Sarah Helm, Kathy DeWitt, Joy Felio and
Emma Hoyt all accomplished important fieldwork and most assisted
with data management. We thank Frances Schuster for map making,
NIGEC, Barnard College and the Pew Fellowship program for supporting student fellows, and J.T. Mates-Muchin and J.D. Lewis for
sharing data. We thank Ernest G. Stillman and William T. Golden for
their long-term support of science in the Black Rock Forest. This
work was funded in part by the Andrew W. Mellon Foundation
(KLG), the Hughes Science Pipeline Project (HR) and the Foundation for Research, Science and Technology, New Zealand (DW).
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