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August 2004 / Vol. 54 No. 8 • BioScience 755

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The role of predation is of major importance to conservationists as the ranges of large carnivores continue
to collapse around the world. In North America, for exam-
ple, the gray wolf (Canis lupus) and the grizzly bear (Ursus
arctos) have respectively lost 53% and 42% of their histori

c

range, with nearly complete extirpation in the contiguous 48
United States (Laliberte and Ripple 2004). Reintroduction of
these and other large carnivores is the subject of intense sci-
entific and political debate, as growing evidence points to the
importance of conserving these animals because they have cas-
cading effects on lower trophic levels. Recent research has
shown how reintroduced predators such as wolves can in-
fluence herbivore prey communities (ungulates) through di-
rect predation, provide a year-round source of food for
scavengers, and reduce populations of mesocarnivores such
as coyotes (Canis latrans) (Smith et al. 2003). In addition, veg-
etation communities can be profoundly altered by herbi-
vores when top predators are removed from ecosystems, as

a

result of effects that cascade through successively lower trophic
levels (Estes et al. 2001). The absence of highly interactive car-
nivore species such as wolves can thus lead to simplified or
degraded ecosystems (Soulé et al. 2003). A similar point was
made more than 50 years ago by Aldo Leopold (1949): “Since
then I have lived to see state after state extirpate its wolves….

I have seen every edible bush and seedling browsed, first to
anemic desuetude, and then to death” (p. 139).

Increased ungulate herbivory can affect vegetation struc-
ture, succession, productivity, species composition, and di-
versity as well as habitat quality for other fauna. Although the
topic remains contentious, a substantial body of evidence in-
dicates that predation by top carnivores is pivotal in the
maintenance of biodiversity. Most studies of these carnivores
have emphasized their lethal effects (Terborgh et al. 1999).
Here our focus is on how nonlethal consequences of preda-
tion (predation risk) affect the structure and function of
ecosystems. The objectives of this article are twofold: (1) to
provide a brief synthesis of potential ecosystem responses to
predation risk in a three-level trophic cascade involving large
carnivores (primarily wolves), ungulates, and vegetation; an

d

(2) to present research results that center on wolves, elk
(Cervus elaphus), and woody browse species in the northern
range of Yellowstone National Park (YNP).

William J. Ripple (e-mail: bill.ripple@oregonstate.edu) is a professor and di-

rector of the Environmental Remote Sensing Applications Laboratory, and

Robert L. Beschta (e-mail: robert.beschta@oregonstate.edu) is a professor

emeritus, in the College of Forestry, Oregon State University, Corvallis, OR

97331. © 2004 American Institute of Biological Sciences.

Wolves and the Ecology of Fear:
Can Predation Risk Structure
Ecosystems?

WILLIAM J. RIPPLE AND ROBERT L. BESCHTA

We investigated how large carnivores, herbivores, and plants may be linked to the maintenance of native species biodiversity through trophic
cascades. The extirpation of wolves (Canis lupus) from Yellowstone National Park in the mid-1920s and their reintroduction in 1995 provided the
opportunity to examine the cascading effects of carnivore–herbivore interactions on woody browse species, as well as ecological responses involving
riparian functions, beaver (Castor canadensis) populations, and general food webs. Our results indicate that predation risk may have profound
effects on the structure of ecosystems and is an important constituent of native biodiversity. Our conclusions are based on theory involving trophic
cascades, predation risk, and optimal foraging; on the research literature; and on our own recent studies in Yellowstone National Park. Additional
research is needed to understand how the lethal effects of predation interact with its nonlethal effects to structure ecosystems.

Keywords: wolves, ungulates, woody browse species, trophic cascades, predation risk

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Trophic cascades
A trophic cascade is the “progression of indirect effects by
predators across successively lower trophic levels” (Estes et al.
2001). In terrestrial ecosystems, top-down and bottom-up
effects can occur simultaneously, although their relative
strength varies, and interactions among trophic levels can be
complex. Here we study top-down processes and associated
trophic interactions that potentially have broad ecosystem ef-
fects. Although our main purpose is to explore nonlethal ef-
fects on ecosystems, we first describe several studies that
emphasize the importance of cascading lethal effects.

Predators obviously can influence the size of prey species
populations through direct mortality, which, in turn, can in-
fluence total foraging pressure on specific plant species or en-
tire plant communities. For example, at the continental scale,
Messier (1994) examined 27 studies of wolf–moose (Alces al-
ces) interactions and generally found that wolf predation
limited moose numbers to low densities (< 0.1 to 1.3 moose per square kilometer [km2], excluding Isle Royale studies), which resulted in low browsing levels in northern North America, especially in areas where wolves and bears both prey on moose. Comparing total deer (family Cervidae) bio- mass in areas of North America with and without wolves, Crête (1999) suggested that the extirpation of wolves and other predators has resulted in unprecedentedly high browsing pressure on plants in areas of the continent where wolves have disappeared.

On a smaller scale, islands provide settings for studying
predator–prey population dynamics. For example, McLaren
and Peterson (1994) studied relationships between wolves,
moose, and balsam fir (Abies balsamea) in the food chain on
Michigan’s Isle Royale. As a result of suppression by moose
herbivory, young balsam fir on Isle Royale showed depressed
growth rates when wolves were rare and moose densities
were high. McLaren and Peterson concluded that the Isle
Royale food chain appeared to be dominated by top-down
control in which predation determined herbivore density
through direct mortality and hence affected plant growth
rates. Terborgh and colleagues (2001) studied forested hilltops
in Venezuela that were isolated by the impounded water of a
large reservoir. When predators disappeared from the
islands, the number of herbivores increased, and the repro-
duction of canopy trees was suppressed because of increased
herbivory in a manner consistent with a top-down theory. On
the islands without predators, Terborgh and colleagues found
few species of saplings represented because of a lack of re-
cruitment, even though many more species of trees made up
the overstory.

Changes in prey behavior due to the presence of predators
are referred to as nonlethal effects or predation risk effects
(Lima 1998). These behavioral changes reflect the need for her-
bivores to balance demands for food and safety, as described
by optimal foraging theory (MacArthur and Pianka 1966).
They include changes in herbivores’ use of space (habitat
preferences, foraging patterns within a given habitat, or
both) caused by fear of predation (Lima and Dill 1990). Such

behaviorally mediated trophic cascades set the foundation for
an “ecology of fear” concept (Brown et al. 1999) and provide
the basis for this study. Ecologists are now beginning to ap-
preciate how predators can affect prey species’ behavior,
which in turn can influence other elements of the food we

b

and produce effects of the same order of magnitude as those
resulting from changes in predator or prey populations
(Werner and Peacor 2003). Interestingly, Schmitz and
colleagues (1997) indicate that the effects of predators on the
behavior of prey species may be more important than direct
mortality in shaping patterns of herbivory.

Predation risk can also have population consequences for
prey by increasing mortality, according to the “predation-
sensitive food” hypothesis (Sinclair and Arcese 1995). This
hypothesis states that predation risk and forage availability
jointly limit prey population size, because as food becomes
more limiting, prey take greater risks to forage and are more
likely to be killed by predators as they occupy riskier sites.
Wolves have been largely absent from most of the United States
for many decades; hence, little information exists on how adap-
tive shifts in ungulate behavior caused by the absence or
presence of wolves might be reflected in the composition
and structure of plant communities.

Prey and plant refugia. Prey refugia are areas occupied by prey
that potentially minimize their rate of encounter with preda-
tors (Taylor 1984). For example, in a wolf–ungulate system,
ungulates may seek refuge by migrating to areas outside the
core territories of wolves (migration) or survive longer out-
side the wolves’ core use areas (mortality) (Mech 1977). The
relative contributions of migration versus mortality in these
ecosystems remain unclear. However, both of these processes
can result in low populations of ungulates in the wolves’ core
use areas and travel corridors, thus creating potential “plant
refugia” by lowering herbivory in areas with high wolf den-
sities (Ripple et al. 2001).

Predation risk effects involving wolves and elk were reflected
in aspen (Populus tremuloides) growth in Jasper National
Park. White and colleagues (1998) reported new aspen growth
(trees 3 to 5 meters [m] tall) following the recolonization of
wolves in the park, with particularly vigorous regeneration in
areas of high predation risk (i.e., near wolf trails). The pop-
ulation dynamics of moose in the presence and absence of
wolves was studied in Quebec by Crête and Manseau (1996).
They found moose densities seven times greater in a region
without wolves compared with the moose–wolf region. In
Grand Teton National Park, Berger and colleagues (2001)
found that the loss of both wolves and grizzly bears allowed
an increase in moose density within the park, followed by an
increase in moose herbivory on willows (Salix spp.).

Historically, aboriginal human hunters in North America
affected the distribution of ungulate species (Kay 1994).
Laliberte and Ripple (2003) used the journals of Lewis and
Clark to assess the influence of aboriginal humans on wildlife
distribution and abundance. They found that areas with
greater human population density had lower species

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diversity and abundance of both large carnivores and ungu-
lates. In today’s ecosystems, in which humans have elimi-
nated large carnivores, predation risk effects may occur
because of human sport hunting; both prey and plant refu-
gia have been documented where elk are hunted by humans.
For example, in Montana, St. John (1995) concluded that elk
adjusted their foraging behavior by browsing far from roads
to avoid human contact and possible predation. As a result,
aspen stands within 500 m of roads were browsed by elk less
than stands farther away. In Colorado, McCain and col-
leagues (2003) found that aspen was heavily browsed and used
year-round by elk on land where sport hunting was excluded.
In surrounding national forest land where hunting was al-
lowed, aspen stands were minimally browsed. In national
parks where both recreational hunting and large carnivores
have been removed, dramatic changes in mammal and plant
populations have been described (White et al. 1998, Soulé et
al. 2003).

Terrain fear factor. The “terrain fear factor” (Ripple and
Beschta 2003) represents a conceptual model for assessing the
relative predation risk effects associated with encounter sit-
uations. This concept indicates that prey species will alter their
use of space and their foraging patterns according to the
features of the terrain and the extent to which these features
affect risk of predation (e.g., avoid sites with high predation
risk; forage or browse less intensively at high-risk sites). On
landscapes with both open and closed habitat structure,
ungulates may use a strategy of hiding in forest cover to
lower predator encounter rates, or they may seek open terrain
to see predators from afar (Kie 1999). In the latter scenario,
the relative level of predation risk at a given site is influenced
both by the probability of a prey animal detecting a preda-
tor (i.e., visibility) and by the probability of the prey escap-
ing if attacked. For example, Risenhoover and Bailey (1985)
found that bighorn sheep (Ovis canadensis) preferred open
habitats and avoided habitats in which vegetation obstructed
visibility. When sheep occasionally used high-risk habitats with
poor visibility, they moved more while foraging, and forage
intake per step was lower than for habitats with good visibility.
Even when high-quality forage occurred at low elevations,
Festa-Bianchet (1988) found that pregnant bighorn sheep
moved away from predators to higher elevations with low-
quality forage.

Altendorf and colleagues (2001) concluded that mule deer
(Odocoileus hemionus) responded to predation risk by bias-
ing their feeding efforts at the scale of both microhabitats
and habitats; the perceived predation risk was lower in open
areas than in forested areas. This matches well with the find-
ings of Kunkel and Pletscher (2001), who found that wolves
were most successful when they could closely approach un-
gulates without detection and that the element of surprise
appeared to be an important factor in their predation success.
Kolter and colleagues (1994) suggested that ibex (Capra ibex)
reduce their predation risk by foraging most often near
“escape terrain” of extremely steep slopes or cliffs, which are

difficult or impossible for wolves and other predators to
negotiate. Caribou (Rangifer tarandus) move to higher ele-
vations to increase the distance between themselves and
wolves traveling in valley bottoms (Bergerud and Page 1987).

All of the behavior changes identified above have the
potential to influence plant composition and structure by
creating local plant refugia at sites whose terrain and landscape
characteristics result in high levels of predation risk. These
refugia typically have a lower percentage of plants browsed or
a smaller amount of the current year’s growth removed by
ungulates than low-risk sites. Furthermore, since factors
affecting predation risk probably occur at specific sites, habi-
tat patches, and other terrain features across larger land-
scapes, ungulates most likely assess predation risk at multiple
spatial scales (Kie 1999, Kunkel and Pletscher 2001).

Predation risk in a dynamic environment. Environmental
variables that may influence the degree of predation risk in-
clude winter weather, wildfire, and the depth and spatial dis-
tribution of snowpacks. Snowpack conditions can greatly
influence ungulates’ access to vegetation (both herbaceous and
woody species) and thus their starvation rates. Variations in
snow depth can also affect the ability of ungulates to escape
predators (Crête and Manseau 1996). For example, wolves
have been found to have higher ungulate kill rates when
snow is deep compared with times when snow is shallow
(Huggard 1993, Smith et al. 2003). Similarly, winter snowpack
accumulation can affect the relationship between wolves,
moose, and vegetation. In years that produced deep snow
cover, moose predation increased and browsing on firs de-
creased, affecting both plant litter production and nutrient dy-
namics (Post et al. 1999). Large snowpack accumulations in
broken terrain may preclude elk foraging and affect herd
distributions, whereas more open landscapes offer opportu-
nities for snow to melt or blow away from foraging areas. Such
open areas also offer good visibility and provide escape ter-
rain with little snow to slow ungulates fleeing from predators.
In mountainous terrain, winters with little snowfall may al-
low ungulates to remain at higher elevations, thus resulting
in reduced levels of browsing on woody species in valley bot-
toms. Conversely, high-snowfall winters are likely to increase
browsing pressure on low-elevation plant communities.

When wildfire resets stand dynamics of upland plant com-
munities (e.g., aspen), combined changes in visibility and es-
cape potential are also likely to occur. For example, fire
typically stimulates prolific aspen suckering and the growth
of dense aspen thickets, reducing visibility and browsing
rates and increasing predation risk, and thus promoting even
more aspen growth and less visibility (Ripple and Larsen
2000, White et al. 2003). When fire leaves behind coarse
woody debris on the ground, predation risk effects are likely
to be more pronounced if the debris serves as an escape im-
pediment (e.g., jackstrawed trees [trees that have fallen in
tangled heaps]). Thus, while both severe winter weather and
wildfire can directly influence ungulate survival through
increased or decreased forage availability, these events also

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shift the relative importance of predation risk in affecting
local and landscape-scale herbivory. Because environmental
factors related to predation risk are episodic, efforts at mod-
eling future ecosystem responses to predator–prey interactions
are likely to remain imprecise. However, in the long term, rela-
tively high ungulate populations may be reduced to lower den-
sities through periodic die-offs caused by lack of forage
(associated with deep snowpacks or extensive wildfire) in
combination with the lethal effects of predation and hunting
(NRC 2002a, Smith et al. 2003).

Ecosystem responses. Ecosystem responses to trophic cascades
can be many and complex (Estes 1996, Pace et al. 1999), but
for simplicity we focus on riparian functions and on beaver
(Castor canadensis) and bird populations. We acknowledge that
trophic cascades can affect many other aspects of ecosystem
structure and function, both abiotic and biotic, including
habitat for numerous species of vertebrates and invertebrates,
food web interactions, and nutrient cycling (Rooney and
Waller 2003).

Although riparian systems typically occupy a small pro-
portion of most landscapes, they have important ecological
functions that affect a wide range of aquatic and terrestrial or-
ganisms as well as hydrologic and geomorphic processes of
riverine systems. For example, riparian plant communities pro-
vide root strength for stabilizing stream banks and hydraulic
roughness during overbank flows, maintain hydrologic con-
nectivity between streams and floodplains, sustain carbon and
nutrient cycling, moderate the temperature of riparian and
aquatic areas, and offer habitat structure and food web sup-
port (NRC 2002b). Thus, where riparian systems are heavily
altered by excessive herbivory, as in periods of wolf extirpa-
tion, the ecological impacts on these systems and their eco-
logical functions can be severe.

Beaver play important roles in riparian and aquatic ecosys-
tems by altering hydrology, channel geomorphology, bio-
chemical pathways, and productivity (Naiman et al. 1986).
Beaver dams flood topographic depressions and floodplains,
creating more habitat for aspen and willow; hence, beaver can
control to some degree the amount of surface water available.
Beaver can also increase plant, vertebrate, and invertebrate
diversity and biomass and alter the successional dynamics of
riparian communities (Naiman et al. 1988, Pollock et al.
1995). The occurrence of predators such as wolves can have
direct consequences for beaver populations, since wolves
have been shown to frequent riparian areas, travel along
stream corridors, and prey on beaver (Allen 1979).

If the presence or absence of wolves in a riparian area has
important effects on ungulate herbivory, then these carnivores
may represent an indirect control on beaver populations.
With wolves present, ungulates may avoid some riparian
areas (Ripple and Larsen 2000, Ripple and Beschta 2003), thus
reducing herbivory on woody browse species (e.g., aspen,
willow, cottonwood) and sustaining the long-term recruitment
of these species as well as providing food for beaver.
Furthermore, risk-sensitive behavior by ungulates may

contribute to relatively high levels of aspen, willow, or cotton-
wood recruitment in portions of a riparian zone where the
capability of ungulates to detect carnivores and escape from
them is low (e.g., tributary junctions, mid-channel islands,
point bars, areas adjacent to high terraces or steep banks, deep
snow) (Ripple and Beschta 2003). Without wolves in the
ecosystem, reduced predation risk may allow ungulate her-
bivory to increase. Where such herbivory is sufficiently severe
and sustained, it may ultimately cause the loss of woody
browse species on which various riparian functions and
beaver depend.

Researchers have recently made connections between the
loss of large carnivores and decreases in avian populations. The
local extinction of grizzly bears and wolves in Grand Teton
National Park caused an increase in herbivory on willow by
moose and ultimately decreased the diversity of Neotropical
migrant birds (Berger et al. 2001). Avian species richness and
abundance were found to be inversely correlated with moose
abundance for sites in and near the park. In the absence of large
carnivores, mesocarnivore release (i.e., an overabundance of
small predators) has been implicated in the decline of bird and
small vertebrate populations throughout North America
(Crooks and Soulé 1999).

The Yellowstone experiment
In the discussion below of recent research results from YNP,
we describe the northern winter range ecosystem, historical
predator–prey–vegetation dynamics, and changes in the
northern range environment since wolf reintroduction in
1995. Not only is the northern range a sufficiently large
ecosystem for assessing trophic cascade effects, the role of elk
relative to woody browse species has been a topic of concern
over many decades.

Northern winter range. The northern winter range comprises
more than 1500 km2 of mountainous terrain, of which
approximately two-thirds occurs within the northeastern
portion of YNP in Wyoming (NRC 2002a). The remainder lies
immediately north of the park and consists of various private
lands and Gallatin National Forest lands in Montana (Lemke
et al. 1998). Nearly 90% of the winter range within YNP lies
between 1500 and 2400 m in elevation, with the remainder
at elevations above 2400 m (Houston 1982). The northern
winter range typically has long, cold winters and short, cool
summers; annual precipitation varies from about 30 cen-
timeters (cm) at lower elevations to 100 cm at higher eleva-
tions. Snowpack water equivalent on 1 April averages only
7 cm at the Lamar Ranger Station (1980 m elevation), in-
creasing to 50 cm or more at higher elevations; snowpack
depths can vary considerably from year to year. Much of the
winter range is shrub–steppe, with patches of intermixed
Douglas fir (Pseudotsuga menziesii) and aspen. Multiple
species of willow, cottonwood, and other woody browse
species are common within riparian zones. Seven species of
ungulates—elk, bison (Bison bison), mule deer, white-tailed
deer (Odocoileus virginianus), moose, pronghorn antelope

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(Antilocapra americana), and bighorn sheep—are found
in northeastern YNP, along with gray wolves, cougars (Felis
concolor), grizzly bears, black bears (Ursus americanus), and
additional smaller predators (table 1).

Yellowstone from the 1800s to 1995. Relatively little is
known about the occurrence of carnivores and ungulates in
northwestern Wyoming in the early 1800s or the effects of
hunting and fire use by Native Americans. Even with the ad-
vent of Euro-American beaver trappers in the mid-1800s,
little information about the biota of the northern range was
systematically recorded. Although YNP was established in
1872, uncontrolled market hunting inside and adjacent to the
park had significant effects on both carnivore and ungulate
populations in the early years of park administration. To
help curtail impacts on wildlife and other resources, in 1886
the US Army assumed responsibility for protecting resources
within the park. Ungulates, bears, and beaver were generally
protected during the period of army administration, which
ended in 1918; however, predators other than bears were
typically killed.

The early 1900s marked an exceptionally important period
in the ecological ledger of YNP’s northern range. When the
National Park Service (NPS) assumed management respon-
sibility in 1918, carnivores other than bears continued to be
hunted. For example, recorded kills included 121 mountain
lions from 1904 through 1925, 136 wolves from 1914 through
1926, and 4350 coyotes from 1907 through 1935 (Schullery
and Whittlesey 1992). This effort ultimately resulted in the ex-
tirpation of wolves in 1926 (figure 1a).

Before 1920, elk populations were probably increasing,
owing to protection efforts by the US Army and the NPS. Al-
though northern range elk populations of more than 25,000
animals (17 elk per square kilometer) were reported in the

early 1900s (Barmore 2003), the accuracy of these estimates
and the role of winter die-offs before the mid-1920s may never
be known (Houston 1982). The annual census of elk on the
winter range began in the mid-1920s (figure 1b) and has

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Table 1. Approximate animal densities for the northern
range of Yellowstone National Park (YNP).

Species Densitya

Carnivores
Grizzly bear (Ursus arctos) Unknown
Black bear (Ursus americanus) Unknown
Cougar (Felis concolor) < 20 Gray wolf (Canis lupus) 50 Coyote (Canis latrans) 200–250

Ungulatesb

Moose (Alces alces) < 0.05 Bighorn sheep (Ovis canadensis) 0.10–0.14 Pronghorn antelope (Antilocapra americana) 0.15 Bison (Bison bison) 0.4–0.5 Mule deer (Odocoileus hemionus) 1.3–2.0 Elk (Cervus elaphus) 8–10

a. Number per 1000 km2 for carnivores, per km2 for ungulates.
b. White-tailed deer (Odocoileus virginianus) have been infrequent-

ly observed in Yellowstone’s northern range, because this habitat is at
the “extreme upper limit of marginal winter range” for this species
(YNP 1997).

Source: Adapted from Smith and colleagues (2003), except for
pronghorn antelope, which was adapted from YNP (1997).

Figure 1. Historical trends for the northern range of
Yellowstone National Park since 1900: (a) relative num-
bers of wolves (Weaver 1978, Smith et al. 2003); (b) an-
nual elk numbers (indicated with diamond shapes) for
the northern range herd (Houston 1982, YNP 1997, Bar-
more 2003, Smith et al. 2003); (c) relative recruitment of
woody browse species (Barmore 2003, Beschta 2003,
Larsen and Ripple 2003); and (d) relative numbers of
beaver (Warren 1926, Jonas 1955, Kay 1990, Smith et al.
2003). The width of the gray tone represents the uncer-
tainty of animal or plant numbers over the period of
record, based on the information for each species in the
literature citations. Elk populations shown in (b) were
not censused during the winters of 1996/1997 and
1997/1998; however, mortality due to winter weather,
and not wolf predation, is thought to be the primary rea-
son for the general decrease in elk numbers following the
winters of 1996/1997 and 1997/1998. The general in-
crease in beaver from about 1900 to 1920 is thought to be
a recovery of depressed populations following heavy trap-
ping pressure in the late 1800s. The increase in beaver
also occurred at a time when predators were increasingly
being removed from the park.

a
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b

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restored
(1995)

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(1968)

Wolves
extirpated

(1926)

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continued until the present, with data missing for some years.
With the removal of predation and associated predation risk
effects following the extirpation of wolves, elk in the north-
ern range had a significant impact on the recruitment of
deciduous woody species. As a consequence, recruitment of
upland aspen and riparian cottonwood soon crashed (figure
1c). This loss of recruitment continued over multiple decades,
even though sport hunting of elk occurred each winter when
some of the elk left the park, and park administrators delib-
erately sought to reduce the elk population from the mid-
1920s to 1968 (figure 1b).

Ripple and Larsen (2000) evaluated aspen overstory
recruitment in YNP over the last 200 years, using increment
core data collected in 1997 and 1998 and aspen diameter
data collected by Warren (1926). Successful aspen overstory
recruitment occurred on the northern range of YNP from the
mid-1700s to the 1920s, after which it essentially ceased.
They found that aspen recruitment ceased during the same
years (1920s) that gray wolves were extirpated from the park.
In a later study, Larsen and Ripple (2003) concluded that the
lack of recruitment was not correlated with indices of climate.

In a study of cottonwoods in the Lamar Valley portion of
the northern range, Beschta (2003) evaluated recruitment over
the last two centuries and found reduced recruitment in the
1920s and 1930s, with little cottonwood recruitment after the
1930s. The recruitment gap occurred independently of fire
history, flow regimes, or other factors affecting normal stand

development. Beschta (2003) concluded that the extirpation
of wolves allowed elk to browse highly palatable cottonwood
seedlings and suckers unimpeded during winter months and
prevented any recruitment from occurring for nearly a half-
century (figure 2). An exception to the general lack of cotton-
wood recruitment in the Lamar Valley occurred adjacent to
the Lamar Ranger Station, where elk-culling operations were
centered from the 1930s to 1968. Apparently the predation risk
from humans at this facility allowed a few young cotton-
woods to establish after 1933 and ultimately to grow above the
browse level of elk.

The NPS initiated efforts in the mid-1920s to reduce the
size of elk herds in the northern range because of concerns
about overgrazing; those efforts continued until the late 1960s
(YNP 1997). From the 1930s to the early 1950s, the elk pop-
ulation on the northern range generally fluctuated between
8000 and 11,000 animals (5.3 to 7.3 elk per square kilometer).
By the 1950s and 1960s, live trapping and shooting of elk by
NPS personnel, in combination with sport hunting of animals
that seasonally migrated outside of park boundaries, reduced
the number of elk to between 4000 and 8000 animals (2.7 to
5.3 elk per square kilometer) (figure 1b). For comparison,
White and colleagues (2003) indicate that more than four
elk per square kilometer is considered a high density in the
Canadian Rockies. Of the nearly 75,000 elk removed from the
northern range herd over the period 1926–1968, approxi-
mately 36% involved culling operations by the NPS and 74%

760 BioScience • August 2004 / Vol. 54 No. 8

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Figure 2. Elk browsing among cottonwood trees in the wintertime along the Lamar River in the northern
range of Yellowstone National Park during the period when wolves had been extirpated. Note the lack of
recruitment of small and intermediate-sized cottonwood trees that has occurred over many decades and
the general lack of vigilance indicated by the elk. Photograph: Yellowstone National Park.

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represented hunting kills outside the park. Seasonal sport
hunting just outside the park boundary may have caused
some elk to remain within the park instead of following
former down-valley migrations (Barmore 2003). However,
by the late 1960s, when the elk population had been re-
duced, recruitment of woody browse species did not occur
(figure 1c).

Following a cessation of culling efforts in 1969, the elk pop-
ulation began to increase rapidly and eventually attained
herd sizes ranging from 12,000 to 18,000 animals (8 to 12 elk
per square kilometer) between the late 1970s and the mid-
1990s (figure 1b). Although the NPS has generally charac-
terized the post-1968 management period as one of “natural
regulation” (NRC 2002a), the gray wolf—a keystone preda-
tor—and its associated lethal and nonlethal effects remained
absent during this period (until its reintroduction in 1995),
thus allowing a continuation of high levels of herbivory on
woody browse species.

Various other studies (Houston 1982, Kay 1990, Romme
et al. 1995, Meagher and Houston 1998) have noted declines
in woody browse species (aspen, willows, and berry-
producing shrubs) during the 20th century. Willows repre-
sent deciduous woody browse species that are commonly
found in riparian areas associated with the streams and rivers
of YNP. As with aspen and cottonwood, widespread losses of
willows have occurred in northern YNP over the past century
(Chadde and Kay 1996, Barmore 2003, Singer et al. 2003).

However, much of the evidence of willow loss is based on com-
parisons of paired historical photographs, often taken widely
spaced in time (Meagher and Houston 1998). Whereas in-
crement cores from existing aspen and cottonwood stands pro-
vide a convenient means of determining when recruitment
declines occurred, similar temporal documentation of loss is
not available for willow stands.

Historical photographs provide evidence of young aspen
and willow thickets on the northern range of YNP in the early
20th century (Houston 1982, Kay 1990, Meagher and Hous-
ton 1998). Houston (1973) attributed the common occurrence
of aspen thickets on the northern range in the 1800s and early
1900s to the occurrence of frequent fires. Historically, elk
may have avoided the interior of aspen thickets because of pre-
dation risk effects resulting from a lack of visibility and in-
creased impediments to escape associated with high stem
densities (Ripple and Larsen 2000). Then and now, postfire
accumulations of coarse woody debris serve as barriers to
browsing as well as impediments to escape (Ripple and Larsen
2001).

Meagher and Houston (1998) commented on the visible
effects of preferential ungulate browsing along the edge of as-
pen thickets (figure 3). This type of risk-sensitive foraging has
also been observed for caribou in Alaska, which skirt willow
thickets to avoid predation by wolves (Roby 1978). The hy-
pothesis that elk typically browse on the edge of aspen thick-
ets to avoid predation by wolves is also supported by empirical

August 2004 / Vol. 54 No. 8 • BioScience 761

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Figure 3. Photograph taken in 1900 near Tower Junction on the northern range of Yellowstone National
Park, showing evidence of elk browsing on the outer stems of a 3- to 5-meter-tall aspen thicket in the fore-
ground and multiple aspen thickets on a distant hillslope (Meagher and Houston 1998). We hypothesize
that dense regeneration after wildfire resulted in high levels of predation risk in the interior of aspen thick-
ets; thus, browsing is evident only along the outer edges of the thicket. Even following the widespread fires of
1988, such aspen thickets are not common on the northern range. Photograph: Yellowstone National Park.

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data from the Canadian Rockies. When elk were under risk
of predation by wolves, the number of elk pellets was higher
on the edge of aspen thickets than in the interior of aspen
patches (White et al. 2003).

Viewed from a perspective of trophic cascades and preda-
tion risk, the plant community responses experienced in
northeastern YNP over the 20th century are consistent with
the expected consequences of extirpating gray wolves. The re-
sultant lack of predation and predation risk allowed elk to for-
age unimpeded on woody browse species, causing
much-simplified plant communities of low stature (figure 4).
Without the presence of this keystone predator, the only ma-
jor limitation to accessing woody browse species each winter
was snow depth. Since valley bottoms in the northern range
typically have relatively shallow snow depths (Barmore 2003),
this situation ensured that woody plants in riparian areas
were heavily affected by browsing (figure 2).

Even though coyotes, bears, and cougars were present in the
park throughout the 20th century, these predators have had
no documented effects on winter patterns of elk herbivory.
Furthermore, upland (aspen) and riparian (willow, cotton-
wood) woody browse species were heavily browsed in spite
of long-term NPS efforts to reduce elk numbers. The poten-
tial long-term sustainability of many woody browse species
in the northern range represents a major ecological concern,
since the pattern of unimpeded browsing resulting from a lack
of predation risk continued from the 1920s to the mid-1990s.

For example, aspen clones that may have existed on the
northern range for thousands of years, if lost, cannot be re-
stored except through seeding events, which are rare (Kay 1990,
Romme et al. 1995, Larsen and Ripple 2003).The persistent
overbrowsing and reduction of woody browse species has also
had consequences for other faunal species (figure 5a). With
fewer aspen and riparian woody plants, the capability of
these plant communities to provide food for avian species is
greatly diminished (Dobkin et al. 2002). For beaver, although
historical details are lacking, the impacts have apparently
been severe. The Yellowstone region abounded in beaver in
the early 1800s, but extensive trapping by Euro-Americans be-
gan in the 1830s and continued through the latter half of the
19th century. The beaver population apparently began to re-
cover by the early 1900s and attained relatively high numbers
by the early 1920s (figure 1d; Warren 1926). However, the
number of beaver underwent a major decline in the late
1920s (Schullery and Whittlesey 1992), with only scattered
colonies of beaver remaining by the early 1950s (Jonas 1955).
Numbers of beaver in the northern range remained low over
the next five decades; during the 1980s and early 1990s,
beaver were essentially absent from streams of the northern
range (Kay 1990, YNP 1997). The loss of beaver populations
appears to represent an ecological chain reaction to behav-
iorally mediated trophic cascades involving elk, following
the extirpation of wolves. According to the NPS (1961), the
decrease in beaver in the northern range, which began in the

762 BioScience • August 2004 / Vol. 54 No. 8

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Figure 4. Flow diagram of predator–prey encounters for wolves and elk in Yellowstone
National Park (YNP) since 1926. Modified from Lima and Dill (1990).

YNP 1926–1995

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late 1920s, resulted from interspecific competition with elk:
Beaver would fell the larger stems of aspen, willow, or cot-
tonwood for food and dam material, while elk would consume
all new shoots. Thus, unimpeded browsing by elk may have
effectively destroyed any food supplies for beaver.

Yellowstone after wolf reintroductions (1995–present). Un-
der the protection of the 1973 federal Endangered Species Act,
an experimental population of wolves was reintroduced into
YNP during the winter of 1995/1996, following a 70-year pe-
riod without their presence (figure 1a). Since the reintro-
duction of 31 wolves into YNP in the mid-1990s, their
numbers have steadily increased. By the end of 2001, the
population of wolves in Yellowstone’s northern range had
grown to 77 animals (Smith et al. 2003). Even with the re-
introduction of wolves and their subsequent increase in
recent years, we are still in the early stages of understanding
how their restoration is influencing ungulates, vegetation,
riparian functions, or other ecological components in north-
ern Yellowstone (figure 5b).

Following the reintroduction of wolves, Ripple and Beschta
(2003) found that predation risk associated with various ter-
rain conditions (and their related fear factors) played a role
in the selective release of willow and cottonwood from the
browsing pressure caused by elk in the Lamar Valley of north-
ern YNP. In 2001 and 2002, they found willow and young cot-
tonwood plants 2 to 4 m in height, which is in stark contrast
with the long-term observations of plants less than 1 m in

height during the decades before wolf reintroduction. Willow
and cottonwood were found to be subject to less browsing
pressure (figure 6) at potentially high-risk sites with limited
visibility (i.e., limited opportunities for prey to see ap-
proaching wolves) or with terrain features that could impede
the escape of prey, such as sites below high terraces or steep
cutbanks and near gullies. As an additional indicator of ri-
parian recovery, several new beaver colonies have recently been
established on the northern range, a rare occurrence over the
last five decades (figure 1d). The number of beaver colonies
on the park’s northern range increased from one in 1996 to
seven in 2003 (YNP files).

Ripple and Beschta (2003) suggested that elk would in-
creasingly forage at sites that allow early detection and suc-
cessful escape from wolves, since the Lamar Valley has a
predominately open habitat structure (figures 3, 4). Had the
woody plant communities in the northern range not been so
thoroughly simplified and degraded by multiple decades of
persistent and unimpeded elk herbivory in the absence of
wolves, the differential plant responses to predation risk
following wolf reintroductions might not have been readily
observed. Conversely, if elk densities in the future are
reduced, with concurrent decreases in overall browsing pres-
sure, we envisage that it will be more difficult to detect
differential plant responses associated with predation risk.
If elk densities become low enough, we expect a more wide-
spread release from browsing of woody plants rather than
release only at high-risk sites.

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Figure 5. Trophic interactions due to predation risk and selected ecosystem responses to (a) wolf extirpation
(1926–1995) and (b) wolf recovery (post-1995) for northern ecosystems of Yellowstone National Park. Solid
arrows indicate documented responses; dashed arrows indicate predicted or inferred responses.

Trophic cascades without wolves Trophic cascades with wolves
Trophic

cascades
model

Predators

Prey

Plants

Other
ecosystem
responses

Elk browse woody species unimpeded
by predation risk

Wolves extirpated (1926–1995)

Decreased recruitment of woody browse
species (aspen, cottonwood, willows,

and others)

Loss of food web
support for

aquatic, avian,

and other fauna

Loss of riparian
beaver

Loss of riparian
functions

Recovery of
riparian functions

Recolonization
of beaver

Recovery of food
web support for
aquatic, avian,

and other fauna

Channels stabilize, recovery of
wetlands and hydrologic

connectivity

Channel incision and widening,
loss of wetlands, loss of hydrologic

connectivity between streams
and floodplains

Increased recruitment of
woody browse species

Elk foraging and movement pat-
terns adjust to predation risk

Wolves restored (post-
1995)

a b

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Conclusions
Can predation risk structure ecosystems? Our answer—based
on theory involving trophic cascades, predation risk, and
optimal foraging, in addition to a developing body of empirical
research—is yes. Although some may find the support for this
answer equivocal, we find it compelling when all the evi-
dence is combined. Predation risk probably affects ecosystems
in both subtle and dramatic ways through various interactions,
many of which are unknown. For example, little is known
about how elk use scent and sound in conjuction with visual
indicators for assessing predation risk. Because many carni-
vores have been extirpated from their original ranges, there
has been little opportunity to study their lethal and non-
lethal effects on prey, alone or in combination with episodic
abiotic events. Ultimately, researchers and managers need to
understand how the interaction of lethal and nonlethal effects
structures the ecosystem. In Yellowstone, the role of lethal ef-
fects may become increasingly important in the future, as the
combined effects of predation by wolves, bears, and hunters,
along with periodic severe winter weather events, may ulti-
mately cause lower elk populations.

The concept of trophic cascades provides a basis for un-
derstanding, perhaps for the first time, the often conflicting
viewpoints regarding interactions between elk (as well as
beaver and other fauna) and vegetation in Yellowstone’s
northern range. Over a period of many decades, the intense
ecological and political debate regarding potential over-
browsing effects of elk on the northern range of YNP has al-
most always centered on numbers of elk (NRC 2002a). In
contrast, our assessment of the broader literature and the Yel-
lowstone research indicates that the extirpation of the gray
wolf—a keystone predator in this ecosystem—is most likely
the overriding cause of the precipitous decline and cessa-
tion in the recruitment of aspen, cottonwood, and willow
across the northern range. This hiatus in recruitment of
woody species is also directly linked to the loss of beaver and
the decline in food availability for other faunal species. It is
important to note that the loss of recruitment occurred de-
spite long-term variations in winter weather, snowpack, and
other climate variables, with or without the occurrence of fire,
and independent of efforts by the NPS to control ungulate
numbers inside the park (pre-1968) or to let them increase
by ceasing control efforts (post-1968).

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Figure 6. Willow along Blacktail Creek in spring 1996 (left) and summer 2002 (right). Following a 70-year
period of wolf extirpation, heavy browsing of willows and conifers is evident in the 1996 photograph. In
2002, after 7 years of wolf recovery, willows show evidence of release from browsing pressure (increases in
density and height). Photographs: left, Yellowstone National Park; right, William J. Ripple.

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In terms of future management of the northern range un-
gulate herds, our assessment suggests that restoration goals
should focus on the recovery of natural processes. In the
case of Yellowstone, the return of wolves represents an example
of active management to recover a lost keystone species.
However, passive restoration of other ecosystem processes and
components as a result of the combined lethal and nonlethal
effects of this restored predator can now play out in ways that
we cannot easily predict and perhaps will not fully understand
for many decades. In addition to restoring large carnivores
such as wolves, it may be important to recover historical un-
gulate migrations as much as possible, especially in situations
where ungulates tend to avoid natural migrations in an effort
to lower their risk of predation or other impacts from humans
and, as a consequence, reside inside park or reserve bound-
aries.

Since much of our discussion has focused specifically on
the northern range of YNP, we are not sure of the extent to
which our conclusions on behaviorally mediated trophic
cascades match what has occurred to ungulates, plants, and
associated ecosystem responses in other portions of North
America where wolves have been extirpated and, in some cases,
reintroduced. In the last decade, wolf recovery efforts have
been initiated in portions of Montana, Idaho, Arizona, New
Mexico, and the upper Midwest. If ecosystem responses
similar to those that have occurred historically or that are
under way on the northern range are documented in other
locations, we may finally understand more fully the obser-
vations and concerns of Aldo Leopold from over half a
century ago.

Acknowledgments
We thank Kevin Crooks, John Kie, Steve Lima, and Dale Mc-
Cullough for reviewing an early draft of this article and for
providing helpful comments. We are also grateful to Paul
Schullery for providing information on historical wildlife
populations.

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Fire as a global ‘herbivore’: the ecology
and evolution of flammable
ecosystems
William J. Bond

1

and Jon E. Keeley

2,3

1
Department of Botany, University of Cape Town, Rondebosch,

South Africa

2
U.S.GeologicalSurvey, Western Ecological ResearchCenter, Sequoia-Kings Canyon National Parks,Three Rivers, CA93271-9651, USA

3
Department of Ecology and Evolutionary Biology, University of California, Los Angeles, CA 90095, USA

It is difficult to find references to fire in general

textbooks on ecology, conservation biology or biogeo-

graphy, in spite of the fact that large parts of the world

burn on a regular basis, and that there is a considerable

literature on the ecology of fire and its use for managing

ecosystems. Fire has been burning ecosystems for

hundreds of millions of years, helping to shape global

biome distribution and to maintain the structure and

function of fire-prone communities. Fire is also a

significant evolutionary force, and is one of the first

tools that humans used to re-shape their world. Here,

we review the recent literature, drawing parallels

between fire and herbivores as alternative consumers

of vegetation. We point to the common questions, and

some surprisingly different answers, that emerge from

viewing fire as a globally significant consumer that is

analogous to herbivory.

Parallels between fire and herbivory

Ecologists and biogeographers generally assume that
plant distribution, abundance and, therefore, community
composition, structure and biomass, are determined
largely by climate and soils. This is implicit in current
attempts to model species range shifts in response to
climate change [1]. However, nearly 50 years ago, Hair-
ston et al. [2] suggested that the properties of ecosystems
are instead determined by the regulation of herbivores by
predators. In the absence of predators, herbivore popu-
lations would proliferate, consuming such large quantities
of vegetation that plant communities would be trans-
formed to those tolerant of herbivory rather than those
best able to compete for resources. Critics claimed that
terrestrial plants are largely inedible so that, even
without predators, herbivores could seldom consume
enough to transform ecosystems [3]. The effects of fire
are, in many ways, analogous to those of herbivory, but
have been missing from the trophic ecology literature.
Although usually treated as a disturbance, fire differs
from other disturbances, such as cyclones or floods, in that
it feeds on complex organic molecules (as do herbivores)
and converts them to organic and mineral products. Fire

Corresponding author: Bond, W.J. (bond@botzoo.uct.ac.za).
Available online 3 May 2005

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0169-5347/$ – see front matter Q 2005 Elsevier Ltd. All rights reserved

differs from herbivory in that it regularly consumes dead
and living material and, with no protein needed for its
growth, has broad dietary preferences. Plants that are
inedible for herbivores commonly fuel fires.

How does fire, unconstrained by low food quality, fit the
predictions of Hairston et al. [2] as an ecosystem consumer
that is unconstrained by predators? Here, we discuss the
ecology of flammable ecosystems, using the term
‘consumer control’ for ecosystemsin which fire orherbivores
significantly alter biomass, the mix of plant growth forms,
and species composition in ecosystems. We contend that
consumer control is important ecologically, biogeo-
graphically and evolutionarily when the consumer is fire.

Fire and consumer control of ecosystems

Polis [3], in a review of the ‘green world’ hypothesis,
argued that terrestrial vegetation is determined largely by
climate, locally modified by low-nutrient soils, with
consumer control by herbivores sometimes occurring but
being localized in space and time. How can the global
importance of consumers (herbivores and fires) versus
resources (climate and soils) in shaping vegetation be
evaluated? A useful alternative to meta-analyses of
experimental studies (often limited in space, time,
taxonomic bias and reportage counts) is to compare
potential versus actual ecosystem properties for a given
locality. If an ecosystem differs greatly from its resource-
limited potential properties, then it is a candidate for
‘consumer control’, be it either by herbivory or fire
(Figure 1). Frequent fires reduce the height of the
dominant plants (Figure 2) and, therefore, the position,
but not necessarily the amount, of leaves and canopy
photosynthesis. Woody plant biomass, rather than primary
productivity, is therefore the more revealing measure of
consumer control by fire. The problem is how to measure
potential biomass, the ‘carrying capacity’ of trees at a site,
against which actual ecosystems can be measured.

Dynamic global vegetation models (DGVMs) can be
used to provide an approximation of climate-limited
potential biomass [4]. DGVMs are complex models,
analogous to global climate models, which ‘grow’ plants
according to physiological principles using climate and soil
physical properties as input [5,6]. The models predict
vegetation responses to global change and can simulate

Review TRENDS in Ecology and Evolution Vol.20 No.7 July 2005

. doi:10.1016/j.tree.2005.04.025

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TRENDS in Ecology & Evolution

0

200
Plant available moisture

T
re

e
b

io
m

a
ss

High

High

Low
Low

Actual

Consumer
control

Climate
potential

Figure 1. Assessing consumer control of tree biomass. The extent of consumer

control of an ecosystem can be measured as the difference between tree biomass at

‘climate potential’ and the actual tree biomass. Large differences between potential

and actual woody biomass suggest significant consumer control of the ecosystem.

‘Climate potential’ can be viewed as the carrying capacity of a site for trees.

Review TRENDS in Ecology and Evolution Vol.20 No.7 July 2005388

potential vegetation for any given location (Figure 2).
According to these simulations (e.g. Figure 3), vast areas
of wooded grasslands in Africa and South America, and
smaller areas of grassy ecosystems and shrublands on all
vegetated continents, have the climate potential to form
forests. Closed forests, which currently cover a quarter of
the land surface on Earth, would more than double in
extent if world vegetation was as ‘green’ as it could be [4].
These simulations contradict current perceptions that
consumer control is of negligible importance in terrestrial
ecosystems [3,7]. The biomes most at variance with
climate potential are C4 grasslands and savannas,
especially in more humid regions, such as Brazilian
cerrados and the wetter regions of Africa. These are the
most frequently burnt ecosystems in the world, burning
several times in a decade and some burning twice a year
[8,9]. Thus, fire is the prime candidate for consumer control
of large parts of the world. Past or future changes in the
extentoftheseecosystems,orspecieswithinthem,cannotbe
understood without understanding the ecology of fire.

0

5000

10000

15000

20000

25000

Zimbabwe 1 Z

B
io

m
a
ss

g
m

–
2

South Africa

Figure 2. Changes in woody biomass in savanna long-term burning experiments. The un

and the shaded bars indicate biomass where fire has been excluded for 35 years or more.

Woody biomass simulated by the Sheffield DGVM for ‘fire off’ is indicated by the filled

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The implications of Figure 3 are even more significant
when it is recognized that the mismatch with climate
potential is based only on biomass and not on changes in
species composition. The simulations cannot identify
those ecosystems in which fires change species compo-
sition without significantly altering the biomass of trees,
such as conifer forests [10] or Australian eucalypt
communities [11]. The full extent of fire-controlled
vegetation, defined as ecosystems that are altered in
structure, composition and functioning when fire is
released or suppressed, is much greater.

Fire and consumer control of species composition

Trophic cascades, measured as large changes in species
composition, are an expected consequence of predator
removal in ecosystems where consumers have the poten-
tial to proliferate in their absence. Although evidence for
trophic cascades in terrestrial ecosystems is disputed [7],
cascading changes in species composition are common-
place where fire is the consumer. For example, in tropical
forests, a single fire can reduce woody plant richness by a
third to two-thirds depending on fire severity and can
have negative impacts on a diverse array of faunal
components [12–15]. Changes in fuel distribution and
microclimate after a tropical forest fire increase the
probability of more fires and conversion of forest to scrub
and grassland [12,15].

By contrast, for ecosystems with a long history of fire,
there is concern over the cascading consequences of
anthropogenic fire suppression. In tall grass prairies,
and comparable grasslands elsewhere, fire suppression
has led to the loss of as many as 50% of the plant species
[16,17]. Small herbaceous plants with high light require-
ments for growth and seedling establishment are the
worst affected. Changes in faunal composition have also
been reported, for example, in dry dipterocarp woodlands,
where fire suppression has resulted in a marked loss of
termite species [18]. Even greater species losses occur
where fire suppression leads to complete biome switches,
such as from savannas to forests [19,20]. There is, as yet,
no global synthesis of species turnover in different

imbabwe 2 Venezuela

Site

North America

shaded bars indicate aboveground woody biomass in frequently burnt treatments

Sites are ranked, from left to right, according to increasing plant available moisture.

squares. Modified, with permission from the New Phytologist Trust, from [4].

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TRENDS in Ecology & Evolution

Bare C3 C4 Ang CropGym

(a)

(b)

Key:

1

2 3

4
5

6 7
8

9

10

Figure 3. A comparison of global biome distribution at climate potential (a) versus

actual vegetation (b). Biomes are represented by the cover of the dominant plant

functional type: C3 grasses or shrubs; C4 grasses or shrubs; Ang, angiosperm trees;

gym, gymnosperm trees (mainly conifers). The numbers indicate sites where fire

has been excluded for several decades. All the higher rainfall sites showed a

successional tendency to form forest following suppression of fire. The map of

potential World vegetation, limited only by climate, was simulated using a DGVM

(using global climate and soil databases). The map of actual vegetation was sourced

from ISLSCP: (ftp://daac.gsfc.nasa.gov/data/inter_disc/biosphere/land_cover/); repro-

duced, with permission from the New Phytologist Trust, from [4].

Review TRENDS in Ecology and Evolution Vol.20 No.7 July 2005 389

ecosystems and under different fire regimes following fire
release or suppression. We would expect a continuum of
responses from near-complete species replacement follow-
ing biome switches to negligible changes in ecosystems
where fires, although predictable, are infrequent. The
Yellowstone fires of 1988, for example, caused no loss or
gain of species in this landscape [21]. Thus, it is not yet
possible to draw a global map to show the extent of
ecosystems whose species composition would change
significantly if fires were suppressed.

The variable nature of fire as a consumer control

Flammable ecosystems include boreal forests, eucalypt
woodlands, shrublands, grasslands and savannas. Why, if
fire is such an influential consumer, is there such a
diversity of growth form mixtures in flammable ecosys-
tems? Fire ecologists have looked first to the diversity of
fire regimes for answers. A fire regime includes the
patterns of frequency, season, type, severity and extent
of fires in a landscape (Box 1). Vegetation consumed and
patterns of fire spread vary across landscapes, and
different fire regimes produce different landscape pattern-
ing and select for different plant attributes. It follows that
changes in fire regimes, within a given landscape, should
have major ecosystem consequences.

Consider the conifer forests of southwestern North
America. In these semi-arid landscapes, forests have long

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been shaped by a fire regime of frequent relatively low
intensity (low flame height and temperature) surface fires.
These forests share attributes with subtropical grasslands
in that fires are ignited by frequent lightning strikes at the
beginning of the monsoon season, when the fuels are at
their driest. However, primary productivity in conifer
forests is lower than in mesic savannas owing to their
greater aridity and this translates into lower fire
frequency, lower fire intensity and greater heterogeneity
in ‘feeding patterns’ of the fire [22]. As a consequence,
opportunities exist for the occasional establishment of
trees that persist to form low-density forests. Fires exhibit
a sort of ‘selective herbivory’, consuming herbaceous
surface biomass but leaving the dominant overstorey
trees untouched. Following human settlement during the
early 20th century, these conifer landscapes have been
managed with a policy of total fire suppression, which is a
fortuitous experiment on how fire controls vegetation
structure, and has resulted in near-total fire exclusion.
Forests that naturally burned at rates of once or twice a
decade have now gone unburned for more than a century
[23], resulting in major shifts in ecosystem structure and
function. Tree density has increased by an order of
magnitude or more, with major losses in the herbaceous
understorey and species diversity. In addition, the absence
of fire has resulted in changes in many ecosystem
components. Of profound management importance is the
fact that fire suppression has lead to fuel accumulation
and this has set the forest on a different trajectory such
that, when fires do occur, they now feed as massive forest-
consuming ‘monsters’, rather than in the manner of
ground-dwelling herbivores.

Most work on fire regimes is constrained to particular
landscapes and ecosystems. There is no global synthesis
on what determines fire regimes in world ecosystems. We
do not yet understand the synergies and relative import-
ance of ignition, dry periods, the properties of vegetation
as fuel, or topographic barriers to fire spread in determin-
ing which fire regimes occur where. This seriously under-
mines our ability to predict the consequences of global
change for fire-affected ecosystems or to interpret past
changes in the distribution of flammable ecosystems.
What is clear is that different fire regimes select for
different plant attributes and similar fire regimes select
for similar attributes. Savanna ecologists worldwide find
similar plant traits with similar fire responses [24].
Ecologists working in Mediterranean-type shrublands
find convergent fire-related plant traits on different
continents [25,26] but these are different from those of
savannas. Transgressing from one fire regime to another
seems to be as difficult as finding commonalities between
insect and mammal herbivory, because the biology of the
‘organisms’ is so different.

Fire and community assembly

Hairston et al. [2] predicted relatively little competition
between plants where herbivores proliferate in the
absence of predators, because plant growth would be
limited more by consumption than by resources. Instead,
community assemblages would comprise species that are
best able to persist and thrive in the face of repeated

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Box 1. Fire regimes

Gill [61] introduced the concept of a fire regime, which we have

modified to include: (i) fuel consumption and fire spread patterns; (ii)

intensity; (iii) severity; (iv) frequency; and (v) seasonality.

Fuel consumption and fire spread
Fires consume a range of fuel types, which has profound

impacts on ecosystems. Surface fires spread by fuels that are

close to the ground, such as grass or dead leaf and stem

material, whereas crown fires burn in the canopies of shrub- and

tree-dominated associations. Ground fires burn soils that are rich

in organic matter. They can be ignited by lightning strikes and

can smolder for long periods until changes in the weather favor

surface or crown fires.

Some forests have a heterogeneous mix of surface fires,

crown fires and unburned patches, which is important to

ecosystem processes such as tree recruitment. For example, in

the mixed conifer forests of the Sierra Nevada in California,

patches of high-intensity fires produce light gaps that are

important for tree regeneration [62]. These gaps also accumulate

fuels at a slower rate and thus have a greater probability of

being missed by fires until saplings reach sufficient size to

withstand them [63].

The ecological importance of fire size varies with the

ecosystem and also with different species in the system. For

example, chaparral shrublands commonly experience large

crown fires that can completely denude tens of thousands of

hectares. This poses no threat to the plant species in these

ecosystems because regeneration is entirely dependent upon

endogenous processes (Box 2). However, mixed conifer forests

in the western USA are potentially more sensitive to fire size.

Historically, these forests have burned with a mix of surface

fires, which left dominant trees alive, and crown fires, which

killed all trees within small patches from a few hundred square

meters to a few hundred hectares. Reproduction of the dominant

trees requires gaps generated by crown fires, but they must be

within dispersal distance of parent trees. When crown fires are

very large, regeneration is negatively impacted.

Intensity
Fire intensity refers to the energy release or, more loosely, to

other direct measures of fire heating or behavior, such as flame

length and rate of spread. Fireline intensity, which is the energy

per length of fire front, is increasingly used as a standard for fire

intensity.

Severity
Although fire intensity is a measure of immense importance to fire

fighters, ecologists are often more interested in fire severity, broadly

defined as a measure of ecosystem impact. In forested ecosystems,

tree mortality is commonly used as a metric for fire severity;

however, other metrics are used in shrublands where all above-

ground plants are consumed.

Frequency
Fire frequency is the occurrence of fire for an area and time period of

interest. There are complications with assessing fire frequency that

involve complex fire behavior at different spatial scales with different

limitations. Fire rotation interval is the time required to burn the

equivalent of a specified area, whereas fire return interval is the time

interval between fires at any one site [10].

Season
Fire season is dictated by the coincidence of ignitions and low fuel

moisture. This is usually the driest time of the year, which varies with

regional climate. In many ecosystems, humans have greatly altered

fire season by providing ignitions outside the natural lightning storm

period.

Review TRENDS in Ecology and Evolution Vol.20 No.7 July 2005390

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defoliation. One of the striking features of the fire ecology
literature is that there are many studies on life-history
traits that enable the persistence of species in a given fire
regime (Box 2), but few on resource acquisition and
competition. The consistency with the predictions of
Hairston et al. [2], that competition will be of minor
importance in consumer-controlled ecosystems, seems to
have gone unnoticed.

The plant traits that are important for fire persistence
are different in communities that experience different fire
regimes. In crown-fire regimes, where all woody biomass
is consumed, there are numerous studies of the mode of
recovery from burning (vegetative sprouting or non-
sprouting), fire-stimulated recruitment, time to first
reproduction and the persistence of seedbanks to the
next fire [20,26,27]. These plant traits, together with the
patterns of fire consumption, especially its frequency, are
widely used for predicting compatible species assemblages
[26,28]. However, community membership is seldom
attributed to competitive interactions with other plant
species, except when those species change the disturbance
regime [29].

In surface-fire regimes, such as savannas, fires feed
selectively, consuming plants in the grass layer but not
trees taller than 2–4 m. The coexistence of trees and
grasses has been attributed to niche differentiation, with
grasses being the more successful competitors for
resources in the soil surface, and trees accessing resources
in deeper soil layers [30]. An alternative idea, consistent
with consumer-controlled ecosystems, is that tree cover is
limited by demographic bottlenecks at different life-
history stages in tree growth [30,31]. Fire would be a
major cause of these bottlenecks in frequently burnt
savannas, reducing seedling establishment and prevent-
ing saplings from emerging from the ‘fire trap’, the flame
zone produced by grass fires. Vertebrate herbivores have
analogous effects, suppressing seedlings by heavy brows-
ing with rare burst of recruitment when plants are
released from herbivory [32]. The niche differentiation
hypothesis predicts no changes in tree cover from fire
suppression (or herbivore exclusion) because tree cover is
limited by resource competition. But many long-term fire
exclusion experiments (Figure 2) show that tree cover is
limited by fire. In these instances, consumer control, rather
than resource competition, determines tree cover [33].

Fire as an evolutionary agent

There are few studies of the evolution of fire-adaptive
traits, and many plant traits have been uncritically
labeled as ‘fire adaptations’ without any rigorous analysis
either as to the functional importance of the trait, or its
phylogenetic origin. For example, post-burn sprouting is
often seen as a ‘fire adaptation’, but sprouting per se is a
widespread trait in angiosperms. Evolutionary interpret-
ationsoftheloss or gain ofsprouting in different fireregimes
make no sense without phylogenetic analysis [34,35].

Among the most compelling new studies are those
exploring the evolution of flammability. In a debate
echoing that over whether plants have evolved to promote
herbivory (and just as controversial), ecologists have
asked whether plants in fire-maintained ecosystems

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Box 2. Life histories shaped by fire

Of the many traits that can be interpreted as being of functional

importance in fire-controlled environments, two have captured most

attention: sprouting and fire-triggered seedling recruitment. Sprouting

is the vegetative regeneration that occurs following the destruction of

living tissues. This can be either from roots or stems following the

death of all aboveground tissues, or along stems where branches

have been killed. Sprouting is a widespread trait in woody species

and is not closely tied to fire-prone environments [35]. One

exception is Pinus, a genus in which sprouting is rare and

apparently derived in crown fire ecosystems [40]. However,

sprouting from basal lignotubers that are produced as a normal

development stage is a combination much more commonly found

in fire-prone Mediterranean-climate ecosystems [25]. Sprouting in

the context of other life-history characteristics represents complex

patterns that have recently been reviewed elsewhere [35].

Many species in fire-prone environments with stand-replacing fires

have seedling recruitment restricted to the first postfire year [20,27]. In

flammable southern hemisphere shrublands, many species produce

serotinous fruits that open following fire and disperse seeds that

readily germinate following the wet season rains [64]. In comparable

shrublands of the northern hemisphere, serotiny is relatively rare. In

both hemispheres, many species produce seeds that are dormant and

accumulate in the soil. Germination is triggered by either heat or

smoke (or charred wood) [65]. Heat-stimulated germination is typically

in hard-seeded species that have a physical seed coat barrier to water

uptake. Germination is triggered by heat shock from fire, or by high soil

temperatures on open sites. There is a marked phylogenetic pattern in

that certain plant families are associated with either one mode or

another; for example, heat-stimulated germination is widespread in

Fabaceae, Cistaceae, Convolvulaceae and Sterculiaceae, and smoke-

stimulated germination is lacking [65]. Heat-stimulated germination is

globally widespread in numerous fire-prone ecosystems. Chemical

stimulated germination is triggered by smoke and/or charred wood. It

has, so far, been found to be important in only three Mediterranean-

climate shrublands, California chaparral [65], South African fynbos [66]

and Australian heathlands [67]. In California, the plant families in which

this germination mode is found are generally not the same as in the

southern hemisphere shrublands, indicating that this trait might have

convergently evolved.

Fire-stimulated flowering is another mechanism for post-burn

seedling recruitment [20]. Flowering occurs in the first postfire year

on resprouts from bulbs or rhizomes, followed by abundant seedling

recruitment in the second postfire year. Most species continue to

flower sporadically in later years, thus there is no obligate dependence

on fire for flowering. One exception is the South African fynbos

geophyte Cyrtanthus ventricosus, which germinates within days of a

fire, regardless of the season, and remains dormant until flowering is

again stimulated by smoke from another fire [68].

Not all species in fire-prone environments have life histories that

have been shaped by fire. In Californian and Mediterranean Basin

shrublands, many species have seedling recruitment that is restricted

to fire-free conditions [69]. They have a suite of reproductive traits,

including seed dispersal and seed germination behavior, which are

quite distinct from species with fire-stimulated seedling recruitment. In

both ecosystems [69,70], these non-fire types are from older lineages

and are derived from taxa that had origins under a different climate. It

has been suggested that these traits are no longer adaptive and

represent historical effects and species sorting processes [70]. An

alternative view is that these life-history syndromes are adapted to

habitats that still exist in fire-prone landscapes, and the coexistence of

fire-type and non-fire types is promoted by natural variability in fire

frequency [69].

Review TRENDS in Ecology and Evolution Vol.20 No.7 July 2005 391

have evolved flammability. Are there benefits for flam-
mable plants that outweigh the costs to survival of
burning more fiercely? Theory predicts that flammability
could, indeed, evolve if fire spread from a flammable plant
to kill its neighbors, and if the progeny of more flammable
mutants were more likely to recruit into the gaps created
[36,37]. In these models, flammability acts as a ‘niche
constructing’ trait [38,39], modifying the local environ-
ment to the benefit of the flammable genotype. This
hypothesis makes the testable prediction that flammable
morphology and fire-stimulated recruitment should be
correlated traits, and there is some support for this
prediction in pines [40]. In Pinus, serotiny (the retention
of seed in cones which open after a fire), a fire-recruitment
trait, is correlated with dead branch retention, a flamm-
ability trait. Plants that retain dead branches are more
likely to carry a fire into the canopy than are plants that
self prune. Schwilk and Ackerly [41] tested whether these
traits showed correlated evolution in pine phylogeny. Using
a set of ‘supertree’ phylogenies, the authors found strong
support for the predicted association between serotiny
and dead branch retention, and also between these
and other ‘fire-embracing’ morphological traits, such as
thin bark, early maturation age and more flammable
foliage, which would be expected in these stand-
replacing fire regimes [40]. It would be intriguing to
explore the evolution of flammability in other taxa and
other ecosystems. Has ‘niche construction’, via the
evolution of flammability of common species, played a
part in the spread of the flammable formations in
which they are contained?

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Studies of trait evolution, and the origins of the woody
flora of savannas, are hampered by our lack of under-
standing of the key traits needed to survive in grass-
fuelled fire regimes. Traits that are common in crown-fire
regimes are rare or absent in savannas [40]. In productive
grassy ecosystems, fires are too frequent to provide safe
sites for seedlings and fire-stimulated seedling recruit-
ment, including serotiny, seems to be an exception. Fires
are too frequent for the evolution of woody non-sprouters
and sprouting is the norm [31,42,43].

Trees that survive anthropogenic fires in tropical forests
tend to be those that have thicker, insulating bark [12].
Although trees in savannas are often thick barked,
regeneration of new plants is perhaps the main obstacle
for maintaining populations. Seedlings and saplings face
frequent and severe fire damage in mesic savannas.
Between fires, seeds have to germinate and seedlings have
to acquire bud and root reserves to resprout to survive the
next fire. Given that fires occur several times in a decade,
seedlings would need to acquire the ability to resprout
rapidly. A recent study in Brazil confirms this conjecture
[44]. The only consistent differences between seedlings of
sister species from forest and savanna habitats was the
greater allocation to coarse roots, associated with sprouting,
in the savanna species and more plastic light responses [44].

Frequent fires also select for a peculiar sapling growth
form, with pole-like stems and swollen underground roots
(Figure 4, [45,46]). The pole-like stem facilitates rapid
bolting towards a height that is out of reach of surface fires
and the roots provide the resources to resprout if the stem
fails to reach a fire-proof size before the next burn.

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Review TRENDS in Ecology and Evolution Vol.20 No.7 July 2005392

Saplings can be trapped in the flame zone for decades. In
longleaf pine, Pinus palustris, from the southeastern
USA, saplings bolt only once and fire-induced sapling
mortality is high [47] presumably because they lack
comparably large root reserves. In the polymorphic
African tree, Acacia karroo, pole-forming saplings occur
in frequently burnt savannas whereas plants form cage-
like architectures where fires do not occur [46] (Figure 4).
The contrasting sapling architectures suggest alternative
evolutionary trajectories when plants are exposed to
selection by different types of consumer (i.e. fire versus
browsers).

Fire and the origin of biomes

The large area occupied by flammable biomes, especially
in the tropics and sub-tropics, has often been attributed to
anthropogenic burning. Although anthropogenic fires

Figure 4. Fire, herbivory and tree architecture in savannas. (a) Grass-fuelled fires are very

and saplings. However, flame heights are too low to cause significant damage to em

architecture, such as this Acacia karroo from a South African savanna. (c) The cage-lik

(d) Shows a young sapling of Terminalia sericea, a common African savanna tree specie

sapling growth above the flame zone while also ensuring the ability of the plant to resp

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have extended areas of flammable vegetation [48,49],
there is now evidence that natural fires occurred long
before humans and that flammable ecosystems pre-date
anthropogenic burning by millions of years [4,50]. Stable
isotope evidence shows that C4 grassy ecosystems, the
most extensive flammable formation worldwide, first
appeared between 6 million years ago (Ma) and 8 Ma
[51]. Cerling et al. [51] hypothesized that this was due to
decreasing atmospheric CO2, but recent studies of
paleoatmospheres do not support their assumption of
low concentrations of CO2 during the late Miocene [52,53].
The spread of grasses has often been attributed to
coevolution with mammal grazers. However it is hard to
see how, by consuming grass, grazers would promote the
spread of grasslands at the expense of forests. By contrast,
grass-fuelled fires are known to promote the spread of
grassy ecosystems by carving holes in forests [29]. Strong

frequent in mesic savannas, posing a problem for the recruitment of tree seedlings

ergent trees. (b) Many trees in frequently burnt savannas have pole-like sapling

e architecture of A. karroo occurs where fires are rare but browsers are common.

s The combination of pole-like stems and reserves stored in swollen roots facilitate

rout after repeated fires.

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Review TRENDS in Ecology and Evolution Vol.20 No.7 July 2005 393

candidates for feeding the fires are the highly productive
and highly flammable C4 grasses of the humid tropics,
which began replacing forests from the late Tertiary
onwards [54–56].

Recent studies in Australia link fire to the origin and
spread of the biomes of that continent. Bowman [11] set
out to explain the distribution of the small fragments of
‘rainforest’ distributed as an archipelago in the vast seas
of eucalypt formations. The rainforests are closed for-
mations of fire-intolerant trees with shade-tolerant under-
storey species. By contrast, the eucalypt formations
support shade-intolerant understoreys of flammable
grasses or shrubs. Similar anomalous forest patches in a
matrix of flammable grasslands are widespread in Africa
and South America [4,57]. Generations of ecologists have
debated the determinants of these alternative biome
types, citing aridity, fire, nutrient-poor soils and anthro-
pogenic burning. Bowman critically evaluated the evi-
dence for all of these for Australia. Both biomes occur
across a wide rainfall gradient with ‘rainforest’
(a misnomer) petering out where rainfall drops below
600 mm yK1. Contrary to popular perceptions, the flam-
mable formations are not ‘arid’. Indeed, eucalypts are less
tolerant of drought than are many rainforest trees sharing
the same climate [11]. Bowman concluded that the only
consistent difference between trees of the two formations
was in their fire response: eucalypts have a remarkable
ability to survive and thrive under frequent fires, whereas
rainforest trees are killed by repeated burning [11,58]. The
implication is that if you could ‘switch off’ fires for a long
enough period, large areas of Australia would support an
entirely different ‘rainforest’ flora. This process is underway
in wet eucalypt formations, where fires have been reduced
and rainforests are invading [59]. However, continent-wide
replacement is unlikely because the rate of rainforest
colonization is slow in drier climates and on nutrient-poor
soils compared with the frequency of fires [59].

Recent molecular systematic studies have found
that the phylogenetic structure of legumes was caused
as much by ecological setting as by tectonic history
and the location of land masses [60]. The implication is
that speciation and subsequent dispersal of some taxa
are confined ecologically, within biomes, as much as by
dispersal problems across oceans. The appearance of
grass-fuelled fires might be an example of an intra-
continental ‘vicariance’ event, segregating taxa into
those that could tolerate fires and those that could
not. New molecular tools, coupled with focused
ecological studies, promise new insights into the
evolutionary history of those parts of the world
where fire uncouples biomes from their climate-limited
potential.

Conclusions

We have shown major similarities between fire and
herbivory and argued for a more-inclusive view of top-
down, or, in this instance, consumer control of biomes. We
believe that the global extent of fire as a consumer, its
many parallels with herbivory, its role in selecting for
particular plant traits and in the evolution of biomes, is
worthy of much wider attention from ecologists.

www.sciencedirect.com

There is an added incentive for greater understanding
of fire as a globally important consumer. Climate change,
habitat fragmentation, the unprecedented transport of
highly flammable plants to novel settings, and the
ubiquitous overlay of human impacts on fire regimes
demand a new level of synthetic understanding for our
peaceful coexistence with this charismatic beast.

Acknowledgements
We thank Jeremy Midgley, C.J. Fotheringham, Ian Woodward, Guy
Midgley, Ross Bradstock, David Keith, Dave Bowman, Malcolm Gill, Alan
Andersen and Herve Fritz for useful discussions. Jeremy Midgley and
anonymous reviewers provided useful comments on the article.

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http://www.sciencedirect.com

• Fire as a global ‘herbivore’: the ecology and evolution of flammable ecosystems

Parallels between fire and herbivory
Fire and consumer control of ecosystems
Fire and consumer control of species composition
The variable nature of fire as a consumer control
Fire and community assembly
Fire as an evolutionary agent
Fire and the origin of biomes
Conclusions
Acknowledgements
References

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Annu

.

Rev. Ecol. Syst. 1998. 29:207–31
Copyright c© 1998 by Annual Reviews. All rights reserved

ROADS AND THEIR MAJOR
ECOLOGICAL EFFECTS

Richard T. T. Forman and Lauren E. Alexander
Harvard University Graduate School of Design, Cambridge, Massachusetts 02138

KEY WORDS: animal movement, material flows, population effects, roadside vegetation,
transportation ecology

ABSTRACT
A huge road network with vehicles ramifies across the land, representing a sur-
prising frontier of ecology. Species-rich roadsides are conduits for few species.
Roadkills are a premier mortality source, yet except for local spots, rates rarely
limit population size. Road avoidance, especially due to traffic noise, has a greater
ecological impact. The still-more-important barrier effect subdivides populations,
with demographic and probably genetic consequences. Road networks crossing
landscapes cause local hydrologic and erosion effects, whereas stream networks
and distant valleys receive major peak-flow and sediment impacts. Chemical ef-
fects mainly occur near roads. Road networks interrupt horizontal ecological
flows, alter landscape spatial pattern, and therefore inhibit important interior
species. Thus, road density and network structure are informative landscape
ecology assays. Australia has huge road-reserve networks of native vegetation,
whereas the Dutch have tunnels and overpasses perforating road barriers to en-
hance ecological flows. Based on road-effect zones, an estimated 15–20% of the
United States is ecologically impacted by roads.

INTRODUCTION

Roads appear as major conspicuous objects in aerial views and photographs,
and their ecological effects spread through the landscape. Few environmental
scientists, from population ecologists to stream or landscape ecologists, recog-
nize the sleeping giant, road ecology. This major frontier and its applications to
planning, conservation, management, design, and policy are great challenges
for science and society.

207
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208 FORMAN & ALEXANDER

This review often refers to The Netherlands and Australia as world leaders
with different approaches in road ecology and to the United States for es-
pecially useful data. In The Netherlands, the density of main roads alone is
1.5 km/km2, with traffic density of generally between 10,000 and 50,000 ve-
hicles per commuter day (101). Australia has nearly 900,000 km of roads for
18 million people (66). In the United States, 6.2 million km of public roads
are used by 200 million vehicles (85). Ten percent of the road length is in
national forests, and one percent is interstate highways. The road density is
1.2 km/km2, and Americans drive their cars for about 1 h/day. Road density is
increasing slowly, while vehicle kilometers (miles) traveled (VMT) is growing
rapidly.

The term road corridor refers to the road surface plus its maintained roadsides
and any parallel vegetated strips, such as a median strip between lanes in a
highway (Figure 1; see color version at end of volume). “Roadside natural
strips” of mostly native vegetation receiving little maintenance and located
adjacent to roadsides are common in Australia (where road corridors are called
road reserves) (12, 39, 111). Road corridors cover approximately 1% of the
United States, equal to the area of Austria or South Carolina (85). However,
the area directly affected ecologically is much greater (42, 43).

Theory for road corridors highlights their functional roles as conduits, bar-
riers (or filters), habitats, sources, and sinks (12, 39). Key variables affecting
processes are corridor width, connectivity, and usage intensity. Network theory,
in turn, focuses on connectivity, circuitry, and node functions (39, 71).

This review largely excludes road-construction-related activities, as well
as affiliated road features such as rest stops, maintenance facilities, and en-
trance/exit areas. We also exclude the dispersed ecological effects of air pollu-
tion emissions, such as greenhouse gases, nitrogen oxides (NOX), and ozone,
which are reviewed elsewhere (85, 135). Bennett’s article (12) plus a series of
books (1, 21, 33, 111) provide overviews of parts of road ecology.

Gaping holes in our knowledge of road ecology represent research oppor-
tunities with a short lag between theory and application. Current ecological
knowledge clusters around five major topics: (a) roadsides and adjacent strips;
(b) road and vehicle effects on populations; (c) water, sediment, chemicals, and
streams; (d ) the road network; and (e) transportation policy and planning.

−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−−→
Figure 1 Road corridor showing road surface, maintained open roadsides, and roadside natural
strips. Strips of relatively natural vegetation are especially characteristic of road corridors (known
as road reserves) in Australia. Wheatbelt of Western Australia. Photo courtesy of BMJ Hussey.
See color version at end of volume.

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ROADS AND ECOLOGICAL EFFECTS 209

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210 FORMAN & ALEXANDER

ROADSIDE VEGETATION AND ANIMALS

Plants and Vegetation
“Roadside” or “verge” refers to the more-or-less intensively managed strip,
usually dominated by herbaceous vegetation, adjacent to a road surface
(Figure 1). Plants on this strip tend to grow rapidly with ample light and with
moisture from road drainage. Indeed, management often includes regular mow-
ing, which slows woody-plant invasion (1, 86). Ecological management may
also maintain roadside native-plant communities in areas of intensive agricul-
ture, reduce the invasion of exotic (non-native) species, attract or repel animals,
enhance road drainage, and reduce soil erosion.

Roadsides contain few regionally rare species but have relatively high plant
species richness (12, 139). Disturbance-tolerant species predominate, espe-
cially with intensive management, adjacent to highways, and exotic species
typically are common (19, 121). Roadside mowing tends to both reduce plant
species richness and favor exotic plants (27, 92, 107). Furthermore, cutting and
removing hay twice a year may result in higher plant species richness than does
mowing less frequently (29, 86). Native wildflower species are increasingly
planted in dispersed locations along highways (1).

Numerous seeds are carried and deposited along roads by vehicles (70, 112).
Plants may also spread along roads due to vehicle-caused air turbulence
(107, 133) or favorable roadside conditions (1, 92, 107, 121, 133). For exam-
ple, the short-distance spread of an exotic wetland species, purple loosestrife
(Lythrum salicaria), along a New York highway was facilitated by roadside
ditches, as well as culverts connecting opposite sides of the highway and the
median strip of vegetation (133). Yet few documented cases are known of
species that have successfully spread more than 1 km because of roads.

Mineral nutrient fertilization from roadside management, nearby agriculture,
and atmospheric NOX also alter roadside vegetation. In Britain, for example,
vegetation was changed for 100–200 m from a highway by nitrogen from traffic
exhaust (7). Nutrient enrichment from nearby agriculture enhances the growth
of aggressive weeds and can be a major stress on a roadside native-plant commu-
nity (19, 92). Indeed, to conserve roadside native-plant communities in Dutch
farmland, fertilization and importing topsoil are ending, and in some places
nutrient accumulations and weed seed banks are reduced by soil removal (86;
H van Bohemen, personal communication).

Woody species are planted in some roadsides to reduce erosion, control
snow accumulation, support wildlife, reduce headlight glare, or enhance aes-
thetics (1, 105). Planted exotic species, however, may spread into nearby natural
ecosystems (3, 12). For example, in half the places where non-native woody
species were planted in roadsides adjacent to woods in Massachusetts (USA),
a species had spread into the woods (42).

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ROADS AND ECOLOGICAL EFFECTS 211

Roadside management sometimes creates habitat diversity to maintain native
ecosystems or species (1, 86, 131). Mowing different sections along a road, or
parallel strips in wide roadsides, at different times or intervals may be quite ef-
fective (87). Ponds, wetlands, ditches, berms, varied roadside widths, different
sun and shade combinations, different slope angles and exposures, and shrub
patches rather than rows offer variety for roadside species richness.

In landscapes where almost all native vegetation has been removed for cul-
tivation or pasture, roadside natural strips (Figure 1) are especially valuable as
reservoirs of biological diversity (19, 66). Strips of native prairie along roads
and railroads, plus so-called beauty strips of woodland that block views near
intensive logging, may function similarly as examples. However, roadside nat-
ural strips of woody vegetation are widespread in many Australian agricultural
landscapes and are present in South Africa (11, 12, 27, 39, 66, 111). Overall,
these giant green networks provide impressive habitat connectivity and disperse
“bits of nature” widely across a landscape. Yet they miss the greater ecological
benefits typically provided by large patches of natural vegetation (39, 41).

In conclusion, roadside vegetation is rich in plant species, although appar-
ently not an important conduit for plants. The scattered literature suggests a
promising research frontier.

Animals and Movement Patterns
Mowing, burning, livestock grazing, fertilizing, and planting woody plants
greatly impact native animals in roadsides. Cutting and removing roadside veg-
etation twice a year in The Netherlands, compared with less frequent mowing,
results in more species of small mammals, reptiles, amphibians, and insects
(29, 86). However, mowing once every 3–5 y rather than annually results
in more bird nests. Many vertebrate species persist better with mowing af-
ter, rather than before or during, the breeding period (86, 87). The mowing
regime is especially important for insects such as meadow butterflies and moths,
where different species go through stages of their annual cycle at different times
(83). Roadsides, especially where mowed cuttings are removed, are suitable
for ∼80% of the Dutch butterfly fauna (86).

Planting several native and exotic shrub species along Indiana (USA)
highways resulted in higher species richness, population density, and nest den-
sity for birds, compared with nearby grassy roadsides (105). Rabbit (Sylvilagus)
density increased slightly. However, roadkill rates did not differ next to shrubby
versus grassy roadsides.

In general, road surfaces, roadsides, and adjacent areas are little used as
conduits for animal movement along a road (39), although comparisons with
null models are rare. For example, radiotracking studies of wildlife across the
landscape detect few movements along or parallel to roads (35, 39, 93). Some
exceptions are noteworthy. Foraging animals encountering a road sometimes

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212 FORMAN & ALEXANDER

move short distances parallel to it (10, 106). At night, many large predator
species move along roads that have little vehicular or people traffic (12, 39).
Carrion feeders move along roads in search of roadkills, and vehicles some-
times transport amphibians and other animals (11, 12, 32). Small mammals
have spread tens of kilometers along highway roadsides (47, 60). In addition,
migrating birds might use roads as navigational cues.

Experimental, observational, and modeling approaches have been used to
study beetle movement along roadsides in The Netherlands (125–127). On wide
roadsides, fewer animals disappeared into adjacent habitats. Also, a dense grass
strip by the road surface minimized beetle susceptibility to roadkill mortality
(126, 127). Long dispersals of beetles were more frequent in wide (15–25 m)
than in narrow (<12 m) roadsides. Nodes of open vegetation increased, and narrow bottlenecks decreased, the probability of long dispersals. The results suggest that with 20–30-m-wide roadsides containing a central suitable habitat, beetle species with poor dispersal ability and a good reproductive rate may move 1–2 km along roadsides in a decade (127).

Adjacent ecosystems also exert significant influences on animals in corridors
(39). For example, roadside beetle diversity was higher near a similar patch
of sandy habitat, and roadsides next to forest had the greatest number of forest
beetle species (127). In an intensive-agriculture landscape (Iowa, USA), bird-
nest predation in roadsides was highest opposite woods and lowest opposite
pastures (K Freemark, unpublished data). Finally, some roadside animals also
invade nearby natural vegetation (37, 47, 54, 60, 63, 127).

The median strip between lanes of a highway is little studied. A North
Carolina (USA) study found no difference in small-mammal density between
roadsides on the median and on the outer side of the highway (2). This result was
the same whether comparing mowed roadside areas or unmowed roadside areas.
Also, roadkill rates may be affected by the pattern of wooded and grassy areas
along median strips (10).

In conclusion, some species move significant distances along roadsides and
have major local impacts. Nevertheless, road corridors appear to be relatively
unimportant as conduits for species movement, although movement rates should
be better compared with those at a distance and in natural-vegetation corridors.

ROAD AND VEHICLE EFFECTS ON POPULATIONS

Roadkilled Animals
Sometime during the last three decades, roads with vehicles probably overtook
hunting as the leading direct human cause of vertebrate mortality on land. In
addition to the large numbers of vertebrates killed, insects are roadkilled in
prodigious numbers, as windshield counts will attest.

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ROADS AND ECOLOGICAL EFFECTS 213

Estimates of roadkills (faunal casualties) based on measurements in short
sections of roads tell the annual story (12, 39, 123): 159,000 mammals and
653,000 birds in The Netherlands; seven million birds in Bulgaria; five million
frogs and reptiles in Australia. An estimated one million vertebrates per day
are killed on roads in the United States.

Long-term studies of roadkills near wetlands illustrate two important pat-
terns. One study recorded>625 snakes and another>1700 frogs annually
roadkilled per kilometer (8, 54). A growing literature suggests that roads by
wetlands and ponds commonly have the highest roadkill rates, and that, even
though amphibians may tend to avoid roads (34), the greatest transportation
impact on amphibians is probably roadkills (8, 28, 34, 128).

Road width and vehicle traffic levels and speeds affect roadkill rates. Am-
phibians and reptiles tend to be particularly susceptible on two-lane roads with
low to moderate traffic (28, 34, 57, 67). Large and mid-sized mammals are espe-
cially susceptible on two-lane, high-speed roads, and birds and small mammals
on wider, high-speed highways (33, 90, 106).

Do roadkills significantly impact populations? Measurements of bird and
mammal roadkills in England illustrate the main pattern (56, 57). The house
sparrow (Passer domesticus) had by far the highest roadkill rate. Yet this species
has a huge population, reproduces much faster than the roadkill rate, and can
rapidly recolonize locations where a local population drops. The study con-
cluded, based on the limited data sets available, that none of the>100 bird and
mammal species recorded had a roadkill rate sufficient to affect population size
at the national level.

Despite this overall pattern, roadkill rates are apparently significant for a few
species listed as nationally endangered or threatened in various nations (∼9–
12 cases) (9, 39, 43; C Vos, personal communication). Two examples from
southern Florida (USA) are illustrative. The Florida panther (Felis concolor
coryi) had an annual roadkill mortality of approximately 10% of its population
before 1991 (33, 54). Mitigation efforts reduced roadkill loss to 2%. The key
deer (Odocoileus virginianus clavium) has an annual roadkill mortality of∼16%
of its population. Local populations, of course, may suffer declines where the
roadkill rate exceeds the rates of reproduction and immigration. At least a dozen
local-population examples are known for vertebrates whose total populations
are not endangered (33, 39, 43).

Vehicles often hit vertebrates attracted to spilled grain, roadside plants, in-
sects, basking animals, small mammals, road salt, or dead animals (12, 32, 56,
87). Roadkills may be frequent where traffic lanes are separated by imperme-
able barriers or are between higher roadside banks (10, 106).

Landscape spatial patterns also help determine roadkill locations and rates.
Animals linked to specific adjacent land uses include amphibians roadkilled

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214 FORMAN & ALEXANDER

near wetlands and turtles near open-water areas (8). Foraging deer are often
roadkilled between fields in forested landscapes, between wooded areas in
open landscapes, or by conservation areas in suburbs (10, 42, 106). The vicinity
of a large natural-vegetation patch and the area between two such patches are
likely roadkill locations for foraging or dispersing animals. Even more likely
locations are where major wildlife-movement routes are interrupted, such as
roads crossing drainage valleys in open landscapes or crossing railway routes
in suburbs (42, 106).

In short, road vehicles are prolific killers of terrestrial vertebrates. Neverthe-
less, except for a small number of rare species, roadkills have minimal effect
on population size.

Vehicle Disturbance and Road Avoidance
The ecological effect of road avoidance caused by traffic disturbance is probably
much greater than that of roadkills seen splattered along the road. Traffic noise
seems most important, although visual disturbance, pollutants, and predators
moving along a road are alternative hypotheses as the cause of avoidance.

Studies of the ecological effects of highways on avian communities in The
Netherlands point to an important pattern. In both woodlands and grasslands
adjacent to roads, 60% of the bird species present had a lower density near a
highway (102, 103). In the affected zone, the total bird density was approxi-
mately one third lower, and species richness was reduced as species progres-
sively disappeared with proximity to the road. Effect-distances (the distance
from a road at which a population density decrease was detected) were greatest
for birds in grasslands, intermediate for birds in deciduous woods, and least for
birds in coniferous woods.

Effect-distances were also sensitive to traffic density. Thus, with an average
traffic speed of 120 km/h, the effect-distances for the most sensitive species
(rather than for all species combined) were 305 m in woodland by roads with
a traffic density of 10,000 vehicles per day (veh/day) and 810 m in woodland by
50,000 veh/day; 365 m in grassland by 10,000 veh/day and 930 m in grassland
by 50,000 veh/day (101–103). Most grassland species showed population de-
creases by roads with 5000 veh/day or less (102). The effect-distances for both
woodland and grassland birds increased steadily with average vehicle speed
up to 120 km/h and also with traffic density from 3000 to 140,000 veh/day
(100, 102, 103). These road effects were more severe in years when overall bird
population sizes were low (101).

Songbirds appear to be sensitive to remarkably low noise levels, similar to
those in a library reading room (100, 102, 103). The noise level at which popula-
tion densities of all woodland birds began to decline averaged 42 decibels (dB),
compared with an average of 48 dB for grassland species. The most sensitive

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ROADS AND ECOLOGICAL EFFECTS 215

woodland species (cuckoo) showed a decline in density at 35 dB, and the
most sensitive grassland bird (black-tailed godwit,Limosa limosa) responded
at 43 dB. Field studies and experiments will help clarify the significance of
these important results for traffic noise and birds.

Many possible reasons exist for the effects of traffic noise. Likely hypotheses
include hearing loss, increase in stress hormones, altered behaviors, interfer-
ence with communication during breeding activities, differential sensitivity to
different frequencies, and deleterious effects on food supply or other habitat at-
tributes (6, 101, 103, 130). Indeed, vibrations associated with traffic may affect
the emergence of earthworms from soil and the abundance of crows (Corvus)
feeding on them (120). A different stress, roadside lighting, altered nocturnal
frog behavior (18). Responses to roads with little traffic may resemble behav-
ioral responses to acute disturbances (individual vehicles periodically passing),
rather than the effects of chronic disturbance along busy roads.

Response to traffic noise is part of a broader pattern of road avoidance by
animals. In the Dutch studies, visual disturbance and pollutants extended out-
ward only a short distance compared with traffic noise (100, 103). However,
visual disturbance and predators moving along roads may be more significant
by low-traffic roads.

Various large mammals tend to have lower population densities within
100–200 m of roads (72, 93, 108). Other animals that seem to avoid roads in-
clude arthropods, small mammals, forest birds, and grassland birds (37, 47, 73,
123). Such road-effect zones, extending outward tens or hundreds of meters
from a road, generally exhibit lower breeding densities and reduced species rich-
ness compared with control sites (32, 101). Considering the density of roads
plus the total area of avoidance zones, the ecological impact of road avoidance
must well exceed the impact of either roadkills or habitat loss in road corridors.

Barrier Effects and Habitat Fragmentation
All roads serve as barriers or filters to some animal movement. Experiments
show that carabid beetles and wolf spiders (Lycosa) are blocked by roads as
narrow as 2.5 m wide (73), and wider roads are significant barriers to crossing for
many mammals (11, 54, 90, 113). The probability of small mammals crossing
lightly traveled roads 6–15 m wide may be<10% of that for movements within adjacent habitats (78, 119). Similarly, wetland species, including amphibians and turtles, commonly show a reduced tendency to cross roads (34, 67).

Road width and traffic density are major determinants of the barrier effect,
whereas road surface (asphalt or concrete versus gravel or soil) is generally
a minor factor (34, 39, 73, 90). Road salt appears to be a significant deterrent
to amphibian crossing (28, 42). Also, lobes and coves in convoluted outer-
roadside boundaries probably affect crossing locations and rates (39).

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216 FORMAN & ALEXANDER

The barrier effect tends to create metapopulations, e.g. where roads divide
a large continuous population into smaller, partially isolated local populations
(subpopulations) (6, 54, 128). Small populations fluctuate more widely over
time and have a higher probability of extinction than do large populations
(1, 88, 115, 122, 123). Furthermore, the recolonization process is also blocked
by road barriers, often accentuated by road widening or increases in traffic. This
well-known demographic threat must affect numerous species near an extensive
road network, yet is little studied relative to roads (6, 73, 98).

The genetics of a population is also altered by a barrier that persists over many
generations (73, 115). For instance, road barriers altered the genetic structure
of small local populations of the common frog (Rana temporaria) in Germany
by lowering genetic heterozygosity and polymorphism (97, 98). Other than the
barrier effect on this amphibian and roadkill effects on two southern Florida
mammals (20, 54), little is known of the genetic effects of roads.

Making roads more permeable reduces the demographic threat but at the
cost of more roadkills. In contrast, increasing the barrier effect of roads re-
duces roadkills but accentuates the problems of small populations. What is the
solution to this quandary (122, 128)? The barrier effect on populations proba-
bly affects more species, and extends over a wider land area, than the effects
of either roadkills or road avoidance. This barrier effect may emerge as the
greatest ecological impact of roads with vehicles. Therefore, perforating roads
to diminish barriers makes good ecological sense.

WATER, SEDIMENT, CHEMICALS, STREAMS,
AND ROADS

Water Runoff
Altering flows can have major physical or chemical effects on aquatic ecosys-
tems. The external forces of gravity and resistance cause streams to carve chan-
nels, transport materials and chemicals, and change the landscape (68). Thus,
water runoff and sediment yield are the key physical processes whereby roads
have an impact on streams and other aquatic systems, and the resulting effect-
distances vary widely (Figure 2).

Roads on upper hillslopes concentrate water flows, which in turn form chan-
nels higher on slopes than in the absence of roads (80). This process leads to
smaller, more elongated first-order drainage basins and a longer total length
of the channel network. The effects of stream network length on erosion and
sedimentation vary with both scale and drainage basin area (80).

Water rapidly runs off relatively impervious road surfaces, especially in storm
and snowmelt events. However, in moist, hilly, and mountainous terrain, such

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ROADS AND ECOLOGICAL EFFECTS 217

Figure 2 Road-effect zone defined by ecological effects extending different distances from a road.
Most distances are based on specific illustrative studies (39); distance to left is arbitrarily half of
that to right. (P) indicates an effect primarily at specific points. From Forman et al (43).

runoff is often insignificant compared with the conversion of slow-moving
groundwater to fast-moving surface water at cutbanks by roads (52, 62, 132).
Surface water is then carried by roadside ditches, some of which connect di-
rectly to streams while others drain to culverts with gullies incised below their
outlets (132). Increased runoff associated with roads may increase the rates and
extent of erosion, reduce percolation and aquifer recharge rates, alter channel
morphology, and increase stream discharge rates (13, 14). Peak discharges or

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218 FORMAN & ALEXANDER

floods then restructure riparian areas by rearranging channels, logs, branches,
boulders, fine-sediment deposits, and pools.

In forests, the combination of logging and roads increases peak discharges
and downstream flooding (62, 132). Forest removal results in lower evapo-
transpiration and water-storage capabilities, but roads alone may increase peak
discharge rates (62). Also, flood frequency apparently correlates with the per-
centage of road cover in a basin (52, 62, 110).

Roads may alter the subsurface flow as well as the surface flow on wetland
soils (116). Compacted saturated or nearly saturated soils have limited perme-
ability and low drainage capacity. Wetland road crossings often block drainage
passages and groundwater flows, effectively raising the upslope water table and
killing vegetation by root inundation, while lowering the downslope water table
with accompanying damage to vegetation (116, 118).

Streams may be altered for considerable distances both upstream and down-
stream of bridges. Upstream, levees or channelization tend to result in reduced
flooding of the riparian zone, grade degradation, hydraulic structural problems,
and more channelization (17). Downstream, the grade change at a bridge results
in local scouring that alters sedimentation and deposition processes (17, 49).
Sediment and chemicals enter streams where a road crosses, and mathemati-
cal models predict sediment loading in and out of reaches affected by stream
crossings (5). The fixed stream (or river) location at a bridge or culvert re-
duces both the amount and variability of stream migration across a floodplain.
Therefore, stream ecosystems have altered flow rates, pool-riffle sequences, and
scour, which typically reduce habitat-forming debris and aquatic organisms.

Sediment
The volume of sediment yield from a road depends on sediment supply and
transport capacity (5). Sediment yield is determined by road geometry, slope,
length, width, surface, and maintenance (5, 51), in addition to soil properties and
vegetation cover (59). Road surfaces, cutbanks, fillslopes, bridge/culvert sites,
and ditches are all sources of sediment associated with roads. The exposed
soil surfaces, as well as the greater sediment-transport capacity of increased
hydrologic flows, result in higher erosion rates and sediment yields (99).

Road dust as a little-studied sediment transfer may directly damage vegeta-
tion, provide nutrients for plant growth, or change the pH and vegetation (109).
Effect-distances are usually<10–20 m but may extend to 200 m downwind (Figure 2). In arid land, soil erosion and drainage are common road problems (61).

Arctic roads also are often sources of dust. Other ecological issues include
change in albedo, flooding, erosion and thermokarst, weed migration, waterfowl
and shorebird habitat, and altered movement of large mammals (129).

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ROADS AND ECOLOGICAL EFFECTS 219

Landsliding or mass wasting associated with roads may be a major sediment
source (13, 117). Some of this sediment accumulates on lower slopes and is
subject to subsequent erosion. The rest reaches floodplains or streams, where it
alters riparian ecosystems, channel morphology, or aquatic habitat. Although
gradual sediment transport and episodic landslides are natural processes affect-
ing streams, elevated levels caused by roads tend to disrupt aquatic ecosystems.
Indeed, logging roads commonly produce more erosion and sediment yield,
particularly by mass wasting, than do the areas logged (45, 51, 104, 117).

Buffer strips between roads and streams tend to reduce sediments reaching
aquatic ecosystems (77, 91). Buffers may be less effective for landslides than
for arresting water and sediment from culverts and roadside ditches. Good
road location (including avoiding streamsides and narrow floodplains for many
ecological reasons), plus good ecological design of roadsides relative to slope,
soil, and hydrology, may be a better strategy than depending on wide buffers to
absorb sediment.

Water from road ditches tends to deposit finer sediment in streams, whereas
landslides generally provide coarser material. Fine sediment increases turbidity
(51), which disrupts stream ecosystems in part by inhibiting aquatic plants,
macro-invertebrates, and fish (14, 16, 31, 99). Coarse deposits such as logs and
boulders help create deep pools and habitat heterogeneity in streams. During
low-flow periods, fine-sediment deposits tend to fill pools and smooth gravel
beds, hence degrading habitats and spawning sites for key fish (13, 14, 31).
During high-discharge events, accumulated sediment tends to be flushed out
and redeposited in larger water bodies.

In short, roads accelerate water flows and sediment transport, which raise
flood levels and degrade aquatic ecosystems. Thus, local hydrologic and erosion
effects along roads are dispersed across the land, whereas major impacts are
concentrated in the stream network and distant valleys.

Chemical Transport
Most chemical transport from roads occurs in stormwater runoff through or
over soil. Runoff pollutants alter soil chemistry, may be absorbed by plants,
and affect stream ecosystems, where they are dispersed and diluted over con-
siderable distances (16, 50, 66, 137, 138). Deicing salt and heavy metals are the
two main categories of pollutants studied in road runoff.

The primary deicing agent, NaCl, corrodes vehicles and bridges, contami-
nates drinking water supplies, and is toxic to many species of plants, fish, and
other aquatic organisms (4, 16, 84). Calcium magnesium acetate (CMA) is a
more effective deicer, less corrosive, less mobile in soil, biodegradable, and
less toxic to aquatic organisms (4, 84, 89). Also, CaCl used to decrease dust
may inhibit amphibian movement (28).

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220 FORMAN & ALEXANDER

Airborne NaCl from road snowplowing may cause leaf injury to trees (e.g.
Pinus strobus) up to 120 m from a road, especially downwind and downslope
(58, 84). Trees seem to be more sensitive to chloride damage than are common
roadside shrubs and grasses. Sodium accumulation in soils, mainly within 5 m
of a road, alters soil structure, which affects plant growth (84). Road salt has
facilitated the spread of three coastal exotic plants as much as 150 km in The
Netherlands (1).

Deicing agents tend to increase the mobility of chemical elements in soil, such
as heavy metals (by NaCl) and Na, Cl, Ca, and Mg (by CMA) (4). This process
facilitates contamination of groundwater, aquifers, and streams. Because of
dilution, the chemical effects of road runoff on surface water ecosystems may
be primarily confined to small streams, particularly where they run adjacent to
roads (36, 84).

Heavy metals are relatively immobile and heterogeneously distributed in
roadsides, especially due to drainage ditch flows (15, 55, 80). Soils adjacent to
the road surface typically contain the greatest mass (136). Elevated concentra-
tions in grass tissue may occur within 5–8 m of a road, although high lead levels
have been found in soil out to 25 m (30, 65, 82). Elevated lead concentrations
were found in tissue of several small-mammal species in a narrow zone by
roads, with higher lead levels by busy roads (48).

Highway roadsides of 5–15-m width next to traffic densities of 11,000–
124,000 veh/day in The Netherlands had somewhat higher heavy-metal accu-
mulations on the downwind side, but no correlation with traffic density was
found (30). All average levels of Pb, Cd, Zn, Cr, Ni, and As in cut grass
(hay) from these roadsides were below the Dutch maximum-acceptable-levels
for livestock fodder and “clean compost.” Only Zn in some roadsides studied
exceeded the maximum for “very clean compost.”

Many other chemicals enter roadsides. Herbicides often kill non-target plants,
particularly from blanket applications in drifting air. For polycyclic aromatic
hydrocarbons from petroleum (136), the preliminary conclusion for the Dutch
highways was that levels in roadside hay “do not seem to give cause for alarm”
(30). Fertilizer nutrients affect roadside vegetation (19, 86, 92), and nitrogen
from vehicular NOX emission altered vegetation up to 100–200 m from a high-
way in Britain (7). Acidic road runoff may have impacts on stream ecosystems
(36, 81). Of the hazardous materials transported on roads, e.g.>500,000 ship-
ments moved each day in the United States, a small fraction is spilled, although
occasional large spills cause severe local effects (85).

Typical water-quality responses to road runoff include altered levels of heavy
metal, salinity, turbidity, and dissolved oxygen (16, 23, 81). However, these
water-quality changes, even in a wetland, tend to be temporary and localized
due to fluctuations in water quantity (23). Road runoff is a major source of heavy

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ROADS AND ECOLOGICAL EFFECTS 221

metals to stream systems, especially Pb, Zn, Cu, Cr, and Cd (16, 50, 64, 137).
Fish mortality in streams has been related to high concentrations of Al, Mn, Cu,
Fe, or Zn, with effects on populations reorded as far as 8 km downstream (81).
Both high traffic volume and high metal concentration in runoff have correlated
with mortality of fish and other aquatic organisms (59). Floodplain soil near
bridges may have high heavy-metal concentrations (138). Although highway
runoff generally has little adverse effect on vegetation or plant productivity, it
may change the species composition of floodplain plant communities, favoring
common species (138).

Overall, terrestrial vegetation seems to be more resistant than aquatic organ-
isms to road impacts (59, 138). Drainage of road runoff through grassy channels
greatly reduces toxic solid- and heavy-metal concentrations (59). Furthermore,
dense vegetation increases soil infiltration and storage. Therefore, instead of
expensive detention ponds and drainage structures to reduce runoff impacts,
creative grassland designs by roads, perhaps with shrubs, may provide both
sponge and biodiversity benefits. The wide range of studies cited above lead to
the conclusion that chemical impacts tend to be localized near roads.

THE ROAD NETWORK

New Roads and Changing Landscape Pattern
Do roads lead to development, or does development lead to roads? This timeless
debate in the transportation community has greater ramifications as environ-
mental quality becomes more important in the transportation–land-use inter-
action (114). For example, new roads into forested landscapes often lead to
economic development as well as deforestation and habitat fragmentation (22).

At the landscape scale, the major ecological impacts of a road network are the
disruption of landscape processes and loss of biodiversity. Interrupting hori-
zontal natural processes, such as groundwater flow, streamflow, fire spread, for-
aging, and dispersal, fundamentally alters the way the landscape works (40, 53).
It truncates flows and movements, and reduces the critical variability in natural
processes and disturbances. Biodiversity erodes as the road network impacts
interior species, species with large home ranges, stream and wetland species,
rare native species, and species dependent on disturbance and horizontal flows.

A new road system in the Rondonia rainforest of Brazil illustrates such
effects. By 1984, road construction and asphalt paving had stimulated a major
influx of people and forest clearing (25). A regular pattern of primary roads
plus parallel secondary roads 4 km apart was imposed on the land. Commonly,
small forest plots of∼5 ha were gradually converted to grass and joined with
neighboring plots to form large pastures (26). Simulation models of this typical
scenario were compared with models of a worst-case scenario and an “innovative

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222 FORMAN & ALEXANDER

farming” scenario with perennial crops and essentially no fire or cattle (25, 26).
Species requiring a large area and having poor “gap-crossability” disappeared
in all model scenarios after road construction. Species with moderate require-
ments for area and gap-crossability persisted only in the innovative farming
scenario. Reestablishing the connectivity of nature with a network of wildlife
corridors was proposed as a solution to maintain the first group of species,
which are of conservation importance.

The closure and removal of some roads in the grid is an alternative ecological
approach. Natural landscape processes and biodiversity are both inhibited by
a rigid road grid. Closing and eliminating some linkages would permit the
reestablishment of a few large patches of natural rainforest (39, 41, 43). Such a
solution helps create a road network with a high variance in mesh size. Large
natural-vegetation patches in areas remote from both roads and people are
apparently required to sustain important species such as wolf, bear, and probably
jaguar (Canis lupus, Ursus, Felis onca) (39, 76). Temporary road closures could,
for example, enhance amphibian migration during the breeding phase (67). In
contrast, road closure and removal could eliminate motorized vehicle use, thus
reducing numerous disturbance effects on natural populations and ecosystems.

A general spatial-process model emphasizes that roads have the greatest eco-
logical impact early in the process of land transformation (39, 41). They dissect
the land, leading to habitat fragmentation, shrinkage, and attrition. Forest road
networks may also create distinctive spatial patterns, such as converting con-
voluted to rectilinear shapes, decreasing core forest area, and creating more
total edge habitat by roads than by logged areas (79, 96).

Forest roads as a subset of roads in general are characterized as being narrow,
not covered with asphalt, lightly traveled, and remote (98). Among the wide
range of ecological effects of roads (39, 95), forest roads have a distinctive set
of major ecological effects: (a) habitat loss by road construction, (b) altered
water routing and downstream peak flows, (c) soil erosion and sedimentation
impacts on streams, (d ) altered species patterns, and (e) human access and
disturbance in remote areas (43, 45, 62, 132). Thus, an evaluation of logging
regimes includes the ecological effects of both the road network and forest spa-
tial patterns (45, 69). In conclusion, a road network disrupts horizontal natural
processes, and by altering both landscape spatial pattern and the processes, it
reduces biodiversity.

Road Density
Road density, e.g. measured as km/km2, has been proposed as a useful, broad in-
dex of several ecological effects of roads in a landscape (39, 43, 44, 95). Effects
are evident for faunal movement, population fragmentation, human access,
hydrology, aquatic ecosystems, and fire patterns.

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A road density of approx. 0.6 km/km2 (1.0 mi/mi2) appears to be the maxi-
mum for a naturally functioning landscape containing sustained populations of
large predators, such as wolves and mountain lions (Felis concolor) (43, 76, 124).
Moose (Alces), bear (Ursus) (brown, black, and grizzly), and certain other pop-
ulations also decrease with increasing road density (11, 43, 72). These species
are differentially sensitive to the roadkill, road-avoidance, and human-access
dimensions of road density. Species that move along, rather than across, roads
presumably are benefitted by higher road density (12, 39).

Human access and disturbance effects on remote areas tend to increase with
higher road density (39, 72, 76). Similarly, human-caused fire ignitions and
suppressions may increase, and average fire sizes decrease (111).

Aquatic ecosystems are also affected by road density. Hydrologic effects,
such as altered groundwater conditions and impeded drainage upslope, are
sensitive to road density (116, 118). Increased peak flows in streams may be
evident at road densities of 2–3 km/km2 (62). Detrimental effects on aquatic
ecosystems, based on macro-invertebrate diversity, were evident where roads
covered 5% or more of a watershed in California (75). In southeastern Ontario,
the species richness of wetland plants, amphibians/reptiles, and birds each
correlated negatively with road density within 1–2 km of a wetland (38).

Road density is an overall index that averages patterns over an area. Its
effects probably are sensitive to road width or type, traffic density, network
connectivity, and the frequency of spur roads into remote areas. Thus network
structure, or an index of variance in mesh size, is also important in understanding
the effect of road density (39, 76, 79, 96). Indeed, although road density is a
useful overall index, the presence of a few large areas of low road density may
be the best indicator of suitable habitat for large vertebrates and other major
ecological values.

TRANSPORTATION POLICY AND PLANNING

Environmental Policy Dimensions
Ecological principles are increasingly important in environmental transporta-
tion policy, and Australia, The Netherlands, and the United States highlight
contrasting approaches. Australian policy has focused on biodiversity, includ-
ing wildflower protection. An enormous network of road reserves with natural-
vegetation strips 10–200 m wide (Figure 1) stands out in many agricultural land-
scapes (66, 92, 111). Public pressure helped create this system, which “helps
to prevent soil erosion” and “where wildflowers can grow and flourish in
perpetuity” (111). Diverse experimental management approaches involve burn-
ing, weed control, planting native species, and nature restoration. Ecological

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224 FORMAN & ALEXANDER

scientists commonly work side by side with civil engineers in transportation
departments at all levels of government.

In contrast, Dutch policy has focused on the open roadside vegetation, road-
kills, animal movement patterns, and nature restoration (1, 21, 29, 86). This
approach reflects the stated national objectives of (a) recreating “nature, in-
cluding natural processes and biodiversity; and (b) enhancing the national eco-
logical network,” mainly composed of large natural-vegetation patches and
major wildlife and water corridors (1, 46). An impressive series of mitigation
overpasses, tunnels, and culverts provide for animal and water movement where
interrupted by road barriers (24, 43, 44, 86). Environmental activities in trans-
portation revolve around a group of environmental scientists in the national
Ministry of Transport who work closely with engineers and policy-makers at
both local and national levels.

In the United States, environmental transportation policy focuses on vehic-
ular pollutants, as well as engineering solutions for soil erosion and sedimen-
tation (85). A few states have built wildlife underpasses and overpasses to
address local roadkill or wildlife movement concerns. A 1991 federal law—the
Intermodal Surface Transportation Efficiency Act (ISTEA)—establishes policy
for a transportation system that is “economically efficient and environmentally
sound,” considers the “external benefits of reduced air pollution, reduced traffic
congestion and other aspects of the quality of life,” considers transportation
in a region-wide metropolitan area, and links ecological attributes with the
aesthetics of a landscape (43). Thus, US transportation policy largely ignores
biodiversity loss, habitat fragmentation, disruption of horizontal natural pro-
cesses, natural stream and wetland hydrology, streamwater chemistry, and re-
duction of fish populations, a range of ecological issues highlighted in the
transportation community in 1997 (85).

Of course, many nations use ecological principles in designing transporta-
tion systems (21, 33, 63, 95, 134), and environmental scientists, engineers, and
policy-makers in Europe have united to “conserve biodiversity and reduce
. . .fauna casualties” at the international level (21; H van Bohemen, personal
communication). The successful removal of lead from petroleum led to less
lead in roadside ecosystems worldwide. Nevertheless, the huge Australian road-
reserve system and the Dutch mitigation system for animal and water flows are
especially ambitious and pioneering.

Spatial Planning and Mitigation
Most existing roads were built before the explosion in ecological knowledge,
and many are poorly located ecologically (43). Yet the Dutch have developed
a promising transportation planning process for the movement of both people
and natural processes across the land (40, 44, 86). In essence, the ecological
network, consisting of large natural-vegetation patches plus major corridors for

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ROADS AND ECOLOGICAL EFFECTS 225

water and wildlife movement, is mapped. The road network is then superim-
posed on the ecological network to identify bottlenecks. Finally, mitigation
or compensation techniques are applied to eliminate designated percentages of
bottlenecks in a time sequence. The earlier such spatial planning begins, the
greater its effect (41).

Compensation is proposed where bottlenecks apparently cannot be over-
come by mitigation. The principle of no-net-loss has been used internationally
for wetlands with varying success (46, 94), whereas no-net-loss of natural pro-
cesses and biodiversity by roads is a concept only beginning to be applied
(1, 24, 46, 86). The loss, e.g. of biodiversity or groundwater flow, is compen-
sated by increasing an equivalent ecological value nearby. Options include
protection of an equivalent amount of high-quality habitat, reestablishment of
another wildlife corridor, or creation of new habitat. Mitigation, on the other
hand, attempts to minimize detrimental ecological impacts and is illustrated by
the varied wildlife passages (tunnels, pipes, underpasses, overpasses) operating
for animal movement (21, 86).

Diverse tunnel designs focus on small and mid-sized animals. Amphibian
tunnels, generally 30–100 cm wide and located where roads block movement
to breeding ponds or wetlands, are widespread in Europe and rare in the United
States (33, 43, 67). “Ecopipes,” or badger tunnels, are pipes∼40 cm in diameter
mainly designed for movement of mid-sized mammals across Dutch roads, and
located where water can rarely flow through (9, 44, 46, 86). In contrast, Dutch
wildlife culverts are∼120 cm wide, with a central channel for water flow
between two raised 40-cm-wide paths for animal movement. “Talus tunnels”
are designed for a mid-sized mammal that lives and moves in rock talus slopes
in Australia (74).

Wildlife underpasses, generally 8–30 m wide and at least 2.5 m high, have
been built for large mammals in southern Florida and scattered locations
elsewhere in the United States, Canada, and France (33, 43, 54, 63, 93, 113).
Wildlife overpasses, also designed for large mammals, range in width up to
200 m and are scarce: approximately 6 in North America (New Jersey, Utah,
Alberta, British Columbia) (33, 43; BF Leeson, personal communication) and
17 in Europe (Germany, France, The Netherlands, Switzerland) (44, 46, 63, 86,
134). The minimum widths for effectiveness may be 30–50 m in the center
and 50–80 m on the ends (33, 46, 86). The two Swiss overpasses of 140-m and
200-m width remind us that ultimately the goal should be “landscape connec-
tors” that permit all horizontal natural processes to cross roads (43, 44).

These mitigation structures are normally combined with fencing and vegeta-
tion to enhance animal crossing (86). Almost all such passages are successful
in that the target species crosses at least occasionally, and most are used by
many other species. Florida underpasses are used by the Florida panther (Felis
concolor coryi), nearly the whole local terrestrial fauna, and groundwater as

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226 FORMAN & ALEXANDER

well (33, 54). Underpasses and overpasses are used by almost all large mammal
species of a region. Yet, little information exists on crossing rates relative to
population sizes, movement rates away from roads, predation rates, home range
locations, and so forth. Nevertheless, mitigation passages are effective in per-
forating road barriers to maintain horizontal natural processes across the land.

The Road-Effect Zone
Roads and roadsides cover 0.9% of Britain and 1.0% of the United States,
while road reserves (Figure 1) cover 2.5% of the State of Victoria, Australia
(12, 85, 131). Yet how much of the land is ecologically impacted by roads with
vehicles?

The road-effect zone is the area over which significant ecological effects
extend outward from a road and typically is many times wider than the road
surface plus roadsides (Figure 2) (39, 95, 101, 134). The zone is asymmetric
with convoluted boundaries, reflecting the sequence of ecological variables,
plus unequal effect-distances due to slope, wind, and habitat suitability on
opposite sides of a road (40, 43). Knowing the average width of the road-effect
zone permits us to estimate the proportion of the land ecologically affected by
roads (43). For example, based on the traffic noise effect-distances of sensi-
tive bird species described above, road-effect zones cover∼10–20% of The
Netherlands (101).

Finally, a preliminary calculation for the United States was made based on
nine water and species variables in Massachusetts (USA), plus evidence from
the Dutch studies (42). An estimated 15–20% of the US land area is directly
affected ecologically by roads. These estimates reemphasize the immensity and
pervasiveness of ecological road impacts. Moreover, they challenge science and
society to embark on a journey of discovery and solution.

ACKNOWLEDGMENTS

Virginia H Dale, Robert D Deblinger, Malcolm L Hunter, Jr, and Julia A Jones
provided terrific reviews, which we deeply appreciate.

Visit the Annual Reviews home pageat
http://www.AnnualReviews.org

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Annual Review of Ecology and Systematics
Volume 29, 1998

CONTENTS
MOLECULAR TRANS-SPECIES POLYMORPHISM, Jan Klein, Akie
Sato, Sandra Nagl, Colm O’hUigín 1

PRINCIPLES OF PHYLOGEOGRAPHY AS ILLUSTRATED BY
FRESHWATER AND TERRESTRIAL TURTLES IN THE
SOUTHEASTERN UNITED STATES, DeEtte Walker, John C. Avise 23

THE FUNCTIONAL SIGNIFICANCE OF THE HYPORHEIC ZONE
IN STREAMS AND RIVERS, Andrew J. Boulton, Stuart Findlay,
Pierre Marmonier, Emily H. Stanley, H. Maurice Valett 59

ENDANGERED MUTUALISMS: The Conservation of Plant-Pollinator
Interactions, Carol A. Kearns, David W. Inouye, Nickolas M. Waser 83

THE ROLE OF INTRODUCED SPECIES IN THE DEGRADATION
OF ISLAND ECOSYSTEMS: A Case History of Guam, Thomas H.
Fritts, Gordon H. Rodda 113

EVOLUTION OF HELPING BEHAVIOR IN COOPERATIVELY
BREEDING BIRDS, Andrew Cockburn 141

THE ECOLOGICAL EVOLUTION OF REEFS, Rachel Wood 179

ROADS AND THEIR MAJOR ECOLOGICAL EFFECTS, Richard T.
T. Forman, Lauren E. Alexander 207

SEX DETERMINATION, SEX RATIOS, AND GENETIC
CONFLICT, John H. Werren, Leo W. Beukeboom 233

EARLY EVOLUTION OF LAND PLANTS: Phylogeny, Physiology,
and Ecology of the Primary Terrestrial Radiation, Richard M.
Bateman, Peter R. Crane, William A. DiMichele, Paul R. Kenrick, Nick
P. Rowe, Thomas Speck, William E. Stein 263

POSSIBLE LARGEST-SCALE TRENDS IN ORGANISMAL
EVOLUTION: Eight “Live Hypotheses”, Daniel W. McShea 293

FUNGAL ENDOPHYTES: A Continuum of Interactions with Host
Plants, K. Saikkonen, S. H. Faeth, M. Helander, T. J. Sullivan 319

FLORAL SYMMETRY AND ITS ROLE IN PLANT-POLLINATOR
SYSTEMS: Terminology, Distribution, and Hypotheses, Paul R. Neal,
Amots Dafni, Martin Giurfa 345

VERTEBRATE HERBIVORES IN MARINE AND TERRESTRIAL
ENVIRONMENTS: A Nutritional Ecology Perspective, J. H. Choat, K.
D. Clements 375

CARBON AND CARBONATE METABOLISM IN COASTAL
AQUATIC ECOSYSTEMS, J.-P. Gattuso, M. Frankignoulle, R.
Wollast 405

THE SCIENTIFIC BASIS OF FORESTRY, David A. Perry 435

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PATHWAYS, MECHANISMS, AND RATES OF POLYPLOID
FORMATION IN FLOWERING PLANTS, Justin Ramsey, Douglas
W. Schemske 467

BACTERIAL GROWTH EFFICIENCY IN NATURAL AQUATIC
SYSTEMS, Paul A. del Giorgio, Jonathan J. Cole 503

THE CHEMICAL CYCLE AND BIOACCUMULATION OF
MERCURY, François M. M. Morel, Anne M. L. Kraepiel, Marc Amyot 543

PHYLOGENY OF VASCULAR PLANTS, James A. Doyle 567

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