Articles
“The Calamity of So Long Life”:
Life Histories, Contaminants,
and Potential Emerging Threats
to Long-lived Vertebrates
CHRISTOPHER L. ROWE
Persistent contaminants are ubiquitous in the environment, often present at concentrations that may jeopardize reproductive fitness only after long
periods of exposure. As the duration of exposure is largely regulated by life span, long-lived species of high trophic status, such as many reptiles, birds,
and mammals, may be at risk of reduced fitness and population decline. Delayed maturation and iteroparity confer the potential for cumulative
effects to be expressed prior to reproduction, and large parental investments in yolk and milk may threaten offspring because of exposure during
critical developmental periods. Long generation times may delay emergence of obvious effects on populations, perhaps eluding early intervention,
while constraining rates at which populations may recover if conditions subsequently improve. Life history theory thus suggests that the suite of traits
that optimized reproductive fitness throughout long-lived species’ evolutionary histories may ultimately put them in peril in the modern,
anthropogenically altered environment.
Keywords: birds of prey, marine mammals, persistent contaminants, reproductive fitness, reptiles
R
eproductive fitness, the probability that an individual will produce viable offspring of equal or greater
replacement value, provides an avenue by which environmental influences on individuals transcend to population
and higher-order phenomena. Life history strategies represent
the suite of traits that optimize reproductive fitness of individuals within the environmental and phylogenetic constraints imposed on the species during its evolutionary history.
A species’ life history strategy is shaped by trade-offs among
alternative, attainable expressions of traits, including (but
not limited to) life span; time to maturation; number of lifetime breeding events; and parental investment in offspring
number, size or energy content, or care (Stearns 1989, Congdon et al. 2001). The combination of traits and the specific
expressions that optimize fitness are unique to a species, yet
have classically been broadly categorized as “r-selected” or
“K-selected” life history strategies (Pianka 1970). While these
categorizations represent theoretical extremes bounding a
vast diversity of trait combinations found in nature, and
thus no realized life history strategy represents an “ideal”
r- or K-selected strategy, the categorizations are conceptually
useful for broadly assessing potential environmental influences
on fitness resulting from a particular life history strategy.
Species that possess r-selected strategies tend to be opportunistic, short-lived, rapid to mature, and semelparous
(breeding only once), producing large numbers of small off-
www.biosciencemag.org
spring. Such species typically exploit rapidly changing environmental conditions and are capable of establishing populations quickly in novel environments (Pianka 1970). Relative
to r-selected species, those employing K-selected life history
strategies are long-lived, slow to mature, iteroparous (breeding multiple times), and typically make large per capita
parental investments in few offspring during a given breeding event (Pianka 1970). This life history strategy is optimized for environmental conditions that on average are
relatively stable over an individual’s long reproductive lifetime,
yet it is resilient to periodic, stochastic fluctuations in habitat quality that may compromise a given breeding event (a
“bet-hedging” strategy; Stearns 1976).
Persistent, synthetic contaminants were developed and
have been used widely only in the past century, yet already their
use has resulted in contamination of habitats on a global
scale (e.g., Schwarzenbach et al. 2006). Synthetic compounds
such as organochlorine pesticides, polychlorinated biphenyls
(PCBs), and brominated fire retardants (brominated diphenyl
ethers, or BDEs) can persist in the environment for decades,
Christopher L. Rowe (e-mail: [email protected]) is an associate professor
in the University of Maryland Center for Environmental Science, Chesapeake
Biological Laboratory, in Solomons, Maryland. © 2008 American Institute of
Biological Sciences.
July/August 2008 / Vol. 58 No. 7 • BioScience 623
Articles
providing continual exposure to resident organisms for all or
substantial portions of their lifetimes. Fortunately, the majority
of natural systems harboring contaminants appear to pose
little risk of acute mortality or other immediately observable
toxic effects to resident species. Most industrial and natural
pollutants in the world’s freshwater systems occur at very
low concentrations, thousands on the order of only parts
per quadrillion (picograms per liter) to parts per trillion
(nanograms per liter), which makes it difficult to identify their
ultimate effects on ecosystems singly and in mixture
(Schwarzenbach et al. 2006). As the toxicological response is
typically influenced by the interaction between the concentration of the contaminant and the duration to which an individual is exposed, long periods of exposure and cumulative
or repeated damage to biological systems may ultimately
result in expression of toxicological effects that would not otherwise occur under a shorter exposure regime. Life history
strategy, by defining the natural life span of individuals and
hence the duration with which they interact with persistent
compounds, can be a critical factor influencing the probability
that effects of low concentrations of contaminants may manifest. Furthermore, physiological and reproductive traits such
as per capita female investment in offspring influence the
probability that compounds such as lipophilic (fat soluble)
contaminants (e.g., PCBs, BDEs, and some pesticides) are
transferred to developing offspring, thereby compromising
embryonic or juvenile health or viability.
In this article I propose that life history traits unique to longlived vertebrates, as well as ecological and physiological traits
that are frequently associated with this life history strategy,
may place them in jeopardy of reduced reproductive fitness
and population stability in habitats contaminated with
persistent synthetic compounds as they occur in the modern
environment. Long life span and delayed maturation in these
species can result in chronic accumulation and expression of
the effects of contaminants before individuals achieve reproductive status. Combined with long life spans, the high
trophic status of many long-lived vertebrates subjects them
to high concentrations of contaminants having biomagnified
in prey items. Upon maturation, large per capita investment
of materials and energy in offspring confers these species
with the propensity to transfer effective concentrations of biologically labile contaminants to offspring during critical
developmental stages. However, despite the numerous ways
in which individuals can be affected, ultimate changes to
population dynamics of long-lived species may be extremely
slow to appear because of characteristically long cohort
turnover times. Emerging at an insidiously slow rate, populations may deteriorate on a temporal scale that delays recognition and responsive initiation of protective measures before
severe changes to demographics manifest. Furthermore, long
turnover times and concomitantly slow population growth
rates constrain the rate at which impacted populations can
recover to prior levels if conditions substantially improve.
Long-lived species are also likely to experience inertia against
rapid genetic adaptation to environmental contaminants,
624 BioScience • July/August 2008 / Vol. 58 No. 7
particularly modern synthetic compounds that have been
present for a relatively short period of time on an evolutionary
scale and hence have exerted selective forces over very few successive generations (if at all). Thus, numerous traits of longlived species that were optimized within the constraints
imposed by phylogeny and abiotic and biotic environmental conditions during these species’ evolutionary histories may
ultimately prove detrimental in modern, chemically altered
environments.
The discussion that follows is largely theoretical and directed
specifically at assessing contaminant effects on reproductive
fitness and population resiliency of species possessing a longlived, K-selected life history strategy. I avoid making direct
comparisons with other life history types (short-lived or
r-selected) because empirical comparisons among species
possessing disparate life history strategies are inherently
confounded by physiological (e.g., capacity for metabolism
of contaminants or clearance from the body; Chappell 1992,
Kannan et al. 2004) and autecological (trophic position)
differences, and interactions among duration of exposure
(astronomical time) and age or ontogenetic status of the
individuals (physiological time). For example, an experiment
comparing the effects of a controlled exposure duration
(1 year, for instance) on species possessing different natural
life spans (1 and 10 years, for instance) would be confounded
by the fact that the short-lived species experiences a full lifetime of exposure, whereas the long-lived species is exposed
for only one-tenth of its lifetime. Although my discussion is
therefore largely grounded in life history theory (a useful
tool for assessing potential environmental influences on
fitness; e.g., Calow et al. 1997, Congdon et al. 2001), it must
be noted that invoking theory requires that assumptions be
made, many of which cannot be rigorously tested empirically.
Life history and exposure regimes
To examine the influence of a life history strategy on responses of individuals and populations to environmental
contaminants, temporal (life span, time to maturation), autecological (e.g., trophic position), and physiological (e.g.,
per capita investment in offspring) phenomena must be considered (Calow et al. 1997, Congdon et al. 2001, Stark et al.
2004). Temporal and ecological features of a life history strategy govern the duration of contaminant exposure, accumulation potential, and the probability that exposure is sufficient
to initiate a response. Physiological traits, while also playing
a role in accumulation potential (e.g., Kelly et al. 2007), govern the types of effects that may be manifest and the severity of those effects.
Before eliciting a toxicological response, a threshold concentration of a compound (target site concentration; Escher
and Hermens 2002) must be present at a specific biological
site of action (a membrane or cellular receptor, for example).
The concentration-effect threshold is a unique function of the
type of contaminant(s) to which an individual is exposed and
the duration of exposure. In toxicological parlance, in which
an ET50 is defined as the duration of exposure resulting in a
www.biosciencemag.org
Articles
specified effect in 50% of the test population, a low concentration of a contaminant will typically have a greater ET50 than
a higher concentration. The occurrence of a large number of
contaminants in the global environment at concentrations that
are unlikely to induce acute, lethal responses (e.g., Schwarzenbach et al. 2006) therefore suggests that, depending on the inherent toxicity and modes of action of the contaminants in
question, considerably long periods of exposure may be required before substantial effects result.
With the exception of transient or migratory species, the
length of time that an individual interacts with contaminants in a given habitat will be largely determined by its natural life span. Long-lived species of high trophic status (such
as some turtles, crocodilians, birds of prey, and marine mammals) subject to contaminants having accumulated or trophically magnified in their prey are most likely to establish high
body burdens of persistent, bioaccumulative contaminants
(particularly lipophilic compounds). In addition to trophic
position, biomagnification models suggest that respiratory
physiology in these taxa confers on them a reduced capacity
to eliminate some moderately lipid-soluble compounds
through aerial respiratory pathways, and thus a greater
propensity for bioaccumulation relative to organisms possessing aquatic respiration (Kelly et al. 2007). For example, the
models presented by Kelly and colleagues (2007) demonstrated biomagnification factors (BMFs; the ratio of lipid
normalized concentration in predator to prey) for endosulfan (an organochlorine pesticide) to be less than 1 for
predatory fish (e.g., no magnification), whereas for reptiles,
sea birds, and marine mammals, BMFs were 4.9, 10, and 22
(e.g., magnification up to 22 times from prey to predator),
respectively.
Contaminant accumulation is also strongly influenced by
metabolic physiology, which regulates clearance rates of the
compounds from the body (Chappell 1992) and biotransformation of parent compounds to sometimes more toxic
metabolites (Kannan et al. 2004). Life span and adult body size
are often (but by no means universally) positively correlated,
whereas basal metabolism (metabolic rate per unit body
mass) tends to scale as an inverse allometric function with
body size (West et al. 2002). Thus, the metabolic rate of larger
animals is typically lower than that of smaller animals possessing similar metabolic physiology (e.g., ecototherms or endotherms). As a result, metabolism-dependent clearance of
accumulated compounds is relatively slow (and half lives of
accumulated contaminants are relatively long) for largebodied animals (Chappell 1992). The relatively low metabolic
rate associated with those long-lived animals that also attain
a large body size may extend the period that contaminants
reside in the body and have the potential to bring about toxic
insult (Chappell 1992) or to be transferred to offspring later
in life.
On the other hand, high metabolic rates in small species
(or in juveniles, which tend to have higher metabolic rates than
adults; Glazier 2005) can lead to rapid biotransformation of
parent compounds to toxic metabolites (Kannan et al. 2004),
www.biosciencemag.org
elevating concentrations of circulating toxic compounds
while reducing storage of the parent compounds. Thus the
influences of metabolic physiology on bioaccumulation are
complex and vary considerably as a result of interactions
among body size and life span (not to mention extrinsic
factors such as temperature; Glazier 2005), themselves not
necessarily being similarly related among taxa or life history
strategies (Speakman 2005).
Because long-lived vertebrates may attain large (e.g., whales)
or relatively small (e.g., many birds and snakes) maximum or
asymptotic size, there is likely to be considerable variation in
patterns of accumulation and metabolism of contaminants
among species that share life history traits. Protracted exposure to persistent contaminants may therefore have various
implications, depending on the kinetics of storage and
metabolism, which may vary with body size or ontogeny.
Production of toxic metabolites may lead to continual delivery of toxic compounds to active sites, whereas long-term storage of accumulated parent compounds may lead to latent
effects for the individual or its offspring during subsequent
remobilization (and metabolism) of stores during times of
fasting (Debier et al. 2006) or reproduction.
Physiological and reproductive traits and
risks of contaminant effects on fitness
Contaminant exposure itself does not ensure that a response
sufficient to compromise reproductive fitness will necessarily occur. Any number of biochemical, cellular, and subtle physiological effects may follow contaminant exposure, yet in
lieu of being expressed in such a way that fitness itself is reduced or negated, higher-order effects on populations are
unlikely (e.g., Forbes et al. 2006). Reproductive fitness of individuals is a vector through which toxic responses transcend the individual and may ultimately bring higher-order,
population-level change, however. Doubtlessly mortality
before attaining the natural life span can reduce fitness by
restricting the total number of lifetime reproductive opportunities. Moreover, when concentrations of contaminants
are insufficient to cause mortality even after long periods of
exposure, it is more likely that effects will be expressed sublethally, potentially affecting fitness through more subtle
mechanisms.
Combined with extended exposure to persistent contaminants, reproductive traits unique to long-lived species may
place them in double jeopardy for reduced fitness in contaminated habitats. Table 1 presents reproductive traits typical
of many long-lived species in the context of how those species
may be compromised by chronic exposure to contaminants.
Some of the mechanisms by which contaminants may operate to reduce fitness are largely theoretical because of the
dearth of long-term data sets that would be required to address them empirically (table 1; e.g., Henriksen et al. 2001).
In particular, the effects of contaminants on growth rates
(size at maturation) and adult mortality schedules in natural
populations are poorly understood for most long-lived species,
as establishing such relationships requires long-term popuJuly/August 2008 / Vol. 58 No. 7 • BioScience 625
Articles
Table 1. Reproductive life history traits associated with long-lived animals and associated potential effects on
fitness and population dynamics as a result of chronic contaminant exposure.
Trait
Contaminant effects that suboptimize strategy
Large per capita parental investment in offspring
Delayed maturation
Maternal transfer of high concentrations of contaminants to offspring
Mortality before reproduction
Chronic accumulation of contaminants that may be transferred to offspring upon
maturation
Reduced size at maturity reducing offspring quantity or quality
Mortality before achieving total potential life long reproductive events
Chronic exposure of offspring to female-derived contaminants (yolk, placenta, milk)
Delayed expression of fitness effects at the population level
Delayed recovery of populations following abatement of fitness effects
Iteroparity
Long embryonic/preparturition/preweaning period
Long cohort turnover time
lation surveys which have generally been prohibitive (Gibbons
1987, Gibbons et al. 2000, Henriksen et al. 2001). On the
other hand, the relationship between per capita parental investment in offspring and exposure of offspring to maternally
derived contaminants is more amenable to study and has
received considerably more empirical treatment.
Long-lived species typically make large per capita investments in few offspring in the form of parental investment in
the embryo (PIE, the direct contributions of energy in support of embryonic development) and parental investment in
care (PIC, direct provisions of energy to the offspring for use
after hatching/parturition and energy expenditures by parents
in rearing a brood; Congdon 1989). Vitellogenesis—yolk
provisioning (figure 1)—or lactation provides a conduit
through which lipid-soluble compounds accumulated by the
female before reproduction or from her proximate diet can
be transferred to offspring (Braune et al. 2001, Skaare et al.
2002, Rauschenberger et al. 2004, Hickie et al. 2007). Offspring
are thus exposed to contaminants before their direct interaction with contaminants in the environment itself. An additional consequence of large investments in offspring (both
PIE and PIC) is a relatively long embryonic or neonatal period during which offspring are sustained by yolk or milk, and
so have extended exposure to maternally derived contaminants
during critical developmental periods. Thus, an additional
temporal aspect of the life history of long-lived species—a
relatively long embryonic or preparturition period—may
impart susceptibility to contaminants via long periods of
exposure to developing offspring (figure 2).
Maternal transfer of contaminants to offspring has been
well documented in long-lived species, including mammals
(Aguilar and Borrell 1994, Henriksen et al. 2001, Skaare et al.
2002, Beckmen et al. 2003, Lie et al. 2003, Schecter et al. 2003,
Hickie et al. 2007), birds (Dirksen et al. 1995, Ewins et al. 1999,
Hebert et al. 2000, Norstrom et al. 2002), and reptiles (Bishop
et al. 1994, Rauschenberger et al. 2004). Transfer of metals
(such as selenium and mercury) has been observed in some
species (e.g., Roe et al. 2004), but most attention has focused
on persistent synthetic compounds, particularly halogenated
compounds such as PCBs and some pesticides. These compounds have received intensive study because their lipophilic
properties provide a route of transfer from a female’s stored
lipids or proximate diet to offspring through yolk or milk, they
626 BioScience • July/August 2008 / Vol. 58 No. 7
Figure 1. Transfer of lipid soluble contaminants to yolk
during vitellogenesis. Contaminants present in the
female’s body fat or proximate diet are transported
to the oocyte via a yolk precursor complex mediated
by the transport protein vitellogenin.
Figure 2. Interactions among several life history traits
characteristic of long-lived species that may constrain
reproductive fitness potential in persistently contaminated environments. Arrows represent routes through
which contaminants may operate to influence fitness.
www.biosciencemag.org
Articles
have a propensity to accumulate to high concentrations in prey
items and in female tissues prior to reproduction, and their
use and release into the global environment is long-standing,
ongoing, and intensive.
The patterns in which contaminants are passed to offspring over a female’s lifetime will influence the relative
exposure of successive cohorts to maternally derived contaminants. In some species (for example, some pinnipeds,
cetaceans, and ursines), a long period of accumulation of
contaminants by females before maturity results in very high
contributions to offspring during the first few breeding
seasons, whereas transfer of contaminants declines in later
breeding events (Aguilar and Borrell 1994, Ylitalo et al. 2001,
Beckmen et al. 2003, Hickie et al. 2007). Contaminant body
burdens in females subsequently decline with age after maturation because of depuration of accumulated contaminants
over successive breeding events, whereas in males, the burdens
increase continually (Aguilar and Borrell 1994, Ylitalo et al.
2001). These species therefore experience the greatest risks of
offspring impairment during a female’s first several breeding
seasons, with risks declining as the females age. In other
cases, researchers observed either no relationship between
female age and contribution to offspring (Bishop et al. 1994,
Ewins et al. 1999) or an increase in the contribution as the
female ages and her body burdens increased (Rauschenberger et al. 2004). Risks to offspring health and fitness in these
species will thus remain relatively constant or will increase
throughout the lifetime of the female.
Patterns in transfer of contaminants reflect their maternal
source, either from somatic stores or from circulating contaminants derived from the female’s diet during vitellogenesis or lactation (“capital” versus “income” reproductive
investments, respectively; e.g., Meijer and Drent 1999). Within
taxa, contributions of energy (and thus lipophilic contaminants) from female somatic stores or from her proximate
diet vary widely; among species of birds, for example, 0% to
100% of energetic investment in offspring may be derived
from female stores (Meijer and Drent 1999). An observed lack
of a relationship between female age (and presumably her
stored contaminant burden) and contribution of contaminants to offspring in some birds and turtles suggests that in
these species or habitats, contaminants present in the
female’s diet are a primary source (e.g., Bishop et al. 1994,
Ewins et al. 1999). However, in other species, dietary contaminants are substantially supplemented with those from
somatic stores, to the extent that dietary contaminants may
be essentially inconsequential (e.g., some sea mammals and
crocodilians; Aguilar and Borrell 1994, Beckmen et al. 2003,
Rauschenberger et al. 2004). Regardless of the proximate
source of contaminants, their delivery to offspring as a result
of large per capita maternal investment of energy and materials ultimately may compromise reproductive success.
There are numerous documented cases of offspring health
or fitness being compromised through maternal transfer of
contaminants. Reptiles, particularly turtles and crocodilians,
have received considerable study because of their typically very
www.biosciencemag.org
long life spans (and thus potential risks of chronic exposure
and effect) and concerns about globally declining reptile
populations attributable in part to habitat contamination
(Gibbons et al. 2000). Much research has examined maternal
transfer of organic contaminants and effects on offspring
health in turtles inhabiting the Great Lakes of North America. In regions of the Great Lakes where contaminants persist
at relatively high concentrations, snapping turtles experienced reduced hatching success attributable to in ovo exposure to maternally derived contaminants (Bishop et al. 1998).
In some instances, successfully hatched individuals displayed
numerous abnormalities that may influence survival probability or future reproductive performance, including less pronounced sexual dimorphism, dwarfism, and missing or
malformed eyes and legs (e.g., Bishop et al. 1991, 1998, de Solla
et al. 1998).
A number of recent studies of alligators residing in contaminated water bodies in Florida have demonstrated strong
relationships between chronic exposure of adults to contaminants and reduced offspring health and fitness. Adults that
had been exposed to synthetic compounds over a lifetime of
feeding in contaminated areas experienced endocrinological
abnormalities (Guillette et al. 1994) and reduced clutch
success (Rauschenberger et al. 2007). Offspring that succeeded in hatching frequently experienced retarded development and growth and reduced reproductive potential
(Guillette et al. 1994, 1995, 1999, Gunderson et al. 2004,
Milnes et al. 2005). Such overall reductions in reproductive
fitness because of chronic contaminant exposure may
explain the declines in alligator populations in the contaminated systems, in stark contrast to the general growth in
alligator populations in many other areas (Guillette et al.
1994, Semenza et al. 1997).
Birds of prey have also received considerable study with
respect to the effects of persistent contaminants, particularly
pesticides (DDT for example), following precipitous declines
in populations of some species in the mid-20th century
(Hickey and Anderson 1968). Whereas the familiar phenomenon of eggshell thinning and subsequent reproductive
failure (Hickey and Anderson 1968) was generally attributed
to dysfunction of the eggshell gland in mature females, subsequent mortality or reduced performance of many of those
individuals that did not succumb to eggshell thinning suggests
that toxicity of maternally derived contaminants was an additional mechanism by which reproductive success was reduced (Dirksen et al. 1995, Fry 1995). Effects on offspring
ranged from embryonic and juvenile mortality to morphological malformations that may hinder feeding (Fox et al.
1991, Tillitt et al. 1992). Although in some areas reproductive
effects associated with contaminants such as DDT and PCBs
declined after the imposition of restrictions on the use of these
compounds (Grier 1982), persistent effects for one or more
decades after cessation of the compounds’ release have been
reported (Fox et al. 1991, Tillitt et al. 1992, Toschik et al.
2005). Recent reports suggest that persistent contaminants
continue to threaten some birds of prey even in areas in
July/August 2008 / Vol. 58 No. 7 • BioScience 627
Articles
which individuals are not restricted to highly contaminated
systems; long-lived fish-eating birds such as albatross, which
have large hunting ranges and thus are not restricted to
local hot spots of contamination, have been found to possess
tissue burdens of contaminants that may be toxic to their offspring (Guruge et al. 2001, Tanabe et al. 2004).
Much recent work has also examined relationships among
female body burdens, maternal transfer, and offspring health
in carnivorous mammals—particularly pinnipeds, cetaceans,
and ursines in the high northern latitudes. Long-distance
atmospheric transport, biomagnification, and localized sources
have resulted in extremely high body burdens of persistent
contaminants in some areas (Muir et al. 1999, Andersen et
al. 2001, Skaare et al. 2002, Lie et al. 2003). For some midmolecular weight compounds, Kelly and colleagues (2007) estimated that biomagnification factors for marine mammals
can exceed those for both reptiles and sea birds, in some
cases by an order of magnitude. Recent studies suggest that
body burdens of persistent, synthetic contaminants in polar
bears and killer whales in some areas exceed thresholds beyond which toxicological effects would be expected (Muir et
al. 1999, Ross 2006). Indeed, polar bears that accumulated relatively high concentrations of synthetic compounds have
been found to express sublethal anomalies in developmental,
immunological, and endocrinological traits (Skaare et al.
2002, Oskam et al. 2003, Lie et al. 2004). High body burdens
in females result in transfer of high concentrations of contaminants to offspring through lactation (e.g., Muir et al.
1999, Henriksen et al. 2001, Neale et al. 2005, Lie et al. 2003).
In some cases, females may transfer contaminant loads sufficient to impair offspring health or survival (Skaare et al.
2002). Such effects on reproductive success have been linked,
albeit inconclusively, to demographic anomalies in polar bear
populations in which the frequency of females older than 16
years having cubs less than one year old was substantially reduced in a population in which body burdens were high relative to other populations (Derocher et al. 2003).
Although much of the preceding discussion focused on contaminant effects on offspring production or health, for which
most data exist, it is important to recognize that population
dynamics of long-lived species are typically more sensitive to
survival or performance of adults relative to earlier life stages
(Crouse et al. 1987, Congdon et al. 1994). The iteroparous,
bet-hedging reproductive strategy characteristic of long-lived
species provides opportunities to compensate for periodic
reproductive failures through multiple breeding attempts
during a lifetime. Thus contaminant effects on offspring production or performance, if occurring only occasionally, would
be expected to have little effect on population dynamics.
However, while the life history strategy is resilient to periodic
environmental fluctuations and failed breeding attempts, it
is tightly constrained by chronic disturbances such that a
sustained reduction in offspring survival potential across an
individual’s lifetime may negate the advantages of this strategy (Congdon et al. 1994). In the case of many modern contaminants that are extremely persistent and cycle through food
628 BioScience • July/August 2008 / Vol. 58 No. 7
webs for very long periods of time, exposure and risks are likely
to be sustained, perhaps over the entire lifetime of an individual. Thus, despite the relatively greater influence of adult
survival on population dynamics under conditions in which
recruitment failure is periodic and stochastic, the persistent
effects of contaminants on reproductive success over much
or all of the life span of individuals may ultimately lead to reduced population sizes.
Temporal inertia
The protracted generation turnover times characteristic of
long-lived species delay any cohort’s influence on population
dynamics (e.g., through recruitment to the adult population) for a considerable time after the cohort is produced.
For example, Congdon and colleagues (1993, 1994) reported
cohort generation times in populations of snapping turtles and
Blanding’s turtles of 25 and 37 years, respectively. Such long
turnover times may provide long-lived species with “temporal inertia” with respect to expression as population-level
change. Under extreme conditions in which recruitment
failure of entire cohorts occurs, it may take decades for the
effects to be realized as substantial changes in the abundance
of reproductive adults. Moreover, under more typical situations in which contaminant exposure is less severe and recruitment to the adult population is only moderately reduced
rather than negated, latent population-level changes will be
much slower to emerge. Thus, the observable effects of contaminants on populations resulting from reduced reproductive success may emerge very subtly in a human frame of
reference, perhaps at a temporal scale that exceeds our ability to recognize the changes as distinct from natural, stochastic population fluctuations (e.g., Pechmann et al. 1991,
Gibbons et al. 2000). As a result, contaminant-induced population declines may not be recognized until considerable
impacts have already occurred. This phenomenon is in many
ways analogous to the “shifting baseline syndrome” (Pauly
1995), which has been attributed to the erosion of commercial fisheries and of other marine resources over the past
century through gradually increasing exploitation and degradation of habitat quality (Jackson 2001).
Long generation times limit not only the rate at which
contaminant effects will be expressed as population-level
change but also the rate at which populations may subsequently recover if conditions improve or if further degradation is to be prevented. For example, in the turtle populations
studied by Congdon and colleagues (1993, 1994), with cohort
generation times of 25 to 37 years, the cessation of contaminant effects on reproductive success would not be predicted
to be expressed at the population level for at least three to four
decades, when recruitment of the first postcontaminantaffected cohorts would have occurred. Crowder and colleagues (1994) predicted such delays in population recovery
using demographic models to assess sea turtle population dynamics. They projected future population trends in response
to regulatory actions requiring the use of turtle-excluding devices on fishing trawls to reduce incidental mortality, a priwww.biosciencemag.org
Articles
mary cause of some declines in sea turtle populations. Even
with the adoption of such devices and enforcement in their
use, the authors projected that more than 70 years would be
required for the population to increase by a single order of
magnitude.
However, in this scenario (Crowder et al. 1994), use of the
exclusion devices, and removal thereby of the primary agent
responsible for population declines, was assumed to be immediate and sustained. Yet in habitats contaminated with
persistent compounds, a legacy of prior inputs remains as the
contaminants continue to cycle through food webs, and thus
abatement of inputs is unlikely to produce an immediate
improvement in conditions and rapid rebound in population
growth rate. For example, despite bans on the use of DDT in
North America and parts of Europe more than 30 years ago,
DDT and its metabolites continue to persist in environmental matrices and wildlife (Simonich and Hites 1995, Harris et
al. 2000). Although reproductive success of some species that
were adversely affected by DDT has improved considerably
since its use has been restricted (Grier 1982), recovery of the
populations has sometimes been very slow. Demographic
models of peregrine falcon populations in California, for example, revealed that despite 15 years of reintroduction efforts,
population growth rates remained negative until approximately 18 years following the ban on DDT (Kauffman et al.
2003).
An additional impediment to the recovery of affected
populations of long-lived species is the relatively slow rate at
which adaptation to contaminants may occur. As genetic
adaptation requires successive generations experiencing
selective pressures, long turnover times may constrain the rate
of evolution of tolerance, unless conditions are exceptionally
severe and drastic reductions in population favor only the most
tolerant individuals. Furthermore, the occurrence of persistent, synthetic contaminants in the environment for only
the past century suggests that relatively few turnovers of generations have occurred for some long-lived species since the
advent and widespread use of these compounds; it is thus unlikely that significant adaptation has occurred for many of
these species.
Delayed manifestation of contaminant effects on populations, combined with a slow rate of subsequent recovery and
limited potential for rapid adaptation to contaminant stresses,
suggests that long-lived species may decline initially at a
nearly imperceptible rate. Once this inertia is overcome and
a significant portion of the populations experience reduced
reproductive fitness, however, population effects may be dramatic and long lasting. Although some bird populations have
recovered or appear to be recovering in some areas after restrictions on contaminants were imposed (Grier 1982), these
successes cannot necessarily be viewed as evidence that contaminants’ threats to long-lived species in general have been
eliminated. Even in regions with restrictions on the release of
contaminants that may compromise reproductive fitness
(PCBs and DDT, for example), long-distance transport and
a legacy of prior release of the contaminants to the environwww.biosciencemag.org
ment present ongoing threats, as evidenced in the last decade
by observations of compromised individual health and fitness
(e.g., Bishop et al. 1998, de Solla et al. 1998, Guillette et al. 1999,
Skaare et al. 2002, Derocher 2003, Milnes et al. 2005).
Summary
Globally, inputs of persistent contaminants to the environment
continue to increase, whether they be synthetic compounds
that are known to affect wildlife, newly developed compounds for agricultural and industrial use, or other persistent
contaminants such as metals. Past experiences with “firstgeneration” synthetic compounds such as DDT demonstrate
contaminants’ potential to affect long-lived species through
subtle physiological mechanisms and reduced reproductive
fitness, in some cases leading to drastic and long-lasting
population declines. Despite the current state of knowledge,
these compounds continue to be released in regions in which
they remain unrestricted. In developing regions in which
agriculture is a primary economic resource, or insecttransmitted diseases pose serious threats to human health,
compounds such as organochlorine pesticides remain in use
or have been restricted in use only recently (e.g., Weissmann
2006). Furthermore, continued development and worldwide
use and release of other synthetic compounds that are
persistent and potentially toxic as well (for example, BDEs
and fluorinated organic compounds) suggest that threats to
the fitness and population status of long-lived species in
much of the world may be ongoing or perhaps may not have
yet appeared.
Life history strategies characterized by long life spans, delayed maturation, large per capita maternal investments in offspring, and long cohort turnover times may be suboptimal
in habitats contaminated with persistent synthetic compounds. Long life spans confer protracted exposure periods
and a propensity to accumulate high concentrations of many
contaminants, particularly in species of high trophic status
that are subject to biomagnification through the food web.
Establishment of high body burdens prior to reproduction and
subsequent maternal transfer of contaminants to embryos and
juveniles through yolk or milk can jeopardize offspring survival or performance. Manifesting at a temporal scale that is
inconsistent with that of human perception, insidiously slow
changes in population size or structure may elude recognition
and delay actions to mitigate further population reductions.
Furthermore, in the event that environmental quality significantly improves, populations that have been compromised
may be exceptionally slow to recover.
Thus the “calamity of so long life” (sensu Hamlet, William
Shakespeare, ca. 1600): a life history strategy for which a long
life span was optimal for attaining reproductive fitness within
historical environmental conditions may prove deleterious
in modern, contaminated habitats in which toxicological
responses are elicited only after long periods of exposure.
Predictions from life history theory, combined with documented reductions in reproductive fitness in contaminated
habitats, historical and ongoing declines of some species,
July/August 2008 / Vol. 58 No. 7 • BioScience 629
Articles
and increasing releases of contaminants worldwide, suggest
that concern for the current and future status of long-lived
species is warranted.
Acknowledgments
This manuscript benefited from suggestions by Karen Eisenreich, David Kidwell, Teresa A. Manyin, and Carys Mitchelmore. Criticism by four anonymous reviewers substantially
improved the manuscript. The Gene Lane Endowment
graciously provided support during the preparation of the
manuscript. This article is dedicated to the memory of David
L. Rowe, who was not long-lived enough.
References cited
Aguilar A, Borrell A. 1994. Reproductive transfer and variation of body load
of organochlorine pollutants with age in fin whales (Balaenoptera
physalus). Archives of Environmental Contamination and Toxicology 27:
546–554.
Andersen M, Lie E, Derocher AE, Belikov SE, Bernhoft A, Boltunov AN,
Garner GW, Skaare JU, Wiig O. 2001. Geographic variation of PCB
congeners in polar bears (Ursus maritimus) from Svalbard east to the
Chukchi Sea. Polar Biology 24: 231–238.
Beckmen KB, Blake JE, Ylitalo GM, Stott JL, O’Hara TM. 2003. Organochlorine contaminant exposure and associations with hematological and
humoral immune functional assays with dam age as a factor in freeranging northern fur seal pups (Callorhinus ursinus). Marine Pollution
Bulletin 46: 594–606.
Bishop CA, Brooks RJ, Carey JH, Norstrom RJ, Ng P, Lean DRS. 1991. The
case for a cause-effect linkage between environmental contamination
and development in eggs of the common snapping turtle (Chelydra s.
serpentina) from Ontario, Canada. Journal of Toxicology and Environmental Health 33: 521–547.
Bishop CA, Brown BP, Brooks RJ, Lean DRS, Carey JH. 1994. Organochlorine contaminant concentrations in eggs and their relationship to body
size, and clutch characteristics of the female common snapping turtle
(Chelydra serpentina serpentina) in Lake Ontario, Canada. Archives of
Environmental Contamination and Toxicology 27: 82–87.
Bishop CA, Ng P, Pettit KE, Kennedy SW, Stegeman JJ, Norstrom RJ, Brooks
RJ. 1998. Environmental contamination and developmental abnormalities in eggs and hatchlings of the common snapping turtle (Chelydra
serpentina serpentina) from the Great Lakes–St. Lawrence River Basin
(1989–1991). Environmental Pollution 101: 143–156.
Braune BM, Donaldson GM, Hobson KA. 2001. Contaminant residues in
seabird eggs from the Canadian Arctic, part I: Temporal trends 1975–1998.
Environmental Pollution 114: 39–54.
Calow P, Sibly RM, Forbes V. 1997. Risk assessment on the basis of simplified life-history scenarios. Environmental Toxicology and Chemistry
16: 1983–1989.
Chappell WR. 1992. Scaling toxicity data across species. Environmental
Geochemistry and Health 14: 71–80.
Congdon JD. 1989. Proximate and evolutionary constraints on energy
relations of reptiles. Physiological Zoology 62: 356–373.
Congdon JD, Dunham AE, van Loben Sels RC. 1993. Delayed sexual maturity and demographics of Blanding’s turtles (Emydoidea blandingii):
Implications for conservation and management of long-lived organisms.
Conservation Biology 7: 826–833.
———. 1994. Demographics of common snapping turtles (Chelydra
serpentina): Implications for conservation and management of long-lived
organisms. American Zoologist 34: 397–408.
Congdon JD, Dunham AE, Hopkins WA, Rowe CL, Hinton TG. 2001.
Resource allocation-based life history trait values: A conceptual basis for
studies of environmental toxicology. Environmental Toxicology and
Chemistry 20: 1698–1703.
630 BioScience • July/August 2008 / Vol. 58 No. 7
Crouse DT, Crowder LB, Caswell H. 1987. A stage-based population model
for loggerhead sea turtles and implications for conservation. Ecology 68:
1412–1423.
Crowder LB, Crouse DT, Heppell SS, Martin TH. 1994. Predicting the
impact of turtle excluder devices on loggerhead sea turtle populations.
Ecological Applications 4: 437–445.
Debier C, Chalon C, Le Boeuf BJ, de Tillesse T, Larondelle Y, Thome JP. 2006.
Mobilization of PCBs from blubber to blood in northern elephant seals
(Mirounga angustirostris) during the post-weaning fast. Aquatic Toxicology
80: 149–157.
Derocher AE, Wolkers H, Colborn T, Schlabach M, Larsen TS, Wiig O. 2003.
Contaminants in Svalbard polar bear samples archived since 1967 and
possible population level effects. Science of the Total Environment 301:
163–174.
de Solla SR, Bishop CA, Van Der Kraak G, Brooks RJ. 1998. Impact of
organochlorine contamination on levels of sex hormones and external
morphology of common snapping turtles (Chelydra serpentina
serpentina) in Ontario, Canada. Environmental Health Perspectives 106:
253–260.
Dirksen S, Boudewijn TJ, Slager LK, Mes RG, Vanschaick MJM, Devoogt P.
1995. Reduced breeding success of cormorants (Phalacrocorax carbo
sinensis) in relation to persistent organochlorine pollution of aquatic
habitats in the Netherlands. Environmental Pollution 88: 119–132.
Escher BI, Hermens JLM. 2002. Modes of action in ecotoxicology: Their role
in body burdens, species sensitivity, QSARs, and mixture effects. Environmental Science and Technology 36: 4201–4217.
Ewins PJ, Postupalsky S, Hughes KD, Weseloh DV. 1999. Organochlorine
contaminant residues and shell thickness of eggs from known-age female
ospreys (Pandion haliaetus) in Michigan during the 1980s. Environmental Pollution 104: 295–304.
Fox GA, Collins B, Hayakawa E, Weseloh DV, Ludwig JP, Kubiak TJ, Erdman
TC. 1991. Reproductive outcomes in colonial fish-eating birds: A
biomarker for developmental toxicants in Great Lakes food chains, II:
Spatial variation in the occurrence and prevalence of bill defects in
young double-crested cormorants in the Great Lakes, 1979–1987.
Journal of Great Lakes Research 17: 158–167.
Forbes VE, Palmqvist A, Bach L. 2006. The use and misuse of biomarkers in
ecotoxicology. Environmental Toxicology and Chemistry 25: 272–280.
Fry DM. 1995. Reproductive effects in birds exposed to pesticides and
industrial chemicals. Environmental Health Perspectives 103: 165–171.
Gibbons JW. 1987. Why do turtles live so long? BioScience 37: 262–269.
Gibbons JW, et al. 2000. The global decline of reptiles, déjà vu amphibians.
BioScience 50: 653–666.
Glazier DS. 2005. Beyond the ‘3/4-power law’: Variation in the intra- and
interspecific scaling of metabolic rate in animals. Biological Reviews
80: 611–662.
Grier JW. 1982. Ban of DDT and subsequent recovery of reproduction in bald
eagles. Science 218: 1232–1235.
Guillette LJ Jr, Gross TS, Masson GR, Matter JM, Percival HF, Woodward AR.
1994. Developmental abnormalities of the gonad and abnormal sex
hormone concentrations in juvenile alligators from contaminated and
control lakes in Florida. Environmental Health Perspectives 102: 680–688.
Guillette LJ Jr, Crain DA, Rooney AA, Pickford DB. 1995. Organization
versus activation: The role of endocrine-disrupting contaminants (EDCs)
during embryonic development in wildlife. Environmental Health
Perspectives 103: 157–164.
Guillette LJ Jr, Woodward AR, Crain DA, Rooney AA, Percival HF. 1999.
Plasma steroid concentrations and male phallus size in juvenile alligators from seven Florida lakes. General and Comparative Endocrinology
116: 356–372.
Gunderson MP, Bermudez DS, Bryan TA, Degala S, Edwards TM, Kools SAE,
Milnes MR, Woodward AR, Guillette LJ. 2004. Variation in sex steroids
and phallus size in juvenile American alligators (Alligator mississippiensis) collected from 3 sites within the Kissimmee-Everglades drainage in
Florida (USA). Chemosphere 56: 335–345.
www.biosciencemag.org
Articles
Guruge KS, Watanabe M, Tanaka H, Tanabe S. 2001. Accumulation status of
persistent organochlorines in albatrosses from the North Pacific and the
Southern Ocean. Environmental Pollution 114: 389–398.
Harris ML, Wilson LK, Elliott JE, Bishop CA, Tomlin AD, Henning KV.
2000. Transfer of DDT and metabolites from fruit orchard soils to American robins (Turdus migratorius) twenty years after agricultural use of DDT
in Canada. Archives of Environmental Contamination and Toxicology
39: 205–220.
Hebert CE, Hobson KA, Shutt JL. 2000. Changes in food web structure
affect rate of PCB decline in herring gull (Larus argentatus) eggs.
Environmental Science and Technology 34: 1609–1614.
Henriksen EO, Wiig O, Skaare JU, Gabrielsen GW, Derocher AE. 2001.
Monitoring PCBs in polar bears: Lessons learned from Svalbard.
Journal of Environmental Monitoring 3: 493–498.
Hickey JJ, Anderson DW. 1968. Chlorinated hydrocarbons and eggshell
changes in raptorial and fish-eating birds. Science 162: 271–273.
Hickie BE, Ross PS, Macdonald RW, Ford JKB. 2007. Killer whales (Orcinus
orca) face protracted health risks associated with lifetime exposure to
PCBs. Environmental Science and Technology 41: 6613–6619.
Jackson JBC. 2001. What was natural in the coastal oceans? Proceedings of
the National Academy of Sciences 98: 5411–5418.
Kannan K, Kajiwara N, Watanabe M, Nakata H, Thomas NJ, Stephenson M,
Jessup DA, Tanabe S. 2004. Profiles of polychlorinated biphenyl congeners,
organochlorine pesticides, and butyltins in Southern Sea otters and
their prey. Environmental Toxicology and Chemistry 23: 49–56.
Kauffman MJ, Frick WF, Linthicum J. 2003. Estimation of habitat-specific
demography and population growth for peregrine falcons in California.
Ecological Applications 13: 1802–1816.
Kelly BC, Ikonomou MG, Blair JD, Morin AE, Gobas F. 2007. Food webspecific biomagnification of persistent organic pollutants. Science 317:
236–239.
Lie E, Bernhoft A, Riget F, Belikov SE, Boltunov AN, Derocher AE, Garner
GW, Wiig O, Skaare JU. 2003. Geographical distribution of organochlorine pesticides (OCPs) in polar bears (Ursus maritimus) in the
Norwegian and Russian Arctic. Science of the Total Environment 306:
159–170.
Lie E, Larsen HJS, Larsen S, Johansen GM, Derocher AE, Lunn NJ, Norstrom
RJ, Wiig O, Skaare JU. 2004. Does high organochlorine (OC) exposure
impair the resistance to infection in polar bears (Ursus maritimus)? Part
II: Possible effect of OCs on mitogen- and antigen-induced lymphocyte
proliferation. Journal of Toxicology and Environmental Health, Part A:
Current Issues 68: 457–484.
Meijer T, Drent R. 1999. Re-examination of the capital and income
dichotomy in breeding birds. Ibis 141: 399–414.
Milnes MR, Bermudez DS, Bryan TA, Gunderson MP, Guillette LJ. 2005.
Altered neonatal development and endocrine function in Alligator
mississippiensis associated with a contaminated environment. Biology of
Reproduction 73: 1004–1010.
Muir D, et al. 1999. Spatial and temporal trends and effects of contaminants
in the Canadian Arctic marine ecosystem: A review. Science of the
Total Environment 230: 83–144.
Neale JCC, Schmelzer KR, Gulland FMD, Berg EA, Tjeerdema RS. 2005. Rapid
communication: Organohalogen levels in harbor seal (Phoca vitulina)
pups increase with duration of nursing. Journal of Toxicology and
Environmental Health, Part A: Current Issues 68: 687–691.
Norstrom RJ, Simon M, Moisey J, Wakeford B, Weseloh DVC. 2002.
Geographical distribution (2000) and temporal trends (1981–2000) of
brominated diphenyl ethers in Great Lakes herring gull eggs. Environmental Science and Technology 36: 4783–4789.
Oskam IC, Ropstad E, Dahl E, Lie E, Derocher AE, Wiig O, Larsen S, Wiger
R, Skaare JU. 2003. Organochlorines affect the major androgenic hormone,
testosterone, in male polar bears (Ursus maritimus) at Svalbard. Journal
of Toxicology and Environmental Health, Part A: 66: 2119–2139.
Pauly D. 1995. Anecdotes and the shifting baseline syndrome of fisheries.
Trends in Ecology and Evolution 10: 430.
www.biosciencemag.org
Pechmann JHK, Scott DE, Semlitsch RD, Caldwell JP, Vitt LJ, Gibbons JW.
1991. Declining amphibian populations: The problem of separating
human impacts from natural fluctuations. Science 253: 892–895.
Pianka ER. 1970. On r and K selection. American Naturalist 104: 592–597.
Rauschenberger RH, Sepulveda MS, Wiebe JJ, Szabo NJ, Gross TS. 2004.
Predicting maternal body burdens of organochlorine pesticides from eggs
and evidence of maternal transfer in Alligator mississippiensis. Environmental Toxicology and Chemistry 23: 2906–2915.
Rauschenberger RH, Wiebe JJ, Sepulveda MS, Scarborough JE, Gross TS. 2007.
Parental exposure to pesticides and poor clutch viability in American
alligators. Environmental Science and Technology 41: 5559–5563.
Roe JH, Hopkins WA, Baionno JA, Staub BP, Rowe CL, Jackson BP. 2004.
Maternal transfer of selenium in Alligator mississippiensis nesting downstream from a coal-burning power plant. Environmental Toxicology
and Chemistry 23: 1969–1972.
Ross PS. 2006. Fireproof killer whales (Orcinus orca): Flame-retardant chemicals and the conservation imperative in the charismatic icon of British
Columbia, Canada. Canadian Journal of Fisheries and Aquatic Sciences
63: 224–234.
Schecter A, Vuk MP, Papke O, Ryan JJ, Birnbaum L, Rosen R. 2003. Polybrominated diphenyl ethers (PBDEs) in US mothers’ milk. Environmental
Health Perspectives 111: 1723–1729.
Schwarzenbach RP, Escher BI, Fenner K, Hofstetter TB, Johnson CA, von
Gunten U, Wehrli B. 2006. The challenge of micropollutants in aquatic
systems. Science 313: 1072–1077.
Semenza JC, Tolbert PE, Rubin CH, Guillette LJ, Jackson RJ. 1997. Reproductive toxins and alligator abnormalities at Lake Apopka, Florida.
Environmental Health Perspectives 105: 1030–1032.
Simonich SL, Hites RA. 1995. Global distribution of persistent organochlorine compounds. Science 269: 1851–1854.
Skaare JU, Larsen HJ, Lie E, Bernhoft A, Derocher AE, Norstrom R, Ropstad
E, Lunn NF, Wiig O. 2002. Ecological risk assessment of persistent organic
pollutants in the arctic. Toxicology 181: 193–197.
Speakman JR. 2005. Body size, energy metabolism and lifespan. Journal of
Experimental Biology 208: 1717–1730.
Stark JD, Banks JE, Vargas R. 2004. How risky is risk assessment: The role that
life history strategies play in susceptibility of species to stress. Proceedings of the National Academy of Sciences 101: 732–736.
Stearns SC. 1976. Life-history strategies: A review of the ideas. Quarterly
Review of Biology 51: 3–17.
———. 1989. Trade-offs in life history evolution. Functional Ecology 3:
259–268.
Tanabe S, Watanabe M, Minh TB, Kunisue T, Nakanishi S, Ono H, Tanaka
H. 2004. PCDDs, PCDFs, and coplanar PCBs in albatross from the
North Pacific and Southern Oceans: Levels, patterns, and toxicological
implications. Environmental Science and Technology 38: 403–413.
Tillitt DE, et al. 1992. Polychlorinated biphenyl residues and egg mortality
in double-crested cormorants from the Great Lakes. Environmental
Toxicology and Chemistry 11: 1281–1288.
Toschik PC, Rattner BA, McGowan PC, Christman MC, Carter DB, Hale RC,
Matson CW, Ottinger MA. 2005. Effects of contaminant exposure on reproductive success of ospreys (Pandion haliaetus) nesting in Delaware
River and Bay, USA. Environmental Toxicology and Chemistry 24:
617–628.
Weissmann G. 2006. DDT is back: Let us spray! FASEB Journal 20: 2427–2429.
West GB, Woodruff WH, Brown JH. 2002. Allometric scaling of metabolic
rate from molecules and mitochondria to cells and mammals. Proceedings of the National Academy of Sciences 99: 2473–2478.
Ylitalo GM, Matkin CO, Buzitis J, Krahn MM, Jones LL, Rowles T, Stein JE.
2001. Influence of life-history parameters on organochlorine concentrations in free-ranging killer whales (Orcinus orca) from Prince William
Sound, AK. Science of the Total Environment 281: 183–203.
doi:10.1641/B580709
Include this information when citing this material.
July/August 2008 / Vol. 58 No. 7 • BioScience 631