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Colloquium
What was natural in the coastal oceans?
Jeremy B. C. Jackson*
Scripps Institution of Oceanography, University of California at San Diego, La Jolla, CA 92093; and Center for Tropical Paleoecology and Archeology,
Smithsonian Tropical Research Institute, Apartado 2072, Balboa, Republic of Panama
Humans transformed Western Atlantic coastal marine ecosystems
before modern ecological investigations began. Paleoecological,
archeological, and historical reconstructions demonstrate incredi-
ble losses of large vertebrates and oysters from the entire Atlantic
coast. Untold millions of large fishes, sharks, sea turtles, and
manatees were removed from the Caribbean in the 17th to 19th
centuries. Recent collapses of reef corals and seagrasses are due
ultimately to losses of these large consumers as much as to more
recent changes in climate, eutrophication, or outbreaks of disease.
Overfishing in the 19th century reduced vast beds of oysters in
Chesapeake Bay and other estuaries to a few percent of pristine
abundances and promoted eutrophication. Mechanized harvesting
of bottom fishes like cod set off a series of trophic cascades that
eliminated kelp forests and then brought them back again as
fishers fished their way down food webs to small invertebrates.
Lastly, but most pervasively, mechanized harvesting of the entire
continental shelf decimated large, long-lived fishes and destroyed
three-dimensional habitats built up by sessile corals, bryozoans,
and sponges. The universal pattern of losses demonstrates that no
coastal ecosystem is pristine and few wild fisheries are sustainable
along the entire Western Atlantic coast. Reconstructions of eco-
systems lost only a century or two ago demonstrate attainable
goals of establishing large and effective marine reserves if society
is willing to pay the costs. Historical reconstructions provide a new
scientific framework for manipulative experiments at the ecosys-
tem scale to explore the feasibility and benefits of protection of our
living coastal resources.
magnitude and exhibit nine major collapses and subsequent
recoveries over 1700 years. These data and shorter records of fish
catches suggest population cycles of 50 to 70 years associated
with alteration of'warm and cold physical regimes (4, 8~. These
cycles exceed the longest instrumental temperature records for
the region and greatly complicate management of fisheries. How
can one determine a sustainable catch against a background of
such extreme natural variation?
Conventional ecological data are clearly inadequate to mea-
sure the ecological impacts of fishing or any other long-term
human disturbance (4, 5, 9~. Most observational records are
much too short, too poorly replicated, and too uncontrolled to
encompass even a single cycle of natural environmental varia-
tion. For example, detailed ecological observations of reef corals
began only in the 1930s. There are a few "before and after"
comparisons of community composition between surveys con-
ducted up to a century ago and the present (104. However, the
longest quantitative time series comprises only a few small
intertidal quadrats on one small island over 30 years (11), and the
longest comparable subtidal records encompass less than 20
years (12~. In both cases, the interval studied is much less than
the generation times of most common coral species and the
intervals between some kinds of major disturbances in coral reef
environments (13~. Several kelp forests and rocky intertidal
communities have been surveyed for about 25 years over scales
of several hectares, so that the data approximate or exceed
generation times of most important species, but not the period-
icity of major climatic cycles (5, 6~. Ecological data for oyster
The persistent myth of the oceans as wilderness brindled reefs, seagrass meadows, level bottoms, and virtually all other
~ ecologists to the massive loss of marine ecological diversity marine communities have similar limitations (14-18~.
caused by overfishing and human inputs from the land over the Paleoecological, archeological, and historical data are the only
pastcenturies.Unriithel980s,coraIreefs~kelpforests~an(Iother means for extending ecological records back long enough to
coastal habitats were discussed in scientific journals and text- document the characteristic variability of marine ecosystems and
books as "natural" or "pristine" communities with little or no the magnitude of earlier anthropogenic change. Here I review
reference to the pervasive absence of large vertebrates or the the transformation of five Western Atlantic coastal ecosystems
widespread effects of pollution. This is because our concept of over the past few centuries as a result of human exploitation and
what is natural today is based on personal experience at the pollution My goals are to demonstrate the extraordinary mag-
expense of historical perspective. Thus, ``natural'' means the way nitude of ecological changes that have been largely forgotten and
things were when we first saw them or exploited them, and to show how awareness of these changes can benefit efforts for
"unnatural" means all subsequent change (1, 2). As in Magritte~s conservation and restoration of coastal ecosystems. My focus is
masterpiece, La Condition Humaine, we see the world through on benth1c communities because extreme overfishing of pelagic
a model of our own creation that organizes and filters under- species such as Atlantic whales, tuna, salmon, and herring is well
standing (3). In the present context, that filter is the sum total known (19, 20). Transformations of benthos are subtler and
of anthropogenic change that took placein the Oceans before we known only to a few specialists. I also focus on ecological
were born. extinction because the magnitude of ecological changes is not
Not all ecological change is anthropogenic, however. Natural generally understood (1, 2, 5, 9), and documentation of actual
conditions in the oceans fluctuate greatly and sometimes sud-
denly on time scales that extend for decades to millennia. Thus,
the filter of individual experience has two components. Changes
caused by humans are the signal and natural variability consti-
tutes the noise that obscures the human footprint (4-6). An
important example of the potential magnitude of natural change
comes from annually layered sediments of the Santa Barbara
Basin (7). Abundances of fish scales of anchovies and sardines
preserved in these sediments fluctuate more than an order of
www.pnas.org/cgi/doi/ 10.1 073/pnas.09 1092898
extinctions of marine species is just beginning (21). More
importantly, too great a focus on species detracts attention from
the transformation and loss of habitats and collapse of natural
ecosystems that drive the processes of extinction.
This paper was presentec] at the National Acaclemy of Sciences colloquium, "The Future of
Evolution," helc] March ~ 6-20, 2000, atthe Arnold and Mabel Beckman Center in Irvine, CA.
*Aciciress reprint requests to: Scripps Institution of Oceanography, University of California
at San Diego, Ea Jolla, CA 92093-0244. E-mail: jbcj~tucsc].eclu.
PNAS 1 May 8, 2001 1 vol. 98 1 no. 10 1 5411-5418
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Representative terms from entire chapter:
green turtles
Caribbean Coral Reefs
Coral reefs are the largest durable biological constructions on
earth. Reefs determine the physical structure of coastlines and
adjacent ecosystems, including seagrass beds and lagoons. Coral
reefs are the most taxonomically diverse marine ecosystems and
provide complex habitat for myriad sessile and mobile organisms
(22, 23~. Recent discoveries of numerous sibling species suggest
diversity is even greater than already described (24~.
Species composition of Caribbean coral communities was
stable for at least 125 thousand years, until the collapse in the
1980s (25-29~. Different environments were dominated by dis-
tinct species assemblages of the corals Acropora, Montastrea,
Diploria, and a few other genera, and the composition of these
assemblages was similar over tens of kilometers of coastline for
tens of thousands of years. Within each habitat, community
membership was more predictable than expected by random
sampling of the habitat-specific species pool. Thus, there was a
clear baseline of coral community composition that serves for
comparison with today.
Western Atlantic reef corals suffered catastrophic mortality in
the 1980s (30-34~. Live coral abundance declined to 1-2% cover
from values of 50~o or more. Dominant framework species of
Acropora and Montastrea were severely affected. Besides overall
reduction in coral abundance there was a shift in life histories of
surviving species (13, 31-33~. Western Atlantic Acropora and
Montastrea are long-lived and reproduce by mass spawning of
gametes that are fertilized and develop in the water column.
These taxa are being replaced by smaller, shorter-lived Agaricia
and Porites with internal fertilization and direct development,
presumably because of selection for shorter life cycles in a regime
of increased human disturbance.
The principal cause of coral mortality was overgrowth by
macroalgae that exploded in abundance after an unidentified
pathogen caused mass mortality of the enormously abundant
grazing sea urchin Diadema antillarum in 1983-1984 (33, 35,
36~. Increasing frequency of coral disease and bleaching were
also major factors (30, 37, 38~. A likely explanation for the
formerly great abundance of Diadema is overfishing of major
fish predators on Diadema and of large herbivorous fishes that
had competed with Diadema for algal food (refs. 33, 36, 39-41;
Fig. 1~.
Overfishing allowed Diadema to increase in abundance and
compensate for loss of herbivorous fishes that ate macroalgae
before overfishing began. Then, when Diadema died out there
were no other large grazers remaining to consume the algae. A
key question is when overfishing began (9~. Jamaican and other
Caribbean reefs were so severely overfished in the 19th century
that northern salt cod were imported en masse to stave off human
starvation (42, 43~. This early overfishing distorted ecological
perspective to the point that reef fishes are described in the best
modern textbook as small "aquarium species" rarely greater than
20-30 cm long (44~. Most species of reef fishes are indeed small
like other animals (45), but this says nothing about size-
frequency distributions of communities of reef fishes before
overfishing (and ecological investigations) began. Indeed, sev-
eral of the earliest European explorers of the Caribbean (46, 47)
carefully described large-scale native and early colonial fisheries
of sharks, groupers, and other large fishes that have rarely been
seen by most ecologists. Remarkably, the same modern textbook
does not mention these species.
The stage for the collapse of Caribbean reef corals was set by
the loss of large fishes sometime in the 19th century (9~. The first
modern study of Caribbean coral reefs in the 1950s (48) de-
scribed coral communities like those in the Pleistocene when
humans were absent from the Americas (25, 49~. Coral commu-
nities did not change noticeably until the epidemic mortality of
Diadema antillarum in the 1980s because ecological redundancy
5412 1 www.pnas.org/cgi/doi/10.1073/pnas.091092898
fish
, \A
477 _-~~ urchins
coral/algae
Fig. 1. Model of the consequences for reef corals of the increase in the sea
urchin Diadema antillarum caused by overfishing of large predatory and
herbivorous fishes and the subsequent mass mortality of Diadema caused by
disease. Reproduced with permission from ref. 41 (Copyright 1994, The Royal
Society). Plane A, pristine condition, with high ratio of corals to macroalgae
because of intense grazing of macroalgae by fishes. Plane B. abundant Dia-
dema grazed macroalgae formerly consumed by herbivorous fishes so the
ratio of corals to macroalgae remained high despite intensive fishing. Plane C,
mass mortality of Diadema caused by infectious disease allowed macroalgae
to proliferate and overgrow corals.
Of herbivores obscured the potential effects of the loss of large
herbivorous fishes for well over a century (9, 33, 50~. Macroalgae
were not able to overgrow corals until the last major herbivore
was lost from the system. Lapointe suggested that nutrient
enrichment might have tipped the competitive balance of mac-
roalgae over corals (51), but this seems unlikely (40, 52, 53~.
In contrast to macroalgal overgrowth, outbreaks of coral
disease are not understood (54~. Climatic variability, humans as
agents of dispersal of pathogens, habitat degradation, and
pollutants have all been invoked as factors that favor increase of
pathogens (55~. However, there is no clear model or mechanism
for how these factors could affect some species and not others,
or consideration of the profound historical changes that previ-
ously affected reef ecosystems. Outbreaks of disease may be
increasing because of the reduction of other species that once
kept specific pathogens in check. In contrast, increasing fre-
quency of severe episodes of coral bleaching is strongly corre-
lated with high sea surface temperatures, and may truly reflect
changes in global climate (56~.
Caribbean Seagrass Meadows
Tropical American seagrasses are less diverse than corals, but
seagrass meadows cover much greater areas than coral reefs (18,
57~. Seagrasses enhance sediment stability, decrease wave en-
ergy, and increase water clarity as well as providing forage,
habitat, and nurseries for diverse and abundant invertebrates
and fishes (57, 58~. The most common Caribbean species are
turtle grass (~Thalassia testudinum), manatee grass (
Seagrasses along the Florida coast experienced mass mortality
in the 1980s because of a wasting disease (60, 61~. Mortality was
positively density dependent and correlated with high temper-
atures and salinities, sulfide toxicity, self-shading, hypoxia, and
infection by the slime mold Labyrinthula sp. Ecologists search for
causes of seagrass mortality in terms of recent chances in
hydrography and pollution (55, 61). However, all of the above
factors except salinity and temperature have changed greatly
because of massive exploitation centuries ago of sea turtles and
manatees that gave the seagrasses their popular names.
Green turtles (Chelonia myda.s) were extraordinarily abundant
when Columbus arrived in the Caribbean (9, 624. Estimates of
adult populations have been calculated, based on the assumption
that population size was regulated by food limitation and by
extrapolating from early hunting data from the Cayman Islands.
Population sizes based on the carrying capacity of turtle grass
range from 16 to 586 million 50-kg adults (62), whereas estimates
based on early hunting data range from 33 to 39 million large
nesting adults (9~. Even the smallest estimate for green turtles
exceeds the highest recorded wildebeest abundances in the
Serengeti (63 )!
What were the effects on seagrass beds of such enormous
numbers of turtles? Blades of turtlegrass grow upward from
the base and can reach 30 cm or more in length (57~. Older,
more distal portions are commonly heavily overgrown by
microorganisms, fungi, algae, and invertebrates, and are bro-
ken off and transported en masse during storms (64~. Green
turtles crop turtlegrass 2-4 cm above the base, and individuals
commonly return repeatedly to the same plots that are main-
tained by continuously cropping grazed areas to feed on more
nutritious new shoots of the turtlegrass (64~. When density of
turtles is comparatively high, individual grazing plots may
merge so that the entire turtlegrass bed is closely cropped (654.
Such close cropping matches Dampier's (46) description of
turtlegrass blades as "six Inches long" (15 cm) when turtles
were abundant, in comparison with much greater lengths
typically observed today (57~. Grazing by green turtles also
reduces 20-fold the flux of detritus and nitrogen to seagrass
sediments and alters their microbial ecology (64, 66-68~. This
happens because turtles (i) consume more of the blades than
fishes and invertebrates, (ii) metabolize cellulose of cell walls
by microbial fermentation in their hindguts, and (iii) disperse
feces and urine over large areas well away from seagrass beds.
In contrast, fishes and invertebrates feeding on turtlegrass
cannot metabolize the cellulose and do not migrate over such
large areas (68~.
Now consider the potential significance of the ecological
demise of green turtles for turtlegrass in Florida Bay. Green
turtles were formerly very abundant in South Florida (69), and
all of the factors identified in seagrass die-offs except changes
in temperature and salinity would have been profoundly
altered by abundant green turtles. Concentration of sulfides in
sediments increases with accumulation of organic material that
may also cause anoxia within sediments and hypoxia of
overlying waters (70), but green turtles greatly decrease ac-
cumulation of organic matter in sediments (68~. Self-shading
is due to the density and foliage height of the leaves, which also
are greatly reduced by green turtles. Finally, infection by slime
molds is positively correlated with density of turtlegrass (61)
and probably depends on the amount of time senescing leaf
tissues are exposed to the environment. Scientific descriptions
of the sites of infections are vague, but leaf segments free of
lesions caused by slime molds for use in experiments were
always obtained from mid-to-basal sections of leaves (71),
which are the youngest portions (57~. In addition, infections
illustrated in photographs occur along the distal portion of the
blade (http://www.floridamarine.org/~. Thus infection begins
on those older portions of leaves that were typically grazed
Jackson
away when turtles were abundant. Elimination of green turtles
is implicated on four counts as the ultimate factor in die-offs
of turtlegrass; a hypothesis that could be tested by manipula-
tive experiments of abundance of green turtles in turtlegrass
beds on an appropriately massive scale.
The demise of green turtles is better documented (9? 69) than
that of manatees (~Trichechus manatus), which feed on manatee
grass and other submerged vegetation (46, 72) and can metab-
olize cellulose as green turtles do (68~. One- to two-ton manatees
were sufficiently abundant along the low-lying and swampy
coasts of Central America and northern South America to merit
extensive and detailed descriptions of their natural history and
how they were commonly hunted (46~. Moreover, the much
better documented and more recent demise of the dugong
(Dugong dugong) in Australia suggests populations of these
enormous relatives of the manatee of about 1 million along the
Australian coast only a century ago (734. Dugongs plow through
seagrass beds in Australia, reducing shoot density and biomass
of by up to 90% (74~. We will likely never know the equivalent
ecological consequences of manatee grazing in pristine seagrass
environments. However, once again Dampier (46) gives us a clue
when he describes manatee grass as "7 or 8 Inches long" (20 cm)
compared with lengths commonly exceeding 20 inches (50 cm)
today (57~.
Chesapeake Bay
Chesapeake Bay is the largest and historically most productive
estuary in North America. During the 20th century, once very
extensive meadows of seagrasses, oyster beds, clams, blue crabs,
and fish declined precipitously, while abundance and production
of phytoplankton, eutrophication, and episodes of hypoxia and
anoxia correspondingly increased (754. Overfishing and increas-
ing runoff of freshwater, nutrients, and sediment from the land
seem the obvious culprits, but physical conditions are extremely
variable (76) and hypoxia was first reported in the 1930s when
modern ecological research was only just beginning (77, 784.
Thus it is impossible to determine the extent of human influence
solely on the basis of modern observations.
The stratigraphic record of sedimentation, pollen, seeds,
diatoms, and geochemistry in sediment cores was used to
reconstruct the ecological history of the northern half of the
watershed over the past 2,000 years (75, 78-80~. Environmental
and biological fluctuations since European settlement exceed all
earlier changes severalfold. Sedimentation rate and concentra-
tions of organic carbon, sulfur, and ragweed pollen increased
suddenly at the end of the 18th century. Diversity of diatom
species, the ratio of benthic to planktonic diatoms, and the
occurrence of seeds of benthic macrophytes gradually declined.
Altogether, the results from the cores demonstrate an ecological
shift in the upper Chesapeake Bay from predominantly benthic
to predominantly planktonic primary production that was well
under way by the early 19th century.
These results were corroborated by more recent observations
of increasing phytoplankton biomass and decreasing submerged
aquatic vegetation over the past 50-75 years (14, 81~. Decline of
the eelgrass Zostera marina was due primarily to wasting disease
caused by the slime mold Labyrinthula sp., the same genus of
pathogen affecting turtlegrass in Florida Bay (14, 82~. Earliest
reports of declines in eelgrass date from the 1890s, but mortality
affecting >90~o of eelgrass populations along the entire East
Coast of North America occurred in the 1930s (14, 824.
Increase in phytoplankton was compounded by massive
overfishing and physical destruction of oyster beds in the 19th
century (lS, 16, 77) in addition to increased loading of
nutrients, especially nitrogen. Like seagrasses, oysters stabilize
the substratum and provide complex habitat for hundreds of
other species (16~. Large oyster beds were a major hazard to
navigation in bays and estuaries from New England to west
PNAS | May 8, 2001 | vo~. 98 ! no. 30 1 54~3
the true ages may range from as much as 40-60 years for some
species (103~.
Age (as opposed to size) of first reproduction for female
groupers is poorly known but is only 6-7 years for the jewfish,
which is the largest species (104), and numbers of eggs spawned
are in the millions (104, 105~. However, groupers reproduce in
spawning aggregations that previously numbered in the tens to
hundreds of thousands and occurred only at specific places and
times of the year (105-107~. As for sea turtles nesting on beaches,
dense spawning aggregations make groupers easy to fish just at
the time when they have the greatest potential to contribute to
future generations (108~.
Approximately 70% of living sharks and rays bear live young,
and hammerheads exhibit placental viviparity (109, 110~. Ages of
maturation typically range from 6 to 18 years, but lemon sharks
take 24 years. Gestation periods are long (6 to 22 months) and
clutch sizes small (2 to 135~. Thus, it is hardly surprising that
sharks exhibit sudden collapse and slow recovery after relatively
few years of intensive fishing (111, 112~.
Collapse of Sessile Ecosystem Engineers. "Ecosystem engineers" are
species that modify, maintain, or create habitats, thereby mod-
ulating availability of resources to other species (113~. Reef-
building corals, seagrasses, oysters, and kelps are among the
most important ecosystem engineers in marine coastal environ-
ments. Their massive physical presence and three-dimensional
complexity help stabilize the physical environment and provide
habitat to thousands of generally smaller associated species (6,
22, 23, 50, 57, 58, 89, 91, 114~. Once-vast populations of
ecosystem engineers have now collapsed along the Western
Atlantic coast from the southern Caribbean to the Gulf of
Maine. The reasons range from complex shifts in competitive
abilities of corals, seagrasses, and kelps after the removal of
keystone consumer species or outbreaks of disease (refs. 18, 33,
61,90-92; Fig.1) to direct physical destruction of oyster beds and
sponge-bryozoan gardens by mechanical dredging and trawling
(15, 16, 98, 99, 114, 115~.
Once, great coral reefs, seagrass meadows, and oyster reefs
were products of growth of dominant framework species and
accumulation of sediments and skeletal debris. Dead skeletons
remain partially intact for various periods after the death of
corals and oysters unless removed by mechanized harvesting,
whereas sea grasses and kelps do not produce such durable
remains, so that three-dimensionality rapidly disappears (90~.
Loss of habitat structure decreases growth and larval recruit-
ment and increases mortality of engineering species (12, 31, 115~.
Diversity and abundance of associated species also drops pre-
cipitously (18, 116~.
Time Lags Between Effects of Overfishing and Collapse of Ecosystem
Engineers. Lengthy time lags between initial harvesting and many
of the resulting ecological consequences are pervasive in tropical
forests (117~. Similarly in the coastal ocean, time lags of decades
to centuries occurred between initial harvesting or destruction of
large vertebrates and subsequent collapse of ecosystem engi-
neers such as reef corals, seagrasses, or kelps (9, 33, 61, 90~.
Similar lags are apparent between increased fluxes of nutrients
and sediments into coastal environments and collapse of reef
corals (118), submerged macrophytes (14, 76), or oysters (15,16,
85~. Of course, oysters were intensively harvested by mining
down the habitat, so their abundance declined much more
rapidly than unfished corals and seagrasses.
One likely explanation for time lags is ecological redun-
dancy, whereby other species take over the ecological role of
species removed by harvesting. This is presumably what hap-
pened after overfishing on coral reefs (ref. 33; Fig. 1) and
extirpation of green turtles in the Caribbean (9~. Ecological
redundancy should increase with taxonomic diversity, which
Jackson
may explain why time lags in the destruction of ecosystem
engineers appear to decrease northward from corals and
seagrasses in the Caribbean to kelps in the Gulf of Maine.
Another important factor is widespread occurrence of thresh-
old effects on human altered ecosystems (33, 119-122~. These
may involve simple thresholds in physiological tolerance to
decreasing light or increasing sediments and nutrients, or more
subtle density-dependent consequences of reduced abundance
on fertilization, recruitment, or the ability to filter large
volumes of water that reduces abundance of phytoplankton.
Such negative feedbacks are exacerbated by the fact that both
over-harvesting and increased nutrients and primary produc-
tion work synergistically to reduce abundance of sessile eco-
system engineers (9, 76, 85, 114~.
1
~ . .
Fishing Down Food Webs. Top carnivores were never an important
part of the human diet on land (123) but are the preferred large
prey in the sea except for green turtles and sirenians. Smaller and
smaller fishes, sea urchins, lobsters, and shrimps are replacing
large fishes, turtles, and sharks as the remnant fisheries in all of
the coastal ecosystems discussed herein (9, 33, 90, 95, 124~.
Free-living animals larger than 1 kg are increasingly rare and
nearly absent on the reefs of Jamaica and many other sites
throughout the Caribbean (33, 125~. The process is reversible,
but only by regulation of fishing.
Farming of the sea, or aquaculture, is a possible alternative to
fishing, but one that carries its own set of potentially harmful
consequences to coastal ecosystems, including eutrophication,
pollution, and the spread of disease (126, 127~. Cultured species
include a wide diversity of algae, oysters, shrimps, and various
fishes from mullets to salmon. Most of the problems of aqua-
culture of algae and herbivorous animals could be alleviated if
goals were broadened to include ecosystem conservation and
management, rather than only to produce food. For example,
benthic algae could be farmed to remove excess nitrogen from
the water column, and oysters and other suspension-feeding
bivalves could be farmed to reduce algal blooms induced by
eutrophication.
Rise of Microbes. Fishing down marine food webs and increasing
pollution from the land are resulting in increasing abundance
and widespread dominance of ecosystem processes by microbes.
Eutrophication is most apparent in bays and estuaries like
Chesapeake Bay (86), but it has extended onto the continental
shelf (88~. Outbreaks of previously rare or unreported toxic
microbes and diseases are another example of the increasing
importance of microbial disruption of coastal ecosystems (30,35,
54, 554.
General Model of Coastal Ecosystem Collapse. I summarized much
of the above in the simple qualitative model in Fig.2 showing the
demise of large animals and ecosystem engineers and the rise of
microbes since European colonization of the Americas. The
model is based on Western Atlantic case studies reviewed in this
paper, but I predict the same general pattern will obtain for the
entire global coastal ocean. They-axes are logarithmic to capture
the orders of magnitude changes in these variables. The time axis
is deliberately general because onset of major changes depends
more on timing of the onset of intensive harvesting or develop-
ment of new fishing technologies than chronological age.
Early ecological extinction of large mobile animals defines the
first major transition in the history of coastal marine ecosystems.
Extirpation of large vertebrates preceded ecological investiga-
tions so that their absence has been uncritically accepted as the
natural "baseline" condition. Their precipitous decline reflects
greater economic desirability, ease of capture, and limited
capacity for increase that is well documented for sea turtles,
manatees, large fishes such as cod and groupers, and sharks. The
PNAS | May S. 2001 | vol. 98 | no. 10 | 5415
n
o
living ha bitat
stnJctum
large
mobile
animals \
then ~
1T
"ha se line"
2
now
3
Fig. 2. Model of the collapse of Western Atlantic coastal ecosystems caused
by overfishing. Arrows indicate the three major ecological transitions dis-
cussed in the text.
second major transformation reflects sudden collapse of sessile
ecosystem engineers (reef corals, seagrasses, and kelps) caused
by indirect effects of overfishing large vertebrates. Ecological
dominance of microbes at the expense of macroorganisms (86)
and increasing frequency of invasions of exotic species (128, 129)
define the third major transition that is increasingly upon us (54,
55, 86, 884.
tool for management, but no more monitoring is required to
know what we have lost. Scientific efforts should be redirected
toward evaluating options for restoration of resources rather
microbes ~ than perpetuating the myth of sustainable fisheries. It is hard to
imagine how increasingly sophisticated and frequent environ-
I mental monitoring and micromanagement could do a fraction of
the good of simply stopping fishing. There is no rational scientific
I ~ basis to continue fishing of wild stocks along the Atlantic coast
of North America or in the Caribbean for the foreseeable future.
(2) Paleoecological, archeological, and historical reconstruc-
tions of coastal marine ecosystems provide the beset evidence for
predicting ecological consequences of establishing very large-
scale marine reserves and other forms of rigorous protection of
fisheries. Formerly pristine conditions of seagrass beds and
oyster reefs of Chesapeake Bay (14-16), or of Caribbean coral
reefs and seagrass beds and the hordes of large animals that lived
upon them (9), seem fantastic and unbelievable today. Scientists,
as well as the general public, set goals and expectations for
marine reserves that are too low because they cannot imagine
how coastal ecosystems used to be only a century ago (1, 2~.
These great changes, and frequently nonlinear transformations
among alternative ecosystem states (31, 33, 119-122), make it
almost impossible to predict the outcomes of complete protec-
tion from fishing and terrestrial inputs based on recent obser-
vations alone. Fortunately, historical records tell us what is
possible. Because few of the large apex predators and herbivores
are extinct, we could restore coastal resources for ecosystem
services and managed harvest.
Why History Matters
Oceans are not wilderness and no Western Atlantic coastal
habitat is pristine. The same is almost certainly true of coastal
oceans worldwide, but this assertion needs rigorous documen-
tation. Neotropical forests are greatly threatened by human
activities and may disappear entirely within this century (1174.
The facts about tropical forests are widely known and much
discussed by governments, international agencies, and the gen-
eral public. By comparison, Neotropical coral reefs are already
effectively "deforested" throughout their entire range, but this
fact received almost no comparable attention until the 1990s (33,
130, 131~. Moreover, human activities leading to the destruction
of coral and oyster reefs, seagrass beds, or kelp forests began
early in the 19th century or earlier, long before comprehensive
scientific study began. In general, we are more aware of the mass
extinction of large vertebrates at the end of the Pleistocene (123)
than what happened in coastal seas only a century ago!
As in geology, the present is not always the key to the past, or
to the future (1324. Understanding what was natural is important
not just for historical curiosity, but for rational management and
conservation of coastal oceans in the future. I conclude with
three basic points that emerge from comparisons of present
conditions with historical baselines.
(1) No wild Atlantic coastal fishery is sustainable at anything
close to present levels of exploitation. Coastal marine ecosys-
tems already have been changed beyond recognition because of
direct and indirect effects of overfishing. Most fishing is unsus-
tainable because (i) inexorable growth of the human population
drives increasing demand, (
1. Pauly, D. (1995) Trends Ecol. Evol. 10, 430.
2. Sheppard, C. (1995) Mar. Poll. Bull. 30, 766-767.
3. Schama, S. (1995) Landscape and Memory (Knopf, New York).
4. McCall, A. D. (1996) Calif Coop. Fish. Invest. 37, 100-110.
5. Dayton, P. K., Tegner, M. J., Edwards, P. B. & Riser, K. L. (1998) Ecol. Appl.
8, 309-322.
6. Dayton, P. K., Tegner, M. J., Edwards, P. B. & Riser, K. L. (1999) Ecol.
AJonogr. 69, 219-250.
Baumgartner, T. R., Soutar, A. & Frerreira-Bartrina, V. (1992) Calif Coop.
Fish. Invest. 33, 24-40.
8. Anderson, P. J. & Piatt, J. F. (1999) Mar. Ecol. Prog. Ser. 189, 117-123.
9. Jackson, J. B. C. (1997) Coral Reefs 16, S23-S32.
10. Davis, G. E. (1982) Bull. Mar. Sci. 32, 608-623.
11. Connell, J. H., Hughes, T. P. & Wallace, C. C. (1997) Ecol. Monogr. 67,
461-488.
12. Hughes, T. P. & Tanner, J. E. (2000) Ecology 81, 2250-2263.
13. Jackson, J. B. C. (1991) BioScience 41, 475-482.
14. Orth, R. J. & Moore, K. A. (1984) Estuaries 7, 531-540.
15. Rothschild, B. J., Ault, J. S., Goulletquer, P. & Heral, M
Prog. Ser. 111, 29-39.
16. Kennedy, V. S. (1996) J. Shellfish Res. 15, 177-183.
17.
(1994) Mar. Ecol. 64.
. Hall, M. O., Durako, M. J., Fourqurean, J. W. & Zieman, J. C. (1999) Estuaries
22, 445-459.
18. Fourqurean, J. W. & Robblee, M. B. (1999) Estuaries 22, 345-357.
19. Cushing, D. H. (1988) The Provident Sea (Cambridge Univ. Press, Cambridge,
U.K.).
20. Botsford, L. W., Castilla, J. C. & Peterson, C. H. (1997) Science 277, 509-515.
21. Carlton, J. T., Geller, J. B., Reaka-Kudla, M. L. & Norse, E. A. (1999)Annu.
Rev. Ecol. Syst. 30, 515-538.
22. Paulay, G. (1997) in Life and Death of Coral Reefs, ed. Birkeland, C. (Chapman
& Hall, New York), pp. 298-353.
23. Reaka-Kudla, M. L. (1997) inBiodiversityII, eds. Reaka-Kudla, M. L., Wilson,
D. E. & Wilson, E. O. (Joseph Henry Press, Washington, DC), pp. 83-108.
24. Knowlton, N. & Jackson, J. B. C. (1994) Trends Ecol. Evol. 9, 7-9.
25. Jackson, J. B. C. (1992) Am. Zool. 32, 719-731.
26. Pandolfi, J. M. & Jackson, J. B. C. (1997) in Proceedings of the Eighth
International Coral Reef Symposium, eds. Lessios, H. A. & Macintyre, I. G.
(Smithsonian Tropical Research Institute, Balboa, Panama), Vol. 1, pp.
397-404.
27. Pandolfi, J. M. & Jackson, J. B. C. (2001) Ecol. Monogr. 71, 49-67.
28. Aronson, R. B., Precht, W. F. & Macintyre, I. G. (1998) Coral Reefs 17,
223-230.
29. Greenstein, B. J., Curran, H. A. & Pandolfi, J. M. (1998) Coral Reefs 17,
249-261.
30. Gladfelter, W. B. (1982) Bull. Mar. Sci. 32, 639-643.
31. Knowlton, N., Lang, J. C. & Keller, B. D. (1990) Smithson. Contr. Mar. Sci. 78.
31, 1-25. 79.
32. Guzman, H. M., Jackson, J. B. C. & Weil, E. (1991) Coral Reefs 10, 1-12. 80.
33. Hughes, T. P. (1994) Science 265, 1547-1551. 82.
34. Bak, R. P. M. & Nieuwland, G. (1995) Bull. Mar. Sci. 56, 609-619. 83
35. Lessios, H. A., Robertson, D. R. & Cubit, J. D. (1984) Science 226, 335-337.
36. Lessios, H. A. (1988) Annul Rev. Ecol. Syst. 19, 371-393.
37. McClanahan, T. R. & Muthiga, N. A. (1998) Environ. Conserv. 25, 12-130.
38. Aronson, R. B. & Precht, W. F. (2000) Hydrobiologia, 1-14.
39. Hay, M. E. (1984) Ecology 65, 446-454.
40. Lewis, S. M. (1986) Ecol. Monogr. 56, 183-200.
41. Jackson, J. B. C. (1994) Proc. Roy. Soc. London Ser. B 344, 55-60.
42. Duerden, J. E. (1901) J. Imp. Agric. Dept. W. Indies, 121-141.
43. Thompson, E. F. (1945) Development and Welfare in the West Indies, Bulletin
18 (Bridgetown, Barbados).
44. Sale, P. F. (1991) in The Ecology of Fishes on Coral Reefs, ed. Sale, P. F.
(Academic, San Diego), pp. 3-15.
45. Brown, J. H. (1995) Macroecology (Univ. of Chicago Press, Chicago).
46. Dampier, W. (1729) A New Voyage Around the World (James and John
Knapton, London); reprinted (1968) (Dover, New York).
47. de Oviedo, G. F. (1535-1557) Historia General y Natural de las Indias, Islas y
Tierra-Firme del Mar Oceano; reprinted (1959) (Atlas, Madrid).
48. Goreau, T. F. (1959) Ecology 40, 67-89.
49. Mesolella, K. J. (1967) Science 156, 638-640.
50. Done, T. J., Ogden, J. C., Wiebe, W. J. & Rosen, B. R. (1996) in Functional
Roles of Biodiversity:A Global Perspective, eds. Mooney, H. A., Cushman, J. H.,
Medina, E., Sala, O. E. & Schulze, E.-D. (Wiley, London), pp. 393-429.
51. Lapointe, B. E. (1997) Limnol. Oceanogr. 42, 1119-1131.
52. Hughes, T. P., Szmant, A. M., Steneck, R., Carpenter, R. & Miller, S. (1999)
Limnol. Oceanogr. 44,1583-1586.
53. Miller, M. W., Hay, M. E., Miller, S. L., Malone, D., Sotka, E. E. & Szmant,
A. M. (1999) Limnol. Oceanogr. 44,1847-1861.
54. Richardson, L. L. (1998) Trends Ecol. Evol. 13, 438-443.
Jackson
55. Harvell, C. D., Kim, K., Burkholder, J. M., Colwell, R. R., Epstein, P. R.,
Grimes, D. J., Hofman, E. E., Lipp, E. K., Osterhaus, A. D. M. E., Overstreet,
R. M., et al. (1999) Science 285, 1505-1510.
56. Aronson, R. B., Precht, W. F., Macintyre, I. G. & Murdoch, T. J. T. (2000)
Nature (London) 405, 36.
57. Zieman, J. C. (1982) The Ecology of the Seagrasses of South Florida: A
Community Profile (U.S. Fish and Wildlife Service, Office of Biological
Services, Washington, DC).
58. Orth, R. J., Heck, K. L., Jr., & van Montfrans, J. (1984) Estuaries 7, 339-350.
59. Jackson, J. B. C. (1972) Bull. Mar. Sci. 23, 313-350.
60. Robblee, M. B., Barber, T. R., Carlson, P. R., Jr., Durako, M. J., Fourqurean,
J. W., Muehlstein, L. K., Porter, D., Yarbro, L. A., Zieman, R. T. & Zieman,
J. C. (1991) Mar. Ecol. Prog. Ser. 71, 297-299.
61. Zieman, J. C., Fourqurean, J. W. & Frankovich, T. A. (1999) Estuaries 22,
460-470.
62. Bjorndal, K. A., Bolten, A. B. & Chaloupka, M. Y. (2000) Ecol. Appl. 10,
269-282.
63. Mduma, S. A. R, Sinclair, A. R. E. & Hilborn, R. (1999) J. Anim. Ecol. 68,
1101-1122.
.. Ogden, J. C. (19803 in Handbook of Seagrass Biology: An Ecosystem Perspec-
tive, eds. Phillips, R. C. & McRoy, C. P. (Garland, New York), pp. 173-198.
65. Williams, S. L. (1988) Mar. Biol. 98, 447-455.
66. Ogden, J. C., Robinson, L., Whitlock, H., Daganhart, H. & Cebula, R. (1983)
J. Exp. Mar. Biol. Ecol. 66, 199-205.
67. Thayer, G. W., Engel, D. W. & Bjorndal, K. A. (1982) J. Exp. Mar. Biol. Ecol.
62, 173-183.
68. Thayer, G. W., Bjorndal, K. A., Ogden, J. C., Williams, S. L. & Zieman, J. C.
(1984) Estuaries 7, 351-376.
69. King, F. W. (1982) in Biology and Conservation of Sea Turtles, ed. Bjorndal,
K. (Smithsonian Inst. Press, Washington, DC), pp. 183-188.
70. Carlson, P. R., Yarbro, L. A. & Barber, T. A. (1994) Bull. Mar. Sci. 54,
733-746.
71. Durako, M. J. & Kuss, K. M. (1994) Bull. Mar. Sci. 54, 727-732.
72. Reynolds, J. E., III, & Odell, D. K. (1991) Manatees and Dugongs (Facts On
File, New York).
73. Lanyon, J. M., Limpus, C. J. & Marsh, H. (1989) in Biology of Seagrasses: A
Treatise on the Biology of Seagrasses with Special Reference to the Australian
Region, eds. Larkum, A. W. D., McComb, A. J. & Shepherd, S. A. (Elsevier,
Amsterdam), pp. 610-634.
74. Preen, A. (1995) Mar. Ecol. Prog. Ser. 124, 201-213.
75. Brush, G. S. (1994) in The Changing Global Environment, ed. Roberts, N.
(Blackwell, Oxford), pp. 398-416.
76. Cronin, T., Willard, D., Karlsen, A., Ishman, S., Verardo, S., McGeehin, J.,
Kerhin, R., Holmes, C., Colman, S. & Zimmerman, A. (2000) Geology 28, 3-6.
77. Wharton, J. (1957) The Bounty of the Chesapeake (Virginia Anniversary
Celebration Corp.), Jamestown 390th Anniversary Historical Booklet 13.
Cooper, S. R. & Brush, G. S. (1991) Science 254, 992-996.
Cooper, S. R. & Brush, G. S. (1993) Estuaries 16, 617-626.
Cooper, S. R. (1995) Ecol. Appl. 5, 703-723.
Harding, L. W., Jr., & Perry, E. S. (1997) Mar. Ecol. Prog. Ser. 157, 39-52.
Muehlstein, L. K. (1989) Dis. Aquat. Organ. 7, 211-221.
. Officer, C. B., Smayda, T. J. & Mann, R. (1982) Mar. Ecol. Prog. Ser. 9,
203-210.
84. Newell, R. I. E. (1988) in Understanding the Estuary: Advances in Chesapeake
Bay Research, eds. Lynch, M. P. & Krome, E. C. (Chesapeake Research
Consortium 129, CBP/TRS 24/88).
85. Ulanowicz, R. E. & Tuttle, J. H. (1992) Estuaries 15, 298-306.
86. Jonas, R. B. (1997) Am. Zool. 37, 612-620.
87. Paerl, H. W., Pinckney, J. L., Fear, J. M. & Peierls, B. L. (1998) Mar. Ecol.
Prog. Ser. 166, 17-25.
88. Turner, R. E. & Rabalais, N. N. (1994) Nature (London) 368, 619-621.
89. Mann, K. H. (1982) Ecology of Coastal Waters: A Systems Approach (Univ. of
California Press, Berkeley).
90. Steneck, R. S. (1997) in Proceedings of the Gulf of Maine Ecosystem Dynamics
Scientific Symposium and Workshop, RARGOM Report 91-1, eds. Wallace,
G. T. & Braasch, E. F. (Regional Association for Research on the Gulf of
Maine, Hanover, NH), pp. 151-165.
91. Elner, R. W. & Vadas, R. L. (1990) Am. Nat. 136, 108-125.
92. Witman, J. D. & Sebens, K. P. (1992) Oecologia 90, 305-315.
93. Bourque, B. J. (1995) Diversity and Complexity in Prehistoric Maritime
Societies: A Gulf of Maine Perspective (Plenum, New York).
94. de la Morandiere, C. (1962-1966) Histoire de la Peche Fran~caise de la Morue
dans l'Amc~rique Septentrionale (Maisonneuve and Larose, Paris), Vols. 1-3.
95. Pauly, D., Christensen, V., Dalsgaard, J., Froese, R. & Torres, F., Jr. (1998)
Science 279, 860-863.
96. Simenstad, C. A., Estes, J. A. & Kenyon, K. W. (1978) Science 200, 403-411.
97. Estes, J. A., Duggins, D. O. & Rathburn, G. B. (1989) Conserv. Biol. 3,
252-264.
98. Dayton, P. K, Thrush, S. F., Agardy, M. T. & Hofman, R. J. (1995) Aquat.
Conserv. Mar. Freshwater Ecosyst. 5, 205-232.
PNAS 1 May 8, 2001 1 vol. 98 1 no. 10 1 5417
99. Watling, L. & Norse, E. A. (1998) Conserv. Biol. 12, 1180-1197.
100. Casey, J. M. & Myers, R. A. (1998) Science 281, 690-692.
101. Committee on Sea Turtle Conservation, National Research Council (1990)
Decline of the Sea Turtles: Causes and Prevention (Natl. Acad. Press, Wash-
ington, DC).
102. Von Buskirk, J. (1994) Copeia 1994, 66-81.
103. Bjorndal, K. A. & Zug, G. R. (1995) in Biology and Conservation of Sea Turtles,
ed. Bjorndal, K. A. (Smithsonian Inst. Press, Washington, OC), 2nd. Ed., pp.
599-600.
104. Bullock, L. H., Murphy, M. D., Godcharles, M. F. & Mitchell, M. E. (1992)
Fish. Bull. 90, 243-249.
105. Carter, J., Marrow, G. J. & Pryor, V. (1990) Proc. Gulf Caribb. Fish. Inst. 43,
65-111.
106. Sadovy, Y. (1990) Proc. Gulf Caribb. Fish. Inst. 43, 43-64.
107. Smith, C. L. (1972) Trans. Am. Fish. Soc. 101, 257-261.
108. Johannes, R. E. (1976) Words of the Lagoon (Univ. of California Press,
Berkeley).
109. Branstetter, S. (1990) in Elasmobranchs as Living Resources: Advances in the
Biology, Ecology, Systematics, and the Status of the Fisheries, eds. Pratt, H. L.,
Jr., Gruber, S. H. & Taniuchi, T. (U.S. Dept. of Commerce, Washington, DC),
NOAA Technical Report NMFS 90, pp. 17-28.
110. Pratt, H. L., Jr., & Casey, J. G. (1990) in Elasmobranchs as Living Resources:
Advances in the Biology, Ecology, Systematics, and the Status of the Fisheries,
eds. Pratt, H. L., Jr., Gruber, S. H. & Taniuchi, T. (U.S. Dept. of Commerce,
Washington, DC), NOAA Technical Report NMFS 90, pp. 97-109.
111. Compagno, L. J. V. (1990) in Elasmobranchs as Living Resources: Advances in
the Biology, Ecology, Systematics, and the Status of the Fisheries, eds. Pratt,
H. L., Jr., Gruber, S. H. & Taniuchi, T. (U.S. Dept. of Commerce, Wash-
ington, DC), NOAA Technical Report NMFS 90, pp. 391-414.
112. Anderson, E. D. (1990) in Elasmobranchs as Living Resources: Advances in the
Biology, Ecology, Systematics, and the Status of the Fisheries, eds. Pratt, H. L.,
Jr., Gruber, S. H. & Taniuchi, T. (U.S. Dept. of Commerce, Washington, DC),
NOAA Technical Report NMFS 90, pp. 473-484.
113. Jones, C. G., Lawton, J. H. & Shachak, M. (1994) Oikos 69, 373-386.
114. Lenihan, H. S. & Peterson, C. H. (1998) Ecol. Appl. 8, 128-140.
115. Lenihan, H. S. (1999) Ecol. Monogr. 69, 251-275.
5418 1 www.pnas.org/cgi/doi/10.1073/pnas.091092898
116. Lenihan, H. S., Peterson, C. H., Byers, J. E., Grabowski, J. H., Thayer, G. W.
& Colby, D. R. (2001) Ecol. Appl., in press.
117. Terborgh, J. (1999) Requiem for Nature (Island Press, Washington, DC).
118. Lewis, J. B. (1984) Coral Reefs 3, 117-122.
119. Sutherland, J. P. (1974)Am. Nat. 108, 859-873.
120. Peterson? C. H. (1984) Am. Nat. 124, 127-133.
121. Knowlton, N. (1992) Am. Zool. 32, 674-682.
122. Paine, R. T., Tegner, M. J. & Johnson, E. A. (1998) Ecosystems 1, 535-545.
123. Diamond, J. (1997) Guns, Germs and Steel: The Fates of Human Societies
(Norton, New York).
124. Steneck, R. S. (1998) Trends Ecol. Evol. 13, 429-430.
125. Munro, J. L. (1983) ICLARM Study Rev. 7, 1-276.
126. Svennevig, N., Reinertsen, H. & New, M., eds. (1999) SustainableAquaculture:
Food for the Future? (Balkema, Rotterdam, The Netherlands).
. Naylor, R. L., Goldburg, R. J., Primavera, J. H., Kaufsky, N., Beveredge,
M. C. M., Clay, J., Folke, C., Lubchenco, J., Mooney, H. & Troell, M. (2000)
Nature (London) 405, 1017-1024.
128. Carlton, J. T. (1989) Conserv. Biol. 3, 265-273.
129. Meinesz, A. (1999) Killer Algae (Univ. of Chicago Press, Chicago).
130. Wilkinson, C. R. (1992) in Proceedings of the Seventh International Coral Reef
Symposium, ed. Richmond, R. (Univ. of Guam Press, Mangilao, Guam), Vol.
1, pp. 11-21.
131. Ginsburg, R. N., compiler (1994) Proceedings of the Colloquium on Global
Aspects of Coral Reefs: Health, Hazards and History 1993 (Rosenstiel School
of Marine and Atmospheric Science, Univ. of Miami, Miami).
132. Ager, D. (1993) The New Catastrophism (Cambridge Univ. Press, Cambridge,
U.K.).
133. Dayton, P. K. (1998) Science 279, 821-822.
134. Paine, R. T. (1992) Nature (London) 3S5, 73-75.
135. Wootton, J. T. (1997) Ecol. Monogr. 67, 45-64.
136. Castilla, J. C. (1999) Trends Ecol. Ecol. 14, 280-283.
137. Murray, S. N., Ambrose, R. F., Bohnsack, J. A., Botsford, L. W., Carr, M. H.,
Davis, G. E., Dayton, P. K., Gotshall, D., Gunderson, D. R., Hixon, M. A., et
al. (1999) Fisheries 24, 11-25.
138. Bowen, W. D. (1997) Mar. Ecol. Prog. Ser. 1S8, 267-274.
139. Boesch, D. F. (2000) Env. Res. Sect. A 82, 134-142.
Jackson
