Here is the complete text of the recently discussed historical overfishing
article, excerpts of which made it into mainstream media.  E-mail me if you
want the pdf version.

Cam

****************************************************************************

Historical Overfishing and the Recent Collapse of Coastal Ecosystems

Jeremy B. C. Jackson,1, 2* Michael X. Kirby,3 Wolfgang H. Berger,1 Karen A.
Bjorndal,4 Louis W. Botsford,5 Bruce J. Bourque,6 Roger H. Bradbury,7
Richard Cooke,2 Jon Erlandson,8 James A. Estes,9 Terence P. Hughes,10 Susan
Kidwell,11 Carina B. Lange,1 Hunter S. Lenihan,12 John M. Pandolfi,13
Charles H. Peterson,12 Robert S. Steneck,14 Mia J. Tegner,1 Robert R.
Warner15 

Ecological extinction caused by overfishing precedes all other pervasive
human disturbance to coastal ecosystems, including pollution, degradation of
water quality, and anthropogenic climate change. Historical abundances of
large consumer species were fantastically large in comparison with recent
observations. Paleoecological, archaeological, and historical data show that
time lags of decades to centuries occurred between the onset of overfishing
and consequent changes in ecological communities, because unfished species
of similar trophic level assumed the ecological roles of overfished species
until they too were overfished or died of epidemic diseases related to
overcrowding. Retrospective data not only help to clarify underlying causes
and rates of ecological change, but they also demonstrate achievable goals
for restoration and management of coastal ecosystems that could not even be
contemplated based on the limited perspective of recent observations alone.

1 Scripps Institution of Oceanography, University of California, San Diego,
La Jolla, CA 92093-0244, USA.
2 Center for Tropical Paleoecology and Archeology, Smithsonian Tropical
Research Institute, Box 2072, Balboa, Republic of Panama.
3 National Center for Ecological Analysis and Synthesis, 735 State Street,
Suite 300, Santa Barbara, CA 93101, USA.
4 Archie Carr Center for Sea Turtle Research and Department of Zoology,
University of Florida, Gainesville, FL 32611, USA.
5 Department of Wildlife, Fish, and Conservation Biology, University of
California, Davis, CA 95616, USA.
6 Department of Anthropology, 155 Pettengill Hall, Bates College, Lewiston,
ME 04240, USA. 
7 Centre for Resource and Environmental Studies, Australian National
University, Canberra, ACT 0200, Australia.
8 Department of Anthropology, University of Oregon, Eugene, OR 97403, USA.
9 U.S. Geological Survey, A-316 Earth and Marine Sciences Building,
University of California, Santa Cruz, CA 95064, USA.
10 Center for Coral Reef Biodiversity, Department of Marine Biology, James
Cook University, Townsville, QLD 4811, Australia.
11 Department of Geophysical Sciences, University of Chicago, 5734 South
Ellis Avenue, Chicago, IL 60637, USA.
12 Institute of Marine Sciences, University of North Carolina at Chapel
Hill, 3431 Arendell Street, Morehead City, NC 28557, USA.
13 Department of Paleobiology, National Museum of Natural History,
Smithsonian Institution, Washington, DC 20560-0121, USA.
14 School of Marine Sciences, University of Maine, Darling Marine Center,
Orono, ME 04573, USA.
15 Department of Ecology, Evolution, and Marine Biology, University of
California, Santa Barbara, CA 93106, USA.
*   To whom correspondence should be addressed. E-mail: jbcj@uc*.ed*
   Deceased. 
------------------------------------------------------------------------
Few modern ecological studies take into account the former natural
abundances of large marine vertebrates. There are dozens of places in the
Caribbean named after large sea turtles whose adult populations now number
in the tens of thousands rather than the tens of millions of a few centuries
ago (1, 2). Whales, manatees, dugongs, sea cows, monk seals, crocodiles,
codfish, jewfish, swordfish, sharks, and rays are other large marine
vertebrates that are now functionally or entirely extinct in most coastal
ecosystems (3-10). Place names for oysters, pearls, and conches conjure up
other ecological ghosts of marine invertebrates that were once so abundant
as to pose hazards to navigation (11), but are witnessed now only by massive
garbage heaps of empty shells.

Such ghosts represent a far more profound problem for ecological
understanding and management than currently realized. Evidence from
retrospective records strongly suggests that major structural and functional
changes due to overfishing (12) occurred worldwide in coastal marine
ecosystems over many centuries. Severe overfishing drives species to
ecological extinction because overfished populations no longer interact
significantly with other species in the community (5). Overfishing and
ecological extinction predate and precondition modern ecological
investigations and the collapse of marine ecosystems in recent times,
raising the possibility that many more marine ecosystems may be vulnerable
to collapse in the near future.

Importance of Historical Data
Most ecological research is based on local field studies lasting only a few
years and conducted sometime after the 1950s without longer term historical
perspective (1, 8, 13). Such observations fail to encompass the life-spans
of many ecologically important species (13, 14) and critically important
environmental disturbances such as extreme cyclones or ENSO (El
Niño-Southern Oscillation) events (8), as well as longer term cycles or
shifts in oceanographic regimes and productivity (15-17). To help address
this problem, we describe ecosystem structure predating modern ecological
studies using well-dated time series based on biological (18, 19),
biogeochemical (20, 21), physical (22), and historical (23) proxies that are
informative over a variety of spatial scales and biogeographic realms (24).
Although proxies vary in precision and clarity of the signals they measure,
the use of multiple proxies that give the same ecological signal greatly
increases confidence in results. Precision in age dating varies from
centuries to a single year, season, or event in the exceptional case of
varved sediments, ice cores, and written historical records (25). Precision
decreases with the amount of biological or physical disturbance to the
sediment analyzed (26).

We exploited data from many disciplines that span the period over which
anthropogenic changes may have occurred. Because our hypothesis is that
humans have been disturbing marine ecosystems since they first learned how
to fish, our time periods need to begin well before the human occupation or
European colonization of a coastal region. Broadly, our data fall into four
categories and time periods:

1) Paleoecological records from marine sediments from about 125,000 years
ago to the present, coinciding with the rise of modern Homo sapiens.

2) Archaeological records from human coastal settlements occupied after
about 10,000 years before the present (yr B.P.) when worldwide sea level
approached present levels. These document human exploitation of coastal
resources for food and materials by past populations that range from
small-scale aboriginal societies to towns, cities, and empires.

3) Historical records from documents, journals, and charts from the 15th
century to the present that document the period from the first European
trade-based colonial expansion and exploitation in the Americas and the
South Pacific (23).

4) Ecological records from the scientific literature over the past century
to the present covering the period of globalized exploitation of marine
resources. These also help to calibrate the older records.

Time Periods, Geography, and Analysis
We recognize three different but overlapping periods of human impact on
marine ecosystems: aboriginal, colonial, and global. Aboriginal use refers
to subsistence exploitation of near-shore, coastal ecosystems by human
cultures with relatively simple watercraft and extractive technologies that
varied widely in magnitude and geographic extent. Colonial use comprises
systematic exploitation and depletion of coastal and shelf seas by foreign
mercantile powers incorporating distant resources into a developing market
economy. Global use involves more intense and geographically pervasive
exploitation of coastal, shelf, and oceanic fisheries integrated into global
patterns of resource consumption, with more frequent exhaustion and
substitution of fisheries. In Africa, Europe, and Asia, these cultural
stages are strongly confounded in time and space, so that their differential
significance is difficult to establish. However, in the Americas, New
Zealand, and Australia the different stages are well separated in time, and
the aboriginal and colonial periods began at different times in the
different regions. Thus, we can distinguish between cultural stages, as well
as between human impacts and natural changes due to changing climate.

The addition of a deep historical dimension to analyze and interpret
ecological problems requires that we sacrifice some of the apparent
precision and analytical elegance prized by ecologists (1, 13, 14).
Paleoecological, archaeological, and historical data were collected for many
purposes, vary widely in methods of collection and quality, and are less
amenable to many types of statistical analysis than well-controlled
experiments. But none of these problems outweighs the benefits of a
historical approach. Clearly, we cannot generate realistic null hypotheses
about the composition and dynamics of ecosystems from our understanding of
the present alone, since all ecosystems have almost certainly changed due to
both human and natural environmental factors (8, 16, 27, 28). Here, we
briefly review long-term human impacts in several key marine ecosystems.
These reconstructions provide insight into the nature and extent of degraded
ecosystems that point to new strategies for mitigation and restoration that
are unlikely to emerge from modern monitoring programs.

Kelp Forests
Kelp forests characterize shallow, rocky habitats from warm temperate to
subarctic regions worldwide and provide complex environments for many
commercially important fishes and invertebrates (29). Northern Hemisphere
kelp forests have experienced widespread reductions in the number of trophic
levels and deforestation due to population explosions of herbivores
following the removal of apex predators by fishing (Fig. 1, A and B). Phase
shifts between forested and deforested states (the latter known as "sea
urchin barrens") result from intense grazing due to increased abundance and
altered foraging patterns of sea urchins made possible in turn by human
removal of their predators and competitors (7, 8, 30-32).
------------------------------------------------------------------------
Fig. 1. Simplified coastal food webs showing changes in some of the
important top-down interactions due to overfishing; before (left side) and
after (right side) fishing. (A and B) Kelp forests for Alaska and southern
California (left box), and Gulf of Maine (right box). (C and D) Tropical
coral reefs and seagrass meadows. (E and F) Temperate estuaries. The
representation of food webs after fishing is necessarily more arbitrary than
those before fishing because of rapidly changing recent events. For example,
sea urchins are once again rare in the Gulf of Maine, as they were before
the overfishing of cod, due to the recent fishing of sea urchins that has
also permitted the recovery of kelp. Bold font represents abundant; normal
font represents rare; "crossed-out" represents extinct. Thick arrows
represent strong interactions; thin arrows represent weak interactions.
[View Larger Version of this Image (56K GIF file)]
------------------------------------------------------------------------


The kelp forest ecosystem of the Northern Pacific arose during the last
20 million years with the evolution of kelps, strongylocentrotid sea
urchins, sea otters, and the extinct Steller's sea cow (6). Sea cows were
widely distributed across the northern Pacific Rim through the Late
Pleistocene. They may have been eliminated from most of their range by
aboriginal hunting at the end of the Pleistocene and in the early Holocene,
because they survived thousands of years longer in the western Aleutian
Islands that were not peopled until about 4000 yr B.P. (6). By the time of
European contact in 1741, sea cows persisted only in the Commander Islands,
the only islands of the Aleutians unoccupied by aboriginal people. European
fur traders killed the last sea cow 27 years later in 1768. We have no idea
to what extent abundant sea cows grazed kelp forests, although their
apparent inability to dive deeply probably limited their grazing to the
surface canopy of kelps and to seaweeds lining the shore (6).

Northern Pacific kelp forests presumably flourished before human settlement
because predation by sea otters on sea urchins prevented the urchins from
overgrazing kelp (30). Aboriginal Aleuts greatly diminished sea otters
beginning around 2500 yr B.P., with a concomitant increase in the size of
sea urchins (31). Fur traders subsequently hunted otters to the brink of
extinction in the 1800s with the attendant collapse of kelp forests grazed
away by sea urchins released from sea otter predation. Legal protection of
sea otters in the 20th century partially reversed this scenario. However,
kelp forests are again being depleted in areas of Alaska because of
increased predation on sea otters by killer whales (33). The whales shifted
their diet to sea otters from seals and sea lions, which are in drastic
decline. 

A similar sequence of events occurred in kelp forests of the Gulf of Maine
(7, 34). Sea otters were never present, but Atlantic cod and other large
ground fish are voracious predators of sea urchins. These fishes kept sea
urchin populations small enough to allow persistence of kelp forests despite
intensive aboriginal and early European hook-and-line fishing for at least
5000 years. New mechanized fishing technology in the 1920s set off a rapid
decline in numbers and body size of coastal cod in the Gulf of Maine (7)
(Fig. 2A and Table 1) that has extended offshore to Georges Bank (35).
Formerly dominant predatory fish are now ecologically extinct and have been
partially replaced by smaller and commercially less important species.
Lobsters, crabs, and sea urchins rose in abundance accordingly (7). Kelp
forests disappeared with the rise in sea urchins due to removal of predatory
fish, and then reappeared when sea urchins were in turn reduced to low
abundance by fishing.
------------------------------------------------------------------------
Fig. 2. Retrospective data showing baselines before ecosystem collapse. (A)
Time series of mean body length of Atlantic cod from kelp forests in the
coastal Gulf of Maine. The earlier five data points are derived from
archaeological records, whereas the last three points are from fisheries
data (113). Vertical bars represent the standard error. Horizontal bars
represent the time range of data for a single interval of observations. (B)
Paleoecological and ecological data showing the percentage of Caribbean
localities with Acropora palmata () or A. cervicornis () as the dominant
shallow-water coral in the Late Pleistocene, Holocene, before 1983, and
after 1983 (114). Percentages of localities are significantly different over
the four time periods for A. palmata (2 = 34.0, P <0.0001, df = 3) and
A. cervicornis (2 = 22.4, P <0.0001, df = 3). Vertical and horizontal
bars
are as in (A). (C) Paleoecological and fisheries data from Chesapeake Bay
showing the ratio in abundance of planktonic to benthic diatoms (dotted
line) (77) and landings of the oyster Crassostrea virginica (solid line)
(80). The planktonic to benthic diatom ratio is a proxy for eutrophication
that shows the relative amount of planktonic to benthic primary production
(77). For over 1200 years this ratio remained fairly constant at about 1:1,
but then increased threefold coincidentally with increased runoff of
sediments and nutrients due to European agriculture after 1750. The ratio
remained at about 3:1 between 1830 and 1930, after which it increased
dramatically to about 8:1. Oyster landings show an initial increase in the
early 19th century, peak in 1884, and subsequent collapse as deep channel
reefs were destroyed by mechanical dredging (80). These data strongly imply
that oysters were able to limit the potential for eutrophication induced by
increased inputs of nutrients between 1750 and 1930 until oyster populations
collapsed as a result of overfishing. [View Larger Version of this Image
(21K GIF file)] 
------------------------------------------------------------------------


Table 1. Retrospective records from coastal ecosystems that offer baselines
that contrast with recent observations. Data source: P, paleoecologial; A,
archaeological; H, historical; F, fisheries; E, ecological. Inferred causes:
1, fishing; 2, mechanical habitat destruction by fishing; 3, inputs.
Abbreviations: BSi, biologically bound silica; DOP, degree of pyritization
of iron; dec., decrease; inc., increase. References after 115 are located on
Science Online (www.sciencemag.org/cgi/content/full/293/5530/629/DC1).
------------------------------------------------------------------------
Parameter of interest    Location    Data source    Proxy    Time of
baseline (yr B.P.)    Baseline observation or estimate    Recent observation
or estimate    Trend    Inferred cause    Ref.
------------------------------------------------------------------------
Kelp forests
Sea Otter    Pacific Ocean    H, E    Area estimates    260    >100,000
individuals    30,000 individuals    >3.3-fold dec.    1    116
Stellar's sea cow    Alaska    H    Herd size    259    <5,000 sea cows    0
Extinction    1    117
Atlantic cod    Gulf of Maine    A    Cod vertebrae    3550    Mean body
length of 1.0 m    Mean body length of 0.3 m    3-fold dec.    1    113
White abalone    California    E    Number per area    30    >2,000 per ha
1.0 ± 0.4 per ha    >2,000-fold dec.    1    118
Coral reefs
Coral    Caribbean Sea    P, E    % sites with A. palmata dominant
125,000     80% of Pleistocene sites    15% of post-1982 sites    5.3-fold
dec.     1    114
Coral    Caribbean Sea    P, E    % sites with A. cervicornis dominant
125,000    63% of Pleistocene sites    0% of post-1982 sites    100% loss
1    114
Coral    Bahamas    P, E    Standardized abundance of A. cervicornis
125,000    12    1    12-fold dec.    1    119
Coral    Belize    P    Relative abundance    3,130    A. cervicornis
dominant    A. cervicornis absent    100% loss    1    45
Coral    Netherlands Antilles    E    Coral cover at 10 m    27    54% coral
cover    31% coral cover    1.7-fold dec.    1    120
Coral    Jamaica    E    Coral cover at 10 m    23    73% coral cover    4%
coral cover    18-fold dec.    1    42
Monk seal    Caribbean Sea    H    Historical reports    >300    Abundant
0    Extinction    1    4, 68
Coral    Moreton Bay    P, E    Acropora dominance in fossil reefs    8,000
Dominated reefs throughout Bay    Only one small Acropora reef left
Decrease    3    121
Tropical and subtropical seagrass beds
Green turtle    Caribbean Sea    E    Biomass estimates    >300
>16.1 ¥ 106 50-kg turtles    >1.1 ¥ 106 50-kg turtles    15-fold dec.   
1    2,
122
Green turtle    Caribbean Sea    H    Hunting, biomass estimates    >300
>3.3 ¥ 107 adult turtles    >1.1 ¥ 106 50-kg turtles    30-fold dec.    1
1,
122
Seagrass beds    Tampa Bay    H    Area    121    30,970 ha    10,759 ha
3-fold dec.    1, 2, 3    123, 124
Dugong    Eastern Australia    H    Herd size    >100    >1.0 ¥ 106
estimated dugongs    14,000 estimated dugongs    >74-fold dec.    1    125,
126
Dugong    Moreton Bay    H    Herd size    107    >104,000 estimated dugongs
500 estimated dugongs    >208-fold dec.    1    125, 127
Oysters and eutrophication in estuaries
Inputs    Chesapeake Bay    P    Sedimentation rate    1,900    0.04 cm
year1    0.2 cm year1    5-fold inc.    3    77
Eutrophication    Chesapeake Bay    P    Total organic carbon    1,900
0.26 mg cm2 year1    2.3 mg cm2 year1    9-fold inc.    2, 3    77
Eutrophication    Chesapeake Bay    P    Centric/pennate diatom ratio
1,450    1:1 ratio    8:1 ratio    8-fold inc.    2, 3    77
Eutrophication    Chesapeake Bay    P    Dinoflagellate cysts (Spiniferites
spp.)    >300    50% relative abundance    80% relative abundance
1.6-fold inc.    2, 3    128
Seagrass beds    Fleets Bay, CB    H    Area    63    273 ha    16 ha
17-fold dec.    1, 2, 3    78
Oyster reefs    Chesapeake Bay    F    Oyster landings    116    6.2 ¥ 105
metric tons year1*    0.12 ¥ 105 metric tons year1    52-fold dec.    2
80
Oyster reefs    Tangier Sound, CB    F    Area    122    44.6 km2    0
100% loss    2    129
Anoxia    Cheapeake Bay    P    Degree of pyritization    1,900    0.32 DOP
0.51 DOP    2-fold inc.    2, 3    77
Seagrass beds    Botany Bay    H    Area of Posidonia beds    58    445 ha
188 ha    2.4-fold dec.    1, 2, 3    130
Eutrophication    Baltic Sea    P    Planktic diatom relative abundance
250    25% of total diatom abundance    80% of total diatom abundance
3-fold inc.    3    131
Eutrophication    Baltic Sea    P    Total organic carbon accumulation rate
138    3.2 gC m2 year1    70 g C m2 year1    22-fold inc.    3    132
Anoxia    Baltic Sea    P    Laminated sediments    100    5% of cores
laminated    90% of cores laminated    18-fold inc.    3    85
Offshore benthic communities
Oyster reefs    Foveaux Strait, NZ    F    Oyster landings    34
127 ¥ 106 oysters/year*    15 ¥ 106 oysters year1     8-fold dec.    2   
93
Oyster reefs     Foveaux Strait, NZ    F    Reef by-catch per station    38
1 in 4 stations had reef by-catch    1 in 7 stations had reef by-catch
2-fold dec.    2    93
Eutrophication    Gulf of Mexico    P    Biologically bound silica    295
0.29% BSi    1.00% BSi    3.4-fold inc.    3    106
Eutrophication    Gulf of Mexico    P    Total organic carbon accumulation
rate    100    2.4 mg C cm2 year1    7.8 mg C cm2 year1    3.3-fold inc.
3    133
Hypoxia    Gulf of Mexico    P    Benthic foraminifera    295    71
Ammonia-Elphidium Index    85 Ammonia-Elphidium Index    1.2-fold inc.    3
134
Eutrophication    Adriatic Sea    P    Eutrophic benthic foraminifera
Nonionella turgida    170    6% relative abundance    38% relative abundance
6-fold inc.    3    135
Eutrophication    Adriatic Sea    P    Coccolithophorids    286    100
cells/g of sediment    1.6 ¥ 106 cells/g of sediment    15,700-fold inc.
3    136
------------------------------------------------------------------------
* Baseline taken from peak in landings.

The more diverse food web of southern California kelp forests historically
included spiny lobsters and large sheephead labrid fish in addition to sea
otters as predators of sea urchins, as well as numerous species of abalone
that compete with sea urchins for kelps (Fig. 1, A and B) (36). Aboriginal
exploitation began about 10,000 yr B.P. and may have had local effects on
kelp communities (37). The fur trade effectively eliminated sea otters by
the early 1800s (38), but kelp forests did not begin to disappear on a large
scale until the intense exploitation and ecological extinction of sheephead,
spiny lobsters, and abalone starting in the 1950s (8, 36) (Table 1 and Fig.
1, A and B). Subsequent fishing of the largest sea urchin species in the
1970s and 1980s resulted in the return of well-developed kelp forests in
many areas that, as in the Gulf of Maine, effectively lack trophic levels
higher than that of primary producers (36, 39).

Coral Reefs
Coral reefs are the most structurally complex and taxonomically diverse
marine ecosystems, providing habitat for tens of thousands of associated
fishes and invertebrates (40). Aboriginal fishing in coral reef environments
began at least 35,000 to 40,000 years ago in the western Pacific (41) but
appears to have had limited ecological impact. Recently, coral reefs have
experienced dramatic phase shifts in dominant species due to intensified
human disturbance beginning centuries ago (1) (Fig. 1, C and D). The effects
are most pronounced in the Caribbean (42) but are also apparent on the Great
Barrier Reef in Australia despite extensive protection over the past three
decades (43). 

Large species of branching Acropora corals dominated shallow reefs in the
tropical western Atlantic for at least half a million years (44-46) until
the 1980s when they declined dramatically (42, 47) (Fig. 2B and Table 1).
Patterns of community membership and dominance of coral species were also
highly predictable (44), so that there is a clear baseline of pristine coral
community composition before human impact.

Western Atlantic reef corals suffered sudden, catastrophic mortality in the
1980s due to overgrowth by macroalgae that exploded in abundance after mass
mortality of the superabundant sea urchin Diadema antillarum that was the
last remaining grazer of macroalgae (42, 47). Early fisheries reports
suggest that large herbivorous fishes were already rare before the 20th
century (48). However, macroalgae were held in check until the last major
herbivore, Diadema, was lost from the system through disease (42, 47).

Corals on the Great Barrier Reef have experienced recurrent mass mortality
since 1960 due to spectacular outbreaks of the crown-of-thorns starfish
Acanthaster planci that feeds on coral (49). The causes of outbreaks are
controversial, but they are almost certainly new phenomena. There are no
early records of Acanthaster in undisturbed fossil deposits, in aboriginal
folklore, or in accounts of European explorers and fishers. Now, in recent
decades, the frequency and intensity of outbreaks have exceeded the
capability of longer lived species to recover as outbreaks have become more
chronic than episodic (50).

One possible explanation for Acanthaster outbreaks is that overfishing of
species that prevy upon larval or juvenile stages of crown-of-thorns
starfish is responsible for massive recruitment of the starfish (51). The
highly cryptic, predator-avoiding behavior of juvenile starfish, their
formidable antipredator defenses as subadults and adults, and the reduction
of some generalized predatory fishes on the Great Barrier Reef all point to
such a "top-down" explanation. Commercial and recreational fishing, as well
as indirect effects of intensive trawling for prawns, are likely
explanations for decreased abundance of predators of crown-of-thorns
starfish (52). Massive recruitment of starfish may also be due to
"bottom-up" increases in productivity due to increased runoff of nutrients
from the land (53). In either case, the explanation is almost certainly
historical and anthropogenic, and cannot be resolved by recent observations
alone. 

Expeditions occurred annually to northern Australia from the Malay
Archipelago throughout the 18th and 19th centuries to harvest an estimated
6 million sea cucumbers each season (54). After European colonization,
industrial-scale fishing developed along the Great Barrier Reef and
subtropical east Australian coast in the early to mid-19th century (55).
Whales, dugongs, turtles, pearl oysters, and Trochus shell were each heavily
exploited only to rapidly collapse, and all have failed to regain more than
a small fraction of their former abundance (55-57). Fishing of pelagic and
reef fishes, sharks, and prawns has continued to the present, although catch
per unit effort has declined greatly (58).

Tropical and Subtropical Seagrass Beds
Seagrass beds cover vast areas of tropical and subtropical bays, lagoons,
and continental shelves (59). Seagrasses provide forage and habitat for
formerly enormous numbers of large sirenians (dugong and manatee) and sea
turtles, as well as diverse assemblages of fishes, sharks, rays, and
invertebrates, including many commercially important species (59-62) (Fig.
1, C and D). Like coral reefs, seagrass beds seemed to be highly resilient
to human disturbance until recent decades when mass mortality of seagrasses
became common and widespread (63-65). Examples include the die-off of
turtlegrass in Florida Bay and the Gulf of Mexico in the 1980s (65) and the
near disappearance of subtidal seagrasses in the offshore half of Moreton
Bay near Brisbane, Australia, over the past 20 to 30 years (63, 64).
Proximate causes of these losses include recent increases in sedimentation,
turbidity, or disease (63-65). However, extirpation of large herbivorous
vertebrates beginning centuries ago had already profoundly altered the
ecology of seagrass beds in ways that increased their vulnerability to
recent events. 

Vast populations of very large green turtles were eliminated from the
Americas before the 19th century (1, 2) (Table 1). Formerly great
populations of green turtles in Moreton Bay, Australia, also were greatly
reduced by the early 20th century (66). Moreover, there are no estimates of
abundances of turtles in Australia at the dawn of European exploitation, so
that reported reductions must be only a small fraction of the total numbers
lost. All turtle species continue to decline at unsustainable rates along
the Great Barrier Reef today (67).

Abundant green turtles closely crop turtlegrass and greatly reduce the flux
of organic matter and nutrients to sediments (59-62, 68). In the near
absence of green turtles today, turtlegrass beds grow longer blades that
baffle currents, shade the bottom, start to decompose in situ, and provide
suitable substrate for colonization by the slime molds that cause
turtlegrass wasting disease (65). Deposition within the beds of vastly more
plant detritus also fuels microbial populations, increases the oxygen demand
of sediments, and promotes hypoxia (65). Thus, all the factors that have
been linked with recent die-off of turtlegrass beds in Florida Bay (65),
except for changes in temperature and salinity, can be attributed to the
ecological extinction of green turtles (27).

European colonists did not exploit tropical American manatees as
systematically as they exploited green turtles, so the data related to
fisheries are poor. We know, however, that manatees were extensively fished
by aboriginal people and by early colonists (68). In Australia, aboriginal
people also harvested dugongs extensively long before European colonization
(3), yet the numbers reported by early colonists were vast. Three- or
four-mile-long herds comprising tens of thousands of large individuals were
observed in Wide Bay in about 1870 (69) and in Moreton Bay as recently as
1893 (70). Widespread colonial exploitation of dugongs for their flesh and
oil along the southern Queensland coast resulted in the crash of the dugong
fishery by the beginning of the 20th century (3) (Table 1). Ironically,
scientists recently reported the "discovery of a large population" of
dugongs in Moreton Bay--a mere 300 individuals (71). Further north, numbers
of dugongs in the vast southern half of the Great Barrier Reef had dwindled
to fewer than 4000 when they were first accurately counted in 1986-87, with
a further 50 to 80% decline in recent years (72). These increasingly
fragmented populations represent the last remnants of the vast herds of the
early 19th century and before.

The ecological implications of these reductions are at least as impressive
as those for green turtles. Moderate sized herds of dugongs remove up to 96%
of above-ground biomass and 71% of below-ground biomass of seagrasses (73).
Their grazing rips up large areas of seagrass beds, providing space for
colonization by competitively inferior species of seagrasses. Dugong grazing
also produces massive amounts of floating debris and dung that are exported
to adjacent ecosystems. The decline in seagrasses in Moreton Bay is
certainly due in large part to the dramatic decline in water quality due to
eutrophication and runoff of sediment (63, 64). Nevertheless, as noted for
green turtles and turtlegrass in Florida Bay, the cessation of systematic
plowing of the bay floor by once abundant dugongs must also have been a
major factor. 

Oysters and Eutrophication in Estuaries
Temperate estuaries worldwide are undergoing profound changes in
oceanography and ecology due to human exploitation and pollution, rendering
them the most degraded of marine ecosystems (74-76) (Fig. 1, E and F). The
litany of changes includes increased sedimentation and turbidity (77);
enhanced episodes of hypoxia or anoxia (74, 75, 77); loss of seagrasses (78)
and dominant suspension feeders (79), with a general loss of oyster reef
habitat (80); shifts from ecosystems once dominated by benthic primary
production to those dominated by planktonic primary production (77);
eutrophication (74-76) and enhanced microbial production (81); and higher
frequency and duration of nuisance algal and toxic dinoflagellate blooms
(82, 83), outbreaks of jellyfish (79), and fish kills (83). Most
explanations for these phenomena emphasize "bottom-up" increases in
nutrients like nitrogen and phosphorus as causes of phytoplankton blooms and
eutrophication (74-76), an interpretation consistent with the role of
estuaries as the focal point and sewer for many land-based, human
activities. Nevertheless, long-term records demonstrate that reduced
"top-down" control resulting from losses in benthic suspension feeders
predated eutrophication.

The oldest and longest records come from cores in sediments from Chesapeake
Bay (77) and Pamlico Sound (84) in the eastern United States and from the
Baltic Sea (85) that extend back as far as 2500 yr B.P. (Fig. 2C and Table
1). A general sequence of ecological change is apparent in all three cases,
but the timing of specific ecological transitions differs among estuaries in
keeping with their unique histories of land use, exploitation, and human
population growth--a difference that rules out a simple climatic
explanation. Increased sedimentation and burial of organic carbon began in
the mid-18th century in Chesapeake Bay, coincident with widespread land
clearance for agriculture by European colonists (77). The main ecological
response was a gradual shift in the taxa responsible for primary production
that began in the late 18th century. Seagrasses and benthic diatoms on the
bay floor declined, while planktonic diatoms and other phytoplankton in the
water column correspondingly increased. However, anoxia and hypoxia were not
widespread until the 1930s when phytoplankton populations and the flux of
organic matter to the bay floor increased dramatically with concomitant loss
of benthic fauna (75, 77) (Fig. 2C and Table 1). Similar changes began in
the 1950s in the Baltic Sea, with widespread expansion of the extent of
anoxic laminated sediments (74, 85), and in the 1950s to 1970s in Pamlico
Sound (84). 

Vast oyster reefs were once prominent structures in Chesapeake Bay (11),
where they may have filtered the equivalent of the entire water column every
3 days (79). Despite intensive harvesting by aboriginal and early colonial
populations spanning several millennia, it was not until the introduction of
mechanical harvesting with dredges in the 1870s that deep channel reefs were
seriously affected (79, 80). Oyster catch was rapidly reduced to a few
percent of peak values by the early 20th century (79, 80) (Fig. 2C and Table
1). Only then, after the oyster fishery had collapsed, did hypoxia, anoxia,
and other symptoms of eutrophication begin to occur in the 1930s (75, 77),
and outbreaks of oyster parasites became prevalent only in the 1950s (80).
Thus, fishing explains the bulk of the decline, whereas decline in water
quality and disease were secondary factors (80). However, now that oyster
reefs are destroyed, the effects of eutrophication, disease, hypoxia, and
continued dredging interact to prevent the recovery of oysters and
associated communities (86). Field experiments in Pamlico Sound demonstrate
that oysters grow well, survive to maturity, and resist oyster disease when
elevated above the zone of summer hypoxia--even in the presence of modern
levels of eutrophication and pollution (87).

Overfishing of oysters to the point of ecological extinction is just one
example in a general pattern of removal of species capable of top-down
control of community structure in estuaries. Dense populations of oysters
and other suspension-feeding bivalves graze plankton so efficiently that
they limit blooms of phytoplankton and prevent symptoms of eutrophication
(88, 89), just as occurs with grazing by zooplankton in freshwater
ecosystems (90). The ecological consequences of uncounted other losses are
unknown. Gray whales (now extinct in the Atlantic), dolphins, manatees,
river otters, sea turtles, alligators, giant sturgeon, sheepshead, sharks,
and rays were all once abundant inhabitants of Chesapeake Bay but are now
virtually eliminated.

Offshore Benthic Communities
Continental shelves cover more of the ocean floor than all previously
discussed environments combined. Commercially important cod, halibut,
haddock, turbot, flounder, plaice, rays, and a host of other ground fishes,
scallops, cockles, and oysters have been fished intensively for centuries
from continental shelves of Europe and North America, and more recently
throughout the world (5, 7, 10, 91). Hook-and-line fishing was replaced by
intensive use of the beam trawl during the 18th century, and industrialized
fishing was further intensified with the advent of large steam- and
diesel-powered vessels and the otter trawl at the end of the 19th century.
Reports of severely depleted fish stocks and shifting of fishing grounds
farther and farther from home ports into the North Sea and the outer Grand
Banks were commonplace by the beginning of the 19th century. Scientific
investigation consistently lagged behind economic realities of depleted
stocks and inexorable exploitation of more-distant fishing grounds. As late
as 1883, Thomas Huxley claimed that fish stocks were inexhaustible (92), a
view discredited by the beginning of the 20th century (5). Today, several
formerly abundant, large fish as well as formerly dense assemblages of
suspension feeders are ecologically extinct over vast areas (7-10, 93).

The Primacy of Overfishing in Human Disturbance to Marine Ecosystems
Overfishing of large vertebrates and shellfish was the first major human
disturbance to all coastal ecosystems examined (Table 1). Ecological changes
due to overfishing are strikingly similar across ecosystems despite the
obvious differences in detail (Fig. 1, A to F). Everywhere, the magnitude of
losses was enormous in terms of biomass and abundance of large animals that
are now effectively absent from most coastal ecosystems worldwide. These
changes predated ecological investigations and cannot be understood except
by historical analysis. Their timing in the Americas and Pacific closely
tracks European colonization and exploitation in most cases. However,
aboriginal overfishing also had effects, as exemplified by the decline of
sea otters (and possibly sea cows) in the northeast Pacific thousands of
years ago. 

There are three important corollaries to the primacy of overfishing. The
first is that pollution, eutrophication, physical destruction of habitats,
outbreaks of disease, invasions of introduced species, and human-induced
climate change all come much later than overfishing in the standard sequence
of historical events (Fig. 3). The pattern holds regardless of the initial
timing of colonial overfishing that began in the Americas in the 16th and
17th centuries and in Australia and New Zealand in the 19th century. The
full sequence of events is most characteristic of temperate estuaries like
Chesapeake Bay. Not all the human disturbances illustrated in Fig. 3 have
affected all ecosystems yet. But wherever these events have occurred, the
standard chronological sequence of human disturbance and modification of
ecosystems is recognizable.
------------------------------------------------------------------------
Fig. 3. Historical sequence of human disturbances affecting coastal
ecosystems. Fishing (step 1) always preceded other human disturbance in all
cases examined. This is the basis for our hypothesis of the primacy of
overfishing in the deterioration of coastal ecosystems worldwide. Subsequent
steps 2 through 5 have not been observed in every example and may vary in
order. [View Larger Version of this Image (18K GIF file)]
------------------------------------------------------------------------



The second important corollary is that overfishing may often be a necessary
precondition for eutrophication, outbreaks of disease, or species
introductions to occur (27). For example, eutrophication and hypoxia did not
occur in Chesapeake Bay until the 1930s, nearly two centuries after clearing
of land for agriculture greatly increased runoff of sediments and nutrients
into the estuary (77). Suspension feeding by still enormous populations of
oysters was sufficient to remove most of the increased production of
phytoplankton and enhanced turbidity until mechanical harvesting
progressively decimated oyster beds from the 1870s to the 1920s (77, 80)
(Fig. 2C). 

The consequences of overfishing for outbreaks of disease in the next lower
trophic level fall into two categories. The most straightforward is that
populations in the lower level become so dense that they are much more
susceptible to disease as a result of greatly increased rates of
transmission (94). This was presumably the case for the sea urchin Diadema
on Caribbean reefs and the seagrass Thalassia in Florida Bay. In contrast,
among oysters disease did not become important in Chesapeake Bay until
oysters had been reduced to a few percent of their original abundance (80),
a pattern repeated in Pamlico Sound (86, 87) and Foveaux Strait, New
Zealand(93). Two factors may be responsible. First, oysters may have become
less fit owing to stresses like hypoxia or sedimentation, making them less
resistant to disease (87). Alternatively, suspension feeding by dense
populations of oysters and associated species on oyster reefs may have
indirectly limited populations of pathogens by favoring other plankton--an
explanation that may extend to blooms of toxic plankton and most other
outbreaks of microbial populations (88).

The third important corollary is that changes in climate are unlikely to be
the primary reason for microbial outbreaks and disease. The rise of microbes
has occurred at different times and under different climatic conditions in
different places, as exemplified by the time lag between events in
Chesapeake Bay and Pamlico Sound (77, 79, 80, 84). Anthropogenic climate
change may now be an important confounding factor, but it was not the
original cause. Rapid expansion of introduced species in recent decades (95)
may have a similar explanation, in addition to increase in frequency and
modes of transport. Massive removal of suspension feeders, grazers, and
predators must inevitably leave marine ecosystems more vulnerable to
invasion (96, 97). 

Synergistic Effects of Human Disturbance
Ecological extinction of entire trophic levels makes ecosystems more
vulnerable to other natural and human disturbances such as nutrient loading
and eutrophication, hypoxia, disease, storms, and climate change. Expansion
and intensification of different forms of human disturbance and their
ecological effects on coastal ecosystems have increased and accelerated with
human population growth, unchecked exploitation of biological resources,
technological advance, and the increased geographic scale of exploitation
through globalization of markets. Moreover, the effects are synergistic, so
that the whole response is much greater than the sum of individual
disturbances (98). This is perhaps most apparent in the rise of
eutrophication, hypoxia, and the outbreak of toxic blooms and disease
following the destruction of oyster reefs by mechanical harvesting of
oysters (79, 80, 86). Other possible examples are outbreaks of seagrass
wasting disease due to the removal of grazers of seagrasses like the green
turtle (27). 

A striking feature of such synergistic effects is the suddenness of the
transition in abundance of different kinds of organisms and community
composition due to threshold effects (99). Ecological diversity and
redundancy within trophic levels is probably the most important reason for
the delay or time lag between the onset of fishing and the subsequent
threshold response (42, 100). The importance of biodiversity in the form of
ecological redundancy is clearly apparent for the delay in the collapse of
kelp forests in southern California compared with Alaska after the
extirpation of sea otters. Sheephead fish, spiny lobsters, and abalone in
the more diverse Californian kelp forests kept sea urchin populations in
check until these predators and competitors of sea urchins had also been
effectively eliminated (8, 36). Similarly, the sea urchin Diadema kept
macroalgae in check long after the extreme overfishing of herbivorous fishes
on Caribbean coral reefs (42).

A second potentially important mechanism for the suddenness of ecosystem
collapse is the elimination of previously unfished refuges that were
protected historically because of distance or expense of access. For
example, reef fishes all around Jamaica in the 1960s rarely reached
reproductive maturity so that the abundant recruits of fishes on Jamaican
reefs at that time must have come from undiscovered populations in Jamaica
or elsewhere (101). But as more and more reefs have been overfished, the
potential sources of such recruits must have effectively disappeared over
wider areas (102). A similar scenario has been proposed for the American
lobster with regard to loss of larvae from deep-water offshore stocks (103)..

Microbialization of the Global Coastal Ocean
Most recent changes to coastal marine ecosystems subsequent to overfishing
involve population explosions of microbes responsible for increasing
eutrophication (74-76, 81), diseases of marine species (104), toxic blooms
(82, 83), and even diseases such as cholera that affect human health (104,
105). Chesapeake Bay (81) and the Baltic Sea (74) are now bacterially
dominated ecosystems with a trophic structure totally different from that of
a century ago. Microbial domination also has expanded to the open ocean off
the mouth of the Mississippi River (106) and to the Adriatic Sea (107).

Nowhere is the lack of historical perspective more damaging to scientific
understanding than for microbial outbreaks. Plans for remediation of
eutrophication of estuaries are still based on the belief that
eutrophication is caused only by increased nutrients without regard to
overfishing of suspension feeders. Even more remarkable is the attribution
of the rise in marine diseases to climate change and pollution (104) without
regard to the pervasive removal of higher trophic levels and the
asynchronous outbreaks of disease in different ecosystems that belie a
simple climatic explanation.

Historical Perspectives for Ecosystem Restoration
The characteristic sequence of human disturbance to marine ecosystems (Fig.
3) provides a framework for remediation and restoration that is invisible
without a historical perspective. More specific paleoecological,
archaeological, and historical data should be obtained to refine the
histories of specific ecosystems and as a tool for management, but the
overall patterns are clear. The historical magnitudes of losses of large
animals and oysters were so great as to seem unbelievable based on modern
observations alone (Table 1). Even seemingly gloomy estimates of the global
percentage of fish stocks that are overfished (108) are almost certainly far
too low. The shifting baseline syndrome is thus even more insidious and
ecologically widespread than is commonly realized.

On the other hand, recognition of these losses shows what coastal ecosystems
could be like, and the extraordinary magnitude of economic resources that
are retrievable if we are willing to act on the basis of historical
knowledge. The central point for successful restoration is that loss of
economically important fisheries, degradation of habitat attractive to
landowners and tourists, and emergence of noxious, toxic, and
life-threatening microbial diseases are all part of the same standard
sequence of ecosystem deterioration that has deep historical roots (27).
Responding only to current events on a case-by-case basis cannot solve these
problems. Instead, they need to be addressed by a series of bold experiments
to test the success of integrated management for multiple goals on the scale
of entire ecosystems. With few exceptions, such as the Caribbean monk seal
and Steller's sea cow, most species that are ecologically extinct probably
survive in sufficient numbers for successful restoration. This optimism is
in stark contrast with the state of many terrestrial ecosystems where many
or most large animals are already extinct (28). Moreover, we now have the
theoretical tools (109) to roughly estimate per capita interaction strengths
of surviving individuals of now rare animals like sea turtles, sirenians,
sharks, and large groupers. We can then use these data to build tentative
models of the consequences of the renewed abundance of these species in
their native environments that can in turn be used to design large-scale,
adaptive experiments for ecosystem restoration, exploitation, and management
(96, 108, 110). 

One obviously timely and overdue experiment is to attempt the amelioration
of eutrophication, hypoxia, and toxic blooms in Chesapeake Bay by massive
restoration of oyster reefs (79). Experiments in Pamlico Sound show that
this is possible (86, 87, 96), and modeling of food webs suggests that even
partial restoration of oysters would reduce eutrophication substantially
(110). Aquaculture of suspension-feeding bivalves like oysters might be
promoted to reverse the effects of eutrophication and to restore water
quality in degraded estuaries. Other important examples include the
restoration of coral reefs and seagrass beds by protection of fishes,
sharks, turtles, and sirenians in very large reserves on the scale of all of
Florida Bay and the Florida Keys--an approach recently advocated for
terrestrial ecosystems (111). Once again, small-scale grazing experiments
with reef fishes (112) show that fishes could reverse the overgrowth of
corals by macroalgae on a massive scale. The potential for reducing diseases
of corals and turtlegrass by restoring natural levels of grazing is unproven
but consistent with historical evidence (27).

In summary, historical documentation of the long-term effects of fishing
provides a heretofore-missing perspective for successful management and
restoration of coastal marine ecosystems. Previous attempts have failed
because they have focused only on the most recent symptoms of the problem
rather than on their deep historical causes. Contrary to romantic notions of
the oceans as the "last frontier" and of the supposedly superior ecological
wisdom of non-Western and precolonial societies, our analysis demonstrates
that overfishing fundamentally altered coastal marine ecosystems during each
of the cultural periods we examined. Changes in ecosystem structure and
function occurred as early as the late aboriginal and early colonial stages,
although these pale in comparison with subsequent events. Human impacts are
also accelerating in their magnitude, rates of change, and in the diversity
of processes responsible for changes over time. Early changes increased the
sensitivity of coastal marine ecosystems to subsequent disturbance and thus
preconditioned the collapse we are witnessing.

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the dominant or codominant corals were derived from 50 studies from Antigua,
Bahamas, Barbados, Belize, Bonaire, Cayman Islands, Colombia, Dominican
Republic, Florida, Haiti, Jamaica, Mexico, Netherlands Antilles, Panama,
Puerto Rico, and U.S. Virgin Islands. Studies contained either
paleoecological data from outcrops of fossil reefs or from sediment cores,
or ecological data. For A. palmata, only localities described as reef crest
or between 0- and 10-m water depth were included (131 localities). For
A. cervicornis, only localities described as forereef, reef slope, or
between 10- and 20-m water depth were included (72 localities). Leeward and
windward environments were not distinguished. The percentage of localities
that contained A. palmata or A. cervicornis as the most abundant coral was
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in the Americas), Holocene (when only aboriginal populations were present),
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115.    We dedicate this paper to the memory of Mia Tegner who died while
diving after this paper was submitted. This work was conducted as part of
the Long-Term Ecological Records of Marine Environments, Populations and
Communities Working Group supported by the National Center for Ecological
Analysis and Synthesis (funded by NSF grant DEB-0072909, the University of
California, and the University of California, Santa Barbara). Additional
support was also provided for the Postdoctoral Associate MXK in the Group.
L.W.B. was also supported by NSF grant OCE-9711448. We thank A. Bolten,
S. Cooper, N. Knowlton, B. Mitterdorfer, E. Sala, and two anonymous
reviewers for discussions and comments on the manuscript.
10.1126/science.1059199
Include this information when citing this paper.

Collections under which this article appears:
Ecology 
Volume 293, Number 5530, Issue of 27 Jul 2001, pp. 629-637.
Copyright © 2001 by The American Association for the Advancement of Science.. 

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