MARINE ECOLOGY PROGRESS SERIES

Mar Ecol Prog Ser Vol. 303: 1–29, 2005 Published November 21

*Email: kemp@hpl.umces.edu © Inter-Research 2005 · www.int-res.com

FEATURE ARTICLE: REVIEW

Eutrophication of Chesapeake Bay: historical trends

and ecological interactions

W. M. Kemp1,*, W. R. Boynton2, J. E. Adolf1, D. F. Boesch3, W. C. Boicourt1, G. Brush4, 

J. C. Cornwell1

, T. R. Fisher1, P. M. Glibert1, J. D. Hagy5, L. W. Harding1, E. D. Houde2, 

D. G. Kimmel1

, W. D. Miller1, R. I. E. Newell1, M. R. Roman1, E. M. Smith6, J. C. Stevenson1

1University of Maryland, Center for Environmental Science, Horn Point Laboratory, Cambridge, Maryland 21613, USA 2University of Maryland, Center for Environmental Science, Chesapeake Biological Laboratory, Solomons, Maryland 20688, USA 3University of Maryland, Center for Environmental Science, Cambridge, Maryland 21613, USA 4Johns Hopkins University, Department of Geography & Environmental Engineering, Baltimore, Maryland 21218, USA 5U.S. Environmental Protection Agency, NHEERL, Gulf Ecology Division, Gulf Breeze, Florida 32561, USA 6University of South Carolina, Department of Biology, Columbia, South Carolina 29208, USA

ABSTRACT: This review provides an integrated synthesis with

timelines and evaluations of ecological responses to eutrophication in Chesapeake Bay, the largest estuary in the USA.

Analyses of dated sediment cores reveal initial evidence of

organic enrichment in ~200 yr old strata, while signs of

increased phytoplankton and decreased water clarity first

appeared ~100 yr ago. Severe, recurring deep-water hypoxia

and loss of diverse submersed vascular plants were first evident

in the 1950s and 1960s, respectively. The degradation of these

benthic habitats has contributed to declines in benthic macroinfauna in deep mesohaline regions of the Bay and blue crabs

in shallow polyhaline areas. In contrast, copepods, which are

heavily consumed in pelagic food chains, are relatively unaffected by nutrient-induced changes in phytoplankton. Intense

mortality associated with fisheries and disease have caused a

dramatic decline in eastern oyster stocks and associated Bay

water filtration, which may have exacerbated eutrophication

effects on phytoplankton and water clarity. Extensive tidal

marshes, which have served as effective nutrient buffers along

the Bay margins, are now being lost with rising sea level.

Although the Bay’s overall fisheries production has probably

not been affected by eutrophication, decreases in the relative

contribution of demersal fish and in the efficiency with which

primary production is transferred to harvest suggest fundamental shifts in trophic and habitat structures. Bay ecosystem

responses to changes in nutrient loading are complicated by

non-linear feedback mechanisms, including particle trapping

and binding by benthic plants that increase water clarity, and

by oxygen effects on benthic nutrient recycling efficiency.

Observations in Bay tributaries undergoing recent reductions

in nutrient input indicate relatively rapid recovery of some

ecosystem functions but lags in the response of others.

KEY WORDS: Eutrophication · Nutrients · Chesapeake Bay

Resale or republication not permitted without 

written consent of the publisher

Chesapeake Bay is a large estuary which has undergone many changes in its ecological properties and

processes in response to nutrient enrichment over the

last 2 centuries. Susceptibility of the Bay to eutrophication arises in part from the long dendritic shoreline

that intimately connects it to its large watershed (covering an area 15 times that of the Bay) which contains

expanding human population centers and extensive

agricultural activities. 

(Satellite image from MODIS,

http://visibleearth.nasa.gov)
Mar Ecol Prog Ser 303: 1–29, 2005

INTRODUCTION

Eutrophication studies have played a central role in

both basic and applied limnology throughout the last

century (Hutchinson 1969, Likens 1972). Until recent

decades, however, few such studies had included

coastal marine systems (e.g. Nixon 1995), which were

often considered invulnerable to effects of nutrient

enrichment because of their well-mixed, well-flushed

nature (Schindler 1981). By the early to mid-1970s,

nutrient-induced increases in algal production and

biomass were being reported for a number of coastal

marine ecosystems (e.g. Ryther & Dunstan 1971, Sönderström 1971, Lehmusluoto 1973, Melvasalo et al.

1975). With an ever-increasing proportion of human

population being centered in coastal watersheds

(Valiela et al. 1992, de Jonge et al. 2002), the problem

of eutrophication has become a central theme of

coastal research and management at regional and

global scales (Jansson 1978, Rosenberg 1985, Gray

1992, Howarth et al. 2000). 

The more recent explosion of coastal eutrophication

research (e.g. Smetacek et al. 1991, Nixon 1995) has

generated a relatively complex conceptual model

of estuarine ecosystem responses to nutrient inputs

(Cloern 2001). Although in most aquatic ecosystems

nutrient enrichment elicits increases in phytoplankton

biomass and decreases in water clarity (e.g. Nielsen et

al. 2002a), estuarine phytoplankton have exhibited a

comparatively greater diversity of bloom-forming species in response to eutrophication (Smith 2003). Nutrient enrichment has contributed to widespread changes

in coastal habitats, including loss of seagrasses (e.g.

Walker & McComb 1992, Short & Wyllie-Echeverria

1996), proliferation of harmful phytoplankton (Smayda

1990) and benthic macroalgae (e.g. Menesguen &

Piriou 1995, Valiela et al. 1997), and depletion of dissolved oxygen in bottom waters (Andersen & Rydberg

1988, Diaz 2001). Indirect evidence suggests that these

habitat effects have altered trophic structures, production, and composition of fish and invertebrate communities through a range of ecological mechanisms

(Caddy 1993). 

Diverse ecological processes tend to ‘buffer’ and

regulate algal growth in coastal environments. These

processes include nutrient assimilation and turbidity

reduction by seagrass beds, control of algal biomass by

herbivores, and suppressed recycling of N and P with

elevated oxygen concentrations. Many of these natural

buffering mechanisms can, however, be compromised

by extreme nutrient enrichment (e.g. Zhang et al.

2003). Once altered, their non-linear nature makes it

difficult for these mechanisms to be reestablished (e.g.

Scheffer & Carpenter 2003, Smith 2003). In addition,

coastal ecosystems may become more vulnerable to

2

Fig. 1. Map of Chesapeake Bay showing water depth and

major tributary systems. Inset map indicates location of 

estuary and watershed
Kemp et al.: Chesapeake Bay eutrophication

effects of nutrient enrichment via increases in temperature and river flow associated with climate change

(e.g. Justi ́c et al. 1996) and by fishery harvest of key

herbivores that otherwise control algal growth (Jackson et al. 2001). While a fully integrated view of coastal

eutrophication is just now emerging, it appears that

the associated dynamics are highly complex.

Recent reviews have emphasized the global nature

of coastal eutrophication, describing broad patterns of

ecological responses (e.g. Nixon 1995, Richardson &

Jørgensen 1996). However, Cloern (2001) suggests

that estuaries can exhibit substantial differences in

magnitude and even trajectory of responses to nutrients,

reflecting complex non-linear and estuary-specific

ecological interactions. Before we can explain differences among estuaries in their responses to nutrient

enrichment, we must review and clarify our understanding of specific systems. Toward that goal, we provide here a synthesis of what we have learned regarding ecological responses to eutrophication in the large

well-studied coastal system of Chesapeake Bay, USA.

ESTUARINE CIRCULATION AND GEOMETRY

Chesapeake Bay is a large estuary located in the USA’s

mid-Atlantic coastal region. The Bay is almost 300 km

long, with a relatively deep (20 to 30 m) and narrow (1 to

4 km) central channel confined by a sill at its seaward

end (Fig. 1). Broad shallow areas flank the central Bay

channel over its entire length (Boicourt et al. 1999), and

depths exceeding 10 m constitute only 24% of the Bay’s

surface area (11 500 km2

), while the mean depth is only

6.5 m (Fig. 2). The slope of depth vs. area curves (Fig. 2)

is steepest for the mid (mesohaline) Bay region compared to the upper (tidal-fresh and oligohaline) and

lower (polyhaline) portions of the estuary. 

An average of 2300 m3 s–1 of freshwater flows from the

Bay’s watershed into its 74.4 km3 water volume, with the

Susquehanna River at the head of the Bay providing

more than half of the flow (Schubel & Pritchard 1986).

This freshwater input sets up stratification, which isolates deep channel waters by suppressing vertical

exchange (Seliger & Boggs 1988, Boicourt 1992). This

‘buoyancy reservoir’ acts over relatively long time scales

(Fig. 3), where winter–spring Susquehanna River flow

controls stratification throughout the summer (Hagy

2002). Strong episodic wind mixing events contribute to

periodic destratification (Goodrich et al. 1987), especially

in the middle reaches of the Bay. Stratification is, however, quickly reestablished from a buildup of the Bay’s

longitudinal salinity gradient (Boicourt 1992). 

River flow drives the estuarine circulation, characterized by a lower-layer counterflow that acts to retain particulate and dissolved materials in the Bay

(Pritchard 1956, 1967). This circulation creates relatively long residence times (90 to 180 d) for freshwater

and nutrients. The combination of the Bay’s long water

residence-time, its stratified water column, and its

narrow central channel isolated by sills and flanked

by wide shallows make this a productive system, with

efficient nutrient use and tendency for depletion of

oxygen from deep waters (Boicourt 1992).

3

a) Upper Bay

0

10

20

30

c) Lower Bay

0

10

20

30

40

0 20 40 60 80 100

% Area or Volume

b) Mid Bay

0

10

20

30

Depth (m)

Area

Volume

Fig. 2. Hypsographic presentation of distributions of % total

estuarine surface area and % total water volume below a

depth plotted vs. that water column depth for 3 regions of

Chesapeake Bay separated by 39.0° N and 37.0° N latitude.

Line defining top of dark shaded area represents estimated

mean depth of 1% surface irradiance (Z1%) in the 1930s, and

line at top of light shaded area represents calculated mean 

Z1% for 1990s (Table 1)
Mar Ecol Prog Ser 303: 1–29, 2005

HISTORICAL PATTERNS OF INPUTS AND

WATER QUALITY

Trends in watershed activities and nutrient loading

Chesapeake Bay’s watershed covers a substantial

area (164 200 km2

) containing diverse ecological and

physiographic features and patterns of human settlement. The ratios of the Bay’s watershed area to its estuarine water area and volume (14.3 and 2.2 m–1, respectively) are large compared to other estuaries (e.g.

Bricker et al. 1999). These features, coupled with the

long (18 800 km) dendritic Bay shoreline, render the

estuary closely connected with its watershed. When

Europeans migrated to this region 4 centuries ago,

they encountered a landscape almost completely covered with temperate forests bordered by wetlands

(Brush et al. 1980). 

During the 400 yr since initial European settlement,

the character of the Bay watershed has undergone

large changes (Curtin et al. 2001). The number of

humans in the watershed has grown exponentially

since colonial times (Fig. 4a), with a 3-fold increase

evident during the last 100 yr. Currently, the drainage

basin contains about 16 million people, but average

population density remains modest (1 person ha–1)

compared to many other coastal areas worldwide (e.g.

Smith et al. 2003a). The initial pattern of land use

change was dominated by progressive land clearing

for agriculture until the mid-1800s when about half the

basin was deforested (Fig. 4b). In the initial phases of

land clearing, farmers needed large tracts of land to

allow long fallow periods; crop rotation and use of

natural fertilizers in the 19th century replaced fallow

farming and allowed cultivation of more erodible and

less fertile lands. Throughout the last 2 centuries urban

land has expanded, while agricultural land has declined, and forested land has gradually increased.

Growth of human population in coastal watersheds is a

major factor contributing to increased nutrient loading

to estuaries throughout the world (Meybeck 1998,

Smith et al. 2003a).

Many rivers link the Bay directly to its watershed.

During the period of record (1890 to 2004), flow of the

largest of these rivers, the Susquehanna (Fig. 4c), has

been highly variable at a range of time scales that

include short drought periods (e.g. early 1953 to 1955),

4

800 1000 1200 1400 1600 1800 2000 2200 2400 2600

7

8

9

10

11

12

13

Stratification, N2 (103 rad2 s-2)

Jan-May Average Susquehanna River Flow (m3 s-1)

Fig. 3. Relationship between January–May average Susquehanna River flow (1986–1999) and April–September average

water column stratification in middle Bay, expressed as average of maximum squared Brunt-Väisälä frequency (N). Indicated line is least-squares regression line (r2 = 0.90). Adapted 

from Hagy (2002)

e) Input Nitrogen Concentration

Year

1950 1960 1970 1980 1990 2000

0

1

2

mg TN l-1

d) Nitrogen Fertilizer Use

0

20

40

kg N yr-1 x 106

c) Susquehanna Flow

400

800

1200

1600

Drought

Tropical Storm Agnes

Wet

Variable wet

Mean Flow m3 s-1 Population x 106 % of Total Land

0

20

40

60

80

100

4

8

12

16

1650 1700 1750 1800 1850 1900 1950 2000

a) Human Population

b) Land-Use

Forest

Agriculture

Developed

0.5

1.5

10

30

Fig. 4. Historical changes in Chesapeake Bay’s watershed in

terms of (a) human population and (b) major land uses from

1600 to 2000; also shown are (c) Susquehanna River flow,

(d) nitrogen fertilizer use in Maryland, (e) concentrations of

total nitrogen at Susquehanna River fall line, 1945 to 2003
Kemp et al.: Chesapeake Bay eutrophication

prolonged droughts (e.g. 1960s), huge flood events

(e.g. a 200 yr storm, tropical storm Agnes, in 1972),

prolonged wet periods (e.g. 1970s), and decades of

extreme variability (e.g. 1990s). While there is no longterm trend evident in flows from the Susquehanna or

other Bay tributaries, there is some suggestion that

interannual variation has increased during the last

several decades (Fig. 4c). In general, year-to-year

variations in river flow cause fluctuations in inputs and

estuarine distributions of freshwater, suspended sediments, and nutrients. These in turn affect stratification

and circulation as well as productivity and organism

abundance (e.g. Hearn & Robson 2001, Hagy et al.

2004). 

Despite the declining proportion of the Bay’s watershed devoted to agriculture since around 1850, use of

commercial fertilizers (Fig. 4d) and import of animal

feed grew dramatically after the 1950s (Fig. 4d). As a

result of growth in fertilizer use and other human

activities in the watershed, nutrient loading to the Bay

has increased. For example, Susquehanna River concentrations of total N (TN) entering the Bay increased

by 2.5-fold from 1945 to 1990, with a small decline during the last decade (Fig. 4e) evidently due to improved

watershed land management (Sprague et al. 2000).

Similar increases in nutrient loading associated with

expanded agricultural activities have been reported for

many coastal regions during the last several decades

(e.g. Eyre & Pont 2003, Kauppila et al. 2003).

Sediment record of eutrophication

Although direct human observations are not available to describe temporal trends in the Bay ecosystem

at multicentury scales, a rich indirect ecological record

exists within the Bay’s sediments. These trends have

been quantified and interpreted in recent studies using

geochemical and paleontological methods applied to

sediment cores collected throughout the estuary (e.g.

Brush & Brush 1994). The history of eutrophication of

Chesapeake Bay during the nearly 400 yr since European colonization has received particular emphasis

(e.g. Brush 1984).

Over the course of recent decades, ~300 dated sediment cores have been collected from Chesapeake Bay

and its tributaries. These cores vary in length from 0.5

to 20 m and contain histories of sedimentation rates,

organic carbon sources, salinity and temperature regimes, O2 conditions, and biological changes evidenced in microfossil assemblages and plant seeds

over the last 12 000 yr (Colman et al. 2002). Time lines

have been established in cores using isotopic tracers

(e.g. 14C, 210Pb) and event horizons (e.g. 137Ce and ragweed pollen). Representative data series for key biotic

and geochemical indices are presented to illustrate

patterns of eutrophication of the Bay during the centuries since European colonization (Fig. 5). The first

signs of human disturbance were evident in the 17th

and 18th centuries (Fig. 5), with rapidly increasing

sedimentation rates (Brush 1989, Cronin & Vann 2003).

Concomitant increases in burial of both total organic

carbon (associated with terrestrial and aquatic plants)

and biogenic silica (associated with planktonic

diatoms) provide early evidence of eutrophication

(Colman & Bratton 2003). 

The sedimentary record also provides key indices of

eutrophication-induced changes in the structure of the

estuarine ecosystem. For example, although rates of

bacterial carbon burial exhibited little change over the

first 3 centuries of record (Fig. 5), a sharp increase was

evident starting in the mid-20th century (Zimmerman

& Canuel 2000), with parallel increases in the ratio of

bacterial carbon to biogenic silica possibly reflecting a

decline in the efficiency at which diatom production is

transferred to upper trophic levels (e.g. Kemp et al.

2001). During the same time period, a pronounced shift

in the ratio of centric (planktonic) to pennate (typically

benthic) diatoms in the preserved record (Fig. 5)

5

||| | | | |

Relative Index Value

1

2

3

4

5

6

7

1600–1770 1770–1870 1870–1910 1910–1950 1950–1975 1975–2000

BSi

Flux

Forams

(Hypoxia)

Diatom Ratio

(Turbidity)

Mo

(H2S)

TOC

Flux

Bacterial C

Fig. 5. Comparison of indicators of eutrophication in sedimentary record of deep channel of mesohaline Chesapeake Bay.

‘Relative index values’ are calculated as ratio of each index

value (rate, concentration, ratio) for each date to mean value

calculated for precolonial (pre-17th century) conditions. Parameters and sources of data used are: total organic carbon

(TOC) mass flux corrected for degradation (Zimmerman &

Canuel 2002; their Fig. 6, cores RD and M3); delivery factors

for bacterial organic carbon (Zimmerman & Canuel 2000, their

Fig. 10); incorporation of biogenic silica (BSi; Colman & Bratton 2003); ratio of centric to pennate diatoms (Cooper 1995,

their Fig. 8); concentration of molybdenum (Mo; Adelson et al.

2001, their Fig. 5, cores 55 and PC-6); and percentage (compared to a baseline of <10%) of total benthic Foraminifera

(Forams) comprised of a hypoxia-tolerant species, Ammonia

parkinsoniana (Karlsen et al. 2000, their Figs. 12 & 13)
Mar Ecol Prog Ser 303: 1–29, 2005

reflects decreases in both water clarity and benthic

algal production (Cooper & Brush 1993). These trends

may have initiated the general shift from a largely

benthic estuarine ecosystem to one dominated by

planktonic processes.

Most sedimentary indicators suggest that the increased occurrence of bottom-water hypoxia and

anoxia in the main-stem Bay is a relatively recent phenomenon. Although a few indices, such as dinoflagellate cyst formation rate (Willard et al. 2003), provide

earlier signs of hypoxia, the majority of reported biotic

and geochemical indicators show that the intense and

recurrent seasonal depletion of O2 is relatively unique

to the last 50 yr. For example, solid-phase molybdenum

formation rates (which are proportional to sulfide

concentrations, Adelson et al. 2001) and the proportion

of total foraminifera fossil abundance comprised of

the hypoxia-tolerant species Ammonia parkinsoniana

(Karlsen et al. 2000) both exhibited steep increases

since the early to mid-20th century (Fig. 5). Similar

trends have been shown for other indices of bottomwater hypoxia, including the degree of pyritization of

iron (Cooper & Brush 1991, 1993) and the ratio of acidvolatile sulfur to chromium reducible sulfur (Zimmerman & Canuel 2002).

PRIMARY PRODUCERS AND BIOGEOCHEMISTRY

Nutrient inputs, concentrations, and fates

Current estimates of total nitrogen (TN) and phosphorus (TP) loads (per estuarine area) to Chesapeake

Bay and its tributaries average about 14 g N m–2 yr–1

and 1.1 g P m–2 yr–1, respectively (Boynton et al. 1995).

These values vary by a factor of 2 between wet and dry

years (Hagy et al. 2004) and are intermediate compared to those of other estuarine systems (Billen et al.

1991, de Jonge et al. 1994, Nixon et al. 1996). Much

higher rates of nutrient inputs (e.g. 100 g N m–2 yr–1) in

urban tributaries (e.g. near Baltimore) promote very

high summer planktonic chlorophyll-a (chl-a) concentrations (>250 μg l–1). As shown for other coastal

systems (e.g. Borum 1996), interannual variations in

loading rates are reflected by nutrient concentrations

in the estuary, with specific relationships between

loading and concentrations often differing among

systems (Fig. 6). 

A mass-balance analysis of sources and sinks

for total N and total P in Chesapeake Bay and 4

major tributaries indicated several consistent patterns

(Boynton et al. 1995). The seasonally varying ratio of

TN:TP watershed inputs to the Bay has been gradually increasing in recent decades (Hagy et al. 2004).

For the whole Bay, diffuse watershed sources represented 60% and 58%, while direct atmospheric

deposition comprised 12% and 6.5% of TN and TP

inputs, respectively. The remainder of the nutrient

inputs are derived primarily from watershed point

sources for TN, while oceanic inputs provide an

additional source (37% of the total) for TP. The dominant loss terms for the estuarine N budget are denitrification (26% of TN inputs) and long-term burial

(35%), while burial is the primary loss term for P.

Losses due to commercial and recreational fisheries

are small for N (9%) and smaller yet for P (5%).

These values vary somewhat among tributary subestuaries, but it is clear that, in contrast to some estuaries, the Bay does not act like a pipe conveying

nutrients directly to the adjacent ocean (Borum 1996,

Nixon et al. 1996). Whereas most of the TN and TP

inputs to the Bay are inorganic forms, N and P

exported to coastal waters are largely organic, indicating active uptake and transformation of nutrient

inputs and a positive net primary production in the

Bay ecosystem (Kemp et al. 1997).

6

b) Potomac River Estuary

a) Maryland Mainstem Chesapeake Bay

1985–1997

Annual Average TN (mg l-1)

TN Load (g N m-2 yr-1)

8 10 12 14 16 18 20 22 24 26

0.5

0.6

0.7

0.8

0.9

1.0

1.1

0.4

0.6

0.8

1.0

1.2

1.4

1.6

1.8

20 25 30 35 40 45 50 55 60

Fig. 6. Relationship between annual mean concentrations of

total nitrogen concentration (TN) and TN loading for (a) main

stem of Chesapeake Bay and (b) Potomac River estuarine 

tributary
Kemp et al.: Chesapeake Bay eutrophication

Phytoplankton biomass and community composition

Increased abundance of phytoplankton is an early

manifestation of eutrophication in many aquatic ecosystems (e.g. Smith 2003). Although rates of diatom

production and burial in Chesapeake Bay sediments

appear to have increased steadily since the early 19th

century (Fig. 5), recent historical analysis of direct

measurements of phytoplankton chl-a (Harding 1994,

Harding & Perry 1997) reveal that surface mixed-layer

concentrations also increased significantly between

1950 and 1994, with the polyhaline regions showing

the largest changes. Extensive monitoring data collected since 1994 have enabled an updating of these

trends (Fig. 7). Chl-a increased 1.5- to 2-fold in oligohaline and mesohaline regions peaking in the 1960s

(Fig. 7a,b) and 5- to 13-fold in the polyhaline region of

the Bay from the 1950s to the 1980s (Fig. 7c). Chl-a has

decreased in the upper oligohaline region since the

1970s (Fig. 7a) in response to relatively high turbidity

(e.g. Fisher et al. 1999, Harding et al. 2002) associated

with elevated river flow (Fig. 4). Although no temporal

trends in chl-a are evident in the main-stem Bay from

1985 to 2004, significant increases and decreases are

apparent in various tributaries (www.eyesonthebay.net). 

This pattern—phytoplankton chl-a increasing between the 1950s and 1980s and unchanged during the

last decade—corresponds to reported trends in N

loading to the Bay during this period (Fig. 4e). Similar

relationships between temporal trends in nutrient

loading and chl-a have been reported for other coastal

environments (e.g. Radach 1992, Cadee 1992). Comparative analyses among estuaries around the world

have also revealed broad positive relationships between nutrient loading or concentration and algal production and/or biomass (e.g. Boynton et al. 1982,

Monbet 1992, Nixon et al. 1996). Primary productivity

of the whole ecosystem (planktonic plus benthic) may,

however, saturate at relatively lower nutrient inputs

(Borum & Sand-Jensen 1996).

As demonstrated in various coastal environments experiencing increased nutrient loading (e.g. Cederwall

& Elmgren 1990, Cadee 1992), direct microscope analyses of Bay waters have revealed general trends of shifts

in phytoplankton community dominance from larger to

smaller cells (e.g. Marshall 1994). Although few historical direct observations are available to examine possible shifts in phytoplankton community structure,

paleobotanical studies reveal increases in the ratio of

centric-to-pennate diatoms during the last 2 centuries

(Cooper & Brush 1991, Cooper 1995). In addition,

analysis of sediment cores indicates that relative abundances of dinoflagellates, cyanobacteria, and small flagellates appear to have increased significantly during

the last half century (Zimmerman & Canuel 2002).

In some instances, eutrophication-induced shifts in

phytoplankton communities involve enhanced growth

of algal species that cause direct harmful effects, including production of toxins, noxious discoloration,

and floating mucilage (e.g. Paerl 1988, Smayda 1990,

Anderson et al. 2002). Although factors causing these

harmful algal blooms (HABs) are complex, many

have been associated with nutrient enrichment (e.g.

Cadee & Hegeman 1986, Lukatelich & McComb 1986,

Smayda 1997). There are numerous instances where

HABs have been reported in Chesapeake Bay and its

tributaries, and several appear to be related directly

7

0

5

10

15

20

a Oligohaline

b Mesohaline

c Polyhaline

V

VI

III

IV

I

II

0

5

10

15

20

0

5

10

15

20

Years

50–59 60–69 70–79 80–89 90–99 00–03

50–59 60–69 70–79 80–89 90–99 00–03

50–59 60–69 70–79 80–89 90–99 00–03

Chl 

a (mg m-3)

Fig. 7. Historical changes in annual mean surface layer concentrations of phytoplankton chlorophyll-a. Data are organized into 6 previously defined segments (numbered consecutively from ocean to land) along axis of main-stem Bay

(Harding 1994), and these are grouped into 3 major salinity 

regions (oligohaline, mesohaline, polyhaline)
Mar Ecol Prog Ser 303: 1–29, 2005

to nutrient inputs. For example, in the oligohaline

Potomac River estuary (Fig. 1), frequency of summer

blooms of the toxic cyanobacterium Microcystis aeruginosa declined sharply in the early 1970s when P removal from sewage was initiated (Sellner et al. 1988,

Jaworski 1990). Although other HAB outbreaks in

the Bay have been more difficult to relate directly to

nutrient enrichment, blooms of both the common Bay

dinoflagellate Prorocentrum minimum and the rarer

mixotrophic dinoflagellate Pfiesteria piscicida appear

to be stimulated by addition of dissolved organic nitrogen (DON) including urea and humic acids (Glibert

et al. 2001, Heil 2005). In addition, ‘brown tide’ blooms

of the small pelagophyte Aureococcus

anophagefferens, which occur in shallow

coastal lagoons (Bricelj & Lonsdale 1997)

including those adjacent to Chesapeake

Bay (Lomas et al. 2001, Trice et al. 2004),

have also been linked to DON enrichment (Berg et al. 1997). Cells of P. minimum and A. anophagefferens inhibit

growth of shellfish (Luckenbach et al.

1993, Bricelj & Lonsdale 1997), while

P. piscicida produces a lethal neurotoxin

(e.g. Burkholder & Glasgow 1997).

Depletion of bottom-water oxygen

Organic matter produced in phytoplankton blooms sinks into deep Bay

waters where it is decomposed in oxygenconsuming processes. The increasing

frequency and magnitude of seasonal

oxygen (O2) depletion from bottom

waters is a phenomenon linked closely

with anthropogenic nutrient enrichment

in Chesapeake Bay and elsewhere (e.g.

Rosenberg 1990, Johannessen & Dahl

1996). Direct measurements, however,

indicate that hypoxia (<2.0 mg O2 l

–1)

occurred occasionally in deep waters of

the main-stem Bay even in the 1930s

(Newcombe & Horne 1938, Newcombe

et al. 1939). Although reports in the

1980s suggested historical trends of

increasing Bay hypoxia (Officer et al.

1984), these patterns were questioned

(Seliger & Boggs 1988), largely because

of the overriding effect of interannual

variations in river flow on stratification

(e.g. Fig. 3). A recent analysis of data

collected between 1950 and 2001 supports both contentions (Hagy et al. 2004).

Despite the significant correlations observed between time-integrated volume of hypoxic

(and anoxic, <0.2 mg O2 l

–1) water and winter–spring

river flow (Fig. 8a), trend analysis revealed that, consistent with patterns in sedimentary records (Fig. 5),

significant increases in severity and spatial extent of

hypoxia have occurred since the 1950s. 

The spatial distribution and seasonal development of

hypoxia in Chesapeake Bay illustrate that O2 depletion arises from interactions between biological and

physical processes (Taft et al. 1980, Kemp et al. 1992).

Hypoxia develops in the Bay’s bottom layer, appearing

first in late spring at the northern limit of the stratified

flow and expanding southward as summer unfolds.

8

a Hypoxia vs. River Flow

10 15 20 25 30 35

0

1

2

3

4

5

6

7

Anoxic Volume (km3) 1000 1400 1800 2200 2600 Annual Hypoxic/Anoxic Volume-Days (km3 d)

Jan–May Average Susquehanna River Flow (m3 s-1)

0

200

400

600

800

1000

1200

Jan–May NO3 Loading (Gg)

Hypoxic Water

y = 0.18x + 424

(r2 = 0.34)

Anoxic Water

y = 0.18x – 82

(r2 = 0.72)

YEARS 1980–2001

y = 0.14x – 0.12

(r2 = 0.33)

YEARS 1950–1979

y = 0.09x – 0.81

(r2 = 0.25)

b Anoxia vs. NO3 Loading

Fig. 8. Trends in volume of low O2 bottom water in mesohaline region of mainstem Chesapeake Bay. (a) Variations in time-integrated summer volume of hypoxic (O2 < 2 mg l–1) water (dashed line, open circles) and anoxic (operationally

O2 ≤ 0.2 mg l–1) water (solid line, closed circles) from 1984 to 2003 vs. mean

winter-spring flow of Susquehanna River, and (b) midsummer volume of anoxic

bottom water vs. winter-spring nitrate loading from Susquehanna River for

earlier years (1950–1979, solid line, filled circles) and for later years (1980–2001, 

dashed line, open circles). Modified from Hagy et al. (2004)
Kemp et al.: Chesapeake Bay eutrophication

The timing of O2 depletion in spring is predicted by

freshwater inflow, which regulates water column stratification and associated rates of O2 replenishment, and

by spring water temperature, which affects respiration

rates (Hagy et al. 2004). Interannual variations in

spring nutrient loading have also been correlated with

rates of organic deposition to sediments (Boynton &

Kemp 2000), which would be expected to promote

increased O2 demand. Initial springtime rates of O2

decline have, however, varied little since 1938 (Hagy

et al. 2004). This suggests that the initial spring decline

in O2 is strongly controlled by physical processes,

while the late spring O2 decline and the extent of summer hypoxia are more closely related to eutrophication

(Hagy et al. 2004).

It is difficult to resolve the relative roles of nutrient

inputs and other factors in the development of summer

hypoxia from a short time series of data. It is clear that

human activities from 1950 to the present have significantly changed N loading to the Bay (Fig. 4d,e).

Since both hypoxia and N loading have increased over

time, the two are correlated. Unexpectedly, however, it

appears that hypoxia has tended to be more severe in

recent years even at equivalent levels of N loading

(Hagy et al. 2004). In fact, there are 2 separate significant relationships between hypoxia and nitrate loading for earlier (1950–1979) and later (1980–2001)

years, with similar slopes but very different intercepts

(Fig. 8b). This implies that the Bay has become less

able to assimilate N inputs without developing

hypoxia, a change that may have arisen from the

degradation of key ecological processes sensitive to

eutrophication effects. Potential mechanisms include

(1) loss of benthic plant biomass due to increased turbidity and loss of oyster biomass, both of which tend to

retain nutrients and organic matter in shallow waters;

(2) increased efficiency of N and P recycling with

marked decreases in denitrification and P precipitation

in response to recent severe and persistent hypoxia.

Redox-induced changes in sediment biogeochemistry

Depletion of O2 from bottom waters causes a decrease in the oxidation-reduction potential (redox

condition), which changes the fundamental nature of

sediment biogeochemistry and nutrient cycling processes (e.g. Stumm & Morgan 1970). Even under normoxic (>3 mg O2 l

–1) conditions, sulfate (SO4) is the

dominant terminal electron acceptor supporting benthic respiration throughout much of the Bay; especially

during summer hypoxia when SO4 reduction rates are

very high (Roden & Tuttle 1993, Marvin-DiPasquale &

Capone 1998). Accumulation of aqueous Fe(II) and

Mn(II) in subpycnocline waters (Gavis & Grant 1986)

indicates that these metals may also serve as respiratory electron acceptors or (as with O2) as oxidants of

reduced sulfur generated from sulfate reduction. Although rapid formation of iron sulfide minerals occurs

under anoxia during summer months, the complete

‘titration’ of iron oxides by reduced sulfur (Cornwell &

Sampou 1995) results in a summer efflux of 65 to 95%

of the reduced sulfur to the overlying water (Roden &

Tuttle 1992), where it exerts a substantial O2 demand

(Kemp et al. 1992). 

In general, low redox conditions result in elevated

effluxes of ammonium (NH4) and phosphate (PO4) from

sediments to overlying water. As has long been

reported for lakes (e.g. Einsele 1936, Mortimer 1941),

release of PO4 from sediments is enhanced under

anoxic conditions in deep regions of Chesapeake Bay

(e.g. Cowan & Boynton 1996). In contrast, Bay sediments in shoal areas above the pycnocline seldom

exhibit high rates of PO4 release (Boynton & Kemp

1985, Reay et al. 1995), except under conditions of

abnormally high pH (Seitzinger 1991). Under anoxic

conditions, the dissolution of near-surface iron oxides

and/or their conversion to iron sulfides (Cornwell &

Sampou 1995) tends to coincide with the onset of high

PO4 effluxes from sediments. In the mesohaline region

of the Bay, effluxes of PO4 from sediments are 5-fold

larger when bottom water O2 is <1.5 mg l–1 (Fig. 9a).

Similarly, the fraction of remineralized NH4 that is

recycled from sediments to overlying water is also controlled by redox conditions (Fig. 9b). Whereas virtually

all regenerated NH4 diffuses into the overlying water

under anaerobic conditions, nitrifying bacteria oxidize

a substantial fraction of this NH4 to nitrate (NO3) when

O2 is present, leading to subsequent reduction to N2

(e.g. Graco et al. 2001). Nitrifying bacteria require

O2, and their activity is inhibited by the elevated

concentrations of hydrogen sulfide that typically accompany anoxia (Henriksen & Kemp 1988, Joye &

Hollibaugh 1995). In the mesohaline Bay, rates of nitrification are high in spring and fall when bottom waters

are oxygenated but are negligible during midsummer

anoxia (Kemp et al. 1990). Consequently, the ratio of

NH4 recycling to total efflux of inorganic N solutes

(NH4 + N2 + NO3) tends to be inversely related to

bottom-water O2 concentrations (Fig. 9b).

The enhanced efficiency of NH4 and PO4 release

from O2-stressed sediments represents an important

biogeochemical feedback mechanism that reinforces

the eutrophication process. Increased phytoplankton

production associated with nutrient enrichment leads

to bottom-water O2 depletion, which in turn increases

the rate and efficiency of nutrient recycling from

sediments to the euphotic zone, thereby amplifying

eutrophication. Negative effects of hypoxia on the

physical-chemical processes that control PO4 recycling

9
Mar Ecol Prog Ser 303: 1–29, 2005

can be reversed within hours with restoration of high

O2 levels, while recovery of nitrifying bacteria from

anoxia effects can take weeks to months. Longer-term

(years to decades) impacts of chronic low O2 conditions

on PO4 and NH4 recycling may arise from anoxiainduced reductions in abundance, size, and activity of

benthic bioirrigating macrofauna (e.g. Schaffner et al.

1992). For example, feeding and burrowing activities

of many benthic macrofaunal species tend to oxygenate sediments, which effectively reduces nutrient

recycling efficiency, particularly for NH4 (e.g. Pelegri

et al. 1994, Mayer et al. 1995).

Declining water clarity and benthic microalgal

production

Water clarity. An important and direct consequence

of eutrophication is decreased water clarity and reduction in light available for benthic primary production

(e.g. Nielsen et al. 2002a). On the other hand, turbidity

in some systems may be controlled more by physical

conditions including watershed soil erosion and resuspension of bottom sediments (e.g. Cloern 1987). While

anthropogenic sediment loading contributed substantially to Chesapeake Bay turbidity in the 18th and

19th centuries (e.g. Roberts & Pierce 1974), watershed

inputs of sediment have declined dramatically since

the 1940s (e.g. Brush 1989). In addition, bottom resuspension associated with sediment dredging and demersal fishing tends to produce only localized and

ephemeral increases in Bay turbidity (e.g. Ruffin 1998).

Consequently, light attenuation in many estuarine

regions such as the middle and lower Bay is largely

controlled by interactions between plankton and suspended sediments (e.g. Gallegos 2001). Sharp declines

in Bay water clarity during spring and summer can be

directly related to algal blooms stimulated by watershed nutrient inputs (Gallegos & Jordan 2002). Reductions in depth distribution of conspicuous benthic

plants (e.g. seagrasses, macroalgae) often provide

direct evidence of eutrophication impacts on water

clarity and benthic primary production (e.g. Giesen

et al. 1990, Nielsen et al. 2002b, Kemp et al. 2004).

Although parallel reductions in benthic microalgal

production certainly occur, these are commonly overlooked (e.g. Rizzo et al. 1992, Fear et al. 2004). 

There are surprisingly few data on water clarity prior

to the 1940s. Limited historical secchi depth (Zsd) measurements collected routinely in plankton studies,

however, reveal dramatic increases in turbidity between 1930 and 1970 (e.g. Heinle et al. 1980, D’Elia et

al. 2003). A large compilation of such Zsd data for the

Patuxent estuarine tributary (Stankelis et al. 2003, W.

Boynton unpubl. data) indicates that water clarity was

better in the late 1930s than in the 1960s and slightly

better in the 1990s compared to the 1960s (Fig. 10).

Although changes in water clarity from the 1930s to

the 1960s were muted in the lower salinity regions

(6 to 10), substantial decreases in Zsd are evident in the

more saline (>10) part of this Bay tributary (Fig. 10).

These data are consistent with anecdotal observations,

which suggest dramatically clearer waters in the lower

Patuxent estuary prior to 1945 (e.g. Klingel 1951).

Benthic primary production. There are few direct

measurements of benthic microalgal production in the

contemporary Bay (e.g. Rizzo & Wetzel 1985, Murray &

Wetzel 1987, Reay et al. 1995) and none prior to 1940.

Compelling indirect evidence, however, indicates that

there has been a fundamental shift in the contribution

of benthic microalgae to total Bay primary production

during the last century. As reviewed above, paleoecological evidence reveals contemporaneous decreases

in benthic diatoms and increases in various indices of

eutrophication during the last century (Fig. 5). Presum10DIP Flux

0

20

40

60

80

100

%

Dissolved Oxygen (mg l-1)

0 8 10 12 2 4 6

20

40

60

80

100

0

a Benthic Recycling of PO4

b N-Recycling Efficiency

y = 100 – 5.18x

(r2 = 0.83)

Fig. 9. Relationships between bottom-water O2 concentrations

and benthic nutrient recycling for (a) effluxes of PO4 from

sediments to overlying water, where histograms indicate

mean and standard error for fluxes at O2 < 1.5 mg l–1 and O2 >

1.5 mg l–1 (W. Boynton et al. unpubl.) and (b) percentage of

total inorganic N efflux from sediments accounted for by DIN

efflux (J. Cornwell et al. unpubl.). Measurements were made

in mesohaline locations in Chesapeake Bay and its tributaries 

during spring and summer
Kemp et al.: Chesapeake Bay eutrophication

ably, nutrient enrichment has promoted an

overall increase in production and biomass

of phytoplankton, which have contributed

to decreases in both water clarity and

growth of benthic diatoms. Although benthic microalgal production may actually

increase with nutrient loading in some

very shallow estuaries (e.g. de Jonge

1990), similar nutrient-induced shifts from

benthic to planktonic microalgal production have been widely reported for other

estuaries (Nielsen et al. 2002b) and for

lakes (Vadeboncoeur et al. 2003). 

Recent analyses suggest that benthic

microalgal production presently contributes less than 10% of the total primary

productivity in Chesapeake Bay (Kemp et

al. 1999). Here, we extend these calculations, which combine data on light attenuation and depth-area geometry (Fig. 2)

with benthic photosynthesis-irradiance

relationships, to estimate benthic microalgal productivity when the Bay was less

eutrophic. Assuming that the relative

changes (from the 1930s to the 1990s) in

Zsd for regions in the Patuxent River estuary (Fig. 10) hold for the entire Bay, we

computed decreases in light reaching the

sediment surface. In this heuristic calculation we also

estimated integrated values for recent and historical

rates of benthic microalgal production in 3 Bay regions

(Table 1). Although we assume no change in benthic

production in the low salinity region, it is estimated

that rates decreased by about 50% in the middle and

lower Bay. If we further assume that phytoplankton

productivity was similar in the 1930s and 1950s and

contemporary Bay-wide mean values for chlorophyllspecific production have not changed (Harding et al.

2002), we estimate that between the 1930s and 1990s

the ratios of benthic-to-planktonic microalgal rates of

primary production have declined from 0.3 to 0.1 and

from 1.4 to 0.1 in the middle and lower Bay, respectively. This marked shift in the polyhaline Bay is

largely attributable to the broad shallow areas flanking

this region’s central channel (Fig. 2). These estimated

declines in benthic-to-planktonic production ratios

would be larger still if we had considered the likely

increases in phytoplankton productivity from the 1930s

to the 1950s as well as the additional loss of water clarity that has probably occurred with declining oyster

and seagrass populations (see below). 

It is, thus, probable that the balance of ecosystem

production was very different in the 1930s, with benthic autotrophs playing a major role. Compared to

plankton, benthic microalgal communities (including

bacteria, meiofauna, and macrofauna) tend to use and

retain nutrients far more efficiently, with relatively

small effluxes of NH4 and PO4 from sediments (Tyler et

al. 2003) and low rates of denitrification (e.g. Sundbäck

et al. 2000). Thus, in systems dominated by benthic

autotrophs nutrient resources generate relatively high

11

Patuxent Secchi Depth Trends

Surface Salinity

6-8 8-10 10-12 12-14

Avg Secchi Depth (m)

0.0

0.5

1.0

1.5

2.0

2.5

1930s

1960s

1990s

7

3 2

9

5 1

8

2

5

6

9

13

Fig. 10. Variations in mean (+SE) Secchi disk depth observations in the Patuxent River estuary during 3 decades (1930s,

1960s, 1990s) along salinity gradient (data from Stankelis et

al. 2003, W. Boynton unpubl). Values above each histogram

indicate number of measurements available for that time 

period and estuarine region

Era Regiona Z1% b Bottom areac < Z1% Benthic GPPd

(m) (106 m2) (109 g C yr–1)

1930s Upper Bay 1.7 101 18

Middle Bay 7.6 828 137

Lower Bay 8.8 1508 210

Total Bay 7.6 2508 396

1990s Upper Bay 1.7 101 18

Middle Bay 4.5 461 77

Lower Bay 5.2 649 99

Total Bay 4.6 1365 217

aUpper Bay defined as main stem area from 39.0–39.5° N; Mid Bay as

37.9–39.0° N; Lower Bay as 37.0–37.9° N

b1990s data from 15 to 20 stations in main stem Bay sampled seasonally

for light attenuation coefficient from 1995–1999. 1930s data estimated

by adjusting 1990s means to reflect historical changes similar to those

observed in the Patuxent River (Fig. 10). Total Bay values based on

area-weighted mean of regional values

cAreas calculated from regional hypsographic data (Fig. 2) from Cronin

& Pritchard (1975)

dCalculated from estimated light attenuation data, an annual average

production vs. irradiance relationship empirically derived from sediment O2-flux data collected in the Middle Bay region from 1994–1995

(Kemp unpubl.), annual mean surface light data (Fisher et al. 2003), and

integrated by multiplying m2 rates by area at 1 m depth intervals then

summing over total depth for each region

Table 1. Summary of annual mean values for depth of 1% light penetration

(Z1%), surface area of bottom shallower than Z1%, and benthic microalgal

gross primary production (Benthic GPP) estimated for 1930s and 1990s in 

3 main-stem regions of Chesapeake Bay
Mar Ecol Prog Ser 303: 1–29, 2005

rates of primary production per unit nutrient inputs.

Mucous excretion by benthic diatoms and depositfeeding invertebrates tends to bind surface sediments

reducing resuspension and turbidity (Madsen et al.

1993), while benthic microalgae production also supports efficient demersal food chains (Miller et al. 1996).

Loss of submersed vascular plants

The shoal areas of Chesapeake Bay, which historically harbored abundant and diverse communities of

seagrasses and other submersed vascular plants, have

experienced dramatic changes in plant density and

distribution during the last 50 yr (Kemp et al. 1983,

Orth & Moore 1983). Many plant species have declined, including at least 15 salt-tolerant freshwater

species in the upper Bay, the seagrass Zostera marina

in polyhaline regions, and the euryhaline seagrass

Ruppia maritima throughout much of the estuary. In

one broad shallow upper Bay area, for example, the

initial decline in plants (1960s) was preceded by prolific growth of the exotic Myriophyllum spicatum,

which itself eventually disappeared along with the

native species by the mid-1970s (Fig. 11a). As noted

elsewhere (e.g. Marba & Duarte 1997), interannual

differences in the Bay’s plant abundance are related to

climatic events, such as the 1972 tropical storm Agnes,

which deposited a large volume of sediments over the

upper Bay (Schubel & Hirschberg 1978). 

Although prior to 1985 there were very few reported

large-scale multispecies losses of submersed plants

from coastal habitats (e.g. Den Hartog & Polderman

1975), such declines of submersed plants in lakes had

already been attributed to eutrophication and associated shading from algal growth (e.g. Jupp & Spence

1977, Phillips et al. 1978). These reports motivated

Bay research of submersed plant responses to nutrient

enrichment. Studies in experimental estuarine ponds

(Kemp et al. 1983, Twilley et al. 1985) showed that

nutrient addition led to significant decreases in submersed plant growth that could be directly explained

by shading from increased phytoplankton and epiphyte biomass (Fig. 11b). Subsequent mesocosm experiments supported these observations for many species

(e.g. Neckles et al. 1993, Neundorfer & Kemp 1993,

Sturgis & Murray 1997, Moore & Wetzel 2000). Significant correlations between plant abundance and water

quality throughout the Bay corroborated experimental

results (Dennison et al. 1993, Stevenson et al. 1993).

Studies elsewhere reported similar results (e.g. SandJensen & Borum 1991, Short et al. 1995), leading to

widespread recognition that submersed plant losses in

response to coastal eutrophication were global in scale

(Duarte 1995, Short & Wyllie-Echeverria 1996).

Submersed plant beds influence many ecological

processes in coastal areas (e.g. Walker et al. 2001). In

Chesapeake Bay these plants provide food and habitat

for diverse animal populations (Lubbers et al. 1990),

including the economically valuable blue crab Callinectes sapidus (Orth & van Montfrans 1990). Submersed plants also affect biogeochemical processes by

enhancing net deposition of suspended particles (Ward

et al. 1984, Rybicki et al. 1997), thereby increasing

water clarity, benthic photosynthesis, and nutrient

assimilation (Kemp et al. 1984). Submersed plants also

release O2 from roots to the surrounding rhizosphere,

stimulating N loss via coupled nitrification-denitrification (Caffrey & Kemp 1990, 1992). For various seagrasses, the trapping of particles and direct assimila12 % Cover

0

20

40

60

80

100

0

0.1

0.2

0.3

1900 1920 1940 1960 1980 2000

Control Low Medium High

0

2

4

6

8

10

Plant

Epi

Phyto

Plants (kg m-2)

Epi (g g-1), Phyto (μg chl a l-1)

Year

a Submersed Plants in Upper Bay

b Nutrient Response

Fig. 11. Variations in submersed vascular plants in upper and

mid Chesapeake Bay. (a) Historical changes in relative abundance of submersed plants at an upper Bay site (Susquehanna

Flats) from 1900 to 2000. Data from 1900 to 1955 based on

seed abundance in dated sediment cores (Brush & Hilgartner

2000); data from 1955 to 1973 are adapted from Bayley et al.

(1978); data from 1974 to 2000 are from Chesapeake Bay Program (www.chesapeakebay.net); (b) variations in biomass of

submersed plants, epiphytic algae (Epi) and phytoplankton

(Phyto) in triplicate experimental estuarine ponds under different nutrient loading rates, from low (0.4 g N and 0.09 g P

m–2 d–1) to high (1.7 g N and 0.4 g P m–2 d–1), plus controls with 

ambient water (adapted from Kemp et al. 1983)
Kemp et al.: Chesapeake Bay eutrophication

tion of solutes from overlying water have been shown

to be important terms in the N and P budgets of these

beds (e.g. Hemminga et al. 1991, Risgaard-Petersen et

al. 1998). Few studies have, however, attempted to

place such rate calculations in the context of estuaryscale nutrient budgets. Drawing from earlier work

(Kemp et al. 1984), we develop here a mass-balance

analysis to illustrate the quantitative importance of

these beds as sinks in the N budget of the entire upper

Bay (Table 2). If submersed plant beds were restored to

historical levels of areal coverage (i.e. most areas <2 m

mean depth, Kemp et al. 2004), they would remove

almost 45% of the current N inputs to the upper Bay

from watershed and atmospheric sources, with most of

this attributable to particle trapping and direct assimilation. Even partial restoration of these plant beds

would, thus, substantially help to mitigate effects of

nutrient loading. 

There is evidence of modest recovery of submersed

plants in the upper Bay (Fig. 11a) and other estuarine

regions during the last 20 yr, especially in tributaries

where nutrient levels have been declining (Glibert

& Magnien 2004). A decade of post-Agnes sediment

stabilization and an extended drought in the late 1980s

(Fig. 4) may have allowed partial reestablishment of

submersed plant species, such as Ruppia maritima and

Zostera marina, which reproduce proficiently by seeds

(e.g. Verhoeven 1979, Harwell & Orth 2002). However,

many Bay regions remain devoid of seagrasses, and

many historically dominant plant species remain

scarce. In any case, it is clear from field and modeling

studies that widespread recovery of submersed plants

in the Bay must begin with reduced nutrient levels and

increased water clarity.

Loss of tidal marshes as nutrient buffers

There has long been scientific interest in understanding patterns of net nutrient exchanges between

tidal marshes and adjacent estuaries (e.g. Nixon 1980,

Dankers et al. 1984, Valiela et al. 2000). Recent studies

suggest that, as with submersed plants, many tidal

marshes are capable of buffering eutrophication

effects by trapping and assimilating nutrients and by

stimulating denitrification at the margins of estuaries

(Bricker & Stevenson 1996). Although early Chesapeake Bay studies of nutrient fluxes in marsh tidal

creeks failed to show clear patterns of nutrient removal

at tidal scales (e.g. Heinle & Flemer 1976, Stevenson

et al. 1977), recent measurements and mass-balance

analyses suggest that Bay marshes can be major nutrient sinks. For example, particulate N and P trapped

in marshes of the tidal fresh Patuxent represent 35%

and 80% of the respective inputs to

the upper estuary (Merrill & Cornwell

2000). Furthermore, rates of denitrification and N assimilation in fringing

tidal marshes have been estimated to

remove almost 80% of the groundwater

N inputs to a smaller Bay tributary

(Stevenson et al. 2002).

During the 17th century, land clearance in the Bay watershed was limited to

small tobacco fields surrounded by thick

vegetation. Consequently, little sediment escaped the land (Stevenson et al.

1999, Brush & Hilgartner 2000). By the

end of the 18th century, however, many

streams were choked with sediment resulting from the increase in agricultural

land and a shift from tobacco to wheat

farming (Cronon 1983). Interestingly,

this led to a large 19th century expansion

of the Bay’s tidal marshes, particularly

in low- and mid-salinity estuarine regions (Froomer 1980). Early maps, for

example, suggest that the areal extent

of marshes on the Bay’s eastern shore

continued to grow until 1900 (peaking

at >40 000 ha) but stabilized thereafter

as sea-level rise accelerated from ~0.5 to

13

Table 2. Estimated influence of submersed vascular plants on nitrogen budget 

for ‘restored’ plant communities in upper Chesapeake Baya

N-sources or sinks N-fluxes Source

(106 kg N yr–1) (% of input)

Nitrogen inputsb

Point discharges 19.2 24 Boynton et al. (1995)

Diffuse loads 55.2 68 Boynton et al. (1995)

Atmospheric deposits 6.2 8 Boynton et al. (1995)

Total input 80.6 1000

Plant nitrogen sinks

Assimilationc 12.1 15 Kemp et al. (1984

Denitrificationd 1.7 2 Caffrey & Kemp (1992)

Sediment N trappinge 22.4 28 Ward et al. (1984)

Total plant sink 36.2 45

aUpper main-stem Bay defined as that region above latitude 38o N excluding

major tributaries. Area of this region is 3.9 × 109 m2, and potential habitat of

submersed plant communities ‘restored’ to historical levels would cover

0.56 × 109 m2 (Kemp et al. 1999, 2004)

bData taken directly from Table 3 in Boynton et al. (1995) for ‘Maryland

Mainstem’

c

Assumes an estimated annual primary production rate of 360 g C m–2 yr–1

(Kemp et al. 1984) and a tissue C:N molar ratio of 19 (Caffrey & Kemp 1990)

dPlant ‘enhancement’ of sediment denitrification taken as difference between rates measured in vegetated and unvegetated sediments (1.8 mmol

N m–2 d–1, Caffrey & Kemp 1990, 1992) for a 4 mo growing season

eBased on plant-enhanced sedimentation of 0.2 cm mo–1 for a 4 mo

growing season, with sediment dry bulk density of 1 g cm–3 and N content

of 0.5% dry wt (Ward et al. 1984)
Mar Ecol Prog Ser 303: 1–29, 2005

3 mm yr–1 between the 17th century and the present

(Kearney 1996, Stevenson & Kearney 1996). Beginning

in the late 20th century, however, many marshes began

showing clear signs that they were unable to keep

pace with sea-level rise (Stevenson et al. 1985, 2000).

Currently, more than 50% of the Bay’s marshes show

evidence of deterioration (Kearney et al. 2002).

Present-day erosion of upper and middle Bay tidal

marshes is significantly reducing nutrient buffering capacity at the estuarine margins (Kearney et al. 1988,

Stevenson et al. 1999). In addition, the release of organic

matter and nutrients from these deteriorating marshes

(Kemp et al. 1997) may actually be accelerating eutrophication in shallow estuarine environments (Stevenson

et al. 2002). For example, severe nighttime hypoxia has

been observed in tidal ponds that were once emergent

marshes (Stevenson et al. 2002). It is, thus, ironic that

many of the Bay’s tidal marshes, which were created in

the 18th century as a consequence of eroding agricultural lands and became effective nutrient and sediment buffers, are now being lost due to both insufficient

upland inputs of sediments and rapidly rising sea level.

FOOD WEBS AND CONSUMER POPULATIONS

Hypoxia and loss of benthic macroinfauna

Although direct measurements of benthic macroinfaunal abundance in deep Chesapeake Bay sediments

are not available prior to 1950, the dominance of laminated sediments in long cores (2 to 3 m) collected in the

mid-Bay channel suggests that bioturbating infauna

have been essentially absent from this region for about

100 yr (Schaffner et al. 1992). A shift of dominant

macroinfaunal species observed in the early 1960s in

deep muddy sediments of a Bay tributary (York River

estuary) was attributed to increasing hypoxic stress

(Boesch & Rosenberg 1981). Regular sampling in the

mesohaline Bay did not begin until the 1970s (Holland

et al. 1977), after severe hypoxia had become well

established as an annual feature. Nonetheless, a significant hypoxia-related decrease in benthic macroinfauna was observed during 1971–1984, primarily

involving the replacement of larger older bivalves by

short-lived opportunistic species (Holland et al. 1987).

In a recent synthesis, Herman et al. (1999) interpreted a significant relationship between benthic

macroinfauna biomass and primary productivity in

shallow normoxic estuarine systems (Fig. 12) to suggest that these benthic animals tend to be food limited.

For the polyhaline lower Bay, measurements of primary productivity and infaunal biomass correspond

well with this relationship, while the mean level of

infaunal biomass in the upper Bay is also consistent

with this relationship when allochthonous inputs

(Kemp et al. 1997) are combined with phytoplankton

productivity (Fig. 12). In the middle seasonally hypoxic

Bay regions, however, benthic biomass is only 10%

of that expected from this region’s high phytoplankton production rates (emphasized by dashed line in

Fig. 12). Bioenergetic analyses indicate that, although

growth of macrobenthos may be limited by food availability in upper and lower Bay regions (e.g. Thompson

& Schaffner 2001), this is not the case in the mid-Bay,

where benthos appear to be degraded (Gerritsen et al.

1994, Hagy 2002). These empirical relationships suggest hypoxia-induced degradation of benthic infaunal

in the main-stem Bay (Hagy 2002) as well as in Bay

tributaries (Llanso 1992, Dauer et al. 2000). The fact

that many of the Bay’s historically important macroinfaunal species have limited tolerances to low O2

(Diaz & Rosenburg 1995) supports this contention.

In general, food-limited benthic communities tend

to increase with eutrophication until hypoxic stress

reverses this trend (Cederwall & Elmgren 1990).

The strong seasonality of macrobenthic biomass

in the middle Bay may also be partially driven by

hypoxia. High summer mortality in sediments below

the pycnocline destroys the large spring recruitment of

opportunistic species in early spring, and this is followed by limited recolonization upon return to normoxia in autumn (Holland et al. 1987). Episodic winddriven lateral oscillations in the Bay’s pycnocline cause

hypoxic deep channel water to well up onto adjacent

14

Annual Phytoplankton Production (g C m-2 yr-1)

Macrobenthic Biomass (g AFDW m

0 100 200 300 400 500 600 700

-2)

0

10

20

30

40

50

60

70

Upper

Bay

Lower

Bay

Mid

Bay

1

2

3

4

6 5

7

8

9 10

11

12

13

14

Fig. 12. Relationship between annual mean values of phytoplankton production and macrobenthic biomass in estuaries

and coastal systems. The Upper Bay value for primary production includes estimates of allochthonous organic matter

inputs from the Susquehanna River. Site locations are as

follows: 1 = Ythan Estuary, 2 = Grevelingen, 3 = Oosterschelde, 4 = Balgzand (1980’s), 5 = Veerse Meer, 6 = Ems Estuary (Wadden), 7 = Lynher Estuary, 8 = Long Island Sound, 9 =

San Francisco Bay, 10 = Balgzand (1970s), 11 = Bay of Fundy,

12 = Westershelde, 13 = Ems Estuary (inner), 14 = Columbia

Estuary (adapted from Herman et al. 1999 and Hagy 2002)
Kemp et al.: Chesapeake Bay eutrophication

shoals (Malone et al. 1986, Sanford et al. 1990), thereby

degrading the shallow-water macrobenthic communities in these otherwise normoxic habitats (Breitburg

1990, Hagy 2002). In addition, infaunal biomass tends

to decline with incipient hypoxia (Holland et al. 1987),

reflecting increased vulnerability to predation for these

new recruits (Pihl et al. 1992). Even in shallow normoxic regions, common midsummer declines in benthos appear due to intensified predation by benthivorous nekton excluded from adjacent deep hypoxic

waters (Kemp & Boynton 1981). 

Zooplankton responses to nutrients

In contrast to benthic invertebrates, there is little evidence of zooplankton community responses to eutrophication in Chesapeake Bay. The Bay’s 2 dominant copepod species Acartia tonsa and Eurytemora affinis play

pivotal roles in transferring phytoplankton production

to fish (Baird & Ulanowicz 1989). These copepods are

consumed directly by bay anchovy, the estuary’s most

abundant fish (Rilling & Houde 1999), and by juvenile

menhaden, white perch, and other forage fish (e.g.

North & Houde 2003). Copepods also experience

significant mortality due to predation by gelatinous

consumers, particularly the planktivorous ctenophore

Mnemiopsis leidyi (Feigenbaum & Kelly 1984). Feeding and bioenergetic studies (Heinle 1966, Heinle et al.

1977) suggest that zooplankton growth in the Bay is

generally not limited by food (White & Roman 1992).

Although copepod biomass may be stimulated by

nutrient additions in mesocosms (Bundy et al. 2003),

such responses generally do not occur in the presence

of planktivorous predators at natural abundances

(Breitburg et al. 1999).

Zooplankton abundance and production in Chesapeake Bay and other estuaries may, however, also be

affected by eutrophication-induced changes in copepod habitats. For example, increases in bottom-water

hypoxia may reduce adult survival and egg hatching

in Acartia tonsa and Oithona colcarva (Roman et al.

1993). Indeed, significantly reduced copepod abundances have been noted in hypoxic waters of the Bay

and other estuaries (e.g. Caumette et al. 1983, Keister

et al. 2000). Although inverse correlations between

zooplankton abundance and nutrient or chl-a concentrations have been reported for highly eutrophic Bay

tributaries (Park & Marshall 2000), such relationships

are not evident for most of the Bay (Kimmel & Roman

2004) or for other shallow estuaries (Lawrence et al.

2004). Hypoxia is a poor predictor of Eurytemora affinis

abundance, but these copepods are strongly related to

interannual variations in the size of oligohaline and

turbid habitats (Roman et al. 2001, Kimmel & Roman

2004) in the upper Bay, where food (detritus and particle-attached bacteria, Crump et al. 1998) is abundant

while salinity and temperature conditions best suit the

physiological needs of these animals (Bradley 1975).

Variations in freshwater inflow to the Bay can also

alter the balance between bottom-up and top-down

controls on zooplankton (Purcell & Decker 2005). Not

only does high flow deliver more nutrients and extend

the habitat of Eurytemora affinis, it also reduces the

distribution and abundance of the medusa Chrysaora

quinquecirrha (Purcell et al. 1999), which feeds heavily

on Mnemiopsis leidyi (Purcell & Cowan 1995). Under

the low flow conditions of 1995, for example, elevated

salinity throughout the Bay promoted large summer

populations of C. quinquecirrha, which controlled the

ctenophores, releasing copepods from predation and

allowing them to increase (Fig. 13). In contrast, high

flow conditions in 1996 stimulated spring abundance

of copepods (primarily E. affinis) by increasing oligohaline habitat (and perhaps stimulating algal growth);

however, low salinity also suppressed summer populations of C. quinquecirrha, thereby allowing M. leidyi

to flourish and graze down copepods (Fig. 13). Such

‘trophic cascades’ involving copepods and jellyfish

might also be initiated by changes in fishing pressure,

as in the Black Sea, where concurrent overfishing and

introduction of non-native M. leidyi may have created

an unbalanced food web (Daskalov 2002, Gucu 2002).

It has been suggested that eutrophication-induced

changes in phytoplankton community structure and O2

conditions of estuaries might lead to shifts in planktonic food webs, favoring pathways dominated by bacteria and gelatinous predators (e.g. Egge & Aksnes

1992, Purcell et al. 2001, Turner 2001). Although bacterioplankton abundance is high in Chesapeake Bay

waters (e.g. Jonas 1997), the ratio of bacterioplankton

to phytoplankton production is lower than expected

from relationships in more oligotrophic systems (e.g.

Ducklow & Shiah 1992). Statistical relationships between hypoxia and jellyfish abundance, reported at

interannual and system-level scales for other eutrophic

coastal regions, are limited to smaller geographic and

temporal scales within the Bay (Purcell et al. 2001).

Loss of oyster beds and benthic filtration

Extensive reefs of the eastern oyster Crassostrea virginica built up along the flanks of Chesapeake Bay and

its tributaries during and after the Holocene transgression. Because the Bay bottom is composed predominantly of fine particles, oysters created a unique hardsubstrate habitat for diverse invertebrate and fish

populations (Coen et al. 1999). Oysters supported a

valuable fishery, with substantial landings from the

15
Mar Ecol Prog Ser 303: 1–29, 2005

1880s to the early 1900s (Fig. 14). Although Bay oyster

growth may have been initially stimulated by anthropogenic nutrient enrichment (Kirby & Miller 2005),

these levels of exploitation proved unsustainable because of excessive removal of adult oysters and their

shell substrate (Kennedy & Breisch 1981). Loss of vertical relief made remaining shell more susceptible to siltation, compounding population decline (Smith et al.

2003b). This fishery-related decline was exacerbated

in the 1950s by outbreaks of 2 major parasitic diseases,

MSX (Haplosporidium nelsoni) and ‘Dermo’ (Perkinsus marinus) (Rothschild et al. 1994). Today, oyster

abundance in the Bay has been reduced to ~1% of 19th

century levels.

Suspension feeding by benthic bivalves, such as oysters, tends to cause a shift in the balance from planktonic to benthic production by reducing concentrations

of phytoplankton and other suspended particulates,

thereby increasing light levels reaching the sediment

surface (e.g. Cohen et al. 1984, Newell & Koch 2004).

The Bay’s 19th century oyster population filtered (during summer months) a water volume equivalent to that

of the upper and middle Bay in ~3.6 d; however, oyster

declines have increased this filtration time to several

hundred days at present (Fig. 14 inset, Newell 1988).

Although this calculation involves many simplifying

assumptions, it illustrates the magnitude of potential

ecosystem effects associated with loss of oysters (e.g.

Gerritsen et al. 1994). Similar calculations for other

coastal systems also suggest large potential impacts of

benthic filter-feeding invertebrates (Cloern 1982, Loo

& Rosenberg 1989). Benthic filter feeders can, thus,

substantially alter trophic dynamics in planktonic and

benthic communities (e.g. Ulanowicz & Tuttle 1992,

Strayer et al. 1999). 

Benthic filtration can also affect nutrient cycling processes. Whereas intense benthic filtration associated

with commercial culture can stimulate nutrient recycling (e.g. Souchu et al. 2001) and growth of benthic

macroalgae (e.g. Rafaelli et al. 1998), experiments in

the Bay suggest that natural oyster reefs may actually

16

a Spring Copepods (April)

0

10

20

30

40

Low Flow 1995

High Flow 1996

b Summer Copepods (July)

0

10

20

30

40

c Mnemiopsis leidyi (July)

0.0

0.5

1.0

1.5

2.0

2.5

3.0

d Chrysaora quinquecirrha (July)

Distance Down Bay (km)

0

10

20

30

40

50

103 ind. m-3 103 ind. m-3 l m-3 ml m-3

Abundance

0-50 50-100 100-150 150-200 200-250 250-300

Fig. 13. Variations in abundance (number of individuals or biovolume) of key plankton consumers along a longitudinal transect from upper to lower main-stem Chesapeake Bay for relatively low flow year (light shaded bars, 1995) and relatively

high flow year (dark shaded bars, 1996). Copepods are mostly

Acartia tonsa and Eurytemora affinis; the ctenophore Mnemiopsis leidyi consumes mostly copepods and the medusa

Chrysaora quinquecirrha consumes copepods and ctenophores; data are binned into 6 regions (0–50, 50–100,

100–150, 150–200, 200–250, 250–300 km) along length of Bay,

with numbers of observations in each bin ranging from 3 to 18 

(M. Roman et al. unpubl.)

Time (yr)

1880 1900 1920 1940 1960 1980 2000

Wet Meat Weight (kg x 106

)

0

10

20

30

40

Maryland

Virginia

Days to Filter Maryland’s

Portion of Chesapeake 

Bay

Time (d)

0

100

200

300

400

500

600

700

228

700

3.6

1880 1988 2003

Fig. 14. Historical record of harvest of eastern oyster (Crassostria virginica) from Chesapeake Bay. Inset shows calculated

changes in time required for oyster population in each year

to filter a volume of water equal to that in the Bay north 

of ~38° N latitude (adapted from Newell 1988)
Kemp et al.: Chesapeake Bay eutrophication

reduce the efficiency by which nutrients are recycled

to phytoplankton. The eastern oyster maintains high

clearance rates that are relatively independent of food

concentration, resulting in the production of copious

particulate N and P as pseudofeces (Newell et al.

2005). Most of this material is deposited initially in

shallow areas, where it supports enhanced production

of deposit-feeding animals and benthic microalgae

(Newell et al. 2005). Increased NH4

+ production in

oyster biodeposits that are degraded in shallow oxygenated sediments may also stimulate N2 loss via coupled nitrification-denitrification (Newell et al. 2002).

To the extent that oyster biodeposits are retained in

estuarine shoals, their production helps to reduce

delivery of planktonic organic matter to the deep

channel where it fuels hypoxia. The demise of oysters

in the Bay has, however, rendered these processes

inoperative. 

Fish and fisheries responses

Chesapeake Bay is a productive fisheries ecosystem,

with commercial landings that exceeded 225 000 metric

tons (196 kg ha–1) until the mid-1990s, when they declined to about 170 000 tons (Houde et al. 1999). Eutrophication and associated hypoxia may have altered

community structure and productivity of fish and invertebrates that sustain Bay fisheries (Breitburg 2002).

Changes in fish populations and fisheries harvest that

occurred during the past 2 centuries are, however, also

tied to many factors, including shifts in fisheries preferences, fishing effort, markets, loss of habitats, declining

water quality, and variable climate/weather.

Conceptually, it has been proposed (Caddy 1993,

2000) that eutrophication effects on fish communities

follow a sequence of 3 stages: (1) nutrient-enhanced

production of demersal and pelagic species (more

food), (2) decline of demersal fish but continued increase in pelagic fish species (benthic habitat loss), and

(3) a general decline in total fish production under conditions of broadly deteriorating water and habitat quality. Based on landings, this sequence has been, in part,

observed in Chesapeake Bay. Fisheries landings are

now dominated by the pelagic species Atlantic menhaden, with catches >150 000 tons during the 1980s

and early 1990s. It is probable that long-term increases

in overall production of menhaden (Luo et al. 2001)

and other Bay fisheries have been stimulated by

increased nutrient loading, as in other ecosystems

(Nixon & Buckley 2002). Although lesions on fish,

HAB-induced fish kills, and fish/crab mortalities from

hypoxic events appear to be increasing, there is only

limited evidence that eutrophication has pushed the

Bay ecosystem into Caddy’s third stage.

The dramatic impact of increased fishing pressure

makes it difficult to detect any effects of eutrophication

on the Bay’s fish populations. For example, as with the

precipitous decline in eastern oyster Crassostrea virginica harvest (Fig. 14), recent declines in blue crab

landings since 1990 likely resulted primarily from

excessive fishing (CBC 2003), not eutrophication. Fishing mortality and climatic factors led to the decline and

failed recruitments of striped bass Morone saxatilis,

and subsequent recovery has occurred by managing

harvest (Richards & Rago 1999). Some species such as

sturgeons, extirpated nearly a century ago by fishing

and habitat loss, probably can no longer reproduce or

rear young in the eutrophic Bay due to lack of summer

habitat with O2 and temperature levels needed for

growth and survival (Niklitschek 2001). Even for dystrophic ecosystems such as the Black Sea and Seto

Inland Sea, it has been suggested that excessive fishing contributed substantially to the destabilization of

food chains and the shifts toward dominance by gelatinous planktonic predators (Daskalov 2002, Gucu 2002,

Nagai 2003).

In general, increases in the ratio of landings of

pelagic to demersal fish species (P:D) have been associated with eutrophication (de Leiva Moreno et al.

2000). In Chesapeake Bay, this ratio increased from

1.90 in the 1960s to 2.66 in the 1990s (Fig. 15). Com17 Relative Landings (%)

a Trophic Guilds in Fishery

b Pelagic:Demersal Fish in Harvest

P:D Ratio

0

10

20

30

40

50

60

70

Planktivore Benthivore Piscivore

60s

70s

80s

90s

0

1

2

3

1960s 1970s 1980s 1990s

Fig. 15. Decadal changes in relative landings (% of total

weight) for fish and invertebrate species harvested in Chesapeake Bay from 1960s to 1990s, (a) organized into 3 feeding

guilds (planktivores, bentivores, piscivores) and (b) organized

as the ratio of planktivores:benthivores (P:D). Data adapted 

from Web site (www.noaa.chesapeakebay.net)
Mar Ecol Prog Ser 303: 1–29, 2005

mercial landings of demersal benthivores declined

from 34.4% to 27.3% of the catch, with concomitant

increases in relative landings of pelagic species. The

largest shift was the progressive increase in pelagic

Atlantic menhaden landings (~38% in the 1960s to

~67% in the 1990s). Declines in demersal blue crab

and oyster landings also contributed to these trends.

Whereas the increase in P:D may be related to parallel

increases in nutrient inputs, phytoplankton biomass,

and hypoxia (e.g. Figs. 4, 7 & 8), shifts in coastal fishing

patterns and markets were also factors. In addition, the

ratio of commercial fishery yield (Houde et al. 1999) to

net primary production (PP, Harding et al. 2002) suggests declining trophic efficiency in the Bay. Fishery

yields were ~2.20 g C m–2 yr–1 (0.54% of PP) in the

1980s compared to ~1.85 g C m–2 yr–1 (0.45%) for the

mid-1990s. The absence of clear evidence of declines

in fish productivity does not necessarily indicate that

there have been no effects of eutrophication on Bay

fisheries. As indicated for the Black Sea, detrimental

effects of eutrophication may not be fully manifested

until a combination of excessive fishing activity, unusual

climate regimes, introductions of alien species, and

nutrient loading overwhelm the ecosystem’s resilience.

SCIENCE AND NUTRIENT-REDUCTION

STRATEGIES

Nutrient limitation for primary production

A question central to the application of scientific

understanding to eutrophication management is the

relative importance of N and P as nutrients limiting

phytoplankton primary production (e.g. Likens 1972).

Whereas estuarine ecosystem responses to nutrient

enrichment are varied and complex, most start with

stimulation of algal (primarily phytoplankton) growth.

In developing a nutrient management strategy, it is

therefore crucial to understand how nutrients limit

algal growth in time and space (Graneli et al. 1990).

Lake phytoplankton are generally P-limited (Schindler

1981), while marine phytoplankton are more often Nlimited (Ryther & Dunstan (1971). P limitation may also

dominate in coastal ecosystems where water exchange

rates exceed P recycling rates (e.g. Smith 1984). Nutrient (N or P) limitation in estuaries tends to vary with

season (e.g. D’Elia et al. 1986, Paasche & Erga 1988)

and salinity regime (Caraco et al. 1987). 

The relative importance of nutrients and light in limiting phytoplankton in the Bay exhibits well-defined

seasonal and regional variations (Fig. 16). (We define

‘light limitation’ as ambient conditions where phytoplankton growth is controlled by insufficient light,

thereby precluding growth responses to nutrient addition.) Such conditions are common in the vertically

mixed upper Bay waters during winter when salinity

is low, water is turbid, and nutrients are replete. P limitation occurs primarily in spring (April–May), when

DIN:PO4 ratios of inflowing river exceed 60:1, and

other resources or conditions are not limiting. For brief

periods in late autumn, P limitation is also evident in

the upper and middle Bay. N limitation is common

throughout the Bay in summer and early autumn when

DIN is depleted from surface waters, while rates of PO4

release from anoxic sediments are high (Fig. 9) and

nutrient inputs from point sources with low DIN:PO4

ratios (<10:1) are more important (Fisher et al. 1992). In

the lower Bay, N and P commonly exhibit colimitation

when both concentrations are low (DIN < 0.5 μmol l–1;

PO4 < 0.1 μmol l–1). 

18

P-Limitation

N+P

Limit

N-Limitation

Main Stem Chesapeake Bay

Month

Distance from Susquehanna R., km

0

50

100

150

200

250

300

Light

Limitation

J F M A M J J A S O N D

Fig. 16. Seasonal and regional variations in nitrogen, phosphorous or light limitation for phytoplankton growth synthesized from bioassay experiments conducted in main stem of

Chesapeake Bay between 1992 and 2002 (Fisher et al. 1992,

1999, 2003). N- and P-limitations were defined as increases in

phytoplankton productivity (PP) and biomass (PB) with N and

P, respectively, additions to experimental containers; N+Plimitation was defined as increases in PP and PB only with

simultaneous additions of N plus P. ‘Light-limitation’ was

defined by the lack of significant responses to any combination of N and/or P addition, combined with a uniform response of increased PP and PB during incubations for controls and all treatments relative to initial ambient conditions
Kemp et al.: Chesapeake Bay eutrophication

These strong seasonal and regional patterns have

important management implications for controlling

eutrophication in Chesapeake Bay (D’Elia et al. 2003).

Reductions in both N and P from rivers and point

sources would be required to reduce

Bay phytoplankton biomass consistently

throughout the year for the whole Bay

(Fig. 16). It is tempting to suggest efficient

nutrient management may be achieved

by varying controls on N and P inputs

seasonally and regionally; however, such

approaches may have limited utility. For

example, the relatively long but variable

residence time of Bay water makes it difficult to time changes in nutrient inputs to

ecological responses in specific regions

and seasons. As noted elsewhere, controlling only P in the upper Bay, where it is

limiting, is likely to cause excess concentrations of the uncontrolled nutrient (N),

which would be transported seaward to

stimulate downstream algal growth (Paerl

et al. 2004). 

Examples of ecosystem responses to

reduced inputs of N and/or P

Although reductions in both N and P

inputs will be required to effect substantial reversal of eutrophication throughout

the Bay, partial nutrient reduction in tributaries has elicited significant ecological

benefits. Two case studies from Chesapeake Bay tributaries, the Potomac and

Patuxent River estuaries, serve to illustrate ecosystem responses to reductions

in P and N inputs. In both systems, pointsource effluents comprise a large fraction

of the total nutrient loading. 

In the Potomac River estuary, improved

sewage treatment in Washington, DC

produced a sharp reduction in pointsource P (and, to a lesser extent, N) loading beginning in the early 1970s (Jaworski

1990). This was followed by substantial

changes in ecosystem characteristics

within the tidal freshwater region of

the estuary. Rapid responses included

decreased algal biomass, higher water

column O2 levels, and increased water

clarity, while submersed vascular plants

exhibited delayed but subsequently extensive recolonization of shoal areas

beginning in 1983 (Fig. 17). Observed

increases in O2 are probably attributable to a combination of decreases in P and organic loading (not shown)

from sewage treatment facilities. The latter would

account for direct reduction in O2 demand, and the for19

TIDAL FRESH

POTOMAC

MESOHALINE

PATUXENT

No Data

0 0 0

1960 1970 1980 1990 2000

Year

1950 1960 1970 1980 1990 2000

Year

Point Source P-Loading Fall Line N-Loading

Phytoplankton Biomass Phytoplankton Biomass

Dissolved Oxygen Dissolved Oxygen

Water Clarity Water Clarity

Submersed Plants Submersed Plants

Cover (ha x 103

) Secchi (m) O2 (mg l-1) Chl a (μg l-1) Load (kg x 103) d-1

0.0

0.4

0.8

1.2

1.6

0.0

0.2

0.4

0.6

0.8

0

2

4

6

8

10

0

20

40

60

80

100

0

2

4

6

8

10

12

0

2

4

6

8

10

12

0.0

0.4

0.8

1.2

1.6

0

1

2

3

4

5

6

0

1

2

3

4

5

0

10

20

30

40

Fig. 17. Time-series plots of nutrient loads, phytoplankton biomass, dissolved

O2, water clarity and submersed vascular plant biomass in tidal fresh portion

of Potomac River estuary and mesohaline portion of Patuxent River estuary.

Data are averaged for 5 or 10 yr intervals. Phytoplankton chl-a concentrations

are from surface waters at both locations, dissolved O2 concentrations are water

column averages for Potomac and bottom layer averages for Patuxent. Water

clarity was measured with a Secchi disk at both sites. Potomac data sources

were as follows: nutrient loads from N. Jaworski (pers. comm.) and Chesapeake Bay Program (CBP, www.chesapeakebay.net), plankton biomass, dissolved O2 and water clarity from N. Jaworski (pers. comm.) and CBP and submersed plants from Carter & Rybicki (1986) and CBP. Patuxent data sources

were: nutrient loads from D’Elia et al. (2003), plankton biomass, O2, water clarity

and submersed plants from Heinle et al. (1980), Stankelis et al. (2003) and CBP
Mar Ecol Prog Ser 303: 1–29, 2005

mer would result in reduced algal biomass (Fig. 17)

retention and decomposition within this estuarine

region. The recovery of submersed plants, which was

their first appearance since their demise in the 1930s

(Carter & Rybicki 1986), was attributed to the

improved water clarity (Carter & Rybicki 1990, Carter

et al. 1994). Resurgent submersed plant beds were

initially (1980) dominated by the non-native species

Hydrilla verticillata, but by 1985 approx. 14 native

plant species were observed in the region (Carter &

Rybicki 1986). Increased benthic suspension feeding

associated with a contemporaneous invasion of the

non-native clam Corbicula fluminea probably contributed to initial recovery of this region of the Potomac. By 1982, however, density of these clams had

declined to 10% of their peak abundance without discernible reductions in water quality (Cohen et al. 1984,

Phelps 1994).

During the same general time period, changes in nutrient loading to the Patuxent River estuary produced a

more complex pattern (Fig. 17). Inputs of N increased

markedly between 1950 and 1985 and, with the application of advanced wastewater treatment, started a decline in the early 1990s that has continued to the present. P inputs decreased abruptly in the late 1980s with

sewage treatment improvements and the banning of

PO4 from detergents (D’Elia et al. 2003). Biomass of

phytoplankton followed trends in nutrient loading

closely, while water clarity and dissolved O2 in deep

waters declined with increasing N loading but have

thus far shown little sign of recovery with the recent

reduced loading (Fig. 17). The decline in submersed

plants began in the late 1960s, with most of the estuary becoming devoid of these plants by the mid-1970s

(Stankelis et al. 2003). Efforts to restore submersed

plants via both planting of shoots and seeds have failed

to produce stands of plants that persist for more than a

year. In general, shoots planted in spring or fall (or

areas seeded in fall) emerged and grew well during

cool periods (<20°C) but became heavily covered by

epiphytic algae in early summer and died by the end of

that summer or by the middle of the next (Stankelis et

al. 2003). Evidently, nutrient load reductions are still

insufficient for the successful expansion of submersed

plants in the lower Patuxent River estuary.

These examples illustrate the propensity of this

estuarine ecosystem for recovery from eutrophication

effects. They also reveal potential weaknesses of

single vs. dual nutrient reduction strategies. In the

Potomac, for example, ecosystem responses to P removal have been confined to the tidal fresh region,

and, as observed elsewhere (Paerl et al. 2004), water

quality conditions have not improved in the lower estuary as unassimilated DIN has been transported seaward to stimulate algal growth in the N-limited region.

Removal of N and P in the Patuxent has resulted in

improved water quality throughout the estuary, although responses of submersed plants and bottom O2

have been limited to date. Even in the Potomac’s tidal

fresh region, changes in concentrations of chl-a and

PO4 suggest a hysteretic pattern, where the same chl-a

levels are associated with lower values of PO4 during

the recovery period (1990s) compared to the degradation (1970s) period (Jones 2000). Nutrient management

strategies should be developed with awareness of the

potential for such non-linear feedback effects (e.g.

Yamamoto 2003).

Strategies for nutrient reductions

Results of scientific research have contributed to a

political and public awareness and commitment to reverse eutrophication in the Chesapeake Bay (Malone

et al. 1993). Documentation of improved water quality

in the tidal Potomac River following upgrading of

wastewater treatment led to growing awareness of

eutrophication problems and the prospects for correcting them. By the early 1980s, however, research results

were suggesting that eutrophication was a far more

pervasive problem throughout the Bay (Boesch et al.

2001). This perspective was translated into a series of

agreements—involving all political jurisdictions within

the watershed—to reduce inputs of both N and P

from point and non-point sources (D’Elia et al. 2003).

Researchers and managers collaborated in developing

detailed nutrient budgets and water quality models to

produce management strategies. Public commitments

to reducing eutrophication were recently codified in

the ambitious Chesapeake 2000 (C2K) agreement

(www.chesapeakebay.net), which calls for reductions

of 48% and 53% for total N and P inputs, respectively

(based on 1985 levels). 

An important component of the C2K agreement is a

plan for restoring 3 major Bay habitats—seagrass

beds, oyster reefs, and tidal marshes. Although largescale recovery of seagrass will ultimately require substantial water quality improvement (e.g. Dennison et

al. 1993, Kemp et al. 2004), this process can be accelerated via regional strategies: (1) in the lower Bay by

direct dispersal of plant seeds collected from existing

Zostera marina beds (Harwell & Orth 2002), (2) in the

upper Bay by protection of existing small plant beds (in

tertiary tributaries) that serve as seed and fragment

sources (e.g. Rybicki et al. 2001), and (3) in the middle

Bay by using existing but ephemeral Ruppia maritima

beds as ‘nursery areas’ for transplanting more stable

perennial species. Initial efforts to increase eastern

oyster biomass by 10-fold (www.oysterrecovery.org)

have produced modest success in isolated areas, but

20
Kemp et al.: Chesapeake Bay eutrophication

Bay-wide restoration will require solving problems of

diseases and shortages of seed stocks and hard substrates (NRC 2004). Restoration of the Bay’s tidal

marshes faces the basic problem of insufficient inputs

of natural sediment to keep pace with sea-level rise.

Recent efforts to restore marshes using channeldredging spoils have generated new techniques to

oxidize and dewater sediments. In any case, even

small progress in the restoration of these habitats could

yield substantial benefits toward recovery from

eutrophication because key ecological processes (e.g.

nutrient assimilation, particle trapping, filter feeding,

denitrification) are enhanced through inherent biological positive-feedback mechanisms.

SYNTHESIS

Symptoms of ecological response to eutrophication have been manifested at a range of time

scales in Chesapeake Bay. Initial signs of

organic enrichment are evident in 200 yr old

sediment strata, while evidence of increased

algal production and decreased water clarity

appear in 100 yr old sediment layers. Severe

hypoxia and loss of submersed plants were first

evident only 4 to 5 decades ago. These changes

in shallow habitats, combined with hypoxiainduced loss of the deep benthic habitats, probably contributed to a gradual increase in the

relative importance of pelagic (compared to

demersal) food chains and declines in associated trophic efficiency (e.g. fishery harvest per

unit primary production) over the last several

decades. Because scientific interest in these

ecological changes tends to lag behind their

occurrence, historical knowledge of responses

to eutrophication depends largely on fairly

imprecise interpretation of biogeochemical indices in sediment strata and on retrospective

analysis of data collected for other purposes. 

A rich but indirect body of evidence suggests

that Bay responses to nutrient enrichment are

complicated by a range of non-linear ecological

feedback mechanisms (Fig. 18; also see Bonsdorff et al. 1997). For example, enhanced particle trapping and sediment binding associated

with benthic plants (seagrass, microalgae) help

to maintain relatively clear water columns,

allowing more light to support more benthic

photosynthesis. As the Bay ‘degrades’ and

becomes more turbid with enrichment, however, these benthic autotrophic communities

decline, allowing more resuspension, decreased

light, and so on. Similarly, nutrient-enhanced

phytoplankton growth and sinking support increased

benthic respiration and severe anoxia, which causes

more efficient benthic recycling of N (inhibition of

nitrification-denitrification) and P (increased solubility

of PO4) to support further production of phytoplankton

(including HABs) in overlying water. 

Although the positive-feedback nature of these

interactions means that they will tend to reinforce and

accelerate the eutrophication process, it also means

that they will reinforce the ‘restoration’ process by

enhancing water quality improvements once they are

initiated (Fig. 18). The relatively rapid recoveries of

some water quality variables with reduced nutrient

21

Degradation Trajectory Restoration Trajectory

More Nutrients

N & P

Less Nutrients

N & P

More Algae Less Algae

& Turbidity & Turbidity

More O2

 in

Deep Water Deep Water

Less O2 in 

Less Light &

Benthic

More Light &

Benthic

Lower Redox & Higher Redox &

NitrificationDenitrificationNitrificationDenitrification

Less N & P

Uptake, More

Resuspension

More N & P

Uptake, Less

Resuspension

More Recycling

of NH4

 + PO4 of NH4 + PO4

Less Recycling

Nutrient Feedback

Water Clarity Feedback

Nutrient Feedback

Production Production

Degraded

Oysters

Eroded

Tidal Marshes

Healthy

Oysters

Expanded

Tidal Marshes

Disease,

Harvest

Sea Level

Rise

Restoration

Sediment

Accretion

Fig. 18. Logic flow diagram illustrating how nutrient addition affects algal

biomass (light gray boxes), which affects both bottom-water oxygen and nutrient recycling (medium gray boxes) as well as water clarity and benthic

primary production (dark gray boxes), and how positive feedback mechanisms (related both to nutrient levels and to water clarity) reinforce this eutrophication cascade regardless of the trajectory. Also illustrated is how the

health of oyster reefs and tidal marshes (light gray boxes) affects these interactions. Along a ‘degradation trajectory’ these interactions reinforce the

decline in water quality and habitat condition; however, along a ‘restoration

trajectory’ they reinforce the improvement of water and habitat quality
Mar Ecol Prog Ser 303: 1–29, 2005

loading in the Potomac and Patuxent tributary systems

(Fig. 17) may reflect these positive-feedback processes, while hysteresis is evident in the delayed

recovery of submersed plant beds. Water filtration

from oyster reefs provides a negative-feedback control

on eutrophication by reducing plankton biomass and

increasing water clarity (Fig. 18). Development of many

of the Bay’s tidal marshes may have actually been promoted by sediment and nutrient loading from cultural

eutrophication. Once established, however, the ability

of these marshes to trap, assimilate, and remove nutrients (Fig. 18) has made them effective buffers, providing a negative-feedback control on further nutrient

inputs from watershed to estuary. Unfortunately, however, this control is now being lost to sea-level rise.

There remain many compelling and relevant scientific questions pertaining to responses of Chesapeake

Bay and other estuaries to nutrient loading. Details of

how interactions among organisms and biogeochemical processes regulate dynamic ecological responses to

nutrient inputs are unresolved. Evidence for hysteresis

in the recovery process suggests the existence of complex mechanisms that are not fully described but that

may be important for managing estuarine resources

(e.g. Jeppesen et al. 1998). Hypothesized eutrophication-induced changes in trophic structure and shifts

between benthic and pelagic food chains need to be

understood better, particularly as they affect the efficiency at which photosynthesis is converted into fisheries production. Future research on coastal eutrophication should be designed to promote both applied

and basic scientific objectives, serving to deepen our

understanding of estuarine ecosystem function and to

develop prudent and effective strategies for managing

valuable coastal resources.

Acknowledgements. We are indebted to many individuals

who contributed discussions, analyses, and technical support

for this paper and the research upon which it is based. These

include R. Batiuk, M. Brooks, J. Caffrey, J. Davis, B. Dennison, R. Lacouture, C. Madden, M. Mallonee, L. Murray, K.

Moore, B. Orth, M. Owens, L. Staver, and J. Testa. Support for

the work was provided over many years by numerous funding

agencies, including US Environmental Protection Agency,

NOAA Maryland Sea Grant College, NOAA Coastal Ocean

Program, National Science Foundation, Maryland Department of Environment, and Maryland Department of Natural

Resources. This is contribution number 3899 from the University of Maryland Center for Environmental Science.

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29

Editorial responsibility: Otto Kinne (Editor-in-Chief), 

Oldendorf/Luhe, Germany

Submitted: December 21, 2004; Accepted: August 21, 2005

Proofs received from author(s): October 25, 2005