Environmental Health Perspectives • VOLUME 109 | SUPPLEMENT 5 | October 2001 699

In the continental United States, estuaries

comprise more than 80% of the coastline

along the Atlantic Ocean and Gulf of Mexico

and more than 10% of the Pacific coast.

These fragile habitats, which are among this

nation’s most important natural resources, are

experiencing declining water quality and

eutrophication (1–4). In a recent comprehensive survey of the trophic status of estuaries in

the continental United States, Bricker et al.

(3) concluded that 84 estuaries, representing

65% of the total estuarine surface area, were

presently showing signs of moderate to high

eutrophic conditions (Table 1). Rapidly

growing and diversifying anthropogenic

inputs associated with agriculture, aquaculture, urbanization, coastal development, and

industrial expansion are a primary cause of

the decline in the quality of natural habitats

in these sensitive waters (1,5,6). 

Within the past three decades, many of

our estuarine and coastal waters have changed

from balanced and productive ecosystems to

ones experiencing sudden trophic changes,

biogeochemical alterations, and a deterioration in habitat quality. Nuisance and sometimes harmful phytoplankton blooms,

accompanied by oxygen depletion, toxicity,

fish kills, and shellfish mortality, are becoming

more common (2,3). The purpose of this general overview is to furnish a working definition for estuarine eutrophication and to

outline some key ecological features associated

with the underlying processes. Although a

comprehensive review of all the facets of

eutrophication are well beyond the scope of

this concise overview, we have incorporated

many key concepts and have supplied a few

selected examples to illustrate different aspects

of estuarine eutrophication. Our overall goal is

to provide readers with a broad working

knowledge of eutrophication processes in the

context of environmental health issues.

A Definition of Eutrophication

The terms eutrophic, mesotrophic, and oligotrophic are adjectives commonly used to

describe the overall state of fertility or “trophic

status” of aquatic ecosystems. These three

broad categories delineate a gradient that ranges

from nutrient-poor, low-biomass systems (oligotrophic) to nutrient-rich, high-biomass habitats (eutrophic). The term eutrophication, in

contrast, describes a process rather than a

trophic state. Nixon (1) proposed that eutrophication be defined as “an increase in the rate of

supply of organic matter to an ecosystem.” This

definition has gained wide acceptance and may

be the most frequently used quantitative measure of eutrophication. Organic matter (OM) is

composed of many heterogenous chemical

compounds. The prevailing chemical element

in OM is carbon (C), and therefore OM is

most easily quantified in terms of units of C.

Hence, OM concentrations may be expressed

as particulate organic carbon (POC) and

dissolved organic carbon (DOC) with units of

μmol CL–1, or alternatively g Cm–3. Rates of

eutrophication are subsequently expressed in

units of C per area (or volume) per unit time. 

Although this working definition of

eutrophication is simple, the actual quantification of rates is a formidable task and the

concentrations of many of the constituents

may be unknown for most ecosystems. In

estuaries, OM occurs in both particulate

(POM; plant debris, detritus, phytoplankton)

and dissolved forms (DOM; humics,

mucopolysaccharides, peptides, lipids). The

distinction between POM and DOM is arbitrary and depends on the methods used to

separate the two fractions (e.g., filtration,

ultrafiltration, centrifugation, dialysis).

Colloids and large proteins, for example, may

be included in either fraction depending on

the separation methodology. Measuring these

components requires specialized analytical

techniques such as C–H–N elemental analysis, dissolved organic carbon analysis, and

ultrafiltration. In addition, the rates of supply

of different forms of OM are highly variable

in natural ecosystems and are strongly influenced by the land use characteristics of the

watershed (rural, agricultural, urban, etc.),

hydrology, and climatology (6–8). 

Sources of OM can come from two major

pathways (1,6). Allochthonous OM originates

outside the estuary and is transported into the

estuary either from watershed runoff and

riverine inflow (usually the primary source) or

from coastal waters through tidal inlets (usually less important). Autochthonous OM is

generated within the system, mostly through

photosynthesis by primary producers or by

benthic regeneration of OM. In estuarine

habitats, the dominant primary producers are

phytoplankton, benthic microalgae, epiphytes,

seagrasses, and other submerged aquatic vegetation. Primary productivity, measured by

Eutrophication is a process that can be defined as an increase in the rate of supply of organic matter

(OM) to an ecosystem. We provide a general overview of the major features driving estuarine

eutrophication and outline some of the consequences of that process. The main chemical

constituent of OM is carbon (C), and therefore rates of eutrophication are expressed in units of C

per area per unit time. OM occurs in both particulate and dissolved forms. Allochthonous OM

originates outside the estuary, whereas autochthonous OM is generated within the system, mostly

by primary producers or by benthic regeneration of OM. The supply rates of limiting nutrients

regulate phytoplankton productivity that contributes to inputs of autochthonous OM. The trophic

status of an estuary is often based on eutrophication rates and can be categorized as oligotrophic

(<100 g C m–2 y–1), mesotrophic (100–300 g C m–2 y–1), eutrophic (300–500 g C m–2 y–1), or

hypertrophic (>500 g C m–2 y–1). Ecosystem responses to eutrophication depend on both export

rates (flushing, microbially mediated losses through respiration, and denitrification) and

recycling/regeneration rates within the estuary. The mitigation of the effects of eutrophication

involves the regulation of inorganic nutrient (primarily N and P) inputs into receiving waters.

Appropriately scaled and parameterized nutrient and hydrologic controls are the only realistic options

for controlling phytoplankton blooms, algal toxicity, and other symptoms of eutrophication in

estuarine ecosystems. Key words: biogeochemistry, coastal, estuary, eutrophic, management,

microbial, phytoplankton, zooplankton. — Environ Health Perspect 109(suppl 5):699–706 (2001).

http://ehpnet1.niehs.nih.gov/docs/2001/suppl-5/699-706pinckney/abstract.html

This article is based on a presentation at the CDC

National Conference on Pfiesteria: From Biology to

Public Health held 18–20 October 2000 in Stone

Mountain, Georgia, USA.

Address correspondence to J.L. Pinckney,

MS 3146, Dept. of Oceanography, Texas A & M

University, College Station, TX 77843-3146 USA.

Telephone: (979) 458-1028. Fax: (979) 845-6331.

E-mail: pinckney@ocean.tamu.edu

We thank S.E. Lumsden and E.B. Örnòlfsdòttir for

their comments on previous drafts of this manuscript.

Funding for portions of this research were provided by

Texas Institute of Oceanography, Texas Parks and

Wildlife Department, National Science Foundation,

North Carolina Sea Grant, and the U.S. Department of

Agriculture. The U.S. government has the right to

retain a nonexclusive, royalty-free license in and to any

copyright covering this article.

Received 8 January 2001; accepted 28 June 2001.

The Role of Nutrient Loading and Eutrophication in Estuarine Ecology

James L. Pinckney,1 Hans W. Paerl,2 Patricia Tester,3 and Tammi L. Richardson1

1Department of Oceanography, Texas A&M University, College Station, Texas, USA; 2Institute of Marine Sciences, University of North

Carolina at Chapel Hill, Morehead City, North Carolina, USA; 3National Ocean Service, Center for Coastal Fisheries and Habitat Research,

Beaufort, North Carolina, USA
Pinckney et al.

700 VOLUME 109 | SUPPLEMENT 5 | October 2001 • Environmental Health Perspectives

either 14CO2 uptake or O2 evolution methods

for determining photosynthetic rates, is fairly

well known for many estuaries. For a few select

estuaries, rates of photosynthesis and estimates

of primary production have been recorded

routinely for several decades and offer useful

databases for assessing long-term trends in

trophic state. Using the known (or estimated)

annual organic C supply (in units of g C m–2

y–1), the trophic status of an estuary can be categorized as oligotrophic (<100), mesotrophic

(100–300), eutrophic (300–500), or hypertrophic (>500) (1). By definition, this annual

estimate should include both allochthonous

and autochthonous OM inputs. However, in

practice, most estuaries are classified solely on

estimates of the autochthonous primary

production of phytoplankton.

DOC and POC

Dissolved organic carbon, the major component in DOM, is one of the largest organic

carbon pools on earth and plays a central role

in the biogeochemistry in estuarine, coastal,

and oceanic environments (7). This broad

class of compounds, although present in high

concentrations in most estuaries, is usually

not considered in most OM budget calculations; nonetheless, it is a large source of OM

for estuaries. In Gulf of Mexico estuaries, for

example, DOC fluxes range from 10 to more

than 300 kilotons C y–1 (Table 2) (8). DOC

is a mixture of a range of sizes of organic

molecules, that are usually characterized by

size or weight. A significant fraction of the

DOC in estuarine waters is composed of colloidal or macromolecular OM (9–11), which

plays an important role in the carbon cycle,

trace metal scavenging (12), and biogeochemical processes (13–15). DOM sources in

coastal systems include production through

phytoplankton exudates (mucopolysaccharides) or phytodetritus (16–18), inefficient

zooplankton feeding (19), fecal pellet decomposition, sedimentary inputs through benthic

exchange (20), urban/agricultural runoff (16),

and sewage inputs. The pool of organic substances is dynamic. Some DOC can be

rapidly recycled by bacteria (21), whereas

other compounds undergo photochemical

degradation into more labile, low molecular

weight DOC that can be used directly by

some phytoplankton (22).

POC suspended in the water column of

estuaries is composed mainly of bacteria, phytoplankton, small zooplankton (ciliates,

rotifers, and other microheterotrophs), fecal

material (feces and pseudofeces), and decaying plant material (detritus) (23–25).

Although larger invertebrates and fish are also

a form of POC, these animals are excluded

either by sample collection or filtration techniques and are usually not considered in estimates of POC. Likewise, surface sediments

contain many of the components listed above

that are delivered by deposition processes.

Phytoplankton POC is a major source of OM

in sediments. Under phytoplankton bloom

conditions, the growth rate of phytoplankton

species and biomass accumulation exceeds the

grazing and export rates. Under these conditions, large amounts of phytoplankton C can

be deposited onto the sediment (24).

Macrophytes, such as the saltmarsh cordgrass

Spartina, seagrasses, and seaweeds can also

contribute large amounts of POC and POM

in the form of detritus (23–25). Sharp chemical gradients and redox conditions in the sediments foster high rates of microbial activity

and biogeochemical cycling (26). Aerobic

decomposition processes consume oxygen at

the sediment–water interface. When oxygen

consumption exceeds the rate of replenishment (by diffusion or mixing processes), the

bottom water may become anoxic. In shallow

estuaries, wind-induced waves or tidal currents may resuspend the deposited POC as

well as the various dissolved organic and inorganic degradation products. Resuspension

events can elevate nutrient concentrations in

the water column, reduce the amount of light

available for photosynthesis, and decrease the

dissolved O2 concentrations due to aerobic

decomposition of the particulate matter (20).

However, strong mixing events can rapidly

oxygenate (within hours) the entire water column and effectively remove all traces of

anoxia (except for mortalities incurred during

the anoxic event).

Causes of Eutrophication

The ultimate cause of estuarine eutrophication

is an increase in organic matter loading. This

loading is accomplished by either (or both) an

increase in the inputs of allochthonous C from

the watershed or by increased phytoplankton

and macrophyte primary production within

the estuary. In the strictest sense of the definition, all estuaries undergo eutrophication periodically (27). Chronic and episodic changes in

hydrodynamics, geomorphology, and climate—including catastrophic events such as

hurricanes, floods, landslides, volcanism, and

earthquakes—are all features that influence

OM input rates. By their very nature, estuarine

ecosystems have evolved over time to buffer

themselves, and in some cases capitalize on,

these pulsing events (27). However, the ecological context of eutrophication involves time

and space scales that transcend ephemeral and

seasonal events. Annual to decadal scales often

are used to assess the long-term changes in

OM loading as well as the ecosystem responses

to increased loading rates.

Although many factors influence estuarine

and coastal primary productivity, the rates of

supply of nutrients are fundamentally important in regulating this process. In estuarine

ecosystems, nitrogen (N) is commonly the

most limiting nutrient for phytoplankton

production (1,28). Biologically available N

can occur in many dissolved organic (DON)

and inorganic (DIN) nitrogen forms.

Allochthonous N that enters from outside the

estuary is commonly referred to as new N

because it represents additional N that can be

used for phytoplankton growth and production. In contrast, autochthonous or regenerated N is nitrogen that is captured and then

recycled in the estuary. 

In N-limited systems, ambient N concentrations govern phytoplankton growth rates

and primary productivity (29). Nutrient

acquisition is a major factor determining the

outcome of competitive interactions and phytoplankton community composition (i.e., the

species that make up the community) (30).

Nutrient uptake and assimilation by phytoplankton varies between species and N source

(31,32). Uptake rates are influenced by the

nutritional state and N-starved individuals

may take up N at higher rates (33,34). In

some species, ammonium inhibits nitrate

and/or urea uptake (35–37). An additional

source of N available to phytoplankton is in

Table 2. Estimates of dissolved organic carbon concentrations and inputs into selected Gulf of Mexico estuaries.a

River/Estuary DOC concentration (μM C) Annual DOC input (× 1010 g C/year)

Nueces/Corpus Christi, TX 560–630 0.5–0.6

San Antonio–Guadalupe, TX 330–480 0.1–0.2

Lavaca Bay, TX 830 3

Colorado, TX 650 2

Galveston Bay, TX 420–480 6.7

Sabine–Neches, TX 530 8.7

Lake Pontchartrain, LA 425–485 2.3

Mobile Bay, AL 424 ± 105 31.7 ± 7

Barataria, LA 558 4–28

Mississippi River, LA 270–330 170–209

aData compiled from Guo et al. (8).

Table 1. Broad characterization of the trophic status of

U.S. estuaries and the relative proportion of the total

estuarine surface area represented by each category.a

Eutrophic Number of Total estuarine

condition estuaries surface area (%)

High 44 40

Moderate 40 25

Low 38 35

Total 122 100

aData compiled from Bricker et al (3).
Estuarine eutrophicaton

the form of low molecular weight DON such

as urea and amino acids (38). 

Nitrogen formulations can also regulate

the cell-size distribution of phytoplankton

communities. The larger species may have the

capacity for more internal storage of nutrients

and become dominant in fluctuating nutrient

regimes (39,40). In general, smaller species

have a higher preference for NH4

+ uptake over

NO3

– than larger phytoplankton species (41).

The implication is that long-term changes in

the sources and concentrations of N compounds in N-limited systems may alter both

the species composition of the phytoplankton

community and the relative size distribution

of phytoplankton cells. Therefore, prolonged

eutrophication can have cascading effects on

the composition of the phytoplankton community, food web relationships, and biogeochemical cycling of impacted ecosystems.

Although N is the primary limiting nutrient in most estuaries, other nutrients may be

important, especially as regulators of phytoplankton community composition. In some

estuaries, phosphate (PO4-

–3

) can also be limiting or colimiting during winter months or

when N concentrations are elevated (42,43).

Occasionally, silicate (Si), which is an essential

nutrient for diatoms, is a limiting nutrient for

diatom growth in some systems (43–45). The

Redfield ratio is used frequently to infer the

potential limiting nutrient. A stoichiometric

molar ratio of 16 N:1 P is accepted as the

“ideal” ratio for phytoplankton growth. Higher

values of the Redfield ratio are interpreted as

indicative of P limitation and, conversely, low

ratios suggest N limitation. However, direct

application of this principle is problematic for

estuarine systems. Measurements of nutrient

concentrations to infer limiting nutrients

neglect the importance of nutrient cycling rates.

Rapid regeneration of nutrients and short

turnover times, which are a known characteristic of estuaries, can mask the “true” nutrient

availability and hence limitation. Furthermore,

measures of water column nutrient concentrations represent what is available for use by phytoplankton. Large amounts of N and other

nutrients may be sequestered and stored in

biomass (i.e., POM) and therefore not considered when budgets are constructed. Nutrient

addition and dilution bioassays, in which natural samples are incubated under various

nutrient concentrations, allow the phytoplankton to function as bioindicators to signal

which nutrient(s) is limiting (46,47). This

experimental approach offers a direct and reliable assessment of the identity and relative

proportions of limiting nutrients.

Nutrient Input Sources

Nutrient input sources can be broadly

grouped into two categories. Point sources

include inputs derived from sewage outfalls,

wastewater treatment plants, industrial wastewater, and stormwater drains. Nonpoint

sources of nutrients are primarily watershed

runoff (riverine inputs), groundwater, and

atmospheric deposition (rainfall and dryfall).

Although the nutrient concentrations from

localized, identifiable point sources are easily

regulated, more diffuse nonpoint sources are

difficult to control. In addition, numerous

surveys have shown nonpoint source loading

to be a dominant source of new N and other

nutrient inputs in many estuarine watersheds,

especially those dominated by agricultural

activities (1,44). Nonpoint nutrient sources

are therefore of major concern for most estuaries. Vegetated buffer zones, located along

the margin of rivers and estuaries, are frequently employed as effective management

practices to control nonpoint source nutrient

inputs derived from watershed runoff.

Estuarine eutrophication is driven generally

by nonpoint sources of nutrients (48). A

regional-scale analysis of fluxes of N from the

North Atlantic coast demonstrated that nonpoint sources of N exceeded sewage inputs (or

point sources) for all regions in both Europe

and North America (49). Overall sewage contributed only 12% of the flux of N from

North America. Non-point sources also dominated P inputs to surface waters of the United

States (50,51). Because of the effort to control

P pollution, nonpoint sources of P have grown

in relative importance since 1980 (52–54). For

the Mississippi River, sewage and industrial

point sources contribute an estimated 10%

(49) to 20% (55) of the total N flux (organic

and inorganic) and 40% of the total P flux

(55). One quarter of the N and P inputs into

the Chesapeake Bay come from wastewater

treatment plants and other point sources

(56,57). For heavily populated watersheds with

relatively small surface areas, wastewater can be

a major source of N to an estuary (58). Even in

some estuaries fed by larger watersheds, wastewater can be the largest source of N if the

watershed is heavily polluted. Wastewater contributes approximately 60% of the N inputs to

the Long Island Sound, New York, largely due

to sewage from New York City (59). 

A rapidly emerging source of nutrients to

estuarine and coastal waters is atmospheric

deposition in the form of rain (wet) and fine

particulates (dry) (60,61). This source provides diverse organic and inorganic chemical

species in both dissolved and particulate

forms. Atmospherically deposited nitrogen

(ADN) is a dynamic mixture of biologically

available dissolved inorganic (NO3

–

, NO2

–

,

NH4

+

; DIN) and organic compounds (amino

acids, organonitrates, urea; DON) (62–65).

The combustion of fossil fuels and emissions

of agricultural and industrial N-containing

compounds into the atmosphere (as gases,

aerosols, and fine particulates) is a highly significant and growing percentage of total N

inputs into estuarine and coastal ecosystems

(6,60,61). Current estimates of the percentage of total (i.e., natural + anthropogenic)

new N inputs attributed to direct ADN range

from 10 to 50%, depending on location and

relative sources (6,61,63). For many estuaries,

ADN is the single most abundant new N

source and warrants scrutiny as a key modulator of eutrophication processes.

Although nutrient concentrations are usually the most common factor limiting primary

production in estuaries, physical factors may

also play a role. In systems with a high concentration of suspended particulates, light penetration may be a principal limiting or colimiting

regulator (66,67) because the rate of photosynthesis is a function of both light quantity (irradiance) and spectral distribution (light quality)

(68). Irradiance levels and light exposure of

phytoplankton are likewise determined by the

vertical mixing rate of the water column (69).

Nutrient Loading

Eutrophication of most estuaries is linked to

the rate at which nutrients (and OM) are added

to the system (Figure 1). Nutrient loading is a

generic term for which the definition varies

depending on the context in which it is used.

In a broad sense, loading is the rate of supply of

Environmental Health Perspectives • VOLUME 109 | SUPPLEMENT 5 | October 2001 701

Phytoplankton

Inputs

 DOM

 POM

 Nutrients

Consumers

Nutrient

regeneration

Nutrient

recycling

Oxygen

dynamics

POM

deposition

Export

 DOM

 POM

 Nutrients

Figure 1. Major estuarine processes related to estuarine eutrophication. DOM and POM refer to dissolved and

particulate organic matter, respectively.
Pinckney et al.

702 VOLUME 109 | SUPPLEMENT 5 | October 2001 • Environmental Health Perspectives

a particular entity to receiving waters; it is

expressed frequently as a rate (e.g., tons N y–1).

Ecosystem responses depend on several critical

physical–chemical characteristics and processes.

Estuary size (surface area), depth, volume,

flushing rate, water residence time, tidal

exchange, vertical mixing, and stratification are

all factors that affect the transport, transformation, retention, and export of nutrients.

Although there are many methods for calculating loading, the simplest approach is to

multiply the nutrient concentration (g N m–3)

by the river discharge rate (m3 s

–1), groundwater flow rate (m3 s

–1), or rainfall amounts (cm

m–2). The accuracy of loading rate estimations

is therefore related to the accuracy of measurements of two variables: flow rate and entity

concentration. Nutrient loading rates, calculated on an annual basis, can be highly variable

from year to year (Table 3). The frequency of

the measurements becomes especially important during episodic events that add large volumes of water to the system (rainstorms,

hurricanes). During these events, nutrient and

OM loading over just a few days can exceed

the total loading for an entire year (Figures 2,

3). Although concentrations of OM or nutrients may be low because of dilution by large

volumes of water, the actual loading rates may

be very high because of the total volume of

water flowing into the estuary. Unfortunately,

during extreme weather events, remote instrumentation (for measuring discharge rates) is

frequently damaged and field collections of

water samples for chemical analyses are postponed because of personnel safety concerns.

Therefore, some of the largest loading events

are frequently missed, or approximated with

low accuracy, during such periods.

The impact of nutrient loading will depend

on how quickly the inputs are transferred

through the estuary. The accumulation of OM

and nutrients in the system can be viewed conceptually as the difference between inputs and

outputs (export) (Figure 1). Nutrients and OM

can be exported from the estuary by tidal flushing or through microbial processes that convert

the combined form into an elemental form

(regeneration). The combined forms of N can

be converted to N2 through coupled nitrification–denitrification (a microbially mediated

process). Ammonium, a primary decomposition product from OM, is first transformed by

microbes to nitrite (NO2

–

) and then nitrate

(NO3

–

) under oxic conditions (nitrification).

Under anoxic conditions, NO3

– and NO2– can

be further transformed to nitrogen gas (N2) by

denitrification. A recent survey of 14 coastal

marine systems showed that denitrification

rates were highly variable but may remove from

3 to 100% of the total dissolved inorganic N

loading (70). Denitrification is one of the primary pathways by which excess N can be

removed from an estuary.

The export rate of OM is also a function

of the residence time of water in the estuary.

The hydraulic residence time of an estuary—

which is the time required to replace the

equivalent amount of freshwater in the estuary by freshwater inputs—varies depending

on many factors, including freshwater input,

circulation, and bathymetry. Mathematically,

this phenomenon is treated as a mass-balance

calculation in which input volume is assumed

to equal output volume. A variety of methods

have been used to calculate residence times

(in units of time) and flushing rates (in units

of volume per unit of time) (71,72). One of

the simplest, but also among the least accurate, is the freshwater fraction method, based

on the ratio of freshwater content of the estuary and the freshwater inflow rate (73).

Another common approach, the tidal prism

method (74), is based on the volume of water

that enters the estuary over a tidal cycle and

the volume of freshwater entering over the

same period. Residence times for Gulf of

Mexico estuaries range from 1 to 350 days,

depending on the estuary (73).

Regeneration and recycling processes

within the estuary are important determinants of system responses and recovery from

increased OM and nutrient loading. The

increase in primary production associated

with eutrophication will increase the concentrations of OM by two main mechanisms:

secretion of mucopolysaccharides by phytoplankton, especially under conditions of

stress, and bacterial degradation of phytoplankton-derived DOM and subsequent

release of DOM and nutrients. DOM (e.g.,

glucose, urea, amino acids) may contain

many of the inorganic nutrients required for

the growth and metabolism of heterotrophic

bacteria in aquatic systems (75–77). The

uptake of NH4

+

, in particular, by heterotrophic bacteria is often significant (78).

The bacterial community may compete with

the phytoplankton for available inorganic

nutrients and, in the presence of mixotrophic

Table 3. Estimated total annual inputs from riverine sources to the Neuse River Estuary, North Carolina

(1994–1998).a

Year

Constituent 1994 1995 1996 1997 1998 Average

Particulate carbon (metric tons C) 2,492 4,137 5,228 1,842 2,564 3,253

Particulate nitrogen (metric tons N) 336 425 571 257 270 372

Nitrate (NO3

–

) (metric tons N) 575 1,515 1,946 1,363 1,940 1,468

Ammonium (metric tons N) 94 159 389 87 133 172

Total DIN (metric tons N) 669 1,674 2,335 1,450 2,073 1,640

Phosphate (PO4

–3) 126 155 169 58 132 128

aData compiled from Paerl et al. (63).

1,200

900

600

300

0

Internal

External

Discharge

10-year mean

Hurricane Fran Hurricane Fran

600

450

300

150

0

Jan Apr Jul Oct Jan Apr Jul Oct Jan Apr Jul Oct

Month (1994–1996)

Discharge (m3/sec) OM loading (metric tons C/day)

Figure 2. Discharge and organic matter loadings for the Neuse River Estuary, North Carolina (1994–1996). Internal:

autochthonous OM loading (primarily phytoplankton productivity); external: allochthonous sources of OM (riverine

inputs). Note the large increase in external OM loading and concurrent high discharge after Hurricane Fran. Following

that event, the major source of OM loading switches from internal sources to external sources. Graph redrawn from

Paerl et al. (63).
Estuarine eutrophicaton

algae and/or heterotrophic flagellates (79), for

the available DOM. 

Another important process in nutrient

regeneration and recycling involves trophic

pathways. In contrast to the “bottom-up”

regulation of phytoplankton biomass and

productivity by nutrients and physical features described above, herbivores and omnivores can exert “top-down” controls on

phytoplankton and other primary producers.

Inorganic N (mostly as NH4

+

) is a product of

grazing by microzooplankton such as heterotrophic flagellates (nanoflagellates and ciliates) (80,81) as well as by larger zooplankton

and benthic invertebrates. Grazing of bacteria

by protozoa may be responsible for a significant fraction of the regeneration of NH4

+ in

marine systems (78). Goldman et al. (82)

estimate that the amount of N regenerated by

heterotrophic grazing of bacteria (i.e., the

microbial loop) is about 10–50% of the bacterial N ingested. Therefore the microbial

loop can play an integral role in mediating

ecosystem responses to eutrophication.

Zooplankton grazers and benthic filter

feeders may consume significant amounts of

phytoplankton biomass and primary productivity in some estuaries. For example, Li and

Smayda (83) reported that, over long time

periods (1973–1990), the phytoplankton

biomass in Narragansett Bay, Rhode Island,

was controlled by zooplankton grazing.

Similarly, Lewitus et al. (84) found that

microzooplankton grazing during the summer

months in the North Inlet Estuary, South

Carolina, was an important regulator of phytoplankton biomass. In Fourleague Bay,

Louisiana, Dagg (85) estimated zooplankton

grazing rates that were nearly equal to the

phytoplankton standing stock in this estuary.

Therefore, the standing stock of phytoplankton in the estuary can be conceptualized as the

product of a dynamic balance between bottom-up and top-down control mechanisms.

Consequences of

Eutrophication

The cascading effects of anthropogenic new N

inputs on estuarine surface waters are conceptualized and summarized in Figure 4.

Nitrogen inputs, as well as P, Fe, Si, and other

potential rate-limiting nutrients, stimulate

phytoplankton growth and modulate community composition. Although primary consumers recycle a portion of phytoplankton C

and N, bloom events triggered by new N

loading may result in the delivery (deposition)

of phytoplankton C to the benthos. Microbial

degradation promotes bottom water anoxia,

which in turn facilitates sediment nutrient

release. Sediment nutrient release events may

further stimulate algal blooms, leading to a

positive feedback loop that perpetuates

blooms and eutrophication. This scenario is

particularly relevant in estuaries that have relatively long residence times (months to years),

where feedback reactions can be important for

sustaining phytoplankton blooms throughout

the optimal growth season in the spring and

summer. Relatively long residence time and

nutrient-sensitive estuaries such as the

Chesapeake Bay in Virginia and Maryland,

the Pamlico Sound in North Carolina, Florida

Bay, and Texas lagoonal systems are susceptible to the detrimental effects of high OM

loading and nutrient input. The continued

addition of new N and OM, especially over

periods of years to decades, sustains and intensifies the eutrophication cycle in estuaries. As

long as OM and nutrient inputs exceed losses,

the estuary will experience eutrophication.

In some estuaries, phytoplankton C loading, especially following blooms, can produce

large-scale bottom water hypoxia (<2 mg O2

L–1) and anoxia (no detectable dissolved O2).

This phenomenon has been well documented

in many estuaries including the Neuse River in

North Carolina (63,86) and Chesapeake Bay

(87,88). Bottom-water hypoxia/anoxia results

from the interaction of excessive C loading

with several nonbiological variables, including

freshwater discharge, vertical stratification,

establishment of a salt wedge, and meteorological conditions (primarily wind) (74,89).

Environmental Health Perspectives • VOLUME 109 | SUPPLEMENT 5 | October 2001 703

Figure 3. River discharge and dissolved inorganic nitrogen (DIN; nitrite, nitrate, ammonium) for the Neuse River

Estuary (1994–1998). Note that loading rates exhibit a regular seasonal pattern with large DIN loadings that occur

during high river discharge. Graph redrawn from Pearl et al. (63).

600

400

200

0

Discharge (m3/sec) DIN loading (metric tons N/day)

25

20

15

10

5

0

Jan

Mar

May

Jul

Sep

Nov

Jan

Mar

May

Jul

Sep

Nov

Jan

Mar

May

Jul

Sep

Nov

Jan

Mar

May

Jul

Sep

Nov

Jan

Mar

1994 1995 1996 1997 1998

Month

10-year mean

Daily discharge

Figure 4. Relationships among nutrient inputs, phytoplankton responses, and oxygen dynamics.

Nutrient inputs

Phytoplankton

community

Form of limiting nutrient

(NO3

–

, NH4

+

, DON)

Nutrient ratios

Grazed phytoplankton species

Water column

carbon recycling

Decreased

O2 depletion

potentials

Oxic conditions

Hypoxia

Anoxia

Physical controls

-Mixing

-Stratification

Increased

O2 depletion

potentials

Carbon deposition

(POC)

Nuisance/toxic

phytoplankton species

-Dinoflagellates

-Cyanobacteria

Nutrient regeneration

Decomposition of POM
Pinckney et al.

704 VOLUME 109 | SUPPLEMENT 5 | October 2001 • Environmental Health Perspectives

Hypoxic and anoxic conditions can persist for

weeks and cover large areas in estuaries that

have relatively long water residence (weeks to

months), low flushing rates, and persistent vertical stratification (63). High NH4

+ concentrations usually arise from anoxic conditions (90).

The anoxic conditions inhibit nitrification and

the NH4

+ released from the sediments cannot

be nitrified to NO3

–

. Benthic release of PO4

–3 is

also enhanced under low dissolved oxygen

concentrations (90). Thus, persistent anoxic

conditions may result in elevated ammonium

concentrations which are used by phytoplankton that capitalize on the readily available

NH4

+

. If these conditions continue, phytoplankton blooms may develop. The tight coupling between anoxia and phytoplankton

blooms makes it difficult to assign cause and

effect relationships without careful consideration of the hydrological, chemical, and biological conditions initiating and controlling these

events. Assessments of conditions during and

following bloom events often provide little

insight into causative factors (91,92).

Phytoplankton C loading can have a direct

impact on the oxygen dynamics of estuaries.

However, nutrient and OM inputs can also

play a role in determining the predominant

species within phytoplankton communities.

The alteration of nutrient concentrations and

ratios may change the phytoplankton community composition significantly, with subsequent cascading impacts at higher trophic

levels of the food web (Figure 4) (93). In estuarine and coastal waters, phytoplankton are

exposed to a range of N compounds in varying ratios and supply rates, depending on

location relative to N inputs. Previous work

has shown that phytoplankton species exhibit

different growth responses to N sources (94).

In particular, nitrate (NO3

–

) and ammonium

(NH4

+

) uptake rates vary spatially and seasonally, suggesting differential community

responses to the N sources (95). These

species-specific responses may play an important role in structuring natural phytoplankton

communities (41) and subsequent biogeochemical alterations (e.g., C and N cycles). 

Phytoplankton-mediated assimilation and

fate of inorganic nutrients and organic matter

(both POM and DOM) are significant components of nutrient transformations in estuarine systems (90,96,97). Microalgal species

composition and abundance are intimately

linked to higher trophic levels through preferential grazing by herbivores (zooplankton,

suspension feeders, deposit feeders) (98–101).

The implication for food webs is that some

phytoplankton blooms may be ungrazed and

deposited in surficial sediments, providing

labile carbon substrates for benthic respiration, hypoxia, and anoxia following large

blooms. Changes in the phytoplankton

species that support primary consumers and

predators may result in the loss of available

food for these consumers. The loss of the type

(phytoplankton species) and abundance of

food resources could result in the displacement of motile species such as fish and crabs

to other, more suitable habitats. Species that

cannot search for food in other areas (e.g.,

oysters, clams, benthic infauna, zooplankton,

etc.) may suffer reduced reproductive fitness

and significant mortalities due to starvation.

Many important commercial and recreational

fisheries are among groups that may be negatively impacted (1,2,48).

The development of harmful algal blooms

(HABs) is a possible consequence of eutrophication (2). Limiting nutrients, such as N, P,

and Si, foster the growth of most photosynthetic phytoplankton species, and therefore

high nutrient concentrations may produce

algal blooms (28,42,45,48). Some of these

blooms may be composed of toxic species (2).

Perhaps a more interesting aspect of eutrophication is that DOM, in addition to dissolved

inorganic nutrients, may foster the growth of

HAB species. Several studies have shown that

OM can stimulate the growth of marine

dinoflagellates (81,102–104). Inputs of terrestrially derived DOM, particularly humic substances, have been implicated as important

factors in blooms of Gymnodinium breve off

Florida (105) and Alexandrium tamarense in

the St. Lawrence Estuary in Canada (106).

Blooms of toxin-producing Gymnodinium

catenatum in southeast Tasmanian waters

appear to be preceded by a rainfall “trigger”

(107) with an associated influx of DOM from

land runoff. Doblin et al. (108) showed that

both the growth rate and biomass of G. catenatum increased when cultures were supplemented with a combination of humic acid and

naturally collected DOM. They concluded the

effect was manifested as a result of an indirect

effect of DOM on micronutrient availability

(Se) and/or N and P. Rates of uptake of DON

compounds by the brown tide organism

Aureococcus anophagefferens off Long Island

exceed rates of NH4

+ uptake (109). Blooms of a

similar species (Aureoumbra lagunensis) are

prevalent in some Texas estuaries (110).

Dinoflagellates (G. breve and G. sanguineum)

are among the major HAB formers in the Gulf

of Mexico (111). The heterotrophic dinoflagellate Pfiesteria piscicida Steidinger &

Burkholder and related Pfiesteria-like dinoflagellates reportedly use organic compounds

(112,113), but the primary mode of nutrition

for the nontoxic flagellated zoospore stage

seems to be other phytoplankton species

(113,114). Even if shifts in phytoplankton

community composition do not result in

HABs, they can still have cascading effects that

ultimately result in major changes in the

trophodynamics and food webs in essential

nursery habitats. Economically important

resources such as fisheries, recreation, and

tourism could be impacted negatively and

irreparably by these ecosystem alterations

(1,2,48). 

Another major group of toxic/nuisance

phytoplankton that occurs in estuaries is

cyanobacteria. Growing frequencies and geographic expansion of toxic cyanobacterial

blooms appears to be a worldwide phenomenon (4,115). The combined effect of

increased water residence time and growing

pollutant inputs has led to longer exposure

periods of opportunistic microorganisms to

growth-promoting conditions. Bloomforming cyanobacteria are often superior

competitors under these conditions and frequently thrive in nutrient-enriched waters

(116). For non–N2-fixing genera such as

Microcystis, Oscillatoria, or Lyngbya, both N

and P are important. Phosphate alone can

stimulate the growth of N2-fixing genera such

as Anabaeana, Anabaenopsis, Aphanizomenon,

Cylindrospermopsis) (117–119). Riverine and

estuarine waters in the southeastern United

States tend to exhibit nutrient-enriched,

N-limited (low N:P ratios) conditions, favoring cyanobacterial (especially N2-fixing genera)

dominance (120). Many cyanobacterial bloom

formers also prefer OM-enriched conditions

(117), including OM originating from urban

wastewater, agricultural runoff, forest humics,

and soil erosion products (118). Substantial

individuality exists among cyanobacterial nuisance taxa in terms of their abilities to exploit

these environmental variables.

Eutrophication Management

Although OM is the currency used to measure

eutrophication, the mitigation of the effects of

eutrophication involves the regulation of inorganic nutrient (primarily N and P) inputs into

receiving waters. These nutrients may be

delivered to the estuary as dissolved inorganic

ions or bound to OM and later released as the

OM is decomposed by microbial processes.

Without question, excessive nutrient inputs

lead to estuarine eutrophication. There is little

doubt that increases in supply rates and

changes in chemical forms and ratios of

growth limiting nutrients play roles in

regional and global expansion of blooms.

While nutrient enrichment can be invoked for

at least partially driving HAB expansion,

extrapolating specific local examples to other

estuaries, as well as regional coastal scales is

often confounded by the complicated interplay of physical, chemical, and biotic controls

on bloom dynamics. Invariably, even nearby

estuaries do not exhibit identical physical (residence time, stratification, advection), chemical (specific nutrient supply and cycling rates),

and biological (community composition,

microbial associations, grazing) characteristics.

Therefore ecosystem sensitivity and response
Estuarine eutrophicaton

to enhanced nutrient loading may manifest

itself quite differently in such systems, often

leading to contrasting trophic states, degrees

of ecosystem resiliency to change, desirability

of habitats, and resources.

From a water quality management perspective, understanding the mechanistic linkages in space and time between man-made

alterations of hydrologic and nutrient regimes

in the watershed is of critical importance for

identifying growth- and bloom-regulating

factors that may be used for developing shortand long-term control strategies to slow

eutrophication (121,122). The first step

involves the identification of relevant phytoplankton taxa, characterizing environmental

controls on their growth and toxin-producing

characteristics, and incorporating this information in hydrodynamic and nutrient loading budgets depicting ecosystem physical

chemical variability and extremes (88,89,

123,124). When these features are known, we

can begin to develop realistic and effective

management alternatives aimed at controlling

algal bloom development and expansion.

Arguably, appropriately scaled and parameterized nutrient and hydrologic controls are

the only realistic options for controlling

bloom, toxicity problems, and eutrophication

in estuarine ecosystems.

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