Eurtrophication Review Articles 

Eutrophication of Freshwater and Coastal Marine Ecosystems 

A Global Problem 

Val H. Smith 

Correspondence address: Dr. Val H. Smith, Department of Ecology and Evolutionary Biology, 1200 Sunnyside Avenue, 

University of Kansas, Lawrence, KS 66045, USA (vsmith@ku.edu) 

DOI: htto://dx.doi.or~10.1065/esor2002.12.142 

Abstract 

Goal, Scope and Background. Humans now strongly influence almost every major aquatic ecosystem, and their activities have dramatically altered the fluxes of growth-limiting nutrients from the 

landscape to receiving waters. Unfortunately, these nutrient inputs 

have had profound negative effects upon the quality of surface waters worldwide. This review examines how eutrophication influences the biomass and species composition of algae in both fleshwater and costal marine systems. 

Main Features. An overview of recent advances in algae-related 

eutrophication research is presented. In freshwater systems, a summary is presented for lakes and reservoirs; streams and rivers; and 

wetlands. A brief summary is also presented for estuarine and coastal 

marine ecosystems. 

Results. Eutrophication causes predictable increases in the biomass 

of algae in lakes and reservoirs; streams and rivers; wetlands; and 

coastal marine ecosystems. As in lakes, the response of suspended 

algae in large rivers to changes in nutrient loading may be hysteretic 

in some cases. The inhibitory effects of high concentrations of inorganic suspended solids on algal growth, which can be very evident in 

many reservoirs receiving high inputs of suspended soils, also potentially may occur in turbid rivers. Consistent and predictable 

eutrophication-caused increases in cyanobacterial dominance of 

phytoplankton have been reported worldwide for natural lakes, and 

similar trends are reported here both for phytoplankton in turbid 

reservoirs, and for suspended algae in a large river. 

Conclusions. A remarkable unity is evident in the global response 

of algal biomass to nitrogen and phosphorus availability in lakes 

and reservoirs; wetlands; streams and rivers; and coastal marine 

waters. The species composition of algal communities inhabiting 

the water column appears to respond similarly to nutrient loading, 

whether in lakes, reservoirs, or rivers. As is true of freshwater ecosystems, the recent literature suggests that coastal marine ecosystems will respond positively to nutrient loading control efforts. 

Recommendations and Outlook. Our understanding of freshwater 

eutrophication and its effects on algal-related water quality is strong 

and is advancing rapidly. However, our understanding of the effects 

of eutrophication on estuarine and coastal marine ecosystems is much 

more limited, and this gap represents an important future research 

need. Although coastal systems can be hydrologically complex, the 

biomass of marine phytoplankton nonetheless appears to respond 

sensitively and predictably to changes in the external supplies of nitrogen and phosphorus. These responses suggest that efforts to manage nutrient inputs to the seas will result in significant improvements 

in coastal zone water quality. Additional new efforts should be made 

to develop models that quantitatively link ecosystem-level responses 

to nutrient loading in both freshwater and marine systems. 

Keywords: Algae; coastal marine ecosystems; estuaries; eutrophication; 

harmful algal blooms; nitrogen; nutrients; lakes; reservoirs; rivers; phosphorus; phytoplankton; shallow lakes; wetlands 

Introduction 

Humans now strongly influence almost every major aquatic 

ecosystem, and their activities have dramatically altered the 

fluxes of growth-limiting nutrients from the landscape to 

receiving waters. One primary cause has been the rapid 

intensification of agriculture (Matson et al. 1997). The global production of agricultural fertilizers alone released <10 

million metric tonnes of nitrogen in 1950, but may exceed 

135 million metric tonnes of N by the year 2030 (Vitousek 

et al. 1997). Substantial N is also applied to croplands in 

the form of animal manures, for which regulatory standards are generally far less stringent than those applied to 

human sewage (Carpenter et al. 1998a). Similarly, a small 

but ecologically very significant portion of agriculturallyapplied phosphorus is exported from the land surface to receiving waters (Bennett et al. 2001, Kauppi et al. 1993, Rekolainen et al. 1995, Sims et al. 1998). Atmospheric loading of 

P to surface waters has also increased (Brunner and Bachofen 

1998). In addition, humans use flowing waters as convenient wastewater disposal systems, and the loading of N and 

P to the world's surface waters is very strongly influenced 

by human population density and land use (Cole et al. 1993, 

Caraco 1995). 

Unfortunately, these nutrient inputs can have profound effects upon the quality of receiving waters (Carpenter et al. 

1998a, Correll 1998). Just as fertilizing an agricultural crop 

or forest can result in enhanced plant growth, the most common effect of increased N and P supplies to aquatic ecosystems is an increase in the abundance of algae and aquatic 

plants~ However, the environmental consequences of excessive nutrient enrichment are more serious and far-reaching 

than nuisance increases in plant growth alone. The degradation of water resources by eutrophication can result in 

loss of the amenities or services that these aquatic resources 

provide (Postel and Carpenter 1997). In addition, the 

eutrophication of both freshwater and coastal marine systems causes a wide array of undesirable symptoms that are 

either directly or indirectly related to the nuisance growth 

of aquatic plants (Table 1). 

Eutrophication accounts for almost one half of the impaired 

lake area and 60% of impaired river reaches within the United 

States alone (U.S. EPA 1996a), and eutrophication-related 

water quality impairment can have very substantial negative 

economic effects (Carpenter et al. 1998b, Corrales and Maclean 

1995). Eutrophic drinking water supply reservoirs, for example, are much more likely to have higher treatment costs; greater 

126 ESPR - Environ Sci & Pollut Res 10 (2) 126 - 139 (2003) 

9 ecomed publishers, D-86899 Landsberg, Germany and Ft. Worth/TX, USA ~ Tokyo, Japan 9 Mumbai, India ~ Seoul, Korea 
Review Articles Eutrophication 

Table 1 : Adverse effects of freshwater and coastal marine eutrophication 

9 increased productivity and biomass of phytoplankton and suspended algae 

9 shifts in phytoplankton composi~on to bloom-forming species, many of which may be toxic, or which may not be consumed effectively by aquatic grazers 

9 increased productivity, biomass, and species composition of attached microalgae (periphyton) 

9 increased productivity, biomass, and species composition of marine macroalgae 

9 changes in productivity, biomass, and species composition of aquatic vascular plants 

9 reduced yields of desirable finfish and shellfish species 

9 reductions in the health and size of marine coral populations 

9 threats to endangered aquatic species 

9 decreases in water column transparency 

9 taste, odor, and filtration problems in drinking water supplies 

9 depletion of deep water oxygen 

9 decreases in the perceived aesthetic value of the water body 

9 negative economic impacts, including decreased property values and reduced recreational uses 

difficulties in meeting standards for disinfection by-products; 

consumer complaints due to objectionable taste and odor; and 

health hazards due to algal toxins in the finished drinking water 

(Cooke and Kennedy 2001). In coastal marine habitats, harmful algal blooms may render shellfish and fin fish toxic, or cause 

massive fish and shrimp kills; by mid-1994, there were 3,164 

reported incidents of human poisoning and 148 deaths in the 

Asia-Pacific region alone (Corrales and Maclean 1995). Economic losses may exceed $1 million U.S. per event, and monitoring efforts may cost up to $50,000 U.S. for each affected 

area (Corrales and Maclean 1995). 

Because of these problems, there have concerted efforts for 

several decades by aquatic ecologists and civil engineers 

worldwide to develop a rational predictive framework to 

prevent and manage freshwater eutrophication (OECD 1982, 

Cooke et al. 1993). The U.S. EPA has recently published a 

series of nutrient criteria technical guidance manuals for lakes 

and reservoirs, rivers and streams, and marine waters (U.S. 

EPA 2002), and parallel attempts are being made worldwide to evaluate and to predict the responses of aquatic ecosystems to changes in nutrient loading. 

As a result of global efforts to solve the eutrophication problem, we now have a firm mechanistic understanding of the 

relative roles of N and P as growth-limiting nutrients; we have 

developed models that can predict nutrient concentrations in 

receiving waters from watershed nutrient inputs and hydrology; and we have developed a series of models that then relate 

waterbody nutrient concentrations to the critical aspects of 

water quality that are perceived by the general public as being 

important and worthy of preservation. In general, an excessive accumulation of algal biomass is the symptom of 

eutrophication that is most easily perceived and most disliked 

by the public. A remarkable unity in the response of algal 

production to eutrophication has in fact been found in most 

aquatic ecosystems that have been studied, and a demonstrated 

success in reducing algal biomass to publicly acceptable levels 

has been the feature held in common by all successful 

eutrophication control efforts (Smith 1998). 

The literature on eutrophication is expanding rapidly, and 

cannot be reviewed comprehensively in this paper. A large 

number of historical and technical reviews have appeared 

during the past several years (e.g., Carlson and Simpson 

1998, Correll 1998, Smith 1998, Smith et al. 19.99, Swedish 

EPA 2000, U.S. EPA 2002), and this information wilt not be 

repeated here. The purpose of this mini-review is to provide 

a brief update of our knowledge of the effects of eutrophication on algal-related water quality in both freshwater 

and coastal marine ecosystems, and to suggest key areas that 

may be fruitful for future research efforts. 

1 Eutrophication of Freshwater Ecosystems 

1.1 Lakes and reservoirs 

Effects of eutrophication on algal biomass. Because phosphorus has been identified as a key growth-limiting nutrient 

for algae in lakes and reservoirs (Schindler 1977, OECD 

1982), phosphorus-based eutrophication classification systems and management schemes have been developed worldwide. An example of the Swedish eutrophication classification system for lakes is shown in Table 2. 

Table 2" Swedish eutrophication classification system for freshwater lakes according to the growing season biovolume of planktonic algae (modified from 

Will6n 2000 

Class: ...... 

mm perL ;mm per 

la Particularly small biomass <0.1 _<0.1 

1 b Very small biomass 0.1-0.5 0.1-0.5 Oligotrophic 

2 Small biomass 0.5-1.5 0.5-2 Mesotrophic 

3 Moderately large biomass 1.5-2.5 2-4 Eutrophy I 

4 Large biomass 2.5-5 4-8 Eutrophy II 

5 Very large biomass >5 >8 Hypertrophy 

Ultraoligotrophic 

ESPR - Environ Sci & Pollut Res 10 (2) 2003 127 
Eutrophication Review Articles 

In general, the growing season average biomass of algae in 

lakes is strongly dependent upon the concentration of total 

phosphorus (TP, lag/L) in the water. However, the relative 

availabilities of nitrogen and phosphorus change consistently 

with cultural eutrophication (Downing and McCauley 1992), 

and the response of algal biomass to TP is strongly modified 

by the total nitrogen to total phosphorus (TN:TP) ratio (cf. 

Forsberg and Ryding 1980, Smith 1982, Canfield 1983, Prairie et al. 1989, McCauley et al. 1989, Faafeng and Hessen 

1993, Seip 1994). 

The global robustness of algal biomass-nutrient relationships 

has been well demonstrated in many cross-sectional analyses of lakes and reservoirs (Sakamoto 1966, Lund 1970, 

Dillon and Rigler 1974, Jones and Bachmann 1976, Ahl and 

Weiderholm 1977, Nicholls and Dillon 1978, Canfield and 

Bachmann 1981, Chow-Fraser et al. 1994, Seip and Ibrekk 

1988, Smith 1990a, Jones et al. 1998, Niirnberg 1996, Mazumder and Havens 1998, Portielje and Van der Molen 1999, 

Seip et al. 1992, 2000, Brown et al. 2000). Information on 

the responses of lake and reservoir water quality to nutrient 

enrichment is now available from most of the geographical 

regions of the world (e.g., Africa: Thornton 1980, Harper 

et al. 1993, Hecky 1993, Kebede et al. 1994, van Ginkel et 

al. 2000, Varis and Fraboulet-Jussila 2002; Australia: Ferris 

and Tyler 1985, Davis 1997; Central and South America: 

Quir6s 1991,1998; Salas and Martino 1991, Jones et al. 

1993; China: Chang 2002, Dickman et al. 2001; Europe: 

Hrb~i&k and Popovsk)~ 1978; Vyhnfilek et al. 1994; Kop~i&k 

et al. 1996, Istvfinovics and Somly6dy 1999; Greece: 

Danielidis et al. 1996; Ireland and the U.K.: Harper and 

Stewart 1987, Traille 1991, Sutcliffe and Jones 1992; Moss 

et al. 1996, Wilson 1998; India: Bhade et al. 2001, Kaul 

1977; Israel: Gophen et al. 1999; Italy: de Bernardi et al. 

1996, Heinonen et al. 2000; Japan: Seip and Satoh 1984, 

Fukushima and Muraoka 1988; Korea: An and Jones 2000, 

2002; Nepal: Aizaki et al. 1987, Jones et al. 1989; New 

Zealand: Pridmore et al. 1985, McBride and Pridmore 1988, 

Duggan et al. 2001; Oman: Victor et al. 2002; Scandinavia: 

Will~n 2000, Rask et al. 2002; Sri Lanka: Piyasiri 2001; 

Thailand: Jones et al. 2000; and Turkey: Tanik et al. 1999). 

In general, comparative studies of freshwater eutrophication 

strongly suggest that efforts to control external nutrient loading to many lakes will tend to achieve similar reductions in 

their average algal biomass, regardless of geographical location (Seip et al. 2000; but for exceptions, see Sas 1989, 

Moss et al. 1996, Carpenter et al. 1999). However, the growing season or yearly average biomass of algae per se is probably not consciously measured or used by the users of a water 

body as their primary index of water quality impairment: 

the maximal concentrations of algal biomass that occur during transient blooms are much more easily detected, and 

even more strongly disliked by the public. Jones et al. (1979) 

reported a linear arithmetic relationship between mean algal biomass (estimated as chlorophyll a) and maximum algal biomass in U.S. lakes: 

Chlm~ . = 1.7 (Chlmoa,) + 0.2, r 2 = 0.58 (1) 

Other investigators have reported a much steeper arithmetic slope (2.6 rather than 1.7; see OECD 1982); however, 

the variance in the relationship of Jones et al. (1979: their 

Fig. 1) increases strongly with the value of Chl ..... implying 

that empirical models based on log-transformed data should 

be stronger and statistically more appropriate. Stadelmann 

et al. (2001) have examined relationships between mean and 

maximum chlorophyll a during the growing season in Minnesota lakes, and they indeed found a very strong log-linear 

relationship (note that the dependent and independent variables are reversed in their analysis), 

Ioglo (Chl .... ) = 0.91 Ioglo (Chlma,) -0.20, r 2 = 0.91 (2) 

Similarly, Will~n (2000) reported an empirical relationship between late-summer peak biomass and growing-season mean 

algal biomass (based upon algal biovolume rather than Chl a), 

Total Volume~gus t = 1.65 (Total VolumeMay.Octo~, ) -0.26, r 2 = 0.75 (3) 

In addition, she reported a strong relationship between annual peak algal biomass and growing-season mean concentrations of total phosphorus, 

log10 (Annual Peak Total Volume) = 1.512 Ioglo (TPMarch.October) 

-1.924, r 2 = 0.76 (4) 

Phosphorus-based models can also be used to quantify the 

response of nuisance algal bloom frequency to reductions in 

nutrient loading. For example, Havens and Walker (2002) 

have developed predictive models relating water column TP 

concentrations to the frequency algal blooms in Lake Okeechobee, Florida (USA), which they used to establish in-lake 

total phosphorus goals that will be required to reduce water quality impairment in this multiple-use lake. Similarly, 

Lathrop et al. (1998) have developed models to predict the 

phosphorus loading reductions needed to reduce the frequency of blue-green algal blooms in Lake Mendota, Wisconsin (USA). 

Although both the average and the peak biomass of algae in 

lakes often respond very sensitively to changes in nutrient availability, several critical site-specific factors can alter the responses 

of algal growth to external N and P loading. For example, 

variations in depth can potentially modify algal growth 

(Petersen et al. 1997), and deep lakes produce a lower algal 

biomass per unit total phosphorus than shallow lakes (Smith 

1990a). For example, the May-October mean biovolume of 

algae in shallow, wind-mixed lakes (mean depth _<3 meters) is 

typically twice as great as the values observed in thermally 

stratified, deep lakes of similar fertility in Sweden (Will~n 

2000). Similarly, Mazumder (1994a) noted higher levels of 

algal biomass (measured as chlorophyll a) in shallow, mixed 

lakes than in stratified lakes. In addition to these cross-system 

differences, lakes with complex physical structure can exhibit 

remarkable spatial variations in water chemistry and algal 

biomass (cf. Phlips et al. 1993). For example, the within-lake 

variation observed in Table Rock, Missouri (USA) is so large 

that it almost matches the magnitude of lake-to-lake variation 

within the entire state (Jones et al. 1997)! 

128 ESPR - Environ Sci & Pollut Res 10 (2) 2003 
Review Articles Eutrophication 

In addition to the above effects of morphometry, the hydraulic residence time of a waterbody potentially can influence the biomass of algae produced by nutrients if the water 

residence time is shorter than the doubling times (ca. 0.5- 

1.5 days) of phytoplankton algae (Reynolds 1997). Significant effects of hydraulic flushing on phytoplankton biomass 

have indeed been observed in the very small fraction of 

waterbodies that have exceptionally short water residence 

times (McBride and Pridmore 1988, Smith 1990a). However, in environments which experience strong seasonality 

in rainfall and hydrology such as southeastern Asia, dramatic changes in water quality also can occur as a result of 

periodic monsoons (Jones et al. 1997). Other physical factors can modify the biological expressions of eutrophication 

as well. For example, high rates of soil erosion in the watershed can result in elevated inorganic turbidity levels that 

can strongly attenuate underwater light and potentially reduce the algal biomass produced at a given concentration of 

total phosphorus (Canfield and Bachmann 1981, Hoyer and 

Jones 1982, Ferris and Tyler 1985, Carlson 1991, Jones et 

al. 1998, Dokulil 1994, Jones et al. 1997). 

Biotic factors can also be extremely important. In shallow 

lakes that have the potential to develop large populations of 

aquatic vascular plants (macrophytes), the presence or absence of these plants can strongly modify the biomass of 

algae produced at a given nutrient supply level (Canfield et 

al. 1983). For example, Portielje and Van der Molen (1999) 

reported significantly lower chlorophyll a concentrations in 

shallow Dutch lakes and ponds with submersed macrophytes 

coverage exceeding 5% of the total surface area. Shallow 

lakes inherently have the capacity to develop large 

macrophyte populations, and also may exhibit strong internal phosphorus loading due to regeneration from the lake 

sediments. Such lakes may not respond to nutrient loading 

management in the manner predicted by simple eutrophication models (Moss et al. 1996, Carpenter et al. 1999). In 

some cases, shallow eutrophic lakes can alternate between 

two equilibrium points, in which one stable state is dominated by macrophytes, and the alternate stable state is dominated by optically thick phytoplankton communities 

(Scheffer et al. 1993, Scheffer 1998). 

The effects of food web structure on algal biomass in lakes 

have also been extremely well characterized during the past 

25 years (Shapiro et al. 1975, Hrbfieek et al. 1978, Carpenter and Kitchell 1993, Mazumder 1994a,b, Carpenter et al. 

1995, Lacroix et al. 1996, Carpenter 2002). In particular, 

lake-to-lake differences in the abundance and structure of 

fish communities can profoundly influence algal biomass by 

modifying the grazing pressure of zooplankton on phytoplankton. Large cladocerans such as Dapbia pulex and Dapbnia pulicaria typically become abundant in lakes having few 

(or no) planktivorous fish, and these large-bodied herbivores are extremely effective grazers. Lakes dominated by 

planktivorous (zooplankton-eating) fish thus produce a significantly higher algal biomass per unit TP than lakes 

dominated by piscivorous (fish-eating) fish, because the intensity of zooplankton grazing on algae is typically lowest 

in planktivore-dominated systems (Carpenter and Kitchell 

1993, Quir6s 1991,1998). However, the effects of food web 

structure on algal biomass can also be modified by the physical structure of the environment (e.g., lake depth: Proulx et 

al. 1996, Bertolo et al. 2000). 

Such trophic interactions also offer the potential for using 

food web manipulations (biomanipulation: Shapiro et al. 

1975) as a means to reduce nuisance phytoplankton biomass 

in nutrient-enriched lakes. The record of biomanipulation 

as a management tool often has been positive, but the actual 

responses to deliberate fish manipulations have been highly 

variable (Gulati et al. 1990, Cottingham and Knight 1995, 

Sarvala et al. 2000, Sendergaard et al. 2000). 

Effects of eutrophication on algal species composition. The 

biomass of phytoplankton is frequently, but not always, the 

primary focus of attention in lake and reservoir management. Algal species composition in many cases may be a 

more important facet of water quality for the users and consumers of freshwater. Bloom-forming blue-green algae 

(cyanobacteria) typically become dominant over other algal 

species in the phytoplankton of eutrophic lakes (Schindler 

1977, Reynolds 1997, Smith 1990a, 2001). These cyanobacteria can form objectionable surface scums (Klemer and 

Konopka 1989, Paerl 1988); can cause summer fishkills 

(Barica 1978, Lathrop et al. 1998, Kann and Smith 1999); 

and can impair the quality of drinking water (Cooke and 

Kennedy 2001, Smith et al. 2003). 

In addition, many cyanobacteria can produce highly hepatotoxic, cytotoxic, and neurotoxic compounds that are hazardous to humans, cattle, domestic animals, and aquatic 

consumers (Fulton and Paerl 1987, Hawser et al. 1992, Carmichael 1997, Chorus and Bartram 1999, Sivonen 2000, 

van Ginkel et al. 2000, Chorus 2001). Mass occurrences of 

toxic cyanobacteria are a global phenomenon, although the 

timing and duration of the bloom-forming season varies with 

geographic location (Sivonen 2000). In Europe alone, surveys using toxicity bioassays or high performance liquid 

chromatography have revealed an unexpectedly high frequency of toxic blooms (Fig. 1). 

The general patterns of response by toxic cyanobacteria to 

eutrophication are nicely summarized in a recent comparative study by Will~n et al. (2000). Both the intensity of late 

100 

E 

-g 

I1) 

13. 

90 

8O 

70 

60 

5O 

40 

3O 

2O 

10 

0 

Fig. 1 : Frequencies of toxic cyanobacterial blooms in 11 European countries (data from Table 2.2.1 in Sivonen 2000) 

ESPR - Environ Sci & Pollut Res 10 (2) 2003 129 
Eutrophication Review Articles 

Table 3: Swedish eutrophication classification s' 

Very small biomass 

rstem for total phosphorus and water-blooming cyanobacteria in late summer (modified from Willan 2000) 

6-12.5 <0.5 <_2. 

2 Small biomass 12.5-25 0.5-1 - 

3 Moderately large biomass 25-50 1-2.5 3-4 

4 Large biomass 50-100 2.5-5 - 

5 Very large biomass >100 >5 4 

summer cyanobacterial blooms and the number of toxin-producing genera increase with eutrophication in Swedish lakes 

(Table 3). However, Will6n et al. (2000) found no trend in 

peak cyanobacterial toxicity (tag/g dry wt. of algae) with lake 

trophic state, and found that maximum cyanophyte toxicity 

was more closely related to species dominance than to season. 

While many past studies of cyanobacteria have focused primarily on the phytoplankton of natural lakes, toxic bluegreen blooms also frequently occur in enriched, turbid Australian reservoirs (Davis 1997); similarly, 10 of 15 turbid 

Kansas reservoirs studied by Dodds (1996) were found to 

contain potentially toxic, microcystin-producing cyanobacterial strains. The empirical study of Smith (1990b) and the 

data shown in Fig. 2 indicate that cyanobacterial dominance 

can increase dramatically with nutrient enrichment in turbid, man-made impoundments; however, this general response is likely to be modified by the presence of high concentrations of non-volatile suspended solids (Smith 1990b). 

lOO 

E 90-1 9 9 

>o 801, 9 

0 9 ~ 70 -1 9 

.a ~ 60-1 

.__. $ 50 -1 

40 -1 

30-1 9 

~, 20-1 9 

i i 

0 50 100 150 200 250 300 

Total Algal Biovolume, rng/L 

Fig. 2: Relationship between cyanobacterial dominance and trophic state 

(as measured by summer algal biovolume) in 39 turbid Kansas, USA reservoirs. Data from Carney (2001) 

Empirical models relating the relative and absolute biomass 

of cyanobacteria to nutrient concentrations (Smith 1985, 

1986, 1990b, Smith et al. 1987), and models to predict the 

response of cyanobacterial blooms to phosphorus loading 

reductions (Lathrop et al. 1998), have been developed. However, a lively debate is currently being waged in the literature concerning the primary mechanisms that are responsible for cyanobacterial dominance in eutrophic lakes. A 

number of potential mechanisms have been proposed in the 

literature (Shapiro 1989, Klemer and Konopka 1989), but 

perhaps the most contentious is the relative role of nitrogen, 

phosphorus (N, P) ratios. Empirical and experimental evidence supporting the N:P ratio hypothesis of cyanobacterial 

dominance has been summarized by Smith and Bennett 

(1999), and the N:P ratio hypothesis has also recently been 

modified to include the effects of food web structure by Elser 

(1999). However, other researchers (e.g., Trimbee and Prepas 

1987, Canfield et al. 1989, Jensen et al. 1994, Reynolds 

1997, Downing et al. 2001) have concluded from their own 

analyses that elevated absolute concentrations of total phosphorus per se, rather than the presence of low TN:TP ratios, result in cyanobacterial dominance. This extremely 

important fundamental question remains unresolved, and is 

a key topic for future eutrophication research. 

1.2 Streamsand rivers 

Although eutrophication research in streams and rivers has 

lagged behind that of lakes and reservoirs, the nutrient enrichment of streams and rivers is nonetheless of great concern (Behrendt 1993, Biggs 2000). In the U.S., concerns about 

hypoxia in the Gulf of Mexico led to a White House-commissioned task group on the effects of reduced nutrient loading on surface water quality in the Mississippi River basin 

(Brezonik et al. 1999). In Europe, investigators have reported 

extensive eutrophication of the Bure River system in the 

United Kingdom (Moss et al. 1989, Moss and Balls 1989); 

K6hler and Gelbrecht (1998) have documented significant 

eutrophication in several major rivers in Germany; and Billen 

et al. (1994) have demonstrated strong differences in algal 

development at different sites in the Oise and Marne river 

system in France. The concept that flowing waters are almost always nutrient-saturated, and the conclusion that 

physical factors such as light-limitation and short hydraulic 

residence times consistently prevent or restrict any potential 

algal responses to nutrient enrichment in rivers, both no 

longer appear to be tenable (Smith et al. 1999). As will be 

reviewed in separate sections below, the nutrient enrichment 

of flowing waters strongly modifies the biomass and community structure of both suspended and benthic algae. 

Several nutrient classification systems have been proposed 

for flowing waters. For example, McGarrigle (1993) concluded that maintaining a mean annual dissolved inorganic 

phosphorus concentration <47 lag/L was necessary to prevent the nuisance growth of attached algae and to preserve 

water quality suitable for salmonid fishes in Irish rivers. The 

U.K. Environment Agency (1998) has used annual mean 

values of soluble reactive phosphorus (SRP, lagP/L) as the 

boundaries for four trophic state categories in streams and 

rivers: oligotrophic (20 lagP/L) mesotrophic (60 tagP/L); 

mesoeutrophic (100 lagP/L); and eutrophic (200 lagP/L). In 

contrast, Dodds et al. (1998) have proposed trophic state 

130 ESPR - Environ Sci & Pollut Res 10 (2) 2003 
Review Articles Eutrophication 

Table 4: Suggested boundaries for trophic classification of temperate streams (Dodds et al. 1998) 

oligotrophic <20 <60 

mesotrophic 20-70 60-200 

eutrophic >70 >200 

<10 <700 <25 

10-30 700-1500 25-75 

>30 >1500 >75 

boundaries for temperate streams using concentrations of 

total nitrogen (TN, lag/L) and total phosphorus, which may 

be more directly compared to those for lakes (Table 4). The 

general issue of establishing nutrient criteria in streams is 

discussed by Dodds and Welch (2000). 

Effects of eutrophication on suspended algal biomass. The 

production of suspended algae per unit total phosphorus is 

often significantly lower in rivers than in lakes and reservoirs (Dokulil 1994, Soballe and Kimmel 1987, Van Nieuwenhuyse and Jones 1996). This reduced response to nutrients may be caused in some systems by the high washout 

rate that the flowing water imposes on suspended algal 

biomass. For example, a very strong inverse relationship was 

observed between suspended chlorophyll a concentrations and 

discharge in the Hungarian Danube (cf. Fig. 4 in Schmidt 1994). 

In addition, the potential for high algal productivity in nutrient-rich, well-mixed systems may not be realized if the presence of suspended soils greatly reduces underwater light availability (Dokuli11994). To my knowledge, light attenuation by 

suspended inorganic particulates has not been explicitly included in previous empirical models linking nutrient concentrations to suspended algal biomass in fluvial ecosystems. 

However, data from the Mississippi River (Fig. 3) do not reveal a consistent reduction in the yield of suspended chlorophyll a at stations containing high concentrations of non-volatile suspended solids (NVSS, mg/L), and this question should 

be explored further in additional river systems. 

An important prediction of Fig. 3 and other empirical nutrient-chlorophyll models for streams and rivers (e.g., Basu and 

Pick 1996, Van Neeuwenhuyse and Jones 1996, Lohman 

and Jones 1999, Champion and Currie 2000) is that reductions in suspended algal biomass should generally occur in 

rivers that experience significant TP reductions following the 

imposition of point and non-point source controls. A strong 

reduction in suspended chlorophyll a (Fig. 4) indeed appears 

J 

o 

8 

1000 

100 

10 

[] NVSS:TP<0.13 

9 NVSS:TP>0.13 

\ 

[3 

[] 

u 41,~ 

[] 

[] 9 

[] 

1 i 

10 100 

Total Phosphorus, ug/L 

1000 

Fig. 3: Relationship between suspended algal biomass and total phosphorus in the Mississippi River. NVSS, TP ratios greater than 0.13 indicate 

high levels of inorganic turbidity, and potential shading inhibition of algal 

growth in the river. Data from Heiskary and Markus (2001) 

180 

16O 

140 

:=- 120 

~. lO0 

~_ 80 

o 

60 

g~ 40 

20 

Gunston Cove, 

Potomac River 

1997 

0 50 100 150 200 

Total Phosphorus, pg/L 

Fig. 4- Response of Gunston Cove in the tidal freshwater Potomac River 

USA, to reductions in point source phosphorus loading. Note that the response appears to be hysteretic. The data shown are estimates of annual 

means derived from LOWESS fits to the entire dataset by Jones (2000) 

to have occurred in the tidal freshwater Potomac River (USA) 

as a result of point source phosphorus loading controls (Jones 

2000). Also of very great interest in Fig. 4 is an apparent hysteresis in this chlorophyll-phosphorus relationship; such 

hysteretic responses to changes in nutrient loading can be seen 

in lakes (e.g., Lake Washington, USA; see Smith 1998), and 

can significantly complicate the recovery of aquatic ecosystems from eutrophication (Carpenter et al. 1999). 

Effects of eutrophication on suspended algal community structure. As in lakes, nuisance blooms of cyanobacteria can seriously degrade the water quality of fluvial ecosystems (Paerl et 

al. 1990, Prygiel and Leitao 1994, Ha et al. 1998, Mitrovic et 

al. 2000). Heiskary and Markus (2001) provide data that allow an initial test of the hypothesis that eutrophication causes 

predictable and directional shifts in the community structure 

of suspended algae in flowing waters. An analysis of their data 

(Fig. 5) suggests that cyanobacterial dominance in large rivers increases strongly and predictably with phosphorus enrichment, just as it does in eutrophic lakes and reservoirs. 

(D =E 

.I3 

e 

-(3 

O 

o 

o4 

90 

80 

70 

60 

50 

40 

30- 

20- 

10 

0 

0 

%BG = 58.9 Iog(TP) - 86.7 

r 2 = 0.72 

100 200 300 

Total Phosphorus, IJg/L 

J 

400 

Fig. 5: Relationship between blue-green algal dominance and total phosphorus in the Mississippi River. Data from Heiskary and Markus (2001) 

ESPR - Environ Sci & Pollut Res 10 (2) 2003 131 
Eutrophication Review Articles 

Phytoplankton dynamics in rivers thus may not be so different from those observed in shallow lakes (Reynolds 1994), 

with traditional biotic interactions between algae and nutrients perhaps being dampened or modified by fluvial conditions, but not totally unexpressed. 

Effects of eutrophication on benthic algal biomass. As is true 

of periphyton in lakes (Eloranta 2000, Herve 2000), benthic 

algae on natural substrates in flowing waters can respond 

sensitively to anthropogenic nutrient enrichment (Welch et 

al. 1992). A recent meta-analysis of nutrient enrichment 

experiments in streams by Francoeur (2001) has confirmed 

that benthic algal biomass can be significantly increased by 

the addition of either N or P, and Borchardt (1996) published a review of the literature on nutrients and stream periphyton communities. Krewer and Holm ( 1986 ) and Homer 

et al. (1990) observed a strong curvilinear dependence of 

periphyton growth on phosphorus in artificial laboratory 

channels, and this general relationship has been confirmed 

for lotic ecosystems by Dodds et al. (1997, 2002). Similar to 

the nitrogen-phosphorus interactions documented in Section 

1.1 above for lakes, and consistent with the strong N, P 

interactions documented in Francoeur's (2001) stream metaanalysis, the biomass of benthic algae growing on natural 

substrates appears to be dependent upon both TN and TP 

concentrations in the stream water (Fig. 6). 

1000 

100 

_9.o 

x: 1 

0 

0.1 

9 TN:TP<10 0 

O TN:TP>10 I~ D 

[] i1, 9 9 

i i 

10 100 1000 

Total Phosphorus, ug/L 

A~ 

10000 

Fig. 6: Relationship between benthic algal biomass and total phosphorus 

in streams. As is true of phytoplankton in lakes (Smith 1982), the yield of 

periphytic algae in flowing waters appears to be reduced when the TN:TP 

ratio is <10, 1 by mass. Note: these data are based only upon algal samples collected from natural substrates, and errors in the database have 

been corrected from the dataset analyzed in Dodds et al. (2002) 

Effects of eutrophication on the community structure of 

benthic algae. Shifts in community structure along nutrient 

supply gradients are a common feature of most aquatic ecosystems, and attached algae in streams are no exception. 

Whether these communities are influenced by exotic species 

(e.g. Hydrodictyon reticulatum: Hall and Cox 1995, Hall 

and Payne 1997), or are undisturbed by exotic invaders, 

nutrient availability often has a strong selective role in determining the species composition of benthic algal communities. Epilithic diatoms can be used as sensitive indicators 

of streamwater total nitrogen and total phosphorus concentrations (Winter and Duthie 2001), and Ch&elat et al. (1999) 

reported marked directional shifts in overall community 

composition as the total biomass of benthic algae increased. 

The proportion of green algae (Chlorophyta), for example, 

increased with benthic chlorophyll a and was highest at 

moderately eutrophic sites. These studies suggest that predictive models similar to those developed for phytoplankton 

communities in lakes, and for suspended algae in rivers (see 

Fig. 5), can be developed for benthic algae. 

1.3 Wetlands 

Effects of eutrophication on algal biomass and community 

structure. Wetlands (areas where water saturation is the dominant factor determining the nature of soil development, and 

structuring the associated plant and animal communities: Wolff 

1993) are a very common feature of many landscapes worldwide. However, as noted by Crosbie and Chow-Fraser (1999), 

the impact of human land use and nutrient loading on the 

water quality of freshwater wetlands has seldom been documented. Given the consistency of algal responses to nutrient 

enrichment in lakes, it might be expected that similar nutrient-algal biomass relationships would be observed for 

wetlands. A plot of data from 22 marshes in the Great Lakes 

Basin of Ontario, Canada (Crosbie and Chow-Fraser 1999), 

confirmed that suspended chlorophyll a in these wetlands responded strongly to increases in average water column total 

phosphorus (TP) concentrations (Fig. 7): 

log10 (Chlmoa,) = 1.14 Ioglo (TP .... ) + 1.17, r 2 = 0.72 (s) 

1000 

lO0 

m2 

10 

.J 

r 

1 

10 

lOOs 

B 

10s 

9 9 9 nmm 

100 

Total Phosphorus, pg P/L 

m m 

1000 

IC 

9 // 

9 mm,,# m 1 f /" 

/. 

m am // 

/ 

m9 // 

9 9 mm m jiBe" 

9 9 . m 

j/" 9 

///" 9 

// 

10 

i 

100 1000 

Observed Chla, pg/L 

Fig. 7: A. Relationship between suspended chlorophyll a and total phosphorus concentrations in 22 marshes near Ontario, Canada. Each data 

point represents the mean of two sampling trips during the growing season. B. Predicted versus observed chlorophyll a concentrations for the 22 

marshes above, using Equation 4 in Table 2 of Lurid (1998) 

132 ESPR - Environ Sci & PoUut Res 10 (2) 2003 
Review Articles Eutrophication 

Moreover, the algal biomass in these 22 marshes covaried 

with the predictions of a new empirical model developed 

for wetlands located on the Swan Coastal Plain of Australia 

(Lund 1998), although the predicted algal biomass consistently exceeded the observed values. This prediction bias may 

be due in part to the original model specification: the empirical chlorophyll-phosphorus model of Lund (1998) was 

developed using a mixture of data from lakes (permanently 

inundated), sumplands (seasonally inundated), and artificial 

wetlands. Nonetheless, the data in Fig. 7 suggest that the 

biomass of suspended algae in wetlands tends to respond 

strongly and predictably to changes in nutrient inputs. 

The possibility also exists that algal communities in nutrient-enriched wetlands experience directional shifts in species composition that parallel those found in lakes, reservoirs, and large rivers. For example, a recent survey of 32 

Kansas (USA) wetland areas by Carney (2002) revealed that 

the planktonic algae were dominated by cyanobacteria in 

the most eutrophic wetlands. Efforts to develop new models 

for suspended and attached algae in both freshwater and 

coastal marine wetlands would be a valuable new area for 

future eutrophication research; such models could provide 

objective, quantitative guidance in the restoration and protection of wetlands from excess nutrient loading. 

2 Eutrophication of Estuarine and Coastal Marine 

Ecosystems 

A very large fraction of the nutrients exported from the land 

surface to streams and rivers ultimately makes its way to 

the sea, and as a result, estuaries receive more nutrient inputs per unit surface area than any other type of ecosystem 

(Howarth 1993). More than half of the world's human population resides within 60 kilometers of the coast, and more 

than 90% of the world's fisheries depend in one way or 

another on estuarine and near-shore habitats (Hobble et al. 

2000). Increased nutrient loading results in a wide variety 

of changes in the structure and function of coastal marine 

ecosystems (cf. Fig. 2 in Schramm 1999), and protecting 

these systems from the many adverse effects of eutrophication 

is extremely important (U.S. EPA 1996b, National Research 

Council 2000, Rabelais and Nixon 2002). As can be seen in 

Table 5, the Swedish EPA has developed a nutrient-based 

classification scheme for coastal marine ecosystems that 

strongly resembles the trophic state categories presented 

earlier in Tables 2--4 for freshwater ecosystems. 

The possibility of applying a limnological approach to the 

management of coastal marine eutrophication was noted in 

Table 5: Swedish eutrophication classification system for coastal marine 

been rounded from the original molar values) 

the literature over two decades ago (Schindler 1981, Lee 

and Jones 1981). However, although freshwater eutrophication management can be considered a success story in ecosystem science (Smith 1998), it has recently been questioned 

in the literature whether the limnological experience can be 

repeated for estuaries and coastal marine ecosystems (Sharp 

2001). Rapidly accumulating evidence suggests that the answer to this critically important question is an increasingly 

confident yes! While estuarine and coastal marine systems 

can often be hydrologically and spatially complex, the 

biomass of phytoplankton in these systems nonetheless tends 

to respond predictably to changes in the water column concentrations of total nitrogen and total phosphorus. 

An analysis of Danish coastal waters by Nielsen et al. (2002) 

is an excellent example of recent successes in coastal zone 

eutrophication modeling. Nielsen et al. (2002) analyzed a very 

large dataset from 162 stations located in 27 Danish fjords 

and coastal waters, and developed empirical models for total 

nitrogen (TN, tag/L), chlorophyll a (Chl, lag/L), suspended 

matter (Wsusp , mg dry wt./L), and Secchi disk transparency 

(SD, m) that should help to provide objective guidance for 

water quality management in Danish coastal waters, 

In (Chl) = 0.963 In (TN) -4.186, r 2 = 0.30 (6) 

In (SD) = -0.444 In (Chl) -0.367 In (Wsusp) + 2,686, R 2 = 0.81 (7) 

In (SD) = -0.894 In (TN) + 6.818, r 2 = 0.48 (8) 

Additional models based upon either total phosphorus or total nitrogen have been developed for coastal marine ecosystems in the Adriatic Sea (Vollenweider et al. 1992); Canada 

(Champion and Currie 2000, Meeuwig et al. 1998); Mexico 

(Contreras and Kerekes 1993); Scandinavia (Hfikanson and 

Wallin 1991, Borum 1996, Meeuwig et al. 2000, Elmgren and 

Larsson 2001, Nielsen et al. 2002); and the USA (Boynton et 

al. 1996, Bowen and Valiela 2001). Efforts have also been 

made to apply the OECD's critical load concept to Norwegian marine waters (Hessen et al. 1992), and Janicki (2001) 

has developed a chlorophyll a-based trophic state index for 

Florida (USA) estuaries that parallels the highly successful 

trophic state index created by Carlson (1977) for lakes. Even 

more importantly, improvements in coastal marine water quality tend to follow reductions in external nutrient loading (Smith 

et al. 1999, Boesch et al. 2001, Elmgren and Larsson 2001), 

as has been documented in many freshwater lakes. 

However, as stressed by Richardson (1996), scientists cannot be expected to identify and recommend courses of acsurface water (modified from Swedish EPA 2000; nutrient concentrations have 

Very low 

Summer TN; H~L : 

<252 

Summer TP, pg/L 

<15 

2 Low 252-308 15-19 

3 Moderate 308-364 19-24 

4 High 364-448 24-31 

5 Very high >448 >31 

August Chloro~y!l a, p~ 

<1.5 

1.5-2.2 

2.2-3.2 

3.2-5.0 

>5.0 

ESPR - Environ Sci & Pollut Res 10 (2) 2003 133 
Eutrophication Review Articles 

tion to combat coastal zone eutrophication in the absence 

of clear goals for environmental conditions desired in a given 

area. Objective water quality targets can be defined by using one or more of the following methods: 

1. By applying quantitative water quality criteria developed 

by local, regional, or federal environmental protection agencies, such as the empirical boundaries for TN, TP, and Chl a 

concentrations in eutrophic waterbodies that have been 

proposed for coastal waters (e.g. H~kanson 1994); 

2. By using historical data, time series data, or reference 

region data to establish reference values that reflect pristine water quality conditions completely free of any effects of human activity (e.g., Table 2b in Swedish EPA 

2000); or 

3. By using robust survey methods that allow local users of 

the water resource to generate quantitative estimates for 

the value of a key water quality variable (the Quality 

Variable of Concern, sensu Reckhow and Chapra 1983 ), 

that represents their perception of the critical point at 

which undesirable water quality is present. 

Once target goals have been established, quantitative relationships between water quality and nutrients can be used 

to calculate the critical water column nutrient concentrations that must be attained in order to reach these target 

conditions. This information then can be linked to macroscale models of nutrient and water fluxes to the coastal zone 

(e.g., Valiela et al. 1997) in order to identify the corresponding nutrient loading rates that must be achieved to obtain 

desirable water quality in the receiving waters being managed (for example, cf. Fig. 9 in Valiela et al. 2000). However, many critical questions still remain regarding the actual development and implementation of measures to protect 

coastal marine ecosystems from eutrophication, and answering these questions will require thoughtful, trans-disciplinary efforts (cf. Elmgren and Larsson 2001). 

3 Conclusions and Outlook 

A remarkable unity is evident in the global response of algal 

biomass to nitrogen and phosphorus availability in lakes and 

reservoirs; wetlands; streams and rivers; and estuaries and 

coastal marine waters. The species composition of algal communities inhabiting the water column appears to respond 

similarly to nutrient loading, whether in lakes, reservoirs, 

or rivers. As is true of freshwater ecosystems, the recent literature suggests that marine ecosystems respond positively 

to nutrient loading control efforts. 

As can be seen in the above review, our understanding of 

freshwater eutrophication and its effects on algal-related 

water quality is extremely strong and is advancing rapidly. 

Unfortunately, however, our understanding of the effects of 

eutrophication on estuarine and coastal marine ecosystems 

is much more limited, and this gap represents an important 

future research need. While estuarine and coastal marine 

systems may often be hydrologically complex, numerous 

empirical analyses that have been performed worldwide indicate that the biomass of marine phytoplankton responds 

sensitively and predictably to changes in the external supplies of nitrogen and phosphorus. These models in turn suggest that efforts to manage nutrient inputs to the seas will 

result in significant improvements in coastal zone water 

quality. Efforts should be made to develop new models that 

quantitatively link ecosystem-level responses to nutrient loading in both freshwater and marine systems. 

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Received: August 29th, 2002 

Accepted: December 4th, 2002 

OnlineFirst: December 6th, 2002 

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