# EFFECTS OF EUTROPHICATION ON

STREAM ECOSYSTEMS

Lei Zheng, PhD and Michael J. Paul, PhD

Tetra Tech, Inc.2 Abstract

This paper describes the effects of nutrient enrichment on the stru cture and function of stream ecosystems. It starts with the curre ntly well documented dir ect effects of nutrient enrichment on algal biomass and the resul ting impacts on stream chemistry. The paper continues with an explanati on of the less well documented in direct ecological effects of

nutrient enrichment on stream structure and function, including eff ects of excess growth on physical habitat, and alterations to aquati c life community structure from the microbial assemblage to fish and mammals. The paper also dicusses effects on the ecosystem level including changes to productivity, respiration, decomposition, carbon and other geochemical cycles. The paper ends by discus sing the significance of these direct and

indirect effects of nutrient enrichment on de signated uses - especially recreational, aquatic life, and drinking water.3 1. Introduction

1.1 Stream processes

Streams are all flowing natural waters, rega rdless of size. To understand the processes that influence the pattern a nd character of streams and reduce natural variation of different streams, several stream classifi cation systems (including ecoregional, fluvial geomorphological, and stream or der classification) have been adopted by state and national programs. Ecoregional classificat ion is based on geol ogy, soils, geomorphology, dominant land uses, and natural vegetati on (Omernik 1987). Fluvial geomorphological classification explains stream and slope pr ocesses through the application of physical principles. Rosgen (1994) classified stream channels in the United States into seven major stream types based on morphologica l characteristics, including entrenchment, gradient, width/depth ratio, and sinuosity in various land forms. These morphological characteristics affect stream ecosystem pro cesses and community structure and functions. Stream order classification (Str ahler 1964) is also widely a pplied for organizing drainage

networks in the United States. These str eam classification systems describe hydrology and material transport, which in turn in fluence physical, chemical, and biological

processes.

Another classification scheme is to classify streams based on nutrient conditions (EPA 2001a). EPA divides the country into 14 leve l III nutrient ecoregions (Omernik 2000)

with common land use characteristics to bett er assess background nut rient concentrations in different geographic regions. This clas sification reflects spatial and geographic

variations that influence nut rient concentrations in str eams (Rohm et al. 2002, Wickham 2005) and natural background nutrient concentr ations should be established for each region (Smith et al. 2003). Dodds (1998, 2006) proposed classifying streams into trophic state classes similar to those develope d for lakes and reservoirs (EPA 2001b).

One of the most important processes in st reams is nutrient cycling. Stream channels receive nutrients from upstream, terrestrial runoff, ground water, and the atmosphere. The proportion of each source is variable de pending on stream geology, elevation, and regional setting. Different landforms (forest vs. agricultural catchments) and spatial and temporal variables also signi ficantly affect nutrient concen trations and loadings into streams (Arheimer and Liden 2000). Internal nutr ient cycling also provides nutrients to streams (Mulholland 1996). Stream biota use nut rients and convert them into biomass; thus, nutrients are important to ecosystem structure and function.

Two major nutrients, nitrogen (N) and phosphorus (P), occur in streams in various forms as ions or dissolved in solution. Aquatic pl ants convert dissolved inorganic forms of nitrogen (nitrate, nitrite, and ammonium ) and phosphorus (orthophosphate) into organic or particulate forms for use in higher troph ic production. The right balance of nitrogen and phosphorus is essential for maintain ing natural biological communities and ecosystem functions in aquatic systems. In freshwater systems, phosphorus and nitrogen4 are limiting nutrients, that is, the levels of these nutrients limit the biological productivity of such systems.

1.2 Limiting nutrients in streams

Stream primary producers, i.e., algae a nd macrophytes, absorb natural energy from sunlight to fix carbon and conve rt inorganic forms of N and P into organic forms through photosynthesis, storing the energy produced in thei r cells. In most streams, either N or P concentrations or both limit this process. Diffe rent algae have been reported to require different N and P concentrations for growt h. One study found that diatoms require less P (0.3-0.6 μg/L P, Bothwell 1988) to saturate gr owth than filamentous green algae (25-50 μg/L P, Bothwell 1989). Nitrogen limitati on has been reported when ambient N concentration was 55 μg/L in a desert st ream in Arizona (Grimm and Fish 1986) and when it was less than 100 μg/L in an Ozark stream (Loh man et al. 1991). Rier and Stevenson (2006) found that algal growth was 90% of maximum rates or higher in nutrient concentrations of 16 μg/L P and 86 μg /L N. The Redfield ratio (molar ratio of 106:16:1 for C:N:P) has been proposed as a community-wide optimum nutrient ratio (Redfield 1958, Borchardt 1996). High ambient or cellular N:P ratios (N:P >20:1) indicate P is limiting growth; low N:P ratios suggest that N is limiting (N:P<10:1). However, levels of nutrient concentrations and ratios for nutrient limitation are also regulated by other abiotic and biotic factors.

Regional differences may determine limiting nutrients for plant growth. Phosphorus used

to be and is still considered the sole limiti ng nutrient in aquatic systems by a number of authors (Huchinson 1957, Correll 1998, Khan and Ansari 2005). With increasing experimental manipulation of nutrient limita tion, especially bioassays using nutrient diffusing substrates and arti ficial streams, N limitation and N and P co-limitation are quite commonly discovered (Grimm and Fisher 1986, Peterson and Grimm 1992). Borchardt (1996) reviewed st udies in North America and concluded that roughly the

northern half of the United States is P lim ited while the Southwest and Missouri Ozarks are N limited. The Pacific Northwest may be lim ited by both N and P. A meta-analysis of 237 nutrient enrichment studies in temperate streams revealed that 16.5% indicated an N response, 18.1% indicated a P response, 23.2% required N and P be added together for a response, 5% had N or P inhibi tion, and 43% had no response to N or P (Francoeur 2001). These proportions have been confirmed by a similar literature review (Tank and Dodds 2003).5 1.3 Eutrophication problems

Eutrophication means “good food”. In freshwat er systems, eutrophication is a process whereby waterbodies receive excess inorganic nutrients, especially N and P, which stimulate excessive growth of plants and al gae. Eutrophication can happen naturally in the normal succession of some freshwater ecosystems. However, when the nutrient

enrichment is due to the ac tivities of humans, sometimes referred to as “cultural eutrophication”, the rate of th is natural process is greatly intensified. Eu trophication wasrecognized as a pollution problem in North American lakes and reservoirs in the mid20th century (Rohde 1969). Alt hough nutrient pollution has lo ng been recognized as a major problem in streams and rivers (USE PA 2000), the concept of eutrophication has been less commonly used with respect to nutrient enrichment problems in streams (Dodds 1998, 2006).

Nutrient enrichment of streams in the Un ited States is widespread (Carpenter 1998, Correll 1998, Smith et al. 1999, 2006). EPA assessed approximately 840,000 river and stream miles nationwide and reported that 10% of assessed rivers and streams had nutrient enrichment problems, which contri buted to 30% of reported water-quality problems in the impaired rivers and str eams (~ 291,000 miles). (USEPA 2002). Nitrate

concentration has more than doubled in the Mississippi River since 1965 and concentrations in many major rivers in the No rtheast have increased by from 3- to 10-fold since the early 1900s (see reviewed by Vitous ek 1997). Smith et al. (1987) found that at

381 riverine sites in the continental United States, the mean total phosphorus concentration was 130 mg/m 3 , which is almost double the threshold value for eutrophication (75 mg/m 3 ) proposed by Dodds (1998) for streams.

1.4 Sources of nutrient enrichment : point and nonpoint sources

Nutrient concentrations in streams and rivers have been strongly correlated with human land use and disturbance gradients. Both N and P enrichment are linked to agricultural

and urban land uses in the watershed. Fluxe s of total N in temperate-zone rivers surrounding the North Atlantic Ocean are high ly correlated with net anthropogenic input of N to their watersheds (Howarth et al. 1996). Total N and nitrate fluxes and

concentrations in rivers are also correlate d with human population density (Cole et al. 1993, Howarth et al. 1996). Nitrogen fertilizatio n is the main source of N in streams and rivers (Goolsby and Battaglin 2001). Simila rly, urbanization generally leads to higher phosphorus concentrations in urban catchment s (see review by Paul and Meyer 2001). Increasing imperviousness, increased runoff from urbanized surfaces, and increased municipal and industrial discharg es all result in increased lo adings of nutrients to urban streams. This makes urbanization second only to agriculture as the ma jor cause of stream impairment, even though the total area cove red by urban land in the United States is minor compared to agricultural area (Paul and Meyer 2001).

Nutrient enrichment of aqua tic systems from anthropogeni c sources includes point and nonpoint sources (Table 1, adapted from Carp enter et al.1998). Both have degraded aquatic systems. Point sources of nutrients include wastewater effluent (both municipal6 and industrial) and storm sewer discharge. In contrast to po int sources of nutrients that are relatively easy to monitor and regulate , nonpoint sources such as livestock, crop fertilizers, and urban runoff exhibit more spatial and temporal variability. Following strong regulation of point source inputs in re sponse to the Clean Water Act, nutrients from nonpoint sources are now the major sour ce of water pollution in the United States (Carpenter et al. 1998).

Table 1. Characteristics of point and nonpoint sources of chemical inputs to receiving waters recognized by statutes of the United States (from Carpenter et al. 1998).

• Point Sources -- Wastewater effluent (municipal and industrial) -- Runoff and leachate from waste disposal sites

-- Runoff and infiltration from animal feedlots -- Runoff from mines, oil fields , unsewered indu strial sites -- Storm sewer outfalls from cities with a population >100,000 -- Overflows of combined storm and sanitary sewers

-- Runoff from construction sites >2 ha

• Nonpoint Sources -- Runoff from agriculture (including return flow from irrigated agriculture) -- Runoff from pasture and range

-- Urban runoff from unsewered and se wered areas with a population <100,000 -- Septic tank leachate and runoff from failed septic systems

-- Runoff from construction sites

-- Runoff from abandoned mines -- Atmospheric deposition over a water surface -- Activities on land that gene rate contaminants, such as logging, wetland conversion, construction, and development of land or waterways

Seasonal variability of nutrient loading into streams is dramatic in both agricultural and woodland streams impacted by fertilization. Fert ilization of timberlands with nitrogen to increase production is a common practice fo r the timber industry in the United States. Typically, concentrations of nitrate-N in str eams increase rapidly during rainstorms after

fertilization in the fall. Peak concentrati ons have ranged from less than 100 μg/L to greater than 10,000 μg/L (exceeds EPA dri nking water level) during high runoff (Anderson 2002). Ambient NO 3 -N concentrations in forest streams are typically elevated

as much as from 2- to 10-fold for the entire winter and spring following a fall fertilization (Bisson et al. 1992). Stream uptake reduces NO 3 -N concentrations to background levels by summer in most cases (Mulholland 1992, Mulholland and Rosemond 1992). Then a secondary NO 3 -N peak is observed in the subseque nt fall period, which indicates that applied nitrogen fertilizer remains availabl e for leaching to streams beyond the spring and summer growing seasons. An extreme example of long-term availability is reported at Fernow Experimental Forest (i n the Appalachians), where NO 3 -N remained elevated relative to control streams 10 years after fe rtilization with ammoni um nitrate at 336 Kg N/ha (Edwards et al. 1991).7 1.5 Consequences of eutrophication

Allan (2004) has summarized the various im pacts of land use on streams and rivers. Eutrophication, as one of the main causes of stream impairment in the United States, imposes severe threats to ecosystem structur e and function. The dire ct impact of nutrient enrichment is to increase autotrophic pr oduction and change species assemblages including proliferation of filamentous algae. Nutrient en richment also accelerates litter breakdown rates by bacteria and fungi. The dram atic changes at lower trophic levels may also lead to “trophic cascading”. As nutrien t concentrations increase and destabilize the primary producer assemblage and water chemis try, macroinvertebrates and fish may shift from sensitive species to mo re tolerant, often non-native species. Changes in the food web may also cause changes in ecosystem function and further alter stream physical habitat and water chemistry, e.g., decreasing dissolved oxygen.

1.6 The objective of this document

The trophic structure of stream ecosystems can be divided into producers and consumers

. Primary producers often refers to algae, moss, fern s, and higher plants. They convert inorganic nutrients into organic forms and ar e the main energy source of streams. Primary producers are consumed by primary consumers (herbivores) or after they die, the organic matter is added to the detrital cycle and either decomposed by decomposers (bacteria and fungi) or ingested by detritivores

. Secondary consumers (predators) rely on energy sources from primary consumers. Nutrient en richment of streams directly changes the nutrient supplies for primary producers, which then affects consumers and nutrient cycling processes, leading to changes in ecosystem structure and function (i.e., eutrophication).

The objective of this document is to provide an understandin g of the effects of nutrient

enrichment on the structure and function of stream ecosystems. Several authors have documented the relationship between stream enrichment and algal biomass (e.g., New England Interstate Water Pollution Contro l Commission 2001, Virginia Water Resources Research Center 2006, Dodds 2006). And EPA’ s nutrient criteria technical guidance manual for streams also provides background on nutrient impacts. However, the ecological effects of enrichment on stream structure and function are less well discussed. To develop ecologically sound nutrient criteria to protect streams, it is critical to understand how ecological proce sses that affect stream nutrient dynamics and ecosystem function will change with elevating nutrient pollution.

Although it is important to understand nutrien t cycling processes, including nutrient exchange between terrestrial, aquatic, and atmospheric systems (Mulholland 1992, Valett
et al. 1994, Boulton et al.1998, McMahon et al. 1 994, Arheimer and Liden 2000), they are beyond the scope of discussion for this document. Readers interested in nutrient cycling should see some of the many reviews on this topic (e.g., Stream Solute Workshop 1990).8 2. Effect on physical habitat and water chemistry

2.1 Physical habitat modification

Nutrient enrichment can substantially change st ream habitat. The direct effect of nutrient

enrichment in streams is excessive accumula tion of filamentous benthic algae during the peak summer growing season, altering flow environment; physical be nthic habitat used by stream invertebrates and vertebrate orga nisms (Welch et al. 1989, Chessman et al. 1992). Filamentous green algae, such as Cladophora, Ulothrix , and Rhizoclonium , favor

nutrient enriched environments. Under suitable enrichment, they can be several meters long in high-velocity areas. When they are ag ing, they often detach and float near the surface (Power 1990). Vaucheria and Chara can trap sediments to form knolls more than a meter in length in the sandy streams of the north central United States. Chain forming diatoms, such as Terpsinoe

, can also form long filaments in flowing water. Cladophora and other filamentous algae are often habitat fo r small invertebrates, such as chironomids, amphipods, and many meiofauna (see review by Dodds and Gudder 1992).

Periphyton and macrophytes can al so alter water velo city in streams. Dodds and Biggs (2002) found that algae attenuate current ve locity more than macrophytes, and of the different types of algae, dense aggregations of diatoms (primarily Cymbella) attenuate velocity more than filamentous green algae orred algae. Benthic aquatic plants can cause an exponential decline in velocity with depth. In urban channelized streams, Cladophora

and Ulothrix attach to concrete irri gation canals, reducing flow rate and capacity. Aquatic plants clogging stream and drainage ditches causes wa ter backup, flooding, and water loss through evaporation from floating or emergent plant surfaces.

Nutrient enrichment can also lead to excessive phytoplankt on growth. Excessive planktonic algal growth in slow movi ng water can reduce light penetration and consequently limit the growth of submerged aquatic plants, decreasing available habitat and shelter for fish and their food organisms (Sand-Jensen et al. 2000). Lack of aquatic plants can also cause erosion and instability of the stream channel.

2.2 Water chemistry

2.2.1 Dissolved Oxygen (DO)

Nutrient enrichment leads to excessive growth of primary producers as well as heterotrophic bacteria and fungi, which increases the meta bolic activities of stream water and may lead to a depletion of dissolved oxygen (Mallin et al. 2006). During the day, photosynthesis by primary producers provides a large amount of oxygen to the water. At night, photosynthesis stops and elevated respir ation by algae and bacteria continues to consume dissolved oxygen, which can deplet e DO. Furthermore, as primary producers die, they are decomposed by bacteria that consume oxygen. Large populations of decomposers consume more dissolved oxygen, which increases the severity of DO depletion.9 DO depletion has been reported in many st reams through diel studies but is more frequently studied in lowland, slow movi ng streams (House and Denison 1997, Kaenel et al. 2000, Sabster et al. 2000, Wilcock and Nagels 2001). Daily oxygen fluctuations in enriched streams at low flow were reporte d to range from a high of approximately 25 mg/L at noon to a low of approximately 3 mg/L at night (W ong and Clark 1976). Decomposition and microbial activities are lik ely enhanced during summer when rising temperatures lead to lower DO saturation. So metimes streams become anoxic (House and Denison 1997). Excessive amounts of nutrients in streams and rivers may also negatively impact the dissolved oxygen levels of downs tream receiving waters. For example, a zone of hypoxia (< 2 mg/L) in the Gulf of Mexico has been linked to hi gh nutrient inputs from the Mississippi River (USEPA 2000, Mallin et al. 2006).

Long-term nutrient enrichment (of streams) l eads to long term declines in average DO concentrations (Vagnetti et al. 2003, Parr and Mason 2003, 2004). One long-term study (Parr and Mason 2004) found a significant decline in DO during1955-1998 in the River Brett, England. That stream has shifted from an autotrophic to a he terotrophic system.

The composition of the aquatic plant commun ity can affect the DO response. Wilcock and Nagels (2001) found that streams domi nated by submerged macrophytes exhibited the greatest amplitude swings in DO a nd pH, and led to DO levels of ~ 86-128% saturation. Parr and Mason (2004) estimated that macrophytes accounted for 45% of community respiration at the study site while sediment accounted for 36%, and suggested that removal of plants would restore DO. Kaen el et al. (2000) suggest ed that an increase in the oxygen concentration after plant cut ting would only be transient in unshaded,

nutrient-rich streams.

2.2.2 pH

Diel changes in pH often occur in regions with low acid neutrali zing capacity and are related to excessive algal gr owth and stream metabolism. The “normal” pH range in streams is from 6 to 9 with 7 being neutral, less than 7 indicating acidic conditions, and greater than 7 indicating ba sic conditions. During photosynt hesis, carbon dioxide (CO 2 ) and water are converted by sunlight into oxygen and carbohydrate. Hydroxyl ions (OH - ) are produced, raising the wate r column pH. In addition, pl ants use a large amount of

dissolved CO 2 for photosynthesis, resulting in lower levels of carbonic acid (H 2 CO 3 ) in the water column. Thus, photosynthesis increas es water column pH. At night, increased respiration from biota increases the release of CO 2 into the water, increasing the production of carbonic acid and hydroxyl ions, which, in turn, increases the acidity.

Extremely high or low pH values in stream s are harmful to aquatic organisms. For example, high pH levels can be toxic to fish and other organisms. High pH levels damage fish gills, eyes, and skin, and affect fish reproduction. High pH levels also increase the toxicity of some substances, such as ammoni a. Low pH levels can make heavy metals in stream sediment more bioavailable (e.g., Al). pH levels also influen ce the availability of some nutrients, further exacerbating enrichment problems.10 2.2.3 Other chemicals

Toxic effects of chemicals released from cer tain cyanobacteria have been reported in lakes; very few studies have found cyanotoxins in streams. Pfiester ia, a toxic substance produced by dinoflagellates that cause fish kills, has also b een reported in coastal rivers associated with nutrient enrichment (B urkholder 1999). A relativ ely new golden alga, Prymnesium parvum , has been reported to be toxic in Texas. The toxin prymnesin affects gill-breathing organisms including fis h, tadpoles, and clams (Rhodes and Hubbs 1992) and has been responsible for an estimated 2.5 million dead fish and millions of dead clams in the Pecos, the Colorado, and Brazos river basins in Texas.

Other chemicals can taint drinking wate r supplies and recr eational waters. 2- methylisoborneol and geosmin are two chemi cals produced by cyanobacteria that can cause taste and odor problems in drinki ng water. Livestock that drink water contaminated with cyanobacteria have died (Dodds and Welch 2000). Humans who drink

or swim in water that contai ns high concentrations of toxins from cyanobacteria may experience gastroenteritis, skin irritation, alle rgic responses, or liver damage (CDC 2004). In 1991, one of the largest recorded riverine blooms of toxic cyanobacteria occurred in the Murray-Darling River Basin in Australia, resulting in a state of emergency being declared to protect water supplies draw n from the river (Oliver et al. 1999).

3. Direct biological responses of streams to eutrophication: primary producers

Although it is generally recognized that nutrient enrichment st imulates algal growth in nutrient limited streams, increased alga l productivity may or may not result in accumulation of algal biomass. A number of factors, i.e., light availability, grazer intensity, and physical disturbance, can affect algal biomass accumulation in streams. The fate of periphyton biomass produced in aquati c ecosystems follows four different paths (Lamberti 1996): (1) accumulation as a standi ng crop of algae, (2) respiration to CO 2 (decomposition), (3) consumption by herbivores (grazing), and (4) exported to suspended matter. The specific fate of biomass will de pend heavily on grazing regimes (Lamberti et al. 1987, 1993). In heavily grazed systems, pr imary production is either consumed or exported, while reduced grazing pressure allo ws accumulation of algal biomass and its decomposition.11 3.1. Responses of algal bioma ss to nutrient enrichment

A number of authors have documented the positive relationship be tween benthic algal biomass and nutrient concentrations (see reviews by ENSR 2001, Virginia WRRC 2006, Dodds 2002, 2006). These studies include both fiel d manipulations of nutrient levels to assess benthic algal growth (Bothwell 1989, Walton et al. 1995, Stevenson et al. 2006, Rier and Stevenson 2006) and large scale surv eys that investigate relationships between nutrient enrichment and periphyton biomass across different streams (Dodds et al .1997, Biggs and Close 1989, Welch et al. 1992, Biggs 2000, Lohman et al. 1992, Chetelat et al. 1999). These studies established that total N and total P in the water column are significantly related to benthi c algal biomass. Empirical models derived from these studies allow one to predic t the mean summer chlorophyll a level in a stream based on nutrient concentrations (e.g., Biggs 2000).

The positive relationship between algal biom ass and nutrient concentrations is also observed in phytoplankton (Jones et al. 1984, van Nieuwenhuyse and Jones 1996, Basu and Pick 1995, Lohman and Jones 1999). Phyt oplankton in streams r eceive less attention because they are low in abundance and fr equency. However, in many slow moving embayments, algal concentrations can be more than 40 times that in the main stem of the river (Reynolds and Descy 1996). Van Nieuwe nhuyse and Jones (1996) used data from 292 temperate streams and found that summer mean sestonic (suspended) chlorophyll concentration showed a strong curvilinear relationship with summer mean total phosphorus concentration (TP), although much of that sest onic chlorophyll was likely due to dislodged benthic algae technically calle d tychoplankton.

A fair number of nutrient enrichment studi es in streams find no relationship between algal biomass and nutrient concentrations . Many hypotheses have been proposed to explain these results. One possible hypothesi s is that nutrients exceeded the maximum concentration to saturate algal growth (e .g., Munn et al. 1989). However, under most circumstances, a number of other factors, such as hydrologic regi me (e.g., Welch et al. 1988, Biggs 1989, 1995), light availability (e.g., Lowe et al. 1986), grazers (e.g., Rosemond 1998) and interactions of thes e factors (Hill et al. 1995) could have

contributed to the lack of correlations between benthic algal biomass and nutrient

concentrations. Resources and disturbances are the main factors (Figure 1, Biggs 1996) that influence algal acc rual in a stream.12 Figure 1. Summary of the disturbance-resource supply-g razer concept for the co ntrol of benthic algal development in streams. The relative balance of “b iomass accrual” and “biomass loss” is depicted by the width of the triangles that make up the central rect angle. The physiognomy of the community likely to dominate each end of the gradient is also shown. After Biggs 1996.

Hydrological events such as scours signifi cantly reduce the eff ects of nutrients on periphyton biomass (Welch et al. 1988, G hosh and Gaur 1994, Bourassa and Cataneo 1998). Biggs (2000) developed a comprehens ive model linking hydrologic regime and nutrients to algal accrual. This model was de veloped using data from New Zealand rivers and streams across a wide range of land us e practices and hydrologi c patterns. Although the relationship between nut rients and algal biomass was relatively weak (R 2 =0.30), including the time of accrual (time since the last scouring flood) increased R 2 values to approximately 70%, indicating that eutrophicatio n effects are stronger under stable flow

conditions.

Grazers significantly reduce the amount of algal biomass in streams (Steinman 1996). Rosemond et al. (1993) mani pulated both nutrient and herbivore abundances in an experimental mesocosm and found that grazer s could moderate or eliminate the observed increase in periphyton biomass with increasi ng nutrient levels. Other studies also found that increased densities of grazers recruite d from nearby areas (Lowe et al. 1986, Welch et al. 1988, Rosemond 1994, Wellnitz et al. 1996, Bourassa and Cataneo 1998, Hillebrand 2002, Roll et al. 2005) could eliminat e increases of algal biomass to nutrient

additions.

Light is a key factor for algal growth a nd consequent algal biomass (Hill 1996). Algal biomass can be 4 or 5 times higher in stream segments with open canopies than at sites with more closed canopy cover (Lowe et al. 1986). L ight effects increase with trophic state and algal biomass, indicating enhancedimportance of light limitation and selfHigh Biomass

Low Biomass Biomass Accrual Biomass Loss

Grazing Disturbance

Fish Substratum Instability

Velocity

Suspended solids

Invertebrates Temperature

Resources

Nutrients

Light

Low growing, tightly adhering

taxa Erect, stalked
and/or filamentous

taxa13 shading at high nutrient supply (Hillebrand 2005). Light and nutrient limitations also interact, so increased light availability or nutrien ts alone sometimes do not guarantee an elevated biomass. In a nitrogen limited syst em, increased light le vel did not increase accumulation of algal biomass without nitrogen additions (Taulbee et al. 2005). Similarly,

several studies have found that nutrient addition alone does not stimulate algal biomass accrual unless light availability is also in creased (Bernhardt and Likens 2004, Mallory and Richardson 2005). These results demonstrate th at for essential resources such as light and nutrients, the magnitude of the respons e to enrichment by one resource depends on the relative availability of another.

Interactions of light, nutrients, and gr azers may confound the relationship between nutrients and algal biomass in streams. In a meta-analysis of experimental manipulation studies, Hillebrand (2005) f ound that increased light ( light enhancement) increases grazer effect size, whereas grazer pr esence reduces light effects, which indicates that high light favors algal growth but increased algal biom ass is easily offset by increased grazer densities. Similarly, Hill et al. (1995) found light-enhanced primary productivity was reflected by increased grazer densities instead of increased algal biomass.

3.2. Responses of algal species comp osition to nutrient enrichment

Algal species composition changes with elev ated nutrient concentrations (Stevenson 1996, Pan et al. 1996, Stevenson and Smol 2001). Because of their small scale, periphytic algae composition receives less public atte ntion, while problematic macroalgae (e.g., Cladophora) and cyanobacteria receive more . Under most circumstances, a diatom dominated algal community represents healt hy, non-enriched stream water quality, while a predominance of filamentous algae may i ndicate problems with nutrient enrichment. Since algae are often the probl em associated with enrichment, a change of taxonomic composition in a stream can show whether nuisance algae are present and can indicate long or short-term changes in point a nd nonpoint source pollution (Lowe and Pan 1996) that cannot be detected by a one-time sampli ng of water chemistry. Thus, algal species composition could be considered an im portant indicator of nutrient pollution.

Diatom species composition is increasingly used as an indicator of environmental conditions, especially nutrient enrichme nt in streams (Stevenson 1999). European scientists initially used diatoms as indicator s of organic pollution a nd developed tolerance values (e.g., Lange-Bertalot 1979) that were adapted in the United States (Lowe 1974, Bahls 1992, Stevenson and Bahls 1999, KDOW 2002). Recent studies have used a variety of algal attributes to develop biological indices (Hill et al. 2000, 2003, Fore and Grafe 2002, Wang et al. 2005). Diatom indicat ors are most sensitive to nutrient enrichment, and several authors (van Da m et al. 1994, Kelly and Whitton 1995, Kelly et al. 1998, Winter and Duthie 2000) have reporte d this sensitivity. A variety of studies have also developed diatom autoecological preferences for nutrients using weighted average regression techniques like those used in paleoecology (Pan and Stevenson 1996, Potapova et al. 2004, Potapova and Charles 2005).

3.3 Responses of macrophytes to nutrient enrichment14 The definition of macrophytes varies. A ccording to Wetzel (1975), macrophytes represent a taxonomically diverse group of a quatic plants includi ng vascular plants, mosses, ferns, and macroalgae. However, pe ople generally exclude macroalgae from this group. Most macrophytes (i.e., emergent, fl oating-leaved, and submerged groups) are rooted in the sediments so they can persist in current flow and absorb nutrients from sediments. Rooted plants live in a more pr edictable light climate but may experience severe self-shading and gene rally grow more slowly than algae. The nutrient requirements of rooted macrophytes are lower than those of microalgae because of their low growth rates, high internal C:N:P ra tios, and nutrient conserving mechanisms. Nutrient limitation is less important because the plants exploit rich nutrient pools in the sediment (Sandjensen and Borum 1991). One group of m acrophytes (e.g., duckweeds), however, are unattached and rely entirely on nutrients from the water column. They are easily affected by current and wind, and ar e most frequently found in backwaters.

Nutrient effects on macrophytes are poorly studied (Chambers et al. 1999). Several

factors have constrained this research. First, light is the most crucial factor regulating species composition and distribution of m acrophytes in streams (Chambers and Kalff 1985). Light availability is easily affected by water clarity, ripari an canopy cover, and water depth. Competition between phytoplankton and periphyton for nutrients and light may also reduce the direct effects of nutri ent enrichment (Sandj ensen and Borum 1991). Second, macrophytes mostly use nutrients from sediments. Any nutrient loadings in stream water have to be incorporated into the sediments before being available to the plants. However, nutrient supply can affect plant attributes.

Most studies of nutrient enrichment on plan ts focus on wastewater effluent. Nutrient

enrichment in streams and rivers leads to increasing plant biomass (Chambers and Prepas 1994, Gucker et al. 2006), declines in plan t richness (Thiebaut and Muller 1998, SanJensen et al. 2000), and incr eases in plant tissue phosphorus (Thiebaut an d Muller 2003). Rattray et al. (1991) found that the size and tissue phosphorus content of plants grown on eutrophic sediments was approximately tw ice that of those grown on oligotrophic sediments, indicating that nut rient supply strongly affects plant stoichiometry. Removal of aquatic plant biomass with its accumu lated phosphorus has been proposed as a eutrophication control stra tegy (Thiebaut and Muller 2003). Reduction of nutrient (particularly N) input from municipal wast ewater sources led to macrophyte biomass declines in the Bow River (A lberta) (Sosiak 2002). Other authors have (e.g., Schneider and Melzer 2003) developed plant based troph ic indices (Trophic Index of Macrophytes)

based on the concentrations of soluble react ive phosphorus in both the water column and the sediment pores of streams in Germany.15 4. Indirect biological responses of streams to eutrophication: microbial cycling

In heterogeneous stream reaches, microbial biomass is patchily distributed and is controlled by a number of f actors, including light, subs trate carbon availability, temperature, nutrients, and current velocity. The main components of stream microbial biomass are bacteria and fungi. Fungi dominate large substrates such as leaves and wood, while bacteria dominate fine organic substr ate such as sand (Findley et al. 2002). The organisms are important components of stream food webs and play a key role in carbon cycling.

4.1 Bacteria

Similar to algae, bacteria are also limited by nutrients in aquatic systems, especially in planktonic forms (Cole 1982). Bacteria can outcompete algae for nutrients because of their higher surface ar ea to volume ratio (Fuhs et al. 1972, Currie and Kalff 1984a). Bacterioplankton have substantially highe r phosphorus requirements, higher phosphorus contents, and higher net consumption of phosphorus than phytoplankton (Wetzel 2001). Therefore, bacterioplankton can outcompete algae for phosphorus under a wide range of phosphorus supply rates.

Stream periphytic biofilms are mainly com posed of algae and bacteria. Bacteria can either inhibit algae by outcompeting it when nutrients are limited (Currie and Kalf 1984b, Grover 2000, Biddanda 2001), or they may inte ract positively with algae by using its photosynthetic products and decomposing dead plant and algal biomass and recycling nutrients. A number of studies (Geesey et al. 1978, Hudson et al. 1992, Hepinstall and Fuller 1994, Rier and Stevenson 2001, Carr et al. 2005) have demonstrated positive interactions between algae a nd bacteria in periphyton, t hough caution should be used when interpreting these correlations due to the complexity of the environment (Findlay et al. 1993, Bott 1996). Some (Rier and Stevenson 2002) speculate the pos itive correlation between bacteria and algae in biofilms to be a result of the co-dependence between the two for space.

Nutrient enrichment tends to increase both al gal and bacteria bioma ss (Carr et al. 2005). Sobczak (1996) found that interactions between bacteria and algae are weakened in the presence of a labile source of allocht honous DOC, under extreme light limitation, or under extremely oligotrophic conditions wher e algae are severely nutrient limited. Similarly, Rier & Stevenson (2001) found only a positive statistical relationship between algae and bacteria in streams when streams with periphyton chlorophyll a values greater

than 5 μ g cm − 2 were included in the analysis. Nutr ient addition generally increases the coupling of algae and bacteria biomass (Rie r and Stevenson, 2001, 2002, Carr et al. 2005).

Changes in bacteria taxonomic compositi on and abundance in response to nutrient enrichment are less studied (see Findley et al. 2003, Olapade and Leff 2005). Changes in taxonomic abundance are highly dependent on se asons, level of DOM, and interactions between nutrients and DOM. However, certai n bacteria are sensitive to nutrient16 enrichment and excessive growth on sensitive insects has been proposed as an indicator of nutrient enrichment (Lemly 1998, 2000).

4.2 Fungi

Similar to bacteria, fungi also play an importa nt role in detrital decomposition in streams. Fungal communities in many streams are al so limited by nutrients (Grattan and Suberkropp 2001, Tank and Dodds 2003). This limitation can be released by nutrient additions that lead to significantly highe r fungal biomass. Experimental enrichment increased aquatic hyphomycete conidia in the wa ter of a treated stream by 4 to 7 times

more than the controlled str eams (Gulis and Suberkropp 2004). Other effects observed in the same study included increased number of fungal species detected on each sampling date, and changes in dominance patterns a nd relative abundances of individual species. Nutrient pollution may also change aqua tic hyphomycete diversity and sporulation (Pascoal et al. 2005b).

Bacteria and fungi also compete with each other for nutrients (Gulis and Suberkropp 2003b). Gulis and Suberkropp (2003b) found that fungi inhibit bacterial growth and reduce bacterial biomass by 2-fold at low nutri ent concentrations, sugge sting that nutrient availability may modify microbial interactio ns. Fungi seem to be a superior competitor than bacteria on leaves. Fungal biomass can be one or two order of magnitudes higher than bacterial biomass in polluted streams (Gulis and Suberkropp 2003a, Pascoal et al. 2005a).

5. Indirect biological responses of streams to eutrophication: herbivores

Nutrient enrichment accelerates autotrophic pr oduction and algal biomass in streams, and consequently changes ecosystem structure at other trophic levels. Long-term (16 years) stream fertilization (P addition) in an arct ic stream ecosystem resulted in a dramatic change in the community struct ure with positive response to fertilization at all trophic levels (e.g., increases in epilith ic algal stocks, insect densit ies, and fish growth). Both top-down and bottom-up control of different trop hic levels have been observed in stream ecosystems (Shurin et al. 2002) and has led to development of “trophic cascade” models for streams (Power 1992).

5.1 Invertebrates

A large number of observations and experi mental manipulations have shown that

invertebrates are food limited in stream s (Hill 1992, Rosemond et al. 1993, Biggs and Lowe 1994, Lamerti 1996, Biggs et al. 2000). These studies examined the effect of algal biomass on growth of invertebrates and f ound that invertebrate abundance could be strongly stimulated by increasing algal availabi lity. Nutrient enrichment, in many cases,

did not increase the total biom ass of algae in streams; ra ther the energy was converted

into increased invertebrate densit y through herbivory (Hill et al. 1995).17 Changes in macroinvertebrate compositi on with nutrient enrichment are more complicated than changes of abundance. Ma yflies, a group of inve rtebrates that are

considered sensitive to envi ronmental pollutants, show hi ghest relative abundance when algal biomass is at intermediate levels (Miltner and Rankin 1998). The abundance of scrapers, a functional group that is closely related to grazer s, is highest when nutrient levels are elevated, indicating positive effects from increased algal av ailability (Miltner and Rankin 1998). Similarly, scrapers (e.g., Ancylus fluviatilis ) and detritivores (e.g., Oligochaeta, Lumbriculidae) have shown significant increa ses in density or biomass on

certain substrata with enrichment even while total macroinvertebrate density or biomass did not (Sabater et al. 2005) . Benthic invertebrate composition has also been shown to shift from chironomid-amphipod to an oligoc haete-gastropod domina ted assemblage in response to decreases in DIN and changes in benthic algal abundance and sediment organic carbon concentrati ons (Chambers et al. 2006).

Enrichment may also alter benthic habitat for macroinvertebrates. In addition to food sources for invertebrates, benthic algae, espe cially macroalgae, are important habitat for macroinvertebrates (Dudley et al. 1986). Some algal species or growth forms are grazerresistant (e.g., Oedogonium spp.) and are good habitat for many invertebrates (Steinman et al. 1992). For example, long-term additions of P in an arctic stream resulted in an increase in moss density. The increase in phys ical habitat associated with this shift was responsible for changes in the macroinve rtebrate community (Lee and Hershey 2000).

5.2 Fish

Fish may benefit from increases in food av ailability when nutrient additions increase primary and secondary production. Enrichment of oligotrophic streams and rivers may result in increased algal biomass, increased benthic invertebrates, and fishes. Long-term enrichment studies in Vancouver Island, Br itish Columbia (e.g., Slaney and Ward 1993, Slaney and Ashley 1997) repor ted increased fish size up to 2 times. Similar studies (Deegan and Peterson 1992, Slavik 2004) also observed increased fish growth rates in long-term fertilization studies in rivers. However, it is diffi cult to assess how much of the primary production from nutrient enrichment flows through to fish, and the effect of nutrient addition on fish is largely unpredic table. van Dam et al. (2002) estimate that periphyton ingestion by many herbivorous a nd omnivorous fish ranges from 0.24 to 112 mg (3dmg fish) -1 d -1 in a fishpond.

One of the consequences of nutrient enrichment may be loss of sensitive fish taxa and increases in tolerant taxa. The strong co rrelation between fish metrics and nutrient pollution (Miltner and Rankin 1998) indicates that nutrient enrichment has contributed to changes in the structure of fish assemblages. While nutrient enrich ment could potentially benefit fish production in the short term, th e ecological consequence of nutrient addition could have severe impacts on stream eco systems (Stockner et al. 2000). The obvious impact at high nutrient loads is the reducti on in DO, which would exclude many sensitive taxa. In addition, excess algal growth would eliminate important feeding and respiration habitat, further reducing su rvivorship. While it is eviden t that some nutrient subsidy18 benefits the growth of select species, the overall impact is nega tive, especially at stressful

nutrient levels.

5.3 Food web structure

Carpenter et al. (1985) adapte d the “cascade of effects” concept developed from topdown marine food webs to lakes and descri bed it as “cascading trophic interactions”. They posited that each trophic level is cont rolled by both predator s (top-down control) and resources (bottom-up control). Changes at one trophic level would alter material cycling and other trophic levels in the food web (trophic casca ding). Several authors have discussed trophic cascading in stre ams (e.g., Lamberti 1996, Biggs 2000).

Long-term fertilization studies have demons trated the cascading effect of nutrient enrichment at several trophic levels. Huntsm an (1948) first recogni zed that fertilizers stimulate downstream algal growth, and lead to increased insect and fish densities. Since then, more quantitative studies (Peterson et al. 1993, Slaney and Ashley 1998) have shown that nutrient additions increase alga l biomass at least at the beginning of the enrichment. Later, top-down forces take effect to control primary consumers and consequently algal biomass. Generally, grazi ng has demonstrated a larger effect than resource limitation in influencing al gal biomass and composition (Steiman 1996, Lamberti 1996, Flecker et al. 2002). However, interactions among different trophic levels could be regulated by many factors, which so metimes lead to unexpected responses to nutrient additions at higher tr ophic levels (Deegan et al. 1997). In either case, plantherbivore interactions are considered central to food web structure and energy flow in aquatic ecosystems (Lamberti 1996).

While nutrient additions affect higher trophic levels, predators also play an important role in influencing nutrient demand and nutrient supply. Flecker et al. (2002) examined nutrient limitation in the presence and absen ce of fishes and found that the response to nitrogen enrichment is significantly greater on substrates accessi ble to natural fish assemblages compared to substrates where grazing fishes are excluded. Many experiments (e.g., Biggs 2000) de monstrate simultaneous and in teractive effects of topdown and bottom-up factors in limiti ng primary producers in streams.

6. Responses of ecosystem functi on to nutrient enrichment

6.1 Primary production and respiration

Production- Three empirical models have been de veloped to simulate algal nutrient uptake kinetics: the Michaelis-Menten mode l, the Monod model, and the Droop model (Borchardt 1996). All three models are sim ilar to each other in form. The MichaelisMenten model emphasizes nutrient uptake kinetics (Dugdale 1967). The Monod model examines algal specific growth rate along external nutrient con centration (Monod 1950), while the Droop model takes into account in tracellular concentration of the limiting nutrient (Droop 1968). These models postulate that algae accrual follows a logistic growth rate and reaches maximum saturation at some concentration.19 The Monod equation has been applied in a numbe r of studies to inve stigate the nutrientalgal production relationship (Rier and Stev enson 2006). The equation is as follows:

μ = μ max * S/(K μ +S),

where μ is the specific growth rate, μ max is the maximum specific growth rate, K μ is the half-saturation constant for growth for the lim iting nutrient, and S is the external nutrient concentration ( μ mole/L).

Experimental approaches have confirmed th at periphytic algal growth increases along nutrient gradients (Horner et al. 1983, Bothwell 1989, Walt on et al. 1995, Rier and Stevenson 2006, Stevenson et al. 2006). Howe ver, while algal biomass accumulation is a reflection of algal production, maximum al gal production does not lead to maximum areal biomass. Bothwell (1989) examined bot h cellular growth rates and maximum areal biomass, and demonstrated that the amount of P needed to saturate cellular growth rates during early stages of colonization was two orders of magnitude lower than concentrations needed to produce maximum areal biomass. Rier and Stevenson (2006) manipulated nutrient concentrations to ex amine the effect on the growth rate of periphyton in artificial streams, and found that saturating concentrations for algal growth

rates were 3 to 5 times lower than concentr ations needed to produce maximum biomass. These growth patterns can be fit in to modified Monod models, demonstrating the exponential growth of algal biom ass along nutrient gradients.

Correlation analysis between nutrient con centrations and algal biomass and algal productivity in streams also indicate that nutrient enrichment e nhances algal production (Dodds et al. 2002, see review by Dodds 2006) . Dodds (2006) plotted algal production against biomass using Bott et al.’s (1985) co mpiled data from the literature and found a positive correlation between algal primary production and algal biomass in streams, though only 24% of the total variance was expl ained. Field observati ons are consistent with experimental manipulations of nutrien t enrichment and algal production. In addition to algal production, macrophyte production is also stimulated by nutrient enrichment (Gucher and Brauns 2006).

Respiration- It is generally accepted that higher production leads to higher respiration. Dodds (2006) demonstrated that production and respiration by algae in streams are strongly correlated with each other with a correlation coeffi cient close to 0.90. Similarly,

nutrient enrichment increases bacterial and fungal production, which in turn increases aerobic respiration from these decomposers. Thus, the tota l respiration in enriched streams is higher than production. High resp iration:production ratios lead to high biological oxygen demand (BOD) in downstrea m water (in nutrient enriched streams). Thus, nutrient enrichment will generally lead to elevated gross primary production and whole-stream community respiration, a nd decreased stream DO concentrations (Chambers and Prepas 1994, Gucker and Brauns 2006).

A good example is blackwater streams in the eas tern United States. Studies of nutrient loading on phytoplankton, bacterioplankton, and respiration in blackwater streams indicate that changes in nut rient loading stimulate two di fferent biological pathways20 (photosynthetic and heterotrophic activity) (Mallin et al. 2004). Nitrogen additions increase chlorophyll a production and significantly stimulate BOD. Combined organicinorganic phosphorus additions significantly stimulate bacterial abundance, ATP, and

BOD on most occasions. In blackwater stream s, nitrogen inputs stimulate phytoplankton growth, which in turn dies and decomposes in deeper, higher order streams, becoming a source of BOD and lowering DO. Phosphorus i nputs directly stimulate bacterial growth, increase BOD, and lower stream DO concentr ations. In some circumstances, hypoxia is expected (Mallin et al. 2006).

6.2 Secondary production and predators

Generally, secondary producers (consumers) in streams are food limited, and their production and biomass are expected to in crease in streams with higher primary production due to N and P enrichment. Elw ood et al. (1981) found that the initial response to stream nutrient enrichment in a Tennessee woodland stream was increased algal standing crop, which was then consumed by a large increase in grazer abundances.

Continuous enrichment of P-limited stream s on Vancouver Island, British Columbia led to substantial increases in secondary produ cers (Perrin et al. 1 987, Slaney and Ward 1993). Benthic invertebrate biomass increased by from 2 to 7-fold and fish size by from 1.4 to 2-fold (Slaney and Ward 1993). Macr oinvertebrates also ingest bacterial production (Fuller et al. 2004). Be nthic invertebrate biomass increased 4.5-fold and fish (cutthroat trout) increased 6.3-fo ld in a stream enriched by ca rbon that elevated bacterial biomass (Warren et al. 1964). In another study, enrichment did not cause a general increase in macroinvertebrate density or bi omass, but altered assemblage composition in the enriched reach (e.g., Sabater et al. 2005).

Secondary production clearly responds to enri chment and the response may be more clear in oligotrophic than eutrophic streams. Fis h, especially herbivores, can obviously benefit from increasing biomass of primary producer s. Carnivorous or omnivorous vertebrates also prosper from the increased bioma ss of primary consumers (e.g., Deegan and Peterson 1992). A recent study (Johnson a nd Wallace 2005, Johnson et al. 2006) on the growth of salamander larvae, top predators in southern Appalachian streams, indicated

that nutrient enrichment stimulates sala mander larval production by increasing detrital quality and quantity. Although some enrichment may benefit fish production, a transition region in enrichment from beneficial to detr imental effects has not been defined to the

extent that it has for lakes and reservoirs (Welch 1992). It probably exists for different physical types of streams and rivers. Tw o recent stream studies have provided independent estimates of target nutrient concen trations that should be maintained in order to ensure acceptable water quality needed for fish growth if fish growth is the primary concern (Stockner et al. 2001, Compton et al. 2005).21 6.3 Decomposition rates

Microbial decomposition rates are affected by a number of environmental factors, i.e., current velocity, feeding by detritivores, sedimentation, and nutrient enrichment. Microbial decomposition rates are strongly affected by funga l and bacterial productivity

and their relative contributions to total mi crobial biomass (Hieber and Gessner 2002). Findlay et al. (2002) found that bacterial a bundance was higher on fine particles, while fungal biomass was significantly greater on larger particle size classes, i.e., on leaves and

wood. The relative contribution of reach-sca le fungal biomass ranged from 10 to 90% of microbial biomass in this study depending on the quantity of leaves and small wood in

the streams. In streams with abundant leaves , fungi dominate the total microbial biomass (98.4 to 99.8%) and cumulative producti on (97.3 and 96.5%) (Gulis and Suberkropp 2003a). Nutrient enrichment generally incr eases bacterial and fungal productivity and biomass, but does not change the roles of each assemblage. Gulis and Suberkropp (2003a) found that the fungal yield coefficient exceeded that of bacteria by a factor of 36 and 27 in low- and high-nutrient tr eatments, respectively.

Nutrient enrichment significantly increases decomposition rate, microbial respiration, fungal and bacterial biomass, and the sporula tion rate of aquatic hyphomycetes associated with decomposing leaf mate rial (Gulis and Suberkropp 2003c). Comparisons between one reference site and two downstream polluted sites with high nutrient concentrations in the Ave River, Portugal indicated that bacter ial production was greater at the two polluted sites than at the reference sites, wh ile highest fungal biomass and production corresponded to the fastest leaf breakdown. Ot her studies also support the contribution of fungi and bacteria to leaf litter breakdown and the enrichment of microbial activity by nutrients (Suberkropp 1995, Weyers and Suberkropp 1996, Royer and Minshall 2001, Gulis and Suberkropp 2003a, b, c).

In addition to microbial biomass in streams, decomposition rates can also be regulated by other factors, especially th e presence of detritivores. S ponseller and Benfield (2001) found that shredder presence and abundance was critical to leaf breakdown rate in Appalachian headwater streams. Increased sedi mentation from agricult ural input in some streams may limit the distribution of shredde rs and thus influence leaf breakdown in these streams. Similarly, Pascoal et al. (2005a) attributed low leaf breakdown rates to low shredder densities in a polluted river.

6.4 Other functional responses

Carbon cycling -Carbon is not considered limiting for periphyton growth in streams

because CO 2 can be dissolved in water and provides sufficient inorganic carbon for use in algal production. However, bact eria are limited by carbon in many streams with relatively low DOC inputs. Increases in primary producti on due to nutrient enrichment increase DOC and the coupling of periphyton and bacteria since bacteria will rely more heavily on organic carbon sources produced by algae a nd other plants. Micr obial decomposition further increases dissolved organic carbon sources. Thus, nutrient enrichment will enhance carbon cycling, especia lly in slow moving streams.22 Nitrogen fixation - Nitrogen fixation by algae and bact eria may be affected by nutrient additions and changes in N:P ratios. N itrogen fixation by cyanobacteria and other microbes is favored by slow moving or stan ding water, high temper atures, low DIN and high DON (Dodds 1995). Nitrogen fixation can pr ovide a substantial amount of N to a stream depleted of nitrogen (Grimm and Petrone 1997). Although not much information

is available about nutrient enrichment on N fixa tion in streams, studies in lakes indicate that decreases in the N:P ratio will generally lead to increases in N fixation rates by cyanobacteria (Smith 1990, Hendzel et al. 1994 ). Grimm and Petrone (1997) found that biomass specific N2 fixation was positively co rrelated with temperature and light, and negatively correlated with di ssolved inorganic nitrogen.

Alkaline phosphatase activity –Aquatic primary producers, es pecially algae, can excrete an enzyme called alkaline phosphata se in response to P defici ency in the water. Various studies (e.g., Peck et al. 2006) have used the concentration of alkaline phosphatase in the water column and in periphyton as an indicator of P limitation. It is considered the most precise indicator of P limitation over epilithic N:P ratios and algal growth on nutrientdiffusing substrates (Bowman et al. 2005). It is speculated that phosphatase levels of approximately 0.003 mmol/mg chl per hour in dicate moderate P limitation, whereas phosphatase levels above 0.005 mmol/mg chl per hour indicate severe P limitation (Steinman and Mulholland 1996).

Silica limitation - Silica limitation is also reported in some streams. Silica is an important component of diatom frustules. A study in eutrophic Lake Okeechobee, Florida indicated that along with N limitation, silica also lim ited microalgal assemblages dominated by diatoms (Zimba 1998). In streams dominated by diatoms, silica concentrations tended to decrease downstream, implying that diatoms absorb silica from the water (Wall et al. 1998). In some cases, nutrient enrichment gene rally increases the pr oduction rate of green filamentous algae while diatom production decr eases. It is possible that diatoms are limited by silica concentrations in these streams. The extent of this limitation is poorly

understood in streams.

6.5. Nutrient enrichment impa cts on designated uses

Nutrient enrichment may directly lead to ex cessive increases in al gae and other primary producers. The indirect effects of enrichment are to increase stream consumer production, change food web structures, and consequently alter ecosystem function. The deleterious impacts of these structural and functional changes may impair aquatic life uses, limit drinking water resources, and degrade recreati onal and aesthetic uses of waters. Various impacts of nutrient enrichment in streams have been summarized in Table 2 below.23 Table 2. Effect of nutrient enrichment on designated uses of streams.
Excessive nutrient levels allow excessive in creases in algae and other primary producers, which, in turn, prevent streams from meeting their designated uses. Th e adverse effects of either high nutrient levels or the nuisan ce growth of primary producers include, for example:

• Impairment of aquatic life uses -- Diel fluctuation of oxygen concentrations and pH values in streams may negatively impact aquatic life (fish and invertebrates). -- Ammonia toxicity (e.g., at a high level > 1 mg/L) may be derived from high nitrogen

concentrations. -- Overgrowth of many algae can lead to alga l blooms, several toxins from which have been found to kill fish and other aquatic life. -- Indirect effect of excess growth on physical habitat availability. -- Enrichment may lead to loss of divers ity and native taxa, changes in biological community structures (algae, a quatic plant, invertebrate, and fish), and eventually loss of ecosystem function.

• Negative impact on water resources -- High nitrate concentration (>10 mg /L) is toxic in drinking water. -- 2-methylisoborneol and geosmin are two ch emicals produced by cyanobacteria that can cause taste and odor problems in drinking water. -- Diatoms and filamentous algae can clog inta ke screens and filters in water treatment plants. -- Decay of algae may lead to taste and odor problems of drinking water. -- Potentially carcinogenic disi nfection by-products (trihalo methanes, THMs) may form during treatment of drinking water from eutrophic waters.

• Degradation of aesthetic and recreational uses -- Many algae can form large clogs of mats either floating ( Lyngbya, Oscillatoria, Hydrodictyon, Spirogyra ) or attached (e.g., Cladophora and Ulothrix ) to substrata, which are unappealing to swimming, fishing, and boating. -- Large odorous algal masses die off and d ecay after the growing season, and their presence inhibits recreation and human health. -- Cyanobacteria, such as Oscillatoria or Lyngbya wollei , can produce chemical compounds that irritate swimmers. -- Fish grown in waters with algal-derive d chemicals can suffer impacts on flavor. -- Skin rashes, nasal irritation, or other health effects may result from skin contact with algal toxins. -- Slippery streams are a threat to swimmers.

-- Cladophora abundance may slow water flow in canals and inhibit navigability.24 7. Control of eutrophication: best management practices

Nutrient enrichment poses serious threat s to stream ecosystems. Managing nutrient loading into streams will reduce not only th e magnitude of maximum algal biomass, but also the frequency and duration of benthic algal problems in streams (Biggs 2000). To better protect and restore streams, control of both point and nonpoint sources of nutrient loadings into streams is essential. Currentl y, many states have proposed plans to develop

nutrient criteria to control nutri ent loading into streams. Contro l of point sources, such as treated wastewater, can be improved with new technology (Arheimer et al. 2004). Still a persistent problem for implementation of criteria will be control of nonpoint sources. It will require innovative technologi es and better understanding of stream ecosystems to decrease nutrient loadings from nonpoint sources into streams. Best management practices should be implemented including riparian buffer and wetland protection, and smart use of fertilizers in ag ricultural and silviculture. Ne w technologies are contributing to some improvement in nutrient pollution fr om nonpoint sources (A rheimer et al. 2004). More cost-effective practices should be developed to better fulfill this goal.25 Literature Cited

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