Producers—Stevenson and Pan (1999) reviewed the uses of diatoms for assessing environmental conditions in rivers and streams. They traced the use of species compositions of algae to infer amount of pollution to work by Kolkwitz and Marsson in the early 1900s, with substantial contributions by Ruth Patrick in the 1940s and 1950s (as cited by Stevenson and Pan 1999). Studies that use algal assemblages as indicators of the extent of pollution rely on the concept that predictable species shifts occur with set amounts of enrichment (e.g., Kelly 2002). Detailed work has been carried out relating nutrients to diatom and other algal assemblages in several places, mostly in temperate, developed countries. The green alga Cladophora has often been associated with eutrophication events (Hynes 1960) and is ubiquitous in nutrient- rich flowing waters (Dodds and Gudder 1992). Large streamers of Cladophora develop under nutrient-rich conditions. These streamers potentially lead to low O2 events at night, alter the community structure, snag fish lures, slow water flow in canals, and clog industrial and domestic water intakes (Dodds and Gudder 1992).
One of the problems with predicting eutrophication effects in streams is that variability caused by flooding can influence autotrophic state. At one extreme, algal biomass might not accrue with ample light and nutrients if floods always scour biomass. On the other end of the spectrum, attached algae might be able to attain impressive biomass in nutrient-poor water because periphyton can use the small amounts of nutrients that continuously flow by. Biggs (2000) developed a comprehensive model linking hydrologic regime and nutrients to accrual of algal biomass. This model was developed with a database from New Zealand rivers and streams across a wide range of land use practices and hydrologic patterns. Regressions considering only dissolved inorganic nutrients could predict algal biomass with r2 values of approximately 30%. Consideration of the time of accrual (time since the last scouring flood) increased r2 values to about 70%. The work of Biggs (2000) supports the proposition that eutrophication effects will be stronger under stable flow regimes. The effects of eutrophication on macrophytes in flowing waters have been poorly studied, and the effects of nutrient reductions on macrophyte biomass are difficult to predict (Chambers et al. 1999). Biomass of macrophytes declined in the Bow River (Alberta) in response to nutrient control (particularly N) from municipal wastewater sources (Sosiak 2002). Sewage effluent led to substantially greater macrophyte biomass in the Saskatchewan River (Saskatchewan), and this was correlated with somewhat decreased dissolved O2 concentration (Chambers and Prepas 1994). In some rivers and streams with reduced water replacement times, phytoplankton blooms can become problematic, with cyanobacterial blooms more likely in excess-nutrient conditions (Smith 2003). Shorter water turnover time (hydraulic residence time) leads to a decreased amount of suspended chlorophyll per unit concentration of P (Søballe and Kimmel 1987). Problems occur with phytoplankton blooms in European and other rivers around the world (Wehr and Descy 1998). In the Murray Darling river system in South Australia, water withdrawals reduce flow to a near standstill in the river, and excess amounts of nutrients, stratification, and warm temperature stimulate algal blooms (Maier et al. 2001). These blooms are commonly dominated by the hepatotoxic Microcystis. Other slow-flowing rivers in the world suffer a similar fate, particularly those with limited quantities of light-intercepting fine sediments.
Microbial heterotrophs—Although enrichment experiments have documented that rates of microbial heterotrophic processing of organic materials can be stimulated by nutrients (as previously discussed), less is known about influences on the heterotrophic microbial community. If the primary source of organic C to a stream or river is leaf material, N and P need to be obtained from the water column, and nutrient enrichment will increase C utilization rates. One study documented that nutrient enrichment causes shifts in fungal taxa associated with decomposing leaf litter (Gulis and Suberkropp 2002). Presumably, some bacteria that decompose organic matter are better competitors for organic nutrients than others, leading to shifts in community structure in response to nutrient enrichment. Future studies are likely to document this effect, given the recent expansion of molecular techniques. Clear increases in the rates of heterotrophic microbial biogeochemical cycling (denitrification) related to nutrient enrichment by agricultural practices have been demonstrated (Kemp and Dodds 2002).
Food web effects—Effects of C and, particularly, N and P loading on animals in streams are less clear. The effects of C on the animal community are obvious, with greater rates of organic C loading leading to dominance by pollution- tolerant invertebrates (such as Tubifex, Limnodrilus, Chironomus), decreases in diversity, and increases in raw abundance (Hynes 1960). With the advent of BOD treatment in sewage and industrial effluents in developed nations, less attention has been paid to the effects of BOD loading. Enrichment effects related to N and P are less well established. Macroinvertebrate assemblage structure has been correlated statistically with P concentration (Miltner and Rankin 1998). Nutrient enrichment can cause increases in invertebrate abundance and alters assemblage structure (Bourassa and Cattaneo 1998). The clearest study to date on the importance of sustained nutrient loading to the food web occurred on the Lawrence River downstream of Montreal, Quebec. This study used the distinctive isotopic signal of 15N to establish that nutrients from the sewage outfall significantly enriched macroinvertebrates and production of both macroinvertebrates and fishes (deBruyn et al. 2003). The sewage was treated for BOD, but stimulated secondary production over fivefold in spite of the small amount of N and P that entered the food web in the sewage plume 10 km down from the sewage outfall.
Control of cultural eutrophication—Given the definition of the trophic state proposed, and the potential effects of autotrophic and heterotrophic eutrophication, what considerations are important in controlling eutrophication? Mechanistic methods are only beginning to be established for linking in-stream nutrient concentrations to watershed activities. Empirical methods have prevailed (e.g., Dodds et al. 1997) until recently. Modeling efforts are beginning to refine nutrient concentration and loading estimates for rivers, but there still is some difficulty in linking models created for small streams with larger river systems (e.g., Alexander et al. 2002). Ultimately, linking land use practices, including both point and nonpoint sources of nutrients, to instream nutrient concentrations will be necessary to control cultural eutrophication that influences autotrophic state, and potentially influences heterotrophic state.
Nutrient control is, on one level, simple. Agricultural practices, atmospheric loading, and human sewage outfall increase inorganic and organic nutrients in rivers and streams. Technology is available to decrease that input (but nonpoint sources of nutrients such as atmospheric deposition and runoff from cropland remain difficult to control). Best management practices of cropland include riparian buffer strips, cropland terracing, and the use of only the necessary amounts of fertilizer. Effluent from human sewage and livestock- handling facilities can be treated with existing tertiary treatment methods (e.g., denitrification facilities, P precipitation) to reduce N and P loads to lotic waters. The effective reduction of BOD into the waters of most developed countries exemplifies the technical ability of water treatment engineers and managers to remove potentially harmful pollutants at acceptable costs. The challenge now is to determine what lengths are necessary to control point and nonpoint source pollution, and to what degree the benefits of nutrient control justify the costs. Determining the reference trophic state provides a starting point for cost–benefit and feasibility analyses of eutrophication control schemes. Nutrient cycles do not occur in isolation, and colimitation of algal and heterotrophic activity is commonly seen in bioassays (Tank and Dodds 2003). We are only beginning to understand the implications of the effects of humans on the stoichiometry of nutrient loading (Turner 2002). Stoichiometric changes could alter algal assemblages and relative rates of material flux (e.g., Woodruff et al. 1999). Changes in stoichiometry could then cascade to higher trophic levels (Frost et al. 2002).
Given the broad definition of eutrophication presented herein, organic C enrichment should be considered, as well as anthropogenic processes causing shifts in the relative heterotrophic and autotrophic states. For example, increased BOD from sewage has definite influences on stream heterotrophic state. In addition, shifts in riparian vegetation, such as loss of riparian forests, might increase the autotrophic state and decrease the heterotrophic state. In systems such as tallgrass prairies, historically dominated by little riparian vegetation, increases in riparian vegetation could alter the fundamental ecosystem and community structure (Dodds et al. 2004). Finally, organic C enrichment might interact with N and P enrichment. The highest rates of C consumption and the greatest biomass of heterotrophic organisms are expected when loading of N, P, and C are simultaneously high. Water retention times might alter nutrient stoichiometry and heterotrophic and autotrophic states by influencing deposition and nutrient processing rates. Small and large impoundments that were not historically present are now a ubiquitous feature on many river networks. Such impoundments could also alter the balance between heterotrophic and autotrophic states because many recalcitrant C-rich particulate organic materials can settle in the reservoir, and plankton with relatively low values of C:N and C: P could dominate reservoir tail waters (Whiles and Dodds 2002). Humans will affect ever more river miles with hydrologic modification, alter the inputs of organic C and its form to lotic waters through alteration of riparian vegetation and input of BOD in sewage from humans and livestock. Increased fertilizer to grow the crops necessary to feed an expanding human population and increases in industrial livestock operations resulting in vast production of animal waste will cause further eutrophication of already affected rivers and streams. These effects will continue to spread into the few relatively pristine watersheds that remain on earth, altering water quality and influencing the biotic integrity of these waters. Understanding the full implications of these effects will require further knowledge of the native trophic state of streams as a baseline. More complete comprehension of how nutrient interactions influence trophic state, and determination of trophic states of medium to large rivers will improve the scientific basis for managing eutrophication of lotic waters.