Controlling Eutrophication along the Freshwater–Marine Continuum: Dual Nutrient (N and P) Reductions are Essential
- 7.9k Downloads
Expanding human activities along the freshwater to marine continuum of coastal watersheds increasingly impact nutrient inputs, nutrient limitation of primary production, and efforts to reduce nutrient over-enrichment and eutrophication. Historically, phosphorus (P) has been the priority nutrient controlling upstream freshwater productivity, whereas nitrogen (N) limitation has characterized coastal waters. However, changing anthropogenic activities have caused imbalances in N and P loading, making it difficult to control eutrophication by reducing only one nutrient. Furthermore, upstream nutrient reduction controls can impact downstream nutrient limitation characteristics. Recently, it was suggested that only reducing P will effectively control eutrophication in both freshwater and coastal ecosystems. However, controls on production and nutrient cycling in estuarine and coastal systems are physically and chemically distinct from those in freshwater counterparts, and upstream nutrient management actions (exclusive P controls) have exacerbated N-limited downstream eutrophication. Controls on both nutrients are needed for long-term management of eutrophication along the continuum.
KeywordsNutrient limitation Nitrogen Phosphorus Eutrophication Freshwater to marine continuum Water quality
Human population growth, urbanization, and agricultural and industrial expansion are causing unprecedented and alarming rates of nutrient over-enrichment and accelerated plant growth in receiving waters worldwide. Key nutrients of concern are nitrogen (N) and phosphorus (P) because the supply rates of these nutrients most often control or “limit” aquatic plant primary production and biomass formation. Maintaining fertility at the base of the food web is essential for productive, resourceful, and healthy aquatic ecosystems. However, nutrient over-enrichment promotes accelerated production of plant-based organic matter (i.e., eutrophication) to the extent that excessive production, including harmful algal blooms, fuels expanding zones of bottom water hypoxia (dead zones), and leads to fisheries habitat destruction, translating into ecological and economic losses of impacted waters (Nixon 1995; Boesch et al. 2001; Rabalais and Turner 2001; Diaz and Rosenberg 2008).
In some aquatic ecosystems, biologically available N (as NH3) can be supplied endogenously through microbial nitrogen (N2) fixation. This capability has been used to support the argument that P will ultimately be the limiting nutrient, since N2 fixation represents an avenue for helping satisfy N requirements under these circumstances (Doremus 1982; Tyrell 1999). Moreover, inorganic and organic forms of N are often more soluble and hence more available than P (Gunnars and Blomqvist 1997; Wetzel 2001), potentially exacerbating P limitation. Indeed, there are numerous examples of freshwater ecosystems in which P input controls have effectively controlled primary production and successfully arrested the unwanted symptoms of eutrophication (Likens 1972; Smith 1990, 2003). However, there are also examples of large lake systems (especially those with relatively small watersheds) in which N plays a key role as a limiting nutrient (Goldman 1988; Lewis and Wurtzbaugh 2008).
Downstream estuarine and coastal waters are physically, chemically, and biologically distinct from freshwater ecosystems and, as a result, their responses to nutrient inputs and over-enrichment can contrast those observed in freshwater ecosystems (Paerl 1997; Smith 2003; Howarth and Marino 2006; Bianchi 2007). Physically, water supplies and circulation are controlled by the interactive effects of basin morphology, freshwater discharge, and tidal mixing and, as a result, hydrodynamic properties of estuaries and coastal systems, including advection and stratification, flushing, and residence times, can differ markedly from freshwater systems. Geochemically, estuarine and coastal ecosystems are rich repositories of minerals derived from watershed geological (erosional) and biological (plant production, microbial and higher trophic level cycling) processes. As discussed below, these differences strongly influence the types and spatiotemporal extent of nutrient limitation, and they form the rationale for reducing both N and P inputs in order to control estuarine and coastal eutrophication.
The Rationale for a Dual Nutrient Strategy
When considering the need for nutrient controls on estuarine and coastal eutrophication, it is important to recognize that there are major distinctions between N and P cycling as these nutrients are transported to and transformed in estuarine and coastal systems. The N cycle contains diverse and abundant gaseous forms in addition to dissolved and particulate forms, while the P cycle is dominated by non-gaseous dissolved and particulate forms. This means that fractions of the N pool can exchange with and escape to the atmosphere, while P is essentially “trapped” in receiving marine waters. Processes controlling N exchange with the atmosphere include ammonification, denitrification, nitrous and nitric oxide production, products of the anammox reaction, and nitrogen (N2) fixation (Codispoti et al. 2001; Rich et al. 2008). Of these processes, the first five represent “losses”, while the latter process represents a “gain” in terms of influx and efflux of N in marine waters. No analogous air–water exchanges exist in the P cycle. The net effect of the microbially mediated dissolved gaseous fluxes on N is loss of N to the atmosphere, while P remains in the system, either as dissolved or as particulate forms.
In a recent study, based on 37 years of data from whole lake (Lake 227; experimental lakes area, Northwestern Ontario, Canada) experimental nutrient addition work, Schindler et al. (2008) proposed that “eutrophication of lakes cannot be controlled by reducing nitrogen input”. They also expanded this argument to estuarine and coastal waters. The Lake 227 data set is among the most extensive and lengthiest in existence, and it has yielded highly significant insights into lake responses to nutrient manipulations (Schindler et al. 1973). However, the conclusions put forth in this paper are based on a somewhat narrow perspective on biogeochemical cycling and eutrophication dynamics and a lack of consideration of the physical–chemical and biological constraints on the ability of N2-fixing microorganisms to meet ecosystem-level nitrogen requirements along the freshwater to marine continuum. Lastly, remote Lake 227 has not been significantly impacted by anthropogenic nutrient enrichment, other than experimental manipulations. In contrast to this isolated “lake in a bubble”, many, if not a great majority, of other freshwater, estuarine, and coastal ecosystems have been radically altered by human “nutrification” over the past century (c.f. Likens 1972; D’Elia et al. 1986; Paerl 1988; Nixon 1995; Vitousek et al. 1997; Boesch et al. 2001), to the extent that they behave differently from the highly controlled Lake 227. While Lake 227 has, over the past few decades, received continual doses of P (but not N), many coastal systems have experienced ever-increasing N loads from rapidly growing human sources, with severe negative impacts on ecosystem structure and function (Boesch et al. 2001; Diaz and Rosenberg 2008; Paerl and Piehler 2008). This has led to a call for N input constraints, which is warranted and scientifically defensible for the reasons provided below.
In their studies, which were initiated in 1969, Schindler at al. (2008) initially administered N and P and conclusively showed that fertilization with these nutrients greatly enhanced primary production and algal biomass (Schindler et al. 1973). The N/P fertilization ratio (by weight) was ∼12. Dual nutrient fertilization was continued until the mid to late 1980s, when N fertilization was initially reduced and then discontinued while P fertilization was maintained. This led to a change in the N/P loading ratio of less than 5 once N loading was discontinued. Under this final scenario, nutrient stoichiometry conditions became increasingly favorable for N2 fixing cyanobacterial populations to develop (c.f. Smith 1990). Indeed, given the lake’s highly favorable N/P ratio, as well as its physically favorable conditions for diazotrophic cyanobacterial growth and bloom potentials (strong vertical stratification, high degree of water column stability, leading to consistent upper water column heating during the summer months) (Paerl et al. 2001), N2-fixing cyanobacterial blooms predictively developed (Schindler et al. 2008). In this physically favorable environment, the maintenance of chemically favorable conditions (low N/P ratios) has resulted in the continuation of the summer blooms.
Schindler et al. (2008) have extrapolated these N2 fixation responses within this chemically and physically ideal lake ecosystem to freshwater systems in general. They also argue that similar biotic responses can be expected in estuarine and coastal systems (i.e., N2 fixation will assume a greater share of N input) without presenting supporting experimental documentation. The “take home message” from Schindler et al. (2008) is that, because N2 fixation can respond to meet ecosystem N requirements in a regime of P enrichment, P ultimately controls eutrophication and the need for N input controls is obviated.
However, this argument for single (P) nutrient input reductions to control eutrophication is idealized and conceptually and technically inapplicable to many freshwater and marine ecosystems. While Lake 227 is far removed from anthropogenically enhanced N and P loading, nutrient loading dynamics in most regions have changed radically over the past four decades (Vitousek et al. 1997; Boesch et al. 2001), with dramatic impacts on the structure and function of downstream ecosystems (Bricker et al. 1999; Elmgren and Larsson 2001; Diaz and Rosenberg 2008; Duarte et al. 2008). In response to the experimental evidence starting in the 1960s that P over-enrichment caused accelerated eutrophication in many lakes and reservoirs (Likens 1972; Schindler et al. 1973), P loading was restricted in many regions, starting with P removal in point source wastewater discharges, with notable desirable results (improvement of water quality on the North American Great Lakes, Lake Washington, Lake Erken, Sweden, Lake Geneva, Switzerland, Lake Biwa, and Japan). In contrast, N loading has increased in a more unabated fashion, in large part because N was not broadly identified as a limiting nutrient in freshwater ecosystems. However, studies in downstream estuarine and coastal ecosystems, beginning in the 1970s, showed many of these systems to be N-limited (Ryther and Dunstan 1971; Hobbie 1974; Hobbie and Smith 1975; D’Elia et al. 1986; Fisher et al. 1988; Elmgren and Larsson 2001). Furthermore, these studies identified the potential eutrophying threat that accelerating N loading could pose (Ryther and Dunstan 1971; D’Elia et al. 1986), a prediction applicable to coastal systems worldwide, where increased frequencies and geographic expansion of harmful algal blooms, hypoxia, habitat degradation, and fisheries losses have been closely linked to increases in N loading (c.f. Paerl 1988, 1997; Boesch et al. 2001; Rabalais and Turner 2001; Diaz and Rosenberg 2008).
There is irrefutable evidence that anthropogenically driven N loading events in estuarine and coastal waters have accelerated primary production, ultimately translating into organic matter enrichment with major negative biogeochemical consequences, including algal blooms, hypoxia, and altered food web structure and function (c.f. Nixon 1995; Boesch et al. 2001). Two widely cited examples include the large spring phytoplankton blooms in Chesapeake Bay (Fisher et al. 1999; Boynton and Kemp 2000; Harding et al. 2002) and the Mississippi discharge region of the northern Gulf of Mexico (Lohrenz et al. 1997; Rabalais 2002) that are largely the result of N-enriched “freshets”. However, N-driven eutrophication, caused by either expanding diffuse nonpoint and/or point source discharges, has also been thoroughly documented in many other geographically diverse estuarine and coastal systems, including regions of the North Sea; Mediterranean Sea and Baltic Sea (European Environmental Agency 2001; OSPAR 2008; Elmgren and Larsson 2001); Danish and Swedish fjords (Conley 2000; Elmgren and Larsson 2001); the Po River Delta (North Adriatic Sea) (Vollenweider et al. 1992); the Nile River Delta (Nixon 2004); The Pearl River Delta, Hong Kong (Yin et al. 2001); Mobile Bay, Alabama (Lehrter 2008); Tampa Bay, Florida (Greening and Janicki 2006); tributaries of the Chesapeake Bay (D’Elia et al. 1986); and the estuaries of the Pamlico Sound system, North Carolina (Hobbie 1974; Paerl et al. 1998, 2004).
The question naturally arises: why does not N2 fixation compensate for N requirements in all ecosystems along the continuum? The answer to this question requires consideration of environmental factors potentially controlling this process that go far beyond the stoichiometric N/P ratio arguments (i.e., low N/P ratios, especially those below 5 will favor dominance by N2 fixers) presented by Schindler et al. (2008) and others before (Doremus 1982; Tyrell 1999). Nitrogen fixation is an anaerobic process, which evolved in prokaryotic microorganisms before Earth had an oxygen-rich atmosphere (Knoll 2003). As such, microbes conducting this process, including photosynthetic bacteria, cyanobacteria, and a range of chemolithitrophic and heterotrophic bacteria, require oxygen-free conditions. These conditions can be provided in deoxygenated bottom sediments, biofilms, and locally anoxic aggregates and in the bottom hypoxic/anoxic waters of well-stratified, productive ecosystems (Paerl 1990). Some filamentous cyanobacterial genera, including those referred to by Schindler et al. (2008), are capable of conducting N2 fixation under ambient oxic conditions, by confining the process to special oxygen-free cells called heterocysts. Heterocystous species are favored by calm, vertically stratified waters, where they can adjust their buoyancy and form aggregates that enable them to grow and fix N2 under optimal conditions (Paerl et al. 2001), conditions experienced in Lake 227 during mid-summer, N-depleted periods.
Appreciable turbulence, as wind and/or tidal mixing, introduces small-scale shear that can disrupt aggregates and filaments, leading to decreased rates of N2 fixation and even destruction and death of filaments and cells (Moisander et al. 2002a). Extensive vertical mixing also forces these diazotrophs to compete with non-diazotrophic phytoplankton, with the latter type often prevailing due to superior nutrient uptake and growth rates (Reynolds 2006). Estuarine and coastal waters are generally exposed to wind and tidal mixing throughout much of the year, rendering these waters suboptimal for N2 fixation (Paerl and Zehr 2000; Paerl et al. 2001). Despite these constraints, some heterocystous species (e.g., Nodularia spp., Aphanizomenon spp, and Anabaeana spp.) as well as buoyant non-heterocystous species (Trichodesmium spp.) are able to take advantage of periods of calm, relatively warm, vertically stratified conditions in some coastal and even open ocean environments. Examples include the Baltic Sea and some tropical and subtropical ocean systems (Kononen et al. 1996; Karl et al. 2002; Paerl and Fulton 2006). However, even under these conditions, much less than the total N requirement of the whole phytoplankton community or ecosystem is met by N2 fixation, in part because physically favorable conditions are ephemeral and highly seasonal (Paerl et al. 2001; Howarth and Marino 2006). Large, highly exposed lake systems likewise have these environmental constraints.
There are additional reasons why N2 fixation may not be able to operate at optimal rates in N-limited systems, especially those away from land masses. These include deficiency of metals required for the synthesis and activity of the nitrogenase enzyme complex mediating this process, most notably iron (Fe). For example, nitrogenase activity in the widespread subtropical/tropic marine diazotroph, Trichodesmium, was shown to be Fe-limited at times in the Western Atlantic Ocean (Paerl et al. 2004). Interestingly, the bioassays used to determine Fe limitation showed P sufficiency at the time, although at other locations (e.g., North Central Pacific and Western Atlantic) P limitation may at times be operative. In addition, selective grazing could exert top-down controls on N2-fixing microorganisms, reducing their influence on ecosystem-level N inputs and budgets (Howarth and Marino 2006).
In and of itself, salinity is not a barrier to the presence or proliferation of N2 fixers (Moisander et al. 2002b). There are diverse marine analogs to the freshwater heterocystous and non-heterocystous diazotrophs. However, many of these analogs observed in N-deficient marine waters are associated with sediments, higher plants (e.g., marsh- and seagrasses), biofilms, and reefs. These are sites of protection from high degrees of turbulence and mixing, and they are relatively rich sites of macro- and micronutrients needed to support N2 fixation (Paerl 1990). It is therefore concluded that the overall paucity of N2 fixation and resultant persistent N limitation in marine waters can be attributed to many factors in addition to P limitation.
There are numerous additional reasons to warrant reducing N loading as a means of controlling eutrophication along the freshwater to marine continuum. Some of the most serious and detrimental symptoms of eutrophication along this continuum can be attributed to excessive N loading. These include toxin-producing blooms of non-N2-fixing cyanobacteria, most notably the buoyant planktonic scum-former Microcystis and filamentous planktonic Planktothrix, as well as filamentous benthic genera such as Lyngbya and Oscillatoria in freshwater and oligohaline habitats, while downstream mesohaline to euhaline waters are more often dominated by eukaryotic bloom taxa, including dinoflagellates, diatoms, cryptophytes, chrysophytes, and prymnesiophytes. In many instances, recent blooms of these nuisance taxa can be attributed to excessive N loading (Granéli and Turner 2008; Paerl 1997; Paerl and Fulton 2006). Only reducing P inputs to systems impacted by these blooms will not solve the eutrophication problem, certainly not in the short term, because the high levels of productivity and organic carbon production promoted by excessive N loading will stimulate “internal” nutrient recycling of P. In effect, excessive N loading makes P more available. In large lake, reservoir, river, and estuarine and coastal systems that have experienced a long history of either natural or anthropogenic P enrichment, N input reductions can be a highly effective way of reducing overall ecosystem productivity/eutrophication, while enabling the system to gradually “wean” itself of stored P.
In conclusion, when one looks beyond the single system interpretation of nutrient limitation and related eutrophication dynamics to the entire freshwater to marine continuum, it turns out that P limitation is only part of the story. It is essential that we examine and evaluate nutrient limitation and its effects on this larger continuum scale because natural and human influences that affect upstream waters can have significant and adverse consequences on downstream waters, especially if they differ in the nutrient(s) that control production and nutrient cycling. There appear to be important inter-nutrient feedbacks operating when single nutrient reductions are imposed on a component of this continuum; these can have negative impacts on downstream estuarine and coastal waters. Recent experiences increasingly point to the importance and long-term efficacy of dual, as opposed to single, nutrient reduction strategies as a means of controlling eutrophication along the entire continuum.
I appreciate the technical assistance and input of A. Joyner, R. Moll, B. Peierls, N. Hall, and M. Wetz. Research findings were partially supported by the North Carolina Department of Environment and Natural Resources (ModMon and FerryMon Programs, 319 Nonpoint Source Management Program); the North Carolina Sea Grant Program; the US EPA-STAR-EaGLe Program; the National Science Foundation, Ecology, Environmental Engineering and Technology; Chemical and Biological Oceanography; and Ecology of Infectious Diseases Programs.
- Bianchi, T. 2007. Biogeochemistry of estuaries, 700. Oxford: Oxford University Press.Google Scholar
- Boynton, W.R., and W.M. Kemp. 2000. Influence of river flow and nutrient loads on selected ecosystem processes: A synthesis of Chesapeake Bay data. In Estuarine science: A synthetic approach to research and practice, ed. J.E. Hobbie, 269–298. Washington, DC: Island Press.Google Scholar
- Bricker, S.B., C.G. Clement, D.E. Pirhalla, S.P. Orlando, and D.R.G. Farrow. 1999. National estuarine eutrophication assessment: Effects of nutrient enrichment in the Nation’s estuaries. Silver Spring: NOAA, National Ocean Service, Special Projects Office, and the National Centers for Coastal Ocean Science.Google Scholar
- Capone, D.G., M. J. Mulholland, and E. J. Carpenter, eds. 2008. Nitrogen in the marine environment, vol. 2. Orlando: Academic Press.Google Scholar
- Conley, D.J. 2000. Biogeochemical nutrient cycles and nutrient management strategies. Hydrobiologia 419: 87–96.Google Scholar
- D’Elia, C.F., J.G. Sanders, and W.R. Boynton. 1986. Nutrient enrichment studies in a coastal plain estuary: Phytoplankton growth in large scale, continuous cultures. Canadian Journal of Fisheries & Aquatic Science 43: 397–406.Google Scholar
- Doremus, C. 1982. Geochemical control of dinitrogen fixation in the open ocean. Biological Oceanography 1: 429–436.Google Scholar
- Elmgren, R., and U. Larsson. 2001. Nitrogen and the Baltic Sea: Managing nitrogen in relation to phosphorus. The Scientific World, Special Edition (S2):371–377. Balkema Publishers.Google Scholar
- European Environmental Agency. 2001. Eutrophication in Europe’s coastal waters. Report 7/2001. Copenhagen: European Environmental Agency.Google Scholar
- Goldman, C.R. 1988. Primary productivity, nutrients, and transparency during the early onset of eutrophication in ultra-oligotrophic Lake Tahoe, California-Nevada. Limnology and Oceanography 33: 1321–1333.Google Scholar
- Granéli, E., and J. Turner, eds. 2008. Ecology of harmful marine algae. Berlin: Springer-Verlag.Google Scholar
- Granéli, E.K., U. Wallström, U. Larsson, and R. Elmgren. 1999. Nutrient limitation of the Baltic Sea area. Ambio 19: 142–151.Google Scholar
- Hobbie, J.E. 1974. Nutrients and eutrophication in the Pamlico River estuary; North Carolina. Report No. UNC-WRRI-74-100, 239. Raleigh: UNC Water Resources Research Institute.Google Scholar
- Hobbie, J.E., and N.W. Smith. 1975. Nutrients in the Neuse River Estuary, NC, Report No. UNC-SG-75-21, UNC Sea Grant Program, North Carolina State University, 183 p.Google Scholar
- Knoll, A.H. 2003. Life on a young planet: The first three billion years of evolution on earth. Princeton: Princeton Univ. Press.Google Scholar
- Kononen, K., J. Kuparinen, K. Mäkelä, J. Laanemets, J. Pavelson, and S. Nõmmann. 1996. Initiation of cyanobacterial blooms in a frontal region at the entrance to the Gulf of Finland, Baltic Sea. Limnology and Oceanography 41: 98–112.Google Scholar
- Kronvang, B., E. Jeppesen, D.J. Conley, M. Søndergaard, S.E. Larsen, N.B. Ovesen, and J. Carstensenc. 2005. Nutrient pressures and ecological responses to nutrient loading reductions in Danish streams, lakes and coastal waters. Journal of Hydrology 304: 274–288. doi: 10.1016/j.jhydrol.2004.07.035.CrossRefGoogle Scholar
- Likens, G.E. 1972. Nutrient limitation. Limnology and Oceanography, Special Volume 2.Google Scholar
- Lohrenz, S.E., G.L. Fahnenstiel, D.G. Redalje, G.A. Lang, X.G. Chen, and M.J. Dagg. 1997. Variations in primary production of northern Gulf of Mexico continental shelf waters linked to nutrient inputs from the Mississippi River. Marine Ecology Progress Series 155: 45–54. doi: 10.3354/meps155045.CrossRefGoogle Scholar
- Moisander, P.H., J.L. Hench, K. Kononen, and H.W. Paerl. 2002a. Small-scale shear effects on heterocystous cyanobacteria. Limnology and Oceanography 47: 108–119.Google Scholar
- Nixon, S.W. 1995. Coastal marine eutrophication: A definition, social causes, and future concerns. Ophelia 41: 199–219.Google Scholar
- Nixon, S.W. 2004. The artificial Nile—The Aswan High Dam destroyed a fishery, but human activities may have revived it. American Scientist 92: 158.Google Scholar
- OSPAR. 2008. Second OSPAR integrated report on eutrophication status of the OSPAR maritime area. London: OSPAR.Google Scholar
- Paerl, H.W. 1990. Physiological ecology and regulation of N2 fixation in natural waters. Advances in Microbial Ecology 11: 305–344.Google Scholar
- Paerl, H.W., and M.F. Piehler. 2008. Nitrogen and marine eutrophication. In Nitrogen in the marine environment, vol. 2, ed. D.G. Capone, M. Mulholland, and E. Carpenter. Orlando: Academic Press.Google Scholar
- Paerl, H.W., and J.P. Zehr. 2000. Nitrogen fixation. In Microbial ecology of the oceans, ed. D. Kirchman, 387–426. New York: Academic Press.Google Scholar
- Paerl, H.W., L. Prufert-Bebout, and C. Guo. 1994. Iron-stimulated N2 fixation and growth in natural and cultured populations of the planktonic marine cyanobacterium Trichodesmium. Applied Environmental Microbiology 60: 1044–1047.Google Scholar
- Paerl, H.W., J.L. Pinckney, J.M. Fear, and B.L. Peierls. 1998. Ecosystem responses to internal and watershed organic matter loading: consequences for hypoxia in the eutrophying Neuse River Estuary, North Carolina, USA. Marine Ecology Progress Series 166: 17–25. doi: 10.3354/meps166017.CrossRefGoogle Scholar
- Paerl, H.W., J.D. Willey, M. Go, B.L. Peierls, J.L. Pinckney, and M.L. Fogel. 1999. Rainfall stimulation of primary production in Western Atlantic Ocean waters: Roles of different nitrogen sources and co-limiting nutrients. Marine Ecology Progress Series 176: 205–214. doi: 10.3354/meps176205.CrossRefGoogle Scholar
- Paerl, H.W., R.S. Fulton, P.H. Moisander, and J. Dyble. 2001. Harmful freshwater algal blooms, with an emphasis on cyanobacteria. The Scientific World 1: 76–113.Google Scholar
- Paerl, H.W., L.M. Valdes, M.F. Piehler, and M.E. Lebo. 2004. Solving problems resulting from solutions: the evolution of a dual nutrient management strategy for the eutrophying Neuse River Estuary, North Carolina, USA. Environmental Science & Technology 38: 3068–3073. doi: 10.1021/es0352350.CrossRefGoogle Scholar
- Rabalais, N.N, and R.E. Turner, eds. 2001. Coastal hypoxia: Consequences for living resources and ecosystems. Coastal and Estuarine Studies 58. American Geophysical Union, Washington, DC, 454 p.Google Scholar
- Reynolds, C.S. 2006. Ecology of phytoplankton: Ecology, biodiversity and conservation. Cambridge: Cambridge University Press.Google Scholar
- Schindler, D.W., et al. 1973. Eutrophication of Lake 227 by addition of phosphate and nitrate: The second, third and fourth years of enrichment 1970, 1971 and 1972. Journal of Fisheries Research Board Canada 30: 1415–1440.Google Scholar
- Schindler, D.W., R.E. Hecky, D.L. Findlay, M.P. Stainton, B.R. Parker, M. Paterson, K.G. Beaty, M. Lyng, and S.E.M. Kasian. 2008. Eutrophication of lakes cannot be controlled by reducing nitrogen input: Results of a 37 year whole ecosystem experiment. Proceedings of the National Academy of Science USA 105: 11254–11258. doi: 10.1073/pnas.0805108105.CrossRefGoogle Scholar
- Vollenweider, R. A., R. Marchetti, and R. Viviani, eds. 1992. Marine coastal eutrophication. New York: Elsevier Science.Google Scholar
- Wetzel, R.G. 2001. Limnology, 3rd ed. Orlando: Academic Press.Google Scholar