freshwater algae of north america || control of nuisance algae

I. INTRODUCTION

Algae play many important and beneficial roles infreshwater environments. They produce oxygen andconsume carbon dioxide, act as the base for the aquaticfood chain, remove nutrients and pollutants fromwater, and stabilize sediments. Excessive algal growths,however, can cause detrimental effects on aquatic sys-tems, endangering the organisms that live in or dependon these systems and hampering or preventing humanuses of the infested waterways.

When we refer to the kinds of problems that algaecause, it is helpful to divide algae into three groupsaccording to their growth habits: microscopic algae(primarily phytoplanktonic), filamentous mat-formingalgae, and the Chara/Nitella group. Each group posesits own unique problems to aquatic systems. This chapter describes the problems caused by each of thesethree groups and then covers the control methods thattypically are used for these algae.

II. PROBLEMS ASSOCIATED WITH ALGAE

Many of the problems that the public associateswith algae occur in more or less static bodies of water(i.e., ponds, lakes, and reservoirs) with long residence

times. Algae also produce excessive or unwantedgrowths in flowing waters such as streams, rivers, andwater delivery systems. Control of algae in these sites ismore typically achieved with watershed managementtechniques that reduce nutrient inputs than with directcontrol methods. Nuisance algae found in irrigationcanals and drainage systems can be and are managedusing direct control techniques, but the options aremore limited than those used in static systems.Although the algae of flowing waters are discussed, theemphasis in this chapter is on the problems caused bythe algae of lakes and ponds.

A. Microscopic Algae

The term “bloom” is typically reserved for exces-sive growths of microscopic, planktonic algae. Any discussion of bloom-forming algae starts with thecyanobacteria (also referred to as blue-green algae)Microcystis, Anabaena, and Aphanizomenon (seeChap. 3) Blooms of these prokaryotic organisms give a characteristic green or yellow-green color to water.Under static conditions, they rise to the surface to formvery distinctive films and windrows of greenish scum(Fig. 1) [for a review of mechanisms regulating buoyan-cy, see Oliver (1994)]. Their notoriety is well deserved.In addition to being indicators of nutrient (particularly

805

24CONTROL OF NUISANCE ALGAE

Carole A. LembiDepartment of Botany and

Plant PathologyPurdue UniversityWest Lafayette, Indiana 47907

I. IntroductionII. Problems Associated with Algae

A. Microscopic AlgaeB. Macrophytic Filamentous AlgaeC. Chara and Nitella

III. Control Methods for Nuisance AlgaeA. Nutrient ManipulationB. Direct Control Methods

Literature Cited

Freshwater Algae of North AmericaCopyright © 2003, Elsevier Science (USA). All rights of reproduction in any form reserved.

phosphorus)-enriched waters, the presence of cyano-bacterial blooms is a very visible symptom of deterio-rated bodies of water. Population crashes (death) andthe microbial decomposition of cyanobacterial cellsresult in the depletion of dissolved oxygen. Anoxic con-ditions can cause fish kills in bodies of water as smallas prairie potholes (Barica, 1975, 1978) or as large asLake Okeechobee, Florida (Jones, 1987). The largestcyanobacterial bloom recorded on Lake Okeechobee,which occurred on June 30, 1986, covered 337 km2,almost 20% of the lake surface (Lamon, 1995).

The most infamous example of the adverse impactof algal blooms on a large body of water was the dete-rioration of the western basin of Lake Erie in the 1960s(Beeton, 1969; Rosa and Burns, 1987). Increases in thequantity of algal biomass (Davis, 1964) and shifts inspecies composition from diatoms to cyanobacterialblooms (Ogawa and Carr, 1969; Munawar andMunawar, 1976; Nicholls et al., 1980; Stoermer, 1988)exacerbated an already deteriorating water quality situation. The resulting oxygen-depleted conditionshastened the demise of native fish species and theirreplacement by invasive species such as alewife andlamprey.

Crashes of cyanobacterial blooms also have anadverse effect on the aquaculture industry. A fish kill inan 8.9 ha aquaculture pond (Boyd et al., 1975) causedby the die-off of a cyanobacterial bloom resulted inoxygen depletion that killed 6800 kg of catfish. Attoday’s market value, the loss would have been worthbetween U.S. $11,000 and $19,000.

Many other cyanobacteria, as well as planktonicchlorophytes, euglenoids, diatoms, synurophytes, anddinoflagellates can bloom in nutrient-enriched waters.An important phenomenon is the occurrence of red

water and red surface scums, caused by blooms ofOscillatoria (Planktothrix) rubescens, generally in largelakes in early spring (Jaag, 1972; Konopka, 1982a, b),and by species of Euglena and Trachelomonas in staticwaters in mid-to-late summer (Lackey, 1968). Thesered water-causing organisms should not be confusedwith red tides, which are marine, are composed primarilyof dinoflagellates, and often produce toxic compounds.Other than being symptoms of highly nutrient-enrichedwaters, the red color-producing euglenoids of fresh-waters are not toxic. The O. rubescens/agardhiicomplex has been reported to produce hepatotoxins(Carpenter and Carmichael, 1995) and may be respon-sible for dermatitis or skin irritation when people come in contact with contaminated water (Gorhamand Carmichael, 1988), but documented reports ofincidents are rare (W. W. Carmichael, personal commu-nication) in comparison to reports of those caused byother cyanobacteria (discussed below).

The presence of blooms of microscopic algae haslong been associated with eutrophication (Schelske andStoermer, 1971; Schindler, 1975, 1977; Reavie et al.,1995). As a result, most trophic classification systems(e.g., Carlson, 1977; Wetzel, 1983; EPA, 1990) arebased on some measure of algal biomass (e.g., chloro-phyll or cell volumes), Secchi disk readings (a measureof transparency that can be affected by algal biomass),and types of algae present. Cooke et al. (1993b) sum-marized the effects of eutrophication as follows:“Symptoms of eutrophication, such as algal blooms(including surface scums), low transparency, rapid lossof volume in reservoirs, noxious odors, tainted fishflesh, impaired potable water supplies, dissolved oxygen depletions, fish kills, and the development ofnuisance or exotic animal populations (e.g., commoncarp) can bring about economic losses in the forms ofdecreased property values, high cost treatments of rawdrinking water, illness, depressed recreation industries,expenditures for management and restoration, and theneed to build new reservoirs.” While these authorsdescribe algal blooms as one of several symptoms ofeutrophication, the presence of the algal blooms them-selves can lead to low transparency, noxious odors,tainted fish flesh, impaired potable water, dissolvedoxygen depletions, and fish kills.

From an economic standpoint, the most importantproblem caused by algal blooms is production of tasteand odor in surface water supplies (Raman, 1985;Hawkins and Griffiths, 1987). The potential for signifi-cant taste and odor problems exists in the United Statesbecause more water for human uses is obtained fromsurface water than from groundwater. Approximately65% of the 399 billion gallons of freshwater with-drawn for all purposes in the United States in 1985

806 Carole A. Lembi

FIGURE 1 A surface scum formed by a cyanobacterial bloom ofMicrocystis.

was obtained from surface water sources (Solley et al.,1988). The cyanobacteria Microcystis, Anabaena,Aphanizomenon, and Pseudanabaena and the goldenflagellates Synura, Mallomonas, and Dinobryon arecommon causes of taste and odor in water supplies, butdiatoms, dinoflagellates, and even some green algaealso cause problems (Palmer, 1962; Nicholls andGerrath, 1985; American Public Health Association,1992). Colonial species of the diatom Stephanodiscuswere reported to cause undesirable odors and clog filterruns at municipal water plants on the Great Lakes(Stoermer, 1988). The fishy tastes and odors producedby Synura spp. are frequently reported in softwater lakesin Ontario, and the haptophyte Chrysochromulina bre-viturrita has been cited as a producer of particularlyoffensive odors in that area (Nicholls et al., 1982). Thetastes and odors produced by cyanobacteria such asAnabaena and Oscillatoria are caused by two com-pounds: 2-methylisoborneol (2-MIB) at concentrationsgreater than 12 ng L–1 and geosmin at concentrationsgreater than 7 ng L–1 (Simpson and MacLeod, 1991).

The cost of treating water for taste and odor prob-lems is high. Copper sulfate (CuSO4) is widely used tocontrol or eliminate potential taste- and odor-causingalgae in water supply reservoirs. Water treatment withactivated carbon is then used to remove undesirabletastes and odors that do occur. Chlorine has no effecton the removal of musty/earthy aromas nor does treat-ment with ozone or aeration (Maga, 1987). The cost totreat Lake Manatee (Florida) Reservoir water with acti-vated carbon where blooms have occurred has exceed-ed U.S. $14,000 a day (Clarke et al., 1997). Additionalexpenses are incurred to replace the activated carbonfilters, which frequently become clogged with humicsubstances, and to treat the lake water with copper sulfate.

Taste and odor problems are typical not only ofsurface waters where blooms occur but also of deeperwaters. In a study of six Kansas lakes, Arruda andFromm (1989) noted that the mostly fishy and grassytastes in the surface waters were caused by algae andcorrelated with the trophic status of the lake. The foultaste of the bottom water (musty, sulfurous, and rottenegg-like) was typical of highly organic, anoxic sedi-ments and overlying water. Even this problem wasrelated to algae because it was probably caused by thedeposition and slow decomposition of organic mattercontributed over time by dying algal blooms. Having todraw from deep waters to avoid infested surface watercan lead to other problems such as the deposition of iron and manganese in pipes and on clothing inwashing machines.

The other major taste and odor problem caused bycyanobacterial blooms is off-flavors in the flesh of

aquaculture-produced fish (particularly catfish) andother animals (Jüttner et al., 1986; Maga, 1987;Martin et al., 1991; Schrader and Blevins, 1993). Someof the causative organisms are species of Lyngbya,Oscillatoria, Aphanizomenon, Anabaena, andPhormidium. Geosmin (Brown and Boyd, 1982; Lovellet al., 1986) and 2-MIB (Martin et al., 1988; van derPloeg et al., 1995) are the primary chemicals that pro-duce off-flavors. In general, once fish have been taintedwith these compounds, they must be moved to cleanwater for several weeks so the off-flavor can dissipatebefore they are marketed. The harvest and transport ofthe fish to a new site for cleansing are expensive andlaborious. Off-flavor problems have been estimated to add $50 million each year to the cost of producingcatfish in the United States (Schrader et al., 1997).

Various cyanobacteria produce toxins that areharmful to humans and animals [for review, seeCarmichael (1997)]. The major genera are Anabaena,Aphanizomenon, Microcystis, Cylindrospermopsis,Nodularia, and Oscillatoria. There are many reportedinstances of livestock, pets, wild animals, and birdsthat have died after drinking tainted water, but, in gen-eral, the adverse effects of cyanobacterial blooms tohumans have been limited to forms of dermatitis andirritation of the mucous membranes. Some evidencethat human gastrointestinal disorders were associatedwith consumption of water from reservoirs with bloomsof cyanobacteria can be found (Carmichael et al.,1985; Carmichael and Falconer, 1993), and cyanobac-terial toxins were implicated in the deaths of 26 peoplein Brazil when contaminated water was used forhemodialysis (Jochimsen et al., 1998). An associationbetween toxins in water supplies and primary liver cancer in China has been suggested (Carmichael,1994). Fortunately, the instances of human poisoningare rare because the unattractiveness and foul odors ofwater in which an algal bloom occurs usually deterpeople from using or drinking the water.

Algal blooms can have adverse impacts on thehealth of organisms other than fish and humans. Highnutrient concentrations in the water column triggeralgal blooms, which reduce light penetration. Lightreduction can severely limit the growth of submersedvascular plants (Spence, 1976; Jupp and Spence, 1977;Jones et al., 1983), thus decreasing habitat and shelteravailable for fish and fish food organisms. As a result,the most eutrophic lakes are those that are dominatedby bloom-forming algae, with little or no submersedvascular plant production (Wetzel, 2001).

In addition to loss of habitat, cyanobacterialblooms may cause a loss of system-level productivity.Some cyanobacterial species are allelopathic to otheralgae that are considered to be food sources for

24. Control of Nuisance Algae 807

zooplankton (Keating, 1976, 1977, 1978), and theythemselves are not significantly grazed by zooplankton(which is one reason why they can dominate aquaticsystems). Cladoceran populations (e.g., Daphnia)decline or disappear when cyanobacteria, particularlythe filamentous forms, predominate (Burns, 1968;Keating, 1976; Infante and Abella, 1985; Rothhaupt,1991). The major reason for lack of predation appearsto be a mechanical interference with feeding when the filamentous forms accumulate in the filtering apparatus. The loss in available energy suppresses zooplankton reproduction. Webster and Peters (1978)showed that as densities of filaments of Anabaenaspp., Aphanizomenon flos-aquae, Oscillatoria tenuis,or Lyngbya spp. increased, the larger-sized cladoceransfiltered at lower rates, increased rejection rates, anddecreased brood sizes. The increased energy expendi-ture of trying to obtain sufficient food appears toincrease the respiration rates of cladocerans (Porter andMcDonough, 1984), further reducing assimilation efficiency, growth, and reproduction. As few as 50cyanobacterial filaments per mL of water have anadverse effect on zooplankton feeding rates (Infanteand Abella, 1985). The presence of filamentouscyanobacteria can cause a shift from large-bodied to small-bodied zooplankters, which feed on othermaterials such as small algae, bacteria, and organicdebris. Some evidence suggests that the adverse effectof cyanobacteria on cladocerans also is due to toxinproduction (Infante and Abella, 1985; Fulton andPaerl, 1987; DeMott et al., 1991).

B. Macrophytic Filamentous Algae

The problems caused by macrophytic filamentousalgae in aquatic systems are primarily due to their ability to form large mats of vegetation (Fig. 2). Thesealgae are typically found in shallow water where theymay be free-floating (e.g., Pithophora, Rhizoclonium,Spirogyra, and Hydrodictyon) or attached (e.g., Clado-phora, Ulothrix, Stigeoclonium, and Oedogonium) tosubstrata, either living (plants and other algae) or non-living (rocks, cement linings, and sediments). The free-floating forms are generally restricted to static waterssuch as ponds and the sheltered littoral zones of lakes.Attached forms occupy a much wider range of habitats.They are found in both static and flowing systems,including the wave-scoured edges of lakes (e.g., Clado-phora in the Laurentian Great Lakes), fast-flowingstreams, and the extensive irrigation systems and aque-ducts of the western United States (Fig. 3).

The genera listed above are all green algae.Filamentous cyanobacteria also can form free-floatingmats. The cells of filaments of Lyngbya wollei (Speziale

and Dyck, 1992) are quite large (cell diameter: 25–64 μm, length: 2–11 μm) for a cyanobacterium, andthe dimensions, coarseness, and even color (dark green)of the filaments may cause the untrained observer tothink that they are handling a filamentous green alga.The mats are dense and can completely cover pondsand shallow areas of lakes. Many species of Oscilla-toria form benthic mats that break free from the bottom and float to the surface (Fig. 4A) when gasbubble accumulation dislodges the mats (Halfen andMcCann, 1975). These growths are typically dark blue-green to black in color and are quite slimy. The matsare often coated with sediments that were deposited onthem while they were still associated with the bottom

808 Carole A. Lembi

FIGURE 2 Filamentous algal mats (Spirogyra) causing obvious aesthetic problems in a lake.

FIGURE 3 Filamentous algal mats (Cladophora) in an irrigationcanal that have broken away from the sides and are floating down-stream. These mats can clog irrigation intakes and pumps. Photocourtesy of Lars Anderson, USDA-Agricultural Research Service.

substratum (Fig. 4B). Several species of Phormidiumform the “black algae” growths that attach to thecement linings of swimming pools (Fitzgerald, 1959;Adamson and Sommerfeld, 1978).

The occurrence of filamentous algae is widespread.The Florida Department of Environmental Protection,Bureau of Aquatic Plant Management, recently sur-veyed 451 public bodies of water in Florida (Schardt,1994). Filamentous algae were the only submersed“plant” grouping observed in more than 50% of thewater bodies (62%) and were the dominant group in 16% of the waters. They represented one of onlyfive plant groupings that showed an increase over theprevious 12 years (increasing by 22%, more than anyof the other plants). The two genera that were singled

out as being predominant were Pithophora spp. and L.wollei. In an unpublished 1995 survey (J. Schardt, per-sonal communication), L. wollei was collected in morethan 218 lakes in Florida. An indication of the severity(or perceived severity) of algal problems in the state is the fact that approximately 90% of the calls fromindividuals or lake associations to aquatic plant management services in Florida seek information onalgae control (J. Williams, The Lake Doctors, WinterSprings, FL; personal communication).

Extensive survey data are less available for otherparts of North America, but filamentous algae are awidespread management problem. In the midwesternUnited States, approximately 60–70% of the total volume of aquatic plant control chemicals applied isfor the control of (primarily) filamentous algae (R.Johnson, Aquatic Control, Seymour, IN; personal com-munication). Cladophora glomerata, Pithophora spp.,and Rhizoclonium spp. are listed as causing problemsin the western United States and Canada, and Clado-phora and Chara are on the list of the 13 most consis-tently problematic aquatic weed species in these regions(Anderson, 1993).

Excessive growths of mat-forming algae, eitheralone or in combination with aquatic vascular plants,impair recreational activities such as swimming, fishing, and boating. Swimming beaches fouled withalgal mats are not only unappealing but also hazardouswhen ladders, rocks, and submerged concrete are coated by slime-producing species such as Spirogyra(Bennett, 1971) and cyanobacteria. Cladophora growthsin the Great Lakes were noted as posing a potentialdanger to young and inexperienced swimmers who might become entangled in the mats and drown(Herbst, 1969). When large odiferous masses arewashed up on the shore they decay and are aestheticallyunpleasant, are a barrier to recreation, and are impli-cated in taste and odor events in drinking water supplies (Brownlee et al., 1984; Painter and Kamaitis,1987; American Public Health Association, 1992). Theloss of recreational and aesthetic values can have a significant economic impact on waterfront properties.Ormerod (1970) reported that the value of real estateon Lake Erie fronted with Cladophora mats averaged80–85% of the value of clean frontage.

There is some evidence that mat-forming cyano-bacteria (e.g., L. wollei, Oscillatoria spp.) also producetoxins similar to those produced by the phytoplanktoniccyanobacteria (Gunn et al., 1992; Carmichael et al.,1997; Onodera et al., 1997). Although there are noreports to date of animals being killed by the toxinsproduced by L. wollei, dogs were reportedly killed bythe toxins from benthic mats of Oscillatoria (Gunn etal., 1992).


FIGURE 4 Mats of macrophytic Oscillatoria. (A) Infestation of free-floating mats that have broken loose from the sediments in theshallow cove of a lake. (B) Close-up photo of the free-floating mats.The light color is due to the sediment deposited on the surface of themats. Some of the mats have been turned upside down and show thenatural dark color of the organism. Photos courtesy of Neil Gerber,Aquatic Management, Bluffton, Indiana.

Macrophytic algae restrict and greatly reduce theefficiency of culture and harvest activities in fish culture ponds (Tucker et al., 1983). They may competewith phytoplankton, thus reducing the base of theaquatic food chain (Boyd, 1982). A 50% reduction infish production in farm ponds was attributed to heavyPithophora growth and concomitant loss of phyto-plankton (Lawrence, 1954). Although reports implicat-ing macrophytic algae as direct causes of fish kills arefew (e.g., Robinson and Hawkes, 1986), excessive algalgrowth must add to the oxygen deficits that result fromrespiration of submersed plant growth and/or phyto-plankton at night, during periods of cloudy weather, orunder snow-covered ice. Fish kills in ponds that aredominated by filamentous algae are not uncommon(personal observation). Oxygen deficits can be stressfulto fish in other ways by causing declines in food con-sumption and growth and by making them more proneto bacterial infections (Boyd, 1982).

Overpopulation, stunted or reduced growth, and adecline in the capture efficiency of forage species bypredatory fish have been attributed to dense vascularplant growth (Colle and Shireman, 1980; Savino andStein, 1982; Shireman et al., 1983), and algal growth isassumed to contribute to this problem. However, theneed for some macrophytic growth to provide protec-tion for young fish and habitat for fish food organismsis also recognized (Barnett and Schneider, 1974; Wileyet al., 1984). The proper balance of macrophyte cover-age and density for optimal fisheries is at this pointsomewhat controversial (Bettoli et al., 1992, 1993;Hoyer and Canfield, 1996a, b; Maceina, 1996), in partbecause of the variability in study sites and interpreta-tions. The possible beneficial or detrimental roles ofmacrophytic algae as fish habitat clearly require furtherstudy (Hinkle, 1986).

Cladophora, Stigeoclonium, Oedogonium, andUlothrix have been cited as presenting serious problems in irrigation canals because they attach toconcrete canal linings, thus reducing both flow rate and capacity. Cladophora can become associated with beds of pondweeds (Elodea spp.) and coontails(Ceratophyllum spp.), which increases resistance to theflow of water (Mitchell et al., 1989). Mats that breakaway from the linings float downstream where theyfoul pump inlets, irrigation siphons, trashracks, andsprinkler heads (Hansen et al., 1984). C. glomerata inparticular is a major problem in water delivery systemsin the western states. For example, in the 620 kmlength of the Salt River Project Canal, which supplieswater and electricity for the cities of Phoenix andTempe, Arizona, the control of aquatic weeds is amajor operation/maintenance task (Corbus, 1982). Theannual budget in the late 1980s for aquatic macrophyte

control (primarily Cladophora) in this system wasapproximately U.S. $1.5 million.

The clogging of rivers, canals, and drainage ditchesby aquatic plants and algae can prevent adequatedrainage so that water backs up, even to the point ofcausing flooding. Another problem associated withwater conveyance or storage systems, particularly inarid parts of the world, is the potential to lose waterthrough evaporation from floating or emergent plantsurfaces. Although considerable data are available toshow that substantial loss does occur with coverage ofvascular plants such as water hyacinth (Brezny et al.,1973), nothing is known of the potential of surface-floating algae to add to the problem.

The impacts of filamentous algae on the dynamicsof food webs in natural systems have not been welldocumented. Most studies that have been conductedhave focused on Cladophora. For example, Clado-phora is not a major food source for the invertebratesor fish that live in lakes (see reviews by Lembi et al.,1988; Dodds and Gudder, 1992), although it is grazedby fish in river systems (Power, 1990). Cladophoraprovides an extensive surface area for colonization byperiphyton and invertebrates, and the limited grazingthat does occur may have more to do with ingestingthese associated organisms than with the filamentousalga itself (Dodds and Gudder, 1992). On the otherhand, dense growths of Cladophora were reported toreduce invertebrate diversity and to have disruptedshoal spawning by walleye, whitefish, and lake trout inthe Great Lakes (Neil, 1975). Filamentous algal matscompete with submersed vascular plants for space andlight. Examples include the replacement of angio-sperms, such as Najas marina, by Spirogyra (Phillips etal., 1978), diverse macrophytes by C. glomerata (Bolasand Lund, 1974), Elodea by Cladophora and Spirogyra(Simpson and Eaton, 1986), and Potamogeton pectina-tus by Cladophora (Ozimek et al., 1991). Phillips et al.(1978) suggested that replacement of submersed vascu-lar plants in lakes undergoing eutrophication may bedue more to the shading by epiphytic and filamentousalgae than to phytoplankton.

The diversity among filamentous algae, for example,the sliminess of Spirogyra (which appears to preventcolonization by periphyton) in contrast to the thick cellwalls of Cladophora (which provide an excellent sub-stratum for periphyton), the summer domination byPithophora in contrast to the spring/fall distribution ofCladophora, and the net-like habit of Hydrodictyon orthe unbranched habit of Spirogyra in contrast to thebranched habit of Cladophora, suggests that much remainsto be learned about the micro-niches that these algaemake available to invertebrate and fish communities andtheir impacts on food webs of static freshwater systems.

810 Carole A. Lembi

C. Chara and Nitella

Charophytes are usually viewed as being beneficialcomponents of aquatic systems, and their reestablish-ment is an important factor in lake restoration (vanden Berg et al., 1998b). Chara and Nitella are consid-ered excellent habitats for littoral invertebrates(Rosine, 1955; Quade, 1969; Allanson, 1973; Hargebyet al., 1994) and fish (Fassett, 1957; Schardt, 1994),and they are a major food source for herbivorouswaterbirds (Hargeby et al., 1994; van den Berg et al.,1998b). Their ability to form low-growing meadows ofvegetation reduces the resuspension of sediments (vanden Berg et al., 1998b). These macroalgae, however,can cause problems in shallow water when theirgrowths reach the surface of the water, thereby pre-venting successful angling, swimming, and boating(Fig. 5). Chara and Nitella are typically named whenweedy submersed species (mostly vascular plants, suchas Ceratophyllum, Myriophyllum, and Elodea) are

listed. For example, Steward (1993) listed Chara andNitella as among the plant groups causing weed prob-lems in the eastern United States, and Anderson (1993)cited these genera for the western United States also.

Charophytes also produce repellent (allelopathic)materials that exclude certain limnetic species of inver-tebrates (Pennak, 1966, 1973) and phytoplankton(Gibbs, 1973; Anthoni et al., 1980; Wium-Andersen et al., 1982). The latter finding may provide a partialexplanation for the lack of epiphytes and clear waterconditions frequently associated with some charophytespecies [Crawford, 1979; Wium-Andersen et al., 1982(but see Chap. 2, Sect. II.F.4)].

Chara is common in regions with hard water (e.g., areas of the Midwest with a limestone bedrock),and Nitella is more characteristic of soft waters (e.g.,granitic regions of the Northeast). In a survey of 451water bodies in Florida (which has regions of both softand hard waters), Schardt (1994) collected Nitellain 64 bodies, and found that it was dominant in 28 ofthem. Chara was found in 52 bodies and was domi-nant in 15 of them.

An interesting characteristic of Chara is that ittends to colonize sites in which vascular plants havebeen controlled (Nichols, 1984). Drawdown, the drain-ing and exposure of shallow or shoreline areas to desic-cation, eliminated water shield (Brasenia schreberi),restricted the spread of parrot feather (Myriophyllumbrasiliense) and water lily (Nymphaea odorata), butenhanced the infestation of Chara vulgaris in aLouisiana reservoir (Lantz et al., 1964). In a survey ofthe effects of drawdowns, C. vulgaris increased in 33cases, decreased in 15 cases, and stayed the same in 44cases (Cooke et al., 1993b). Invasion or expansion byChara has also been documented after dredging (Bornet al., 1973; Nichols, 1984), mechanical harvesting(Anonymous, 1990), and the application of herbicidesfor the control of vascular plants (C. A. Lembi, personal observations).

Opening of sites disturbed by weed control activi-ties to light is the major reason cited for the invasionby Chara (Born et al., 1973), and recent studies seemto confirm that irradiance is a major factor regulatingcharoid distribution (Steinman et al., 1997). Althoughsome evidence suggests that as eutrophication pro-ceeds, charophyte populations may be reduced becauseof their sensitivity to “toxic” levels of phosphorus (P)(Forsberg, 1965), other studies show that increased Plevels do not have an adverse effect on charoid growth(Blindow, 1988). Melzer et al. (1977) suggested thatincreased P concentrations play an indirect role in thedisappearance of Chara, primarily by causing anincrease in phytoplankton growth and turbidity, whichin turn shades out charoid growths. The restoration of


FIGURE 5 A solid stand of Chara infests this pond. Although thevegetation has not formed a surface canopy, the underwater growthlimits swimming and fishing success.

a Chara community in one system was achieved byreducing P concentrations, which resulted in higherwater transparencies (Simons et al., 1994). However,light may not be the only factor critical to Chara estab-lishment, particularly in mixed plant communities. vanden Berg et al. (1998a) experimentally observed thatChara was negatively impacted by shading from sagopondweed (Potamogeton pectinatus), but the fact thatChara dominates sago pondweed in some clear waterlakes suggests that light is probably not a key factor in that domination. The more efficient use by Charaof carbon (HCO3–) at low concentrations, which aretypical within Chara meadows, has been suggested as apossible reason for its dominance (van den Berg et al.,1998b).

III. CONTROL METHODS FOR NUISANCE ALGAE

Management practices for nuisance algae are divided into two major categories: nutrient manipula-tion and direct control techniques. Nutrient manipula-tion, particularly reduction of nutrient inputs, shouldbe viewed as the best approach for long-term control ofalgal problems. There are situations for which signifi-cant nutrient reduction is impractical or ineffective;under these conditions, direct control of the algal biomass may be the only alternative available. Directcontrol methods should only be viewed as temporarysolutions and should be coupled with longer-termstrategies for reducing nutrient inputs.

A. Nutrient Manipulation

It has long been known that inputs of nutrients,particularly P, stimulate algal growth. Many studieshave shown a strong correlation between total phos-phorus (TP) and planktonic algal biomass (Dillon andRigler, 1974; Jones and Bachmann, 1976; Carlson,1977; Schindler, 1978; Prepas and Trew, 1983). Thepositive relationship between chlorophyll-a concentra-tions (or shallower Secchi disk transparencies) and TPis a commonly used tool to predict water quality andtrophic status (Vollenweider, 1969; Dillon and Rigler,1974; Dillon et al., 1988).

Some lakes are nitrogen (N)-limited. For example,a number of Florida lakes are surrounded by rich phos-phate-containing deposits and soils and therefore maybe N-limited. In a study of 223 Florida lakes, 27%were considered N-limited (Canfield, 1983). Also, studies of lakes in the semiarid and mountainousregions in the western United States indicate that theimportance of N may be equal to or greater than thatof P in limiting phytoplankton growth (Elser et al.,

1990; Reuter et al., 1993). N limitation, as well as Plimitation, has been implicated in the regulation of filamentous algal growth (Spencer and Lembi, 1981;O’Neal et al., 1985; Dodds and Gudder, 1992).

In addition to the amount of phytoplankton bio-mass that is produced with P or N additions, anotherconsideration is species composition, particularly inrelation to nutrient ratios. When sufficient silicon (Si)and N are available in relation to P (high Si:P and N:Pratios), diatom growth appears to be favored (Tilmanand Kiesling, 1984). These conditions are typical ofspring periods in temperate lakes following turnoveror when sediment deposition occurs with spring rains.In late spring or early summer, green algae may domi-nate over diatoms as Si concentrations decrease [lowerSi:P ratios (Sommer, 1983)].

From a management standpoint, the most criticalnutrient ratios are low N:P or Si:P. These generallyoccur under conditions of excessive P loading, and it isunder these circumstances that N-fixing cyanobacteria(Anabaena, Aphanizomenon, and others), which fixatmospheric N2 when the water becomes N-depleted,become dominant (Schindler, 1977; Smith, 1983), particularly during the summer months. For example,the shift from high Si:P and N:P ratios to low ratioswas concomitant with the shift in dominance fromdiatoms to cyanobacteria in the Great Lakes in the1960s (Schelske and Stoermer, 1971; Schelske, 1975).For this reason the emphasis on nutrient removal forgenerally improving water quality has been placed on Prather than on N.

Another approach for maintaining a high N:P ratio is to increase N rather than to decrease P. In fact,several researchers (Leonardson and Ripl, 1980; Smith,1983) suggested that N removal could be counterpro-ductive and that N addition might actually be helpfulin increasing populations of green algae or diatoms inrelation to cyanobacteria. Barica et al. (1980) added Nto ponds with low N:P ratios to see if it could reducethe incidence of cyanobacterial blooms, particularlyblooms of Aphanizomenon that were causing fish killswhen they crashed. The initial N:P ratios in theseponds were around 4 to 5 [Schindler (1977) found that cyanobacteria dominate when N:P ratios droppedfrom 15 to 5]. When low amounts of N were added(0.1 g N m–3 d–1) prior to bloom formation (but notduring the bloom) or when high amounts (1 g N m–3

d–1) were added during the bloom, a shift from Apha-nizomenon to green algae and cryptomonads occurred.The technique worked, but it was not considered to bea realistic approach because N would have to be addedover several to many weeks. Although similar resultswere reported by Stockner and Shortreed (1988), the general consensus is that, when possible, it is much

812 Carole A. Lembi

better to reduce P concentrations than to elevate Nconcentrations. In fact, increasing the N:P ratio stimu-lated the growth of non-nitrogen fixing cyanobacteriasuch as Lyngbya, Oscillatoria, and Chroococcus inmesocosms placed in Lake Okeechobee (Havens andEast, 1997).

In an analysis of the literature, Elser et al. (1990)suggested that both P and N potentially limited algalgrowth. They found little support for P alone as acausative factor. However, they recommended thatefforts should concentrate on P reduction because it iseasier to achieve from a technical standpoint than Nreduction. Clearly where N fixation by planktoniccyanobacteria is a response to N reduction, P should be the more reliable means to lower algal biomass. The same general approach is used for the control offilamentous algae and Chara, although the situationregarding specific nutrient ratios and target amounts isfar from clear.

There are three general approaches for achieving Preduction: decrease external P loading, suppress inter-nal P loading, and increase P output from the system.External inputs of P can be decreased with diversionand advanced wastewater treatment, with detentionbasins and wetlands, and by the initiation of otherwatershed management techniques. Internal P loadingcan be suppressed with alum applications, dredging,and aeration. P-laden waters from the site can bereleased with hypolimnetic withdrawal.

1. Diversion and Advanced Wastewater Treatment

These two techniques (used together) are the mostfrequently used methods to reduce external loading.Diversion is achieved primarily through sewage collec-tion systems, and the water is then subjected to tertiarytreatment in which P is removed by alum (aluminumsulfate), lime (calcium carbonate), or iron (ferric chlo-ride). There have been a number of successes (for casehistories, see Cooke et al., 1993b). Probably the bestexample is Lake Washington in Seattle, Washington(Edmundson and Lehman, 1981; Edmundson, 1994),in which 88% of the lake’s external loading was diverted from 1964 to 1967. TP declined from a meanannual concentration of 64 μg L–1 prior to diversion to 21 μg L–1 5 years after diversion. Chlorophyll-adecreased from 36 to 7 μg L–1 by 1969. Secchi diskdepth increased from 1 to 3.1 m. Further reductions inalgal biomass were attributed to increased populationsof Daphnia [following a decline in planktonicOscillatoria, which negatively impacts Daphnia feeding(Infante and Abella, 1985)] and a decrease in a plank-tivorus crustacean (Neomysis mercedis) population.The condition of the lake in the late 1970s was17 μg L–1 TP, 3 μg L–1 chlorophyll-a, and 7 m Secchi

disk depth, and it had clearly shifted from a eutrophicto a meso- or oligomesotrophic state.

During the 1970s, significant reductions in P load-ing to Lake Erie also were achieved through legislationthat upgraded sewage treatment to include chemicalprecipitation of P and reduced the allowable levels of phosphates in laundry detergents (PhosphorusManagement Strategies Task Force, 1980). Declines inphytoplankton biomass averaged about 5% per yearover the period from 1970 to 1985 (Nicholls andHopkins, 1993) and were correlated with the resultantreduction in P loading (Nicholls et al., 1977, 1980).Interestingly, phytoplankton populations continue todecline due to removal by zebra mussels (Nicholls and Hopkins, 1993). Filamentous algal growths alsorespond to nutrient diversion. For example, Clado-phora biomass and tissue P concentrations at sevensites in Lake Ontario steadily decreased from 1972 to1983 in response to P control programs introduced inthe early 1970s (Painter and Kamaitis, 1987).

Diversion and treatment work best where there aredistinct point sources of nutrient inputs. They are lesssuccessful at sites impacted by nonpoint sources or inwhich significant concentrations of nutrients have beenstockpiled in the sediments and are a major source ofinternal loading.

2. Detention Basins and Wetlands

Discharge of domestic wastewater and urban runoffinto detention basins (also called retention ponds) ornatural or constructed wetlands is often recommendedfor improving water quality before release into a riveror lake (Mitsch and Gosselink, 1993; Olson, 1993;Etnier and Guterstam, 1997). Many local and somestate ordinances now mandate the construction ofretention ponds in new housing developments, industri-al parks, and similar sites. These ponds mostly serve assettling basins for sediments and associated nutrientsand other pollutants (Walker, 1987; Robbins et al.,1991). Of course, these sites themselves become idealenvironments for the development of algal blooms andmats and are in large part the cause for the substantialincrease in the number of companies offering aquaticplant and algal management services in recent years (C. A. Lembi, personal observation). Although algae andother aquatic plants in retention ponds serve as a filtra-tion system for nutrients, urban residents frequentlycomplain about having to look at scummy water!

Highly vegetated wetland areas also act as settlingbasins; in addition, they provide biological filtering,uptake and storage, and transformation (e.g., denitrifi-cation) of nutrients (Mitsch and Gosselink, 1993).Although the P storage capability can be lost in tem-perate areas in winter when the plants are no longer


taking up P and nutrients are released, wetlands dotend to store considerable P in the summer, which is thecritical time for algal blooms to occur in downstreamsites (Cooke et al., 1993b).

3. Watershed Management

The importance of a broad watershed managementprogram to reduce both point and nonpoint sources of fertilizers and other pollutants is gaining increasedrecognition at local, state, and federal levels. In agricul-tural areas, the promotion of best management prac-tices (BMPs) has resulted in widespread acceptance ofpractices that reduce erosion of nutrient-laden soils(Scholze, 1994; EPA, 1998). Such practices include no-til and conservation tillage, vegetated filter stripsand grass waterways, lowering of fertilizer applicationrates, and proper handling of animal manures. Theadoption of BMPs in the United States is voluntaryalthough cost-sharing programs are available throughfederal agencies such as the Farm Services Agency andthe Natural Resources Conservation Service. Section303(d) of the Clean Water Act calls for the implemen-tation of total maximum daily loads into streams andlakes that have low water quality, and the Clean LakesProgram provided assistance in watershed managementand improving water quality in lakes prior to 1995.Clearly, there is general recognition of the importanceof watershed management in improving water quality,and the erosion of sediments into waterways has beenconsiderably lessened. There are, however, still areaswhere the implementation of programs has been slow.An example is the discharge of animal wastes intorivers in North Carolina and Maryland watersheds,which appears to have resulted in fish-killing blooms of the estuarine dinoflagellate Pfiesteria piscida(Burkholder et al., 1997).

4. Alum

In many situations, reduction of external P loadingdoes reduce algal growth. This is particularly true inwater where most of the P loading to the photic zone isfrom external sources and in deep stratified lakeswhere P released from the anaerobic bottom sedimentsand hypolimnion does not reach the photic zone. Inshallow lakes, on the other hand, significant quantitiesof P can be released from the sediments and reach the photic zone (Wetzel, 1990; Cooke et al., 1993b).Therefore, reduction of external P loading may nothave much of a short-term impact on phytoplanktongrowth in these sites. Resuspension of sediments is con-sidered a potential source of nutrients for phytoplank-ton production in many shallow lakes (Carper andBachmann, 1984; Riley and Prepas, 1984; Hansen etal., 1997; Havens and James, 1997). Stauffer and Lee

(1973) calculated that all of the summer algal bloomsin Lake Mendota, Wisconsin, could be accounted forby internal loading of P from the lower waters and sediments to the photic zone. Therefore, steps to reduceinternal P cycling in many of these lakes may be moreeffective in reducing algal growth than the reduction ofexternal P inputs. The methods used to reduce internalP loading include chemical treatment with alum, theremoval of sediments by dredging, and aeration.

Alum (Al2[SO4]3) is used to lower P availabilitythrough P precipitation and to retard P release from thelake sediments (P inactivation). When added to water,alum and P form aluminum phosphate and a colloidalaluminum hydroxide floc to which certain P fractionsare bound (Cooke et al., 1993b). The floc settles to thesediment and continues to sorb and retain P within the lattice of the molecule, thereby preventing furtherrelease of P. Sodium aluminate (AlNaO2), which is agood buffering material, is added to alum treatments tomaintain pH values between 6 and 8 (Kortmann andRich, 1994) because a severe shift in pH can be detri-mental to fish populations. In addition, alum is not recommended for use in waters with an acidic pH orlow alkalinity because of the potential for aluminumtoxicity to fish at pH values below 5.5. Iron salts alsocan be used to inactivate P (Kortmann and Rich,1994), and treatments with calcium salts (Ca(OH)2 andCaCO3) have successfully reduced P loading from bottom sediments in Canadian lakes (Prepas et al.,1990; Babin et al., 1994).

There are numerous examples of success with alum treatments in shallow lake systems, and manytreatments last from 2 to 15 years (Welch et al., 1988;Smeltzer, 1990; Cooke et al., 1993a; Jacoby et al.,1994; Welch and Cooke, 1995). In some lakes, internalloading has been significantly reduced for up to 20years (Welch and Cooke, 1999). Holz and Hoagland(1999) reported improved water clarity, decreasedchlorophyll-a concentrations, reduced cyanobacterialbiomass and abundance, increased Daphnia biomassand abundance, and increased usable fish habitat in ashallow (mean depth = 4 m), alum-treated lake inNebraska.

To ensure success and long-lasting effects, reduc-tion of internal P cycling must be accompanied by areduction in external P loading. Factors that can leadto failure of an alum treatment include continued highexternal P loading (Welch et al., 1988; Barko et al.,1990), redistribution of alum floc to the lake center by wind mixing (Garrison and Knauer, 1984), and Precycling from senescing rooted macrophytes or frommacrophytes that expand their range due to improvedwater clarity (Welch et al., 1988; Welch and Cooke,1999). There also is evidence that cyanobacteria newly

814 Carole A. Lembi

recruited from the sediments can transport P into thewater column, even in alum-treated lakes (Perakis etal., 1996).

5. Dredging

Dredging offers a more permanent solution tointernal P loading in shallow lakes than alum treamentbecause sediments, the actual source of the P loading,are removed from the system. Dredging, however, ismuch more expensive than alum. According to Cookeet al. (1993b), dredging costs nearly 30 times morethan alum initially although over the long term (repeatalum treatments every 10 years), the cost differential isonly 5 times greater if totaled over 50 years. The costof alum treatment averages about U.S. $700 per ha andthe cost of dredging is about U.S. $20,000 per ha(Cooke et al., 1993b). As with all nutrient reductionapproaches, external loading must be reduced or elimi-nated to achieve long-term results.

6. Aeration

P is released from sediments under anoxic condi-tions. The function of aerators (other than to improvehabitat for fish) is to oxygenate the water column, orportions of the water column, and the upper layers ofthe sediments, thereby preventing the occurrence oflow-O2 conditions. In theory, oxidized forms of P arenot released into the photic zone to encourage phyto-plankton blooms.

There are two major methods for aerating pond orlake water (Cooke et al., 1993b; Kortmann and Rich,1994). The first is artificial circulation. This methodoxygenates the whole water column. Air is pumpedfrom a compressor on shore through a tube to aweighted diffuser unit that is placed on the bottom. Airbubbles pass from the diffuser into the water and areoften visible as a surface “boil” (Fig. 6A). This methoddestroys or prevents thermal stratification; therefore, itis not feasible in sites where deep cold water is neces-sary to maintain coldwater fish populations. It is, however, a good solution to potential oxygen depletionproblems for warmwater fish species.

The second method is termed “hypolimnetic” aera-tion. This method maintains stratification because thewater is removed from the hypolimnion, oxygenated at the surface, and then returned to the bottom.Hypolimnetic aeration is used in deep lakes to over-come anoxia, improve coldwater fisheries habitat, andcontrol sediment P release.

The impacts of either type of aeration method onalgal blooms have been difficult to document. The sedi-ment–water interface in many shallow lakes mayalready be oxygenated, in which case aeration will nothave an impact. Cooke et al. (1993b) summarized data

from a number of aerated lakes and observed that phytoplankton content decreased in less than half ofthe lakes examined. Cyanobacterial blooms, however,decreased and green algal populations increased in themajority of cases. This shift was attributed to severalfactors. For example, aeration may increase the carbondioxide concentration in the water, thus lowering thepH and favoring green algal development. The turbu-lence created may disrupt the ability of the cyanobacte-ria to form surface scums, which normally shade outother, potentially competitive, algae. In those cases inwhich cyanobacterial blooms were not affected, watermixing and aeration may have been incomplete.

There is presently no evidence to suggest that aeration has an impact on filamentous algae or Chara.Frodge et al. (1991) reported that Pithophora matsgrowing among vascular plant canopies were associatedwith high concentrations of P in the surface water. Thistrend was thought to be the result of the conversion ofiron-bound P to OH–-bound P at high pH values (>10),


FIGURE 6 Aeration. (A) The surface boil from an underwater circu-lating aerator. (B) A fountain has been attached to the aerator toimprove aesthetics. Photos courtesy of Neil Gerber, AquaticManagement, Bluffton, Indiana.

so that the P was not precipitated even at high dis-solved oxygen levels. Thus, high pH values, often associated with high photosynthetic rates of dense vegetation, can potentially offset aeration effects.

The presence of a fountain in a body of water doesnot mean that the water is being aerated. Aerators arespecifically designed pieces of equipment that move airinto the water column, not spray water into the air. Afountain can be attached to an aerator for aestheticpurposes (Fig. 6B). Although fountains may cause some surface circulation and aeration, they do little toprevent nutrient cycling or fish kills.

7. Hypolimnetic Withdrawal

The principle behind hypolimnetic withdrawal isthe pumping or siphoning of bottom waters that have ahigh P content into receiving waters. The technique hasnot been used frequently. Although evidence suggeststhat it can be successful in reducing the P content of lake water (Cooke et al., 1993b), few studies pro-vide convincing data that algal blooms are reduced.Replacement of water to maintain depth must comefrom a source with a low P content. Another problemwith this technique is the potential for damage down-stream caused by releasing anoxic, nutrient-laden, polluted waters.

8. Summary of Nutrient Manipulation Methods

Nutrient manipulation techniques, particularlythose that regulate P inputs or internal cycling, can suc-cessfully reduce the incidences and severity of algalblooms. In some instances, a single technique is notsufficient. Water supply lakes for the city of St. Paul,Minnestoa, were infested with blooms of Anabaenaand Aphanizomenon (Walker et al., 1989). Chemicals(powdered carbon and potassium permanganate) wereadded at the water treatment plant to reduce taste andodors, and copper sulfate was applied every week during the growing season. These approaches wereunsuccessful. It was only when the lakes were subjectedto a multimethod approach that included the reductionof external and internal P concentrations by using ironchloride to inactivate P, the construction of detentionponds to reduce P loadings from runoff from urbanwatersheds, and hypolimnetic aeration that some success was achieved.

The St. Paul example illustrates the complexities in-volved in nutrient manipulation procedures. Probablythe greatest impediment to initiation of a nutrientremoval plan is the watershed analysis and water qualitytesting (and financial outlay). This analysis is necessaryto determine which approach or combination ofapproaches is most likely to succeed. It is important forwater management agencies and property owner asso-

ciations to accumulate and allocate sufficient resourcesfor a thorough lake and watershed monitoring pro-gram before implementing nutrient manipulation techniques. Once the management approach has beenchosen, however, the science is now to the point wheresuccesses far outnumber failures.

One outcome of nutrient control techniques toreduce algal populations can be increased colonizationby submersed vascular plants. Submersed vascularplants obtain the majority of their nutrients from thesediments rather than from the water (Carignan andKalff, 1980; Barko and Smart, 1980, 1981). Therefore,the techniques described above have little impact onsubmersed vascular plant growth. When shading byphytoplankton or filamentous algal mats is removed,these plants can become established or reestablished.The presence of submersed plants is advantageous inmany instances, but sometimes it has serious conse-quences. For example, the improved clarity of LakeWashington has led to invasion by Eurasian watermil-foil (Myriophyllum spicatum) in shallow areas, andthere are other examples of similar shifts in popula-tions (Spencer and King, 1984; van Donk et al., 1990).Madsen (1996) recorded that Secchi disk values doubled (1.5 to 3 m) in Lake St. Clair, Michigan,between 1967 and 1995, as a result of the introductionof zebra mussels. Macrophytic plant range expandedfrom 60 to 95% of the lake over this period, andEurasian watermilfoil range expanded from 20 to 44%of the lake. Hence, management plans may need toconsider which is the better alternative: a eutrophic,algal-dominated lake vs. relatively clean water withabundant vascular plant growth. The latter problem issomewhat ameliorated when native vascular speciesand Chara, which tend to have shorter growth habitsand provide valuable fish habitat, colonize an area. It isexacerbated, however, when the colonizer is an invasivespecies such as Eurasian watermilfoil or hydrilla(Hydrilla verticillata). The stems of these species growup through the water column to form a canopy of vegetation at the surface (Barko et al., 1986; Smith andBarko, 1990). This growth form not only prevents use of the water but can shade out stands of native vegetation. Such shifts among populations further illus-trate the complexity of dealing with algal and aquaticplant management issues.

B. Direct Control Methods

The goal of direct control methods is to remove thealgal biomass as quickly, efficiently, and cost effectivelyas possible. Although choices have to be made aboutwhich technique to use, the extensive watershed and in-lake monitoring that should precede nutrient manipula-

816 Carole A. Lembi

tion does not have to be conducted. On the other hand,the efficacy of direct control methods often causes theuser to overlook the need to initiate a long-term, nutrient management program. Certainly direct controltechniques have their place in an overall managementplan, but they should not be viewed as the onlyapproach to solving noxious algal blooms or extensivefilamentous algal mat growth.

The major methods of direct control of algal biomass are harvesting, biomanipulation, biologicalcontrols, allelochemicals, and algicides.

1. Harvesting

Harvesting methods can range from hand-pullingor raking to use of large mechanized harvesting equip-ment (Fig. 7). The vegetation is gathered and preferablymoved away from the site so that it cannot wash backinto the water. This is obviously not a technique thatwill remove phytoplankton, but it can be used withsome success for the removal of floating filamentousalgal mats and charophytes.

Hand harvesting or raking of filamentous algalmats and Chara is considered difficult because thesegrowths fragment very easily. The tremendous amountsof biomass (and associated water) make hand laborexhausting and time-consuming. However, probablymore hand harvesting occurs than one would expect.Pond and lake property owners, for example, frequentlyclean off beach and dock areas with rakes and otherhandheld devices. The beaches of Lake Ontario have attimes been hand raked of Cladophora growth washedup on shore by municipal workers. Hand harvestingsometimes is encouraged prior to algicide treatment.This is particularly true in late summer when largeamounts of mat material have accumulated. The death

of an excessive amount of biomass (which leads todecomposition and associated bacterial growth, whichuses the oxygen in the water) can lead to oxygen deple-tion and fish kills. Removing at least some of the bio-mass prior to treatment can help prevent severe oxygendepletion situations.

Most mechanical harvesting activities are directedat the removal of rooted submersed vascular plants.There are, however, reports of mechanical harvestingused successully for Chara control (Conyers and Cooke,1982; Cooke et al., 1993b), particularly in shallowwater where the harvesting blades can cut at the sediment–water interface. Chara is almost invariablycollected along with submersed vascular vegetationwhen the two grow intermixed.

Some evidence suggests that populations of phyto-plankton, filamentous mat-forming algae, and Characan increase after intensive mechanical harvesting ofsubmersed vascular plants (Neel et al., 1973; Nichols1973; Cooke and Kennedy, 1989; Anonymous, 1990).Although the increases have not been clearly associatedwith harvesting, the opening up of areas to light andthe potential increase in nutrients after harvest maymake algal growth more likely to occur.

Another method of harvesting has been used inirrigation systems in the western United States. Racksare inserted at intervals along the canal to collect thealgal mats that slough off the sides of the canal andfloat downstream. The racks are removed periodically,cleaned of algae and other debris, and returned to thecanal (Fig. 8).

A major consideration in harvesting is to ensurethat the collected vegetation does not wash back into


FIGURE 7 A mechanical harvester. Photo courtesy of United MarineInternational, Div. of Liquid Waste Technology, Inc., Somerset,Wisconsin.

FIGURE 8 A rack to collect floating material taken from an irriga-tion canal in California. The majority of vegetation is Cladophora.Photo courtesy of Lars Anderson, USDA-Agricultural ResearchService.

the body of water. Even though it may appear that thealgal mats have dried out once they are exposed to theair and sun, the underlying portions of the mats maystill be viable. Akinetes found in Pithophora mats thatwere exposed to the drying effects of the sun stillshowed 80% viability 136 days after initial stranding(Lembi et al., 1980).

Little use has been made of harvested aquatic vege-tation. It generally has little value as food for livestockor humans, and the energy costs to dry and pellet thevegetation can be prohibitive (National Academy ofSciences, 1976; Joyce, 1993). Some research has beenconducted on the potential use of filamentous algaesuch as Pithophora and Cladophora to make paper,and the protein content of Hydrodictyon, Spirogyra,and Pithophora was reported to be 18–26%, which iscomparable to the protein content in some cyanobacte-ria and vegetables (Khan et al., 1996); however, themajor (but sporadic) use of harvested algal mats atpresent is as mulch and fertilizer for gardens.

2. Biomanipulation

The observation that the relationship between algalgrowth and P concentrations is not perfect [in fact, in astudy of 66 lakes by Schindler (1978), the regressionstatistic explained only 48% of the variance in algalproductivity] led in part to the theories behind bio-manipulation. Much research in the past (reviewed byCooke et al., 1993b) has indicated that the type of zoo-plankton present in a body of water can have an effecton phytoplankton populations. The type of zooplank-ton is affected in turn by the types of fish that are present. The potential for zooplankton (and fish) tohave an effect on phytoplankton populations, irrespec-tive of P content in the water, may explain why, insome instances, phytoplankton populations are loweror higher than those predicted by the P content of thelake.

The term “biomanipulation” was coined byShapiro et al. (1975). It is also referred to as top-downfeeding and involves manipulating the components ofthe trophic cascade (Paine, 1980; Carpenter et al.,1985). The premise of biomanipulation, as elucidatedby Shapiro (1980) and Carpenter et al. (1985, 1987), isthat top predators, such as piscivorous fish, can influ-ence the abundance of planktivorous fish, which inturn can determine the abundance, size structure, andproductivity of zooplankton and phytoplankton (Fig. 9).For example, planktivorous fish tend to feed on large-bodied zooplankton, which results in domination bysmall-bodied zooplankton. Because it is the large-bodied zooplankton (some species of Daphnia) thatfeed most effectively on algae, their reduction is typi-cally accompanied by a relatively high phytoplankton

biomass. The key to success in biomanipulation is theaddition of piscivorous fish, which theoretically shouldreduce the numbers of planktivorous fish, which inturn enhances the development of populations of large-bodied Daphnia species and a decline in phytoplanktonpopulations. The actual manipulation involves theintroduction, where necessary, of piscivorous fish and/or the removal of planktivorous fish. An examplewould be the removal of planktivorous fish from a siteby rotenone treatment and the stocking of piscivorousfish to eliminate any planktivorous fish that might be introduced later. Observations of fish–zooplankton–phytoplankton relationships and the successful manipulation of all or part of the trophic cascade inenclosures, ponds, and lakes (Spencer and King, 1984;Carpenter et al., 1985, 1987; Elser and MacKay, 1989;Gulati, 1990; Mazumder et al., 1990; Quirós, 1995)have provided evidence that the principle is essentiallyvalid.

However, the practice of biomanipulation has beenmarked by inconsistencies that suggest that aquatic systems are much more difficult to manipulate thanoriginally anticipated. Analyses of data in which bio-manipulation did not provide the expected resultsinclude those cited by McQueen et al. (1989),McQueen (1990), Vanni and Findlay (1990), Badgeryet al. (1994), and Noonan (1998). McQueen et al.(1989) suggested that the various trophic levels dependon nutrients and energy flow, which is essentially a bottom-up process rather than top-down. In less pro-ductive oligotrophic lakes, a top-down cascade may

818 Carole A. Lembi

FIGURE 9 A flowchart illustrating the connections between food-chain levels as determined by the biomass of piscivores. Biomanipula-tion attempts to decrease the mass of planktivores in order toincrease the numbers of large-bodied zooplankton, which, in turn,graze on phytoplankton.

extend all the way to phytoplankton, but in nutrientenriched eutrophic systems the results are less clearbecause bottom-up forces are large relative to top-down forces. In other words, more phytoplankton populations can be supported by high P concentrationsthan can be effectively grazed by zooplankton.Benndorf (1989) concluded that the long-term successof top-down manipulation requires a reduction ofexternal P loading. Another complication is that even if large-bodied zooplankton populations increase, thegelatinous (Porter, 1973) or large-celled ungrazable orundigestible algae can still proliferate.

The conditions under which biomanipulation willwork are still unclear. DeMelo et al. (1992) indicatedthat a careful analysis of the data does not support the eutrophic/oligotrophic differences proposed byMcQueen (1990) and others. Differences in the N andP requirements for the growth and reproduction ofDaphnia further complicate the situation. For example,Daphnia growth and reproduction are strongly sup-pressed when they are fed P-limited algae having a highC:P ratio (Sommer, 1992; Urabe et al., 1997; MacKayand Elser, 1998). When zooplankton that have high Pneeds ingest phytoplankton with high N:P ratios, thereis a disproportionate release of unused N to the system,which in turn can affect algal community structurebecause of changing nutrient ratios in the water(Urabe, 1993; Steinman, 1996).

Biomanipulation can have effects on componentsof the aquatic ecosystem that are not directly involvedin the trophic cascade. For example, Spencer and King(1984) reported that zooplankton successfully reducedphytoplankton densities in ponds with no fish or withdense populations of largemouth bass (a piscivore), butthe resulting clear water stimulated dense growths ofCladophora spp. and the submersed vascular plantsElodea canadensis and Potamogeton spp. This kind ofa shift is frequently the outcome of any control ornutrient manipulation technique that reduces phyto-plankton growth. Biomanipulation as a tool is stillexperimental and should not be recommended withoutconsiderable analysis of the composition of the varioustrophic levels and regulating environmental factors.

3. Biological Controls

The use of one organism to control unwantedorganisms has been widely studied in aquatic plantmanagement. Most of the research has focused on thecontrol of aquatic vascular plants, but there are studiesthat suggest potentially useful agents for algae control.

Reports that cyanophages (viruses) lyse cyanobac-terial cells date to the 1960s (Safferman and Morris,1963; Safferman et al., 1969). The first describedcyanophage was named LPP-1 for its ability to lyse

cells of Lyngbya, Phormidium, and Plectonema(Safferman and Morris, 1963). Since then, other cyano-phages have been isolated and evaluated as possiblebiocontrol agents (Padan et al., 1971; Stewart and Daft,1977; Martin and Benson, 1988; Phlips et al., 1990;Monegue and Phlips, 1991). Although most studies havebeen conducted in the laboratory, a few have shownsuccessful results in experimental field enclosures andponds (Martin et al., 1978; Desjardins and Olsen,1983). It is generally agreed that cyanophage applica-tion is most effective when it is applied before the hostpopulations are well established (Desjardins and Olsen,1983; Monegue and Phlips, 1991), but its impact on established cyanobacterial populations is poorlyknown. Eukaryotic algae, including the green algaChlorella, are also susceptible to viruses (Van Etten etal., 1991), but their biocontrol potential is unknown.

Although cyanobacterial and eukaryotic algal populations clearly are affected by phages in natureand although lysis can be induced in the laboratoryand in small-scale tests, no virus has been developedfor control purposes. Extensive field testing under avariety of environmental conditions has not been con-ducted, and the information needed on amount of ino-culum, application technique, and the environmentalfactors conducive for replication and lysis is not avail-able. The selectivity of viruses to one or a few speciesfurther limits the broad application of the technique.Other microorganisms that have shown activity onplanktonic algae include bacteria (Shilo, 1967; Burnhamand Fraleigh, 1983; Walker and Higginbotham, 2000)and fungi (Redhead and Wright, 1978; Canter andJaworski, 1979; Kudoh and Takahashi, 1990). Clearly,all of these organisms play a role in natural succession-al patterns in lakes, but their potential as biologicalcontrols remains unexplored. Even so, this approachshows promise for the future.

A wide variety of insect and other invertebrategrazers, including snails, caddisfly larvae, mayfly larvae, chironomid larvae, and shrimp have reducedbenthic algal growths (Fulton, 1988; Steinman, 1996),but their effect on prolific growths of filamentous algae(outside of river and stream systems) is apparently minimal, and none has been investigated as a potentialbiological control agent. The crayfish Oronectesimmunis significantly reduced stands of Chara (andsubmersed vascular plants) in a New York lake (Letsonand Makarewicz, 1994), but the treatment was moreexpensive and a higher stocking density was requiredcompared with treatments with grass carp (discussedbelow). When the vegetation was removed, the crayfishthemselves became subject to predation. Therefore, it wasdifficult to maintain sufficient crayfish densities to con-sume vegetation regrowth without additional stocking.


A number of fish species have been investigated for their potential to control algae and aquatic vascularplants. The silver carp (Hypophthalmichthys molitrix)and bighead carp (Aristichthys nobilis) consume phyto-plankton and zooplankton (Dimitrov, 1984; van derZweerde, 1993). Many tilapia (Tilapia spp.) are herbiv-orous (Hauser et al., 1976; Smith, 1985). Some are filter-feeders that consume phytoplankton; others feedon macrophytes including filamentous algae andChara. The distribution of tilapia is restricted by tem-perature (they are native to India, Africa, and SouthAmerica); they do not survive in waters colder than10°C. Therefore, in the United States their use has beenconfined to the South and to sites that receive heateddischarge (Crutchfield et al., 1992). In addition, theiruse has many disadvantages, such as the ability toswitch to animal food when they have eliminated plantand algal growth, a high reproductive potential, andinterference with native fish species.

The most successful and widely used biologicalcontrol agent for aquatic vascular plants and somealgae has been the grass carp (Ctenopharyngodonidella; Fig. 10) (see reviews by Cooke et al., 1993b; vander Zweerde, 1993). This fish is native to northernChina and was introduced into the United States in the1960s. It was originally introduced into Arkansas butis now used in at least 35 states for aquatic weed control (Sanders et al., 1991). Because there are nopredators in its native range, grass carp do not showtypical avoidance behaviors and are extremely suscept-ible to predation. Therefore, fish must be at least 20 to25 cm long (~450 g) when stocked to avoid predationby largemouth bass and other native predators. Underideal conditions grass carp can grow to a weight of atleast 23 kg within 5–10 years. Recommended stocking

densities vary from region to region. Recommendationsfrom the Indiana Department of Natural Resources, forexample, suggest 37 grass carp vegetated ha–1 if main-tenance of some vegetation is desired and 74 fish ha–1

if elimination of vegetation is desired. The grass carpsurvives in cold water and begins to feed regularly at about 14°C. Feeding peaks at about 20–26°C anddecreases when the water temperature reaches about33°C.

The concern about the potential for grass carp to reproduce and crowd out native species led to thedevelopment of sterile, triploid grass carp. Even withthis precaution, grass carp must not be introduced inareas in which the elimination of vegetation in wetlandareas would destroy valuable habitat for waterfowl andother animal life.

Young grass carp up to 50 mm (about 2 in) inlength feed mostly on zooplankton. After that theyshift to a diet of filamentous algae, duckweed, and sub-mersed vascular plants. There has never been any evi-dence of the fish shifting to an animal diet once theyexceed a length of about 100 mm. The fish clearly haspreferences for the kinds of plants it eats. Numerouslists of preferred plant species have been published(Fowler and Robson, 1978; Cassani and Caton, 1983;Shireman et al., 1983a; Pine and Anderson, 1991; Sanderset al., 1991; Cooke et al., 1993b), and in almost everycase, Chara and Nitella are listed as preferred or highlypreferred plants. Bauer and Willis (1990) reported thatgrass carp introduced at 49 fish ha–1 almost totallyeliminated Chara (the dominant macrophyte) in 2 yearsin two small South Dakota lakes.

The effectiveness of grass carp for the control offilamentous algal mats is less clear. Some of the refer-ences listed above claim good control of relativelycoarse species such as Cladophora and Pithophora;others indicate no or only weak control of filamentousalgae. In some cases, only high stocking densities ofmore than 123 fish ha–1 have succeeded in controllingfilamentous algae. Spirogyra seems to be least pre-ferred, probably because its slimy nature prevents effec-tive ingestion; however, it too can be eaten if no othervegetation is available (Lembi et al., 1978). In mixedpopulations of vegetation, grass carp will clearly con-sume soft-bodied vascular plants such as pondweeds,elodea, and naiads, and Chara (even though Chara canbe coated with a hard coat of calcium carbonate, fishappear to be attracted to it) in preference to filamen-tous algae. When filamentous algae are the only plantmaterial present, grass carp will feed on it, probably toavoid starvation.

Problems with grass carp include lack of consistencyand predictibility. Use of grass carp does not workunder all circumstances, and visible control may not be

820 Carole A. Lembi

FIGURE 10 The grass carp (Ctenopharyngodon idella), a widelyused biological control agent for certain macrophytic algae and sub-mersed vascular plants.

achieved until several years after introduction, particu-larly in heavily infested areas. Increased turbidity dueto increased phytoplankton populations has been notedin some situations (Shireman et al., 1985; Maceina et al., 1992), and overstocking can result in the totalelimination of all vegetation, a situation that is notconsidered beneficial for fish and other animal life in natural bodies of water. In fact, considerable contro-versy in the sportfishing industry has erupted over thepotential for elimination of vegetation cover by grasscarp, particularly because the effect of removing weedbeds on angling success is still being debated (Bain,1993; Bettoli et al., 1993; Killgore et al., 1998). Othercontrol methods (e.g., mechanical harvesting or the use of chemicals) can be used for weed beds in certainareas of a lake so that some vegetation can be selectivelyretained. Unfortunately, there is little ability, once grasscarp have been introduced, to dictate where they willgraze and how much vegetation they will consume overtime. Stocking rates can be reduced to avoid elimina-tion of plant populations, but because the grass carpwork so slowly and because it is so difficult to predicttheir impact on the vegetation, herbicides/algicides orother methods may have to be used if the fish cannotprovide adequate control. Finally, grass carp may consume desirable native species (including Charaand Nitella) and leave less desirable species, such as the invasive weed Eurasian watermilfoil (Fowler andRobson, 1978).

Another method of biological control is the use ofwaterfowl, specifically geese or swans. Filamentousalgae are consumed by waterfowl, and charophytes area favored food of herbivorous ducks, coots, and swans(Martin et al., 1961; Hargeby et al., 1994). A pair ofswans reportedly will keep a 0.4 ha pond free of submersed vegetation, as will 7–20 geese or ducks ha–1

(Holm and Yeo, 1981). Unfortunately, there are manyproblems associated with the presence of waterfowl.Birds that are introduced for aquatic weed control areusually rendered flightless. Therefore, their diet ofaquatic vegetation must be supplemented to provideadequate nutrition, they must be protected from preda-tors, and lake managers must be willing to toleratetheir aggressiveness during the breeding season. Aswith free-living waterfowl, their waste materials can litter the banks and stimulate phytoplankton blooms.

An unusual form of biological control is the use of competitive plants. Doyle and Smart (1998) showedthat established plantings of the flowering plants pickerelweed (Pontederia cordata) and Americanpondweed (Potamogeton nodosus) reduced the bio-mass of L. wollei by 50% and prevented the formationof floating mats. The effect was attributed to shadingand possibly to competition for nutrients in the sedi-

ments. The main drawback of this method is that itonly works in very shallow areas where the floweringplants can root. In a larger sense, the water user mustbe willing to accept the shift from an algal infestationto dense stands of flowering plants. This fact alonenegates the potential benefits of shading by free-float-ing flowering plant species such as water hyacinth(Eichhornia crassipes) or duckweed (Lemna spp.), bothof which tend to be weedy. On the other hand, thepremise of using competitive submersed plants, possi-bly those that have been selected or genetically modi-fied to produce shortened stems with minimal canopyformation, should be explored further.

The use of biological agents shows great promisefor the control of weedy algae and plants, but it hasdrawbacks. More research is necessary, particularly onthose organisms that might be relatively selective, suchas phages. The introduction of nonselective agents,such as grass carp, has been used successfully in somesituations, but it also has the potential to cause adverseecological impacts in others.

4. Allelochemicals

Allelopathic chemicals are chemicals produced by plants that have either an adverse or beneficial effect (usually adverse) on other plants (Rice, 1984).Although this area has not received much attention incontrolling nuisance algae, there is evidence that allelo-chemicals may be useful.

A number of cyanobacteria have been investigatedfor their potential to produce allelochemicals thatinhibit the growth of other cyanobacteria or algae(Mason et al., 1982; Flores and Wolk, 1986). Allelo-chemicals from some fungi (Redhead and Wright,1978) and terrestrial vascular plants (mostly phenyl-propanoids) (Della Greca et al., 1992) have been shownto have algicidal activity. Studies have been conductedon allelochemicals produced by aquatic vascular plants,but most of the bioassays have been conducted usingvascular plant species, such as duckweed or lettuceseedlings, rather than algae (Elakovich and Wooten,1989; Sutton and Porter, 1989; Wooten and Elakovich,1991). One exception is the study by Gross et al.(1996) in which extracts from Eurasian watermilfoilinhibited cyanobacteria. Although some potential algicidal allelochemicals have been identified, the majorconstraints to further development is the expense ofculturing the organisms and extracting sufficientamounts of the allelochemicals for application. Analternative is to synthesize the active chemical, but this is also costly, particularly in view of the relativecheapness and availability of copper sulfate. Thus, the financial incentive for industry to develop thesecompounds for the aquatic market is lacking.


A cheaper method of allelopathic control may bethe “bale of hay” technique. For years, farmers haveapplied straw or hay to ponds to reduce algal growth.This method has been tested and substantiated in aseries of experimental studies in England (Welch et al.,1990; Gibson et al., 1990; Pillinger et al., 1992, 1994;Newman and Barrett, 1993). In the laboratory, rottingbarley (Hordeum) straw inhibited the growth of severalplanktonic (including Microcystis) and filamentousalgae. Additions to a canal over a 3-year period resultedin reduced biomass of C. glomerata after 2 years. Theapparent mechanism is through the release of quinonecompounds (Pillinger et al., 1994).

Attempts to replicate these effects in NorthAmerican waters have had mixed results. Nicholls(1996) showed decreases in chlorophyll a in pondstreated with barley in Ontario, but tests using barley,wheat (Triticum), and rye (Secale) in ponds withPithophora in North Carolina were unsuccessful (Kay,1997). Discrepancies among results may be due to differences in straw concentrations, target species, envi-ronmental conditions, and length of exposure. Becausebarley is not grown in many parts of the United Statesand thus is not widely available, other forages shouldbe tested. Laboratory and small-scale field testing suggests that alfalfa (Medicago) hay may be effectivefor filamentous algae control (Marencik and Lembi,1998), but additional research is needed to determine ifits rapid breakdown in water could cause oxygendepletion problems. Further research in this area iswarranted.

5. Algicides

A common method of controlling algal infestationsis the use of chemicals (algicides). Of the algicides, themost commonly used compounds are copper-contain-ing products. Other products, such as diquat and themono (N,N-dimethylalkylamine) salt of endothall, areregistered for algae control, but their use is relativelyminor compared to that of the copper-containing products.

Of the copper products, copper sulfate (CuSO4) isthe most widely used algicide for controlling algal populations in water supplies, recreational lakes, andreservoirs (Elder and Horne, 1978; Effler et al., 1980;Raman, 1985). It has been used since at least 1905 andprobably was used earlier (Moore and Kellerman,1905; Murphy and Barrett, 1993). In the late 1960sand early 1970s, more than 9 million kg of CuSO4

were applied annually in waters in the United States(Fitzgerald, 1971). Approximately 68% of all waterarea in the United States treated with a chemical product in 1992 was treated with an algicide (un-published industry data). CuSO4 was applied to 70%

of this area, and the remainder was treated mostly with other copper compounds (chelated copper com-pounds). Even though algicides were used on 68% ofthe area treated, they accounted for just 20% of thetotal sales ($32.5 million) of all aquatic chemicals (her-bicides and algicides) due to their relatively low cost.

Copper is effective on a wide range of algae(Maloney and Palmer, 1956). The toxic agent is freecupric ion (Cu2+), and toxic cupric ion activities rangefrom greater than 10–6 to 10–11 M for species ofdiatoms, dinoflagellates, microscopic green algae, andcyanobacteria (McKnight et al., 1983). The fact thatcyanobacteria are more sensitive to copper than someof the eukaryotic algae (Whitton, 1973; Swain et al.,1986) accounts for its widespread success and accept-ance. Diatoms are probably next in sensitivity followedby the green algae (Swain et al., 1986; Havens, 1994).A copper concentration of 25–40 μg L–1 effectivelycontrolled A. flos-aquae in shallow eutrophic lakes inManitoba (Whitaker et al., 1978), a dose that is con-siderably lower than the typical doses of 125–250 μgL–1. Suppression of nitrogen fixation by Anabaena andAphanizomenon was observed after copper additionsof only 5–10 μg L–1 (Horne and Goldman, 1974), lead-ing to the suggestion that maintaining low doses toinhibit N2 fixation would be an alternative to a singlelarge dose. This approach has not been practicalbecause the short residence time of copper in the watercolumn mandates almost continual copper application.

Certain microscopic green [e.g., Oocystis (Meadoret al., 1993; personal observations)] and euglenoidplanktonic (Hawkins and Griffiths, 1987) algae are rel-atively tolerant to copper treatments. For example, themicroscopic green algae Ankistrodesmus, Scenedesmus,and Pandorina may require copper concentrations ashigh as 500 μg L–1 for control (Copper Sulfate FineCrystals product label).

Among mat-forming green algae, Spirogyra andOedogonium are very susceptible to copper (Whitton,1970; Francke and Hillebrand, 1980), whereasPithophora and Hydrodictyon are considerably moretolerant (Table I). In addition to inherent tolerance orsusceptibility, mat structure also may dictate relativetolerance to copper (or other exogenously appliedmaterials). Mat structure appears to be governed inpart by branching pattern. Pithophora filaments, whichare branched, produce intertwined, tighter mats thanfilaments of Spirogyra and Oedogonium, which areunbranched (Table I). It may be more difficult for copper to penetrate the extremely dense, massive matsthat are produced by Pithophora (Lembi et al., 1984)than the loose, less tangled mats formed by Spirogyraand Oedogonium. Of all mat-forming species, thecyanobacteria are the most tolerant to copper, which

822 Carole A. Lembi

seems unusual given the susceptibility of planktoniccyanobacteria. Both an inherent tolerance (at least 6-fold greater than that of Pithophora; Table I) plus the presence of thick slime and a coating of sedimentmake Oscillatoria mats extremely difficult to control.Likewise, Lyngbya (which also produces sheaths) produces thick, dense mats which probably add to itstolerance. Unfortunately, it is likely (although not welldocumented) that the elimination of susceptible species(both microscopic and mat-forming) has led to theirreplacement by tolerant species.

The mechanisms by which copper affects algaeappear to vary. The list of reported copper effects[taken from Gledhill et al. (1997); see references there-in] indicates that it inhibits photosynthesis (see alsoKallqvist and Meadows, 1978), disrupts electron trans-port in photosystem II (see also Cedeno-Maldondo andSwader, 1974), reduces pigment concentrations, affectsthe permeability of the plasma membrane and induceslosses in cations, inhibits nitrate uptake, restrictsgrowth, affects cell motility, and affects the distributionof proteins, lipids, sterols, sterol esters, and free fattyacids in the cell. Although this list was developed pri-marily for marine algae, there is no reason to think thatthe same effects should not be expected in freshwateralgae. In addition, copper has been reported to inhibitP uptake (Peterson et al., 1984) and to precipitate proteins in the cell (Murphy and Barrett, 1993). All ofthese presumed modes of action suggest that copper isa general algal cell toxicant. When applied at the recommended dosage (250 μg L–1 copper), copper actsrapidly, usually within a period of hours.

A number of mechanisms have been proposed toaccount for the differential tolerance among algae andinclude (Lage et al., 1996) intracellular accumulationof copper in polyphosphate bodies, storage of copperin membrane-bound vesicles, excretion into the medium of organic compounds that bind copper, intra-cellular chelation of copper by organic compounds likephytochelatins, and efflux of the copper. In addition,copper accumulation in the cell walls of some algaeand higher plants has been reported (Pearlmutter andLembi, 1986; Allan and Jarrell, 1989). As noted above,mat structure also should be considered a factor inspecies tolerance.

Water chemistry, particularly pH and alkalinity (ameasure of bicarbonates, carbonates, and hydroxides),plays an important role in copper toxicity (McKnightet al., 1983). Below neutral pH, Cu2+ is the major copper species; above neutral pH the major forms ofcopper are the copper carbonate complexes and mala-chite and tenorite. The various complexes and precipi-tants formed above neutral pH values effectivelyprevent Cu2+ from being taken up by target organisms.At high pH and alkalinity, the concentration of solubleCu2+ in the water is extremely low and possibly ineffec-tive for algal control (Button et al., 1977). In low alka-linity, acidic waters, the recommended dose (250 μg L–1

Cu2+) of CuSO4 will kill algae, but given the relativelyhigh amounts of soluble Cu2+, particularly sensitive fishspecies, such as trout, can also be killed. This phenom-enon is even true for the chelated copper products (discussed below), and all currently registered copperproducts have a statement similar to the one on theCutrine-Plus label: “Do not use in water containingtrout if the carbonate hardness of the water does notexceed 50 mg L–1.”

As total alkalinity increases, so must the CuSO4

dosage to overcome the precipitation problem. Toxicityto fish is essentially nonexistent at high alkalinitiesbecause of the low concentrations of Cu2+ in the water[as little as 0.5% of the total dissolved copper has beencalculated to be present as free cupric ion (Wagemannand Barica, 1979)]. For example, at the low alkalinityof 18.7 mg L–1 (as CaCO3 + HCO3), the LC50 (con-centration that will kill 50% of the population) forbluegill (Lepomis) is 884 μg L–1 copper (3.5 mg L–1

CuSO4) (Herbicide Handbook, 1994). At the moderateto high alkalinity of 166 mg L–1, the LC50 is 7300 μgL–1 copper (29.2 mg L–1 CuSO4). The upper legal limitfor copper use in water is 1000 μg L–1 copper (4 mg L–1

CuSO4); this concentration (which is seldom recom-mended) could kill fish in low alkalinity waters. How-ever, there is almost a 30-fold safety factor between theLC50 and the typical recommended dosage [250 μg L–1

copper (1 mg L–1 CuSO4)] in high alkalinity waters.


TABLE I Taxa, Filament Morphologies, Mat Structures, and Susceptibilities to Coppera of Filamentous Mat-Forming Algae

Taxon Morphology Mat structure EC50b

ChlorophytaSpirogyra Unbranched Loose 1

Oedogonium Unbranched Loose 3

Hydrodictyon Net-like Moderate 48Pithophora Branched Dense 46

CyanobacteriaOscillatoria Unbranched Dense, slime + 290

associated sedimentLyngbya Unbranched Dense 1630

aLembi, 2000; data on Lyngbya from Hallingse and Phlips (1996).bEC50 = concentration of Cu2+ in μg L–1 required to reduce biomass(dry weight) of alga by 50% under laboratory (not field) conditions.

Although direct copper toxicity is seldom a prob-lem to fish, the depletion of oxygen during algal deathand decomposition can cause fish kills. Treatmentsmust never be made to bodies of water that have heavyalgal infestations.

Chelation with organic compounds stabilizes soluble copper and theoretically retards its precipita-tion and adsorption (McKnight et al., 1983). This principle is the basis for the formulation of the chelatedforms of copper as substitutes for copper sulfate. Mostcommercial chelated formulations are variations ofethanolamine complexes. One of the presumed advan-tages of using a copper chelate is that a longer persist-ence of active copper in water should increase algalcontact and control. In a comparative study of copperchelates and copper sulfate used at the same copperdose, Masuda and Boyd (1993) found that chelatesslowed the loss of total copper (from an initial level of500 to 100 μg L–1) from the water from 4.3 to 6.3 days,but they concluded that this advantage did not com-pensate for the greater cost of the chelated formulation.On the other hand, the chelated copper products pro-vide flexibility in application because they are formu-lated either as granules or as liquids, whereas CuSO4 ispackaged as a solid material only. The combination ofliquid formulations with other liquid aquatic herbicidesis useful for commercial applicators to control a broadspectrum of species that include both algae and vascular plants.

Phytoplanktonic blooms are usually treated bypumping the copper compounds (dissolved in water ina tank mounted in the boat or airboat) through a boomand trailing hoses into the water (Fig. 11A). The hosescan be adjusted to deliver the compound to differentdepths in the water column. For example, hoses thatdisperse the copper in the upper meter of water (or surface treatments with a spray) are effective for thecontrol of cyanobacteria that have formed surfacescums. The hoses can be set lower in the water columnto deliver copper to filamentous algal mats that are still lying on the bottom and have not floated to thesurface. Once the mats have floated to the surface, spottreatments directly on the mats with a spray can ensurethat contact with the cells is maximal (Fig. 11B). Thisability to place the treatment directly on or near thetarget organism allows the applicator to selectivelytreat some areas and not others, thus reducing both theamount of copper needed and the volume of water thatcomes into contact with the chemical. Other methodsof application include slow dispersal through a burlapsack or drip or single high dose applications in flowingwater systems such as irrigation and drainage canals,.

The persistence time of copper in water is relativelyshort. Button et al. (1977) found that 95% of the

copper sulfate distributed over a lake surface dissolvedin the upper 1.8 m of the water column, but the totalcopper concentration was at pretreatment levels within24 h. Tucker and Boyd (1978) made 10 applications of 0.84 kg ha–1 CuSO4 at 2-week intervals to pondswithout causing an appreciable increase in total copperconcentration. Wagemann and Barica (1979) reporteda half-life of total dissolved copper from 1 to 7 days.Anderson and Dechoretz (1984) reported a half-life ofabout 1 day with all of the copper gone from the watercolumn at 14–28 days. In other bodies of water, copperhas persisted for up to 30 days after treatment (Elderand Horne, 1978; Whitaker et al., 1978; McKnight,1981; Hawkins and Griffiths, 1987).

The implications of a short residence time in waterare several. First of all, copper effects, whether on target or nontarget organisms, are temporary. In mostcases, algal populations rebound, although not neces-sarily at the same densities or with the same species.Although this is a problem from the standpoint that

824 Carole A. Lembi

FIGURE 11 Algicide applications. (A) A unit set up to deliver chemi-cal through trailing hoses. (B) A unit set up to deliver a spray directlyto the algal mats. B courtesy of Neil Gerber, Aquatic Management,Bluffton, Indiana.

repeated treatments may be needed during a single season to provide adequate control, the short persist-ence reduces the exposure time of nontarget organisms,including humans.

One of the major concerns with the use of copperis its ultimate fate. The copper complexes, as well asdecaying copper-containing algae, fall to the bottomwhere the copper is readily adsorbed onto sediments.Copper, as a heavy metal, persists in the sediment forprolonged periods of time (Frank, 1972; Brown, 1978).The key issue is that copper will accumulate in sedi-ments and be toxic to benthic organisms and then serveas a source of copper to the water after treatment isdiscontinued. The evidence is somewhat conflicting.Sanchez and Lee (1978) noted that copper-enrichedsediments in Lake Monona, Wisconsin (treated over 50 years to 1950), were not interacting with the morerecent sediments deposited or with overlying waters.The copper content of the water was no different from that of local hardwater lakes which had not beentreated, and they concluded that there were no long-term adverse effects resulting from the copper treat-ments. Ankley et al. (1993) studied sediment and porewater from Steilacoom Lake, Washington, which hadbeen “grossly contaminated” by copper because ofcopper sulfate treatments. Extracted copper concentra-tions in sediments ranged from 0.6 to 3.0 μmol g–1 dryweight, but pore water and overlying water concentra-tions were less than the analytical detection limit of7 μg L–1. Toxicity tests showed no effects of the wateron the amphipod Hyalella azteca.

Probably the most negative report of copper treat-ment effects on sediments is from a string of intercon-nected lakes in southern Minnesota that were treatedover a period of 58 years (Hanson and Stefan, 1984).Effects included elevated concentrations of copper inthe sediments, fish kills due to oxygen depletion or possibly copper toxicity, increased internal P cycling,rapid recovery of algal populations within 7–21 days,shifts of game fish to rough fish, disappearance ofmacrophytes, and reductions of benthic macroinverte-brates. Some conditions improved when the sedimentshad been dredged and the use of copper discontinued.It is difficult to evaluate the results of this studybecause factors other than copper applications mayhave caused some of these effects. Fortunately, suchextreme effects have not been observed in most copper-treated lakes. If they did occur in large numbers, copper should have been banned long ago. In fact,Sanchez and Lee (1978) concluded that 50 years ofcopper treatments in Lake Monona had not resulted inthe loss of the excellent sport fisheries supported bythat lake. Nevertheless, the Minnesota study shouldserve as a warning that a program that includes

long-term, whole-lake treatments with copper must bescrutinized and continually monitored for potentialdeleterious effects.

An additional concern is the sensitivity of zoo-plankton to copper. The LC50 for planktonic crus-taceans is 60–90 μg L–1 copper; for rotifers it is1100–1700 μg L–1 (Demayo et al., 1982). Other studieson invertebrates indicate LC50 values ranging from 10 to 130 μg L–1 copper (McIntosh and Kevern, 1974;Winner, 1985; Meador et al., 1993). Concentrations aslow as 8 μg L–1 copper have caused significant effectson cladocerans in life cycle toxicity tests (Belanger etal., 1989), and Hedtke (1984) found large reductionsin zooplankton biomass (along with that of snails, total macroinvertebrates, and midges) when laboratorymicrocosms were treated with 30–270 μg L–1 copper.Most of these concentrations are well within the rangeof normal use dosages. Studies using lake mesocosmsalso showed reductions in zooplankton populations(Moore and Winner, 1989; Havens, 1994) as well asspecies shifts (Moore and Winner, 1989; Winner et al.,1990; Havens, 1994). Both phytoplankton and zoo-plankton communities were more sensitive to copper inthe spring than in the summer or fall (Winner et al.,1990), an observation that was attributed to differ-ences in levels of copper-complexing compounds in thewater during the various seasons.

Effects of copper on zooplankton communities inlakes have been somewhat mixed. Effler et al. (1980)found no effects of a low-level CuSO4 treatment onseveral zooplankton populations in a lake; however,McKnight (1981) found decreases of populations ofBosmina, Tetramastix, and Keratella following treat-ment of Mill Pond, Massachusetts. Long-term effects ofcopper treatments on zooplankton in lakes have notbeen adequately studied, although the short-term persistence of copper within the water column shouldallow most zooplankton populations to recover. Theloss of zooplankton populations, even if temporary,may explain why algal populations can recover, some-times to levels higher than original levels. The loss ofgrazing impact on phytoplankton due to copper effectson zooplankton has been suggested by the work ofMcKnight (1981), Taub et al. (1989), and Havens(1994). In addition, bacterial biomass has been reported to rapidly recover or even increase followingcopper treatment (Effler et al., 1980; Havens, 1994;Dionigi and Champagne, 1995), and in some cases this may be due to loss of grazing pressure from zoo-plankton (Havens, 1994).

Copper is a trace element that is required for thesurvival of many plant and animal species, includinghumans. The low mammalian toxicity of copper whendiluted in water, its short persistence time in water,


and its lack of bioaccumulation in fatty tissues are thereasons that it is the only one of the algicides/herbicidesregistered in the United States for which the use ofwater following treatment at normal doses is notrestricted. This includes use of the water for drinking,swimming, fishing, livestock watering, and irrigation,although a 24-hour waiting period after treatment isdesirable just to be cautious. Copper-containing com-pounds have not been reported to induce cancers inhumans or experimental animals (Sunderman, 1978).The only warning on EPA-approved products is thepotential toxicity to trout under softwater conditions.

6. Summary of Direct Control Methods

Methods for the direct control of algae are avail-able, but none of them ensure that algae problems willbe solved other than in the short term. Even grass carp,which live for up to 16 years, become less efficientfeeders after about 5 years and restocking is requiredfor long-term control.

The adverse effect of copper on food webs must be considered, and appropriate long-term ecosystemstudies must be undertaken. These concerns are offsetin part by the overall safety of copper to humans, bythe short persistence of copper in the water column,and by our ability to place copper directly on or in thevicinity of the target algae. However, lakes and reser-voirs will be better protected if repeated whole-lakecopper treatments can be avoided. A program of water-shed and water quality management through nutrientmanipulation is the best alternative, particularly forphytoplanktonic blooms.

There will still be a role for the use of copper sulfate, at least in the United States. Copper treatmentsmay be needed periodically or at certain sites where filamentous mat-forming algae or Chara are problemsin lakes and reservoirs despite nutrient reductionefforts. Residents on small lakes and ponds that receivelarge inputs of nutrients from nonpoint sources willcontinue to require direct control techniques, particu-larly because they may have few financial or politicalresources to minimize these inputs. The ability to main-tain irrigation systems in the western states free ofalgae and provide maximal water delivery rates tourban and rural users will still require copper applica-tions for the foreseeable future.

Regulatory agencies in several regions of theUnited States have indicated an interest in eliminatingthe use of copper for algae control. Copper is not used to any great extent in Canada because of thatcountry’s very cautious approach to water quality (H.Vandermeulen, Fisheries & Oceans Canada, personalcommunication). However, the immediate benefits andfavorable economics of copper sulfate usage suggest

that this compound will continue to be used in theUnited States. Unfortunately, the ready availability andefficacy of the copper compounds reduce incentivesneeded for development of alternative control methods,such as organically based algicides and allelochemicalsor viral and bacterial biological control organisms.

The short-term nature of direct control techniquesplus the difficulties inherent in regulating nutrientinputs in every situation dictate that algae will posewater quality problems for many years to come. Theseproblems will continue to pose a challenge to all of us who have an interest in maintaining or restoringhealthy aquatic ecosystems.

ACKNOWLEDGMENTS

I thank Erik C. Brockman for his tremendous assis-tance in the literature search for this chapter. I alsoacknowledge the contributions and insights of numer-ous associates, both academic and those who deal withalgal management on a day-to-day basis. The construc-tive comments of reviewers Alan D. Steinman andHerb Vandermeulen are greatly appreciated.

LITERATURE CITED

Adamson, R. P., Sommerfeld, M. R. 1978. Survey of swimming poolalgae of the Phoenix, Arizona, metropolitan area. Journal ofPhycology 14:519–521.

Allan, D. L., Jarrell, W. M. 1989. Proton and copper adsorption tomaize and soybean root cell walls. Plant Physiology89:823–832.

Allanson, B. R. 1973. The fine structure of the periphyton of Charasp. and Potamogeton natans from Wytham Pond, Oxford, andits significance to the macrophyte–periphyton metabolic modelof R. G. Wetzel and H. L. Allen. Freshwater Biology3:535–542.

American Public Health Association. 1992. Standard methods for theexamination of water and wastewater, 18th ed. Greenberg, A.E., Clesceri, L. S., Eaton, A. D., Eds. American Public HealthAssociation, Washington, DC.

Anderson, L. W. J. 1993. Aquatic weed problems and management inthe western United States and Canada, in: Pieterse, A. H.,Murphy, K. J., Eds., Aquatic weeds, the ecology and manage-ment of nuisance aquatic vegetation. Oxford University Press,Oxford, UK, pp. 371–391.

Anderson, L. W. J., Dechoretz, N. 1984. Laboratory and field investi-gations of a potential selective algicide, PH4062. Journal ofAquatic Plant Management 22:67–75.

Ankley, G. T., Mattson, V. R., Leonard, E. N., West, C. W., Bennett,J. L. 1993. Predicting the acute toxicity of copper in freshwatersediments: Evaluation of the role of acid-volatile sulfide.Environmental Toxicology and Chemistry 12:315–320.

Anonymous. 1990. Environmental assessment, aquatic plant man-agement (NR 107) program, 3rd ed. Wisconsin Department ofNatural Resources, Madison.

Anthoni, U., Christophersen, C., Madsen, J. O., Wium-Andersen, S.,

826 Carole A. Lembi

Jacobsen, N. 1980. Biological active sulfur compounds from thegreen alga Chara globularis. Phytochemistry 19:1228–1229.

Arruda, J. A., Fromm, C. H. 1989. The relationship between tasteand odor problems and lake enrichment from Kansas lakes inagricultural watersheds. Lake and Reservoir Management5:45–52.

Babin, J., Prepas, E. E., Murphy, T. P., Serediak, M., Curtis, P. J.,Zhang, Y., Chambers, P. A. 1994. Impact of lime on sedimentphosphorus release in hardwater lakes: The case for hyper-eutrophic Halfmoon Lake, Alberta. Lake and ReservoirManagement 8:131–142.

Badgery, J. E., McQueen, D. J, Nicholls, K. H., Schaap, P. R. 1994.Biomanipulation at Rice Lake, Ontario, Canada. Lake andReservoir Management 10:163–173.

Bain, M. B. 1993. Assessing impacts of introduced aquatic species: Grasscarp in large systems. Environmental Management 17:211–224.

Barica, J. 1975. Summerkill risk in prairie ponds and possibilities ofits prediction. Journal of the Fisheries Research Board ofCanada 32:1283–1288.

Barica, J. 1978. Collapses of Aphanizomenon flos-aquae bloomsresulting in massive fish kills in eutrophic lakes: Effect of weather.International Vereinigung für Theoretische und AngewandteLimnologie Verhandlungen 20:208–213.

Barica, J., Kling, H., Gibson, J. 1980. Experimental manipulation ofalgal bloom composition by nitrogen addition. CanadianJournal of Fisheries and Aquatic Sciences 37:1175–1183.

Barko, J. W., Smart, R. M. 1980. Mobilization of sediment phospho-rus by submersed freshwater macrophytes. Freshwater Biology10:229–238.

Barko, J. W., Smart, R. M. 1981. Sediment-based nutrition of sub-mersed macrophytes. Aquatic Botany 10:339–352.

Barko, J. W., Adams, M. S., Clesceri, N. L. 1986. Environmental fac-tors and their consideration in the management of submersedaquatic vegetation: A review. Journal of Aquatic PlantManagement 24:1–10.

Barko, J. W., James, W. F., Taylor, W. D., McFarland, D. G. 1990.Effects of alum treatment on phosphorus and phytoplanktondynamics in Eau Galle Reservoir: A synopsis. Lake andReservoir Management 6:1–8.

Barnett, B. S., Schneider, R. W. 1974. Fish populations in dense sub-merged aquatic plant communities. Hyacinth Control Journal12:12–14.

Bauer, D. L., Willis, D. W. 1990. Effects of triploid grass carp onaquatic vegetation in two South Dakota lakes. Lake andReservoir Management 6:175–180.

Beeton, A. M. 1969. Changes in the environment and biota of theGreat Lakes, in: Eutrophication, causes, consequences, correc-tives. National Academy of Sciences, Washington, DC, pp.150–187.

Belanger, S. E., Farris, J. L., Cherry, D. S. 1989. Effects of diet, waterhardness, and population source on acute and chronic coppertoxicity to Ceriodaphnia dubia. Archives of EnvironmentalContamination and Toxicology 18:601–611.

Benndorf, J. 1989. Food-web manipulation as a tool in water-qualitymanagement. Aqua (JWSRT) 38:296–304.

Bennett, G. W. 1971. Management of lakes and ponds. VanNostrand Reinhold, New York.

Bettoli, P. W., Maceina, M. J., Noble, R. L., Betsill, R. K. 1992.Piscivory in largemouth bass as a function of aquatic vegetationabundance. North American Journal of Fisheries Management12:509–516.

Bettoli, P. W., Maceina, M. J., Noble, R. L., Betsill, R. K.1993.Response of a reservoir fish community to aquatic vegetationremoval. North American Journal of Fisheries Management13:110–124.

Blindow, I. 1988. Phosphorus toxicity in Chara. Aquatic Botany32:393–395.

Bolas, P. M., Lund, J. W. G. 1974. Some factors affecting the growthof Cladophora glomerata in the Kentish Stour. Water Treatmentand Examination 23:25–51.

Born, S. M., Wirth, T. L., Brick, E., Peterson, J. O. 1973. Restoringthe recreation potential of small impoundments: The MarionMillpond experience. Bulletin No. 71, Wisconsin Department ofNatural Resource Technology, Madison, 20 p.

Boyd, C. E. 1982. Water quality management for pond fish culture.Elsevier, Amsterdam.

Boyd, C. E., Prather, E. E., Parks, R. W. 1975. Sudden mortality of amassive phytoplankton bloom. Weed Science 23:61–66.

Brezny, D., Mehta, I., Sharma, R. K. 1973. Studies of evapotranspi-ration of some aquatic weeds. Weed Science 21:197–204.

Brown, A. W. A. 1978. Herbicides in water, in: Brown, A. W. A., Ed.,Ecology of pesticides. John Wiley, Sons, New York, pp.426–431.

Brown, S. W., Boyd, C. E. 1982. Off-flavor in channel catfish fromcommerical ponds. Transactions of the American FisheriesSociety 111:379–283.

Brownlee, B. G., Painter, D.S., Boone, R. J. 1984. Identification oftaste and odour compounds from western Lake Ontario. WaterPollution Research Journal of Canada 19:111–118.

Burkholder, J. M., Mallin, M. A., Glasgow, H. B., Larsen, L. M.,McIver, M. R., Shank, G. C., Deamer-Melia, N., Briley, D. S.,Springer, J., Touchette, B. W., Hannon, E. K. 1997. Impacts to acoastal river and estuary from rupture of a large swine wasteholding lagoon. Journal of Environmental Quality 26:1451–1466.

Burnham, J. C., Fraleigh, P. D. 1983. Predatory myxobacteria: Lyticmechanisms and prospects as biological control agents forcyanobacteria (blue-green algae), in: Lake restoration, protec-tion and management. EPA-440/5-83-001, pp. 249–256.

Burns, C. W. 1968. The relationship between body size of filter-feed-ing Cladocera and maximum size of particle ingested.Limnology and Oceanography 13:675–678.

Button, K. S., Hostetter, H. P., Mair, D. M. 1977. Copper dispersal ina water supply reservoir. Water Research 11:539–544.

Canfield, D. E. 1983. Prediction of chlorophyll a concentration inFlorida lakes: The importance of P and N. Water ResearchBulletin 19:255–262.

Canter, H. M., Jaworski, G. H. M. 1979. The occurrence of a hyper-sensitive relation in the planktonic diatom Asterionella formosaHassal parasitized by the chytrid Rhizophydium planktonicumCanter emend., in culture. New Phytology 82:187–202.

Carignan, R., Kalff, J. 1980. Phosphorus sources for aquatic weeds:Water or sediments? Science 207:987–989.

Carlson, R. E. 1977. A trophic state index for lakes. Limnology andOceanography 22:363–369.

Carmichael, W. W. 1994. The toxins of cyanobacteria. ScientificAmerican 270:78–86.

Carmichael, W. W. 1997. The cyanotoxins. Advances in BotanyResearch 27:211–256.

Carmichael, W. W., Evans, W. R., Yin, Q. Q., Bell, P., Moczydlowski,E. 1997. Evidence for paralytic shellfish poisons in the fresh-water cyanobacterium Lyngbya wollei (Farlow ex Gomont) comb.nov. Applied and Environmental Microbiology 63:3104–3110.

Carmichael, W. W., Falconer, I. R. 1993. Diseases related to fresh-water blue-green algal toxins, and control measures, in:Falconer, I. R., Ed., Algal toxins in seafood and drinking water.Academic Press, San Diego.

Carmichael, W. W., Jones, C. L. A., Mahmood, N. A., Thiess, W. C.1985. Algal toxins and waterbased diseases. CRC CriticalReviews in Environmental Science and Technology 15:275–313.


Carpenter, E. J., Carmichael, W. W. 1995. Taxonomy of cyanobacte-ria, in: Hallegraeff, G. M., Anderson, D. M., Cembella, A. D.,Eds., Manual on harmful marine microalgae. IOC Manuals andGuides No. 33, UNESCO 1995k pp. 373–380.

Carpenter, S. R., Kitchell, J. F., Hodgson, J. R. 1985. Cascadingtrophic interactions and lake ecosystem productivity. BioScience35:635–639.

Carpenter, S. R., Kitchell, J. F., Hodgson, J. R., Cochran, P. A., Elser,J. J., Elser, M. M., Lodge, D. M., Kretchmer, D., He, X., vonEnde, C. 1987. Regulation of lake primary productivity by foodweb structure. Ecology 68:1863–1876.

Carper, G. L, Bachmann, R. W. 1984. Wind resuspension of sedi-ments in a prairie lake. Canadian Journal of Fisheries andAquatic Sciences 41:1763–1767.

Cassani, J. R., Caton, W. E. 1983. Feeding behavior of yearling andolder hybrid grass carp. Journal of Fish Biology 22:35–41.

Cedeno-Maldondo, A., Swader, J. A. 1974. Studies on the mecha-nism of copper toxicity in Chlorella. Weed Science 5:433–449.

Clarke, R. A., Stanley, C. D., MacLeod, B. W., McNeal, B. L. 1997.Relationship of seasonal water quality to chlorophyll a concen-tration in Lake Manatee, Florida. Lake and ReservoirManagement 13:253–258.

Colle, D. E., Shireman, J. V. 1980. Coefficients of condition forlargemouth bass, bluegill, and redear sunfish in Hydrilla-infest-ed lakes. Transactions of the American Fisheries Society109:521–531.

Conyers, D. L., Cooke, G. D. 1982. A comparison of chemical andmechanical methods for macrophyte control. Lake Line 2:8.

Cooke, G. D., Kennedy, R. H. 1989. Water quality management forreservoirs and tailwaters. Report 1, in: Reservoir water qualitymanagement techniques. Techical Report E-89-1, U. S. ArmyCorps of Engineers, Washington, DC.

Cooke, G. D., Welch, E. B., Martin, A. B., Fulmer, D. G., Hyde, J. B.,Schrieve, G. D. 1993a. Effectiveness of Al, Ca, and Fe salts forcontrol of internal phophorus loading in shallow and deeplakes. Hydrobiologia 253:323–335.

Cooke, G. D., Welch E. B., Peterson, S. A., Newroth, P. R. 1993b.Management and restoration of lakes and reservoirs. Lewis,Boca Raton, FL.

Corbus, F. G. 1982. Aquatic weed control with endothall in a SaltRiver Project Canal. Journal of Aquatic Plant Management20:1–3.

Crawford, S. A. 1979. Farm pond restoration using Chara vulgarisvegetation. Hydrobiologia 62:17–31.

Crutchfield, J. U., Jr., Schiller, D. H., Herlong, D. D., Mallen, M. A.1992. Establishment and impact of redbelly tilapia in a vegetatedcooling reservoir. Journal of Aquatic Plant Management30:28–35.

Davis, C. C. 1964. Evidence for the eutrophication of Lake Erie fromphytoplankton records. Limnology and Oceanography9:275–283.

Della Greca, M., Monaco, P., Pollio, A., Previtera, L. 1992.Structure–activity relationships of phenylpropanoids as growthinhibitors of the green alga Selenastrum capricornutum.Phytochemistry 31:4119–4123.

Demayo, A., Taylor, M. C., Taylor, K. W. 1982. Effects of copper onhumans, laboratory and farm animals, terrestrial plants, andaquatic life. CRC Critical Reviews in Environmental Scienceand Technology 12:183–255.

DeMelo, R., France, R., McQueen, D. J. 1992. Biomanipulation: Hitor myth? Limnology and Oceanography 37:192–207.

DeMott, W. R., Zhang, Q-X., Carmichael, W. W. 1991. Effects oftoxic cyanobacteria and purified toxins on the survival andfeeding of a copepod and three species of Daphnia. Limnologyand Oceanography 36:1346–1357.

Desjardins, P. R., Olsen, G. B. 1983. Viral control of nuisancecyanobacteria (blue–green algae). California Water ResourcesCenter Contribution No. 185, pp. 1–35.

Dillon, P. J., Nicholls, K. H., Locke, B. A., de Grosbois, E., Yan, N.D. 1988. Phosphorus–phytoplankton relationships in nutrient-poor soft-water lakes in Canada. International Vereinigung fürTheoretische und Angewandte Limnologie Verhandlungen23:258–264.

Dillon, P. J., Rigler, F. H. 1974. The phosphorus–chlorophyll rela-tionship in lakes. Limnology and Oceanography 19:767–773.

Dimitrov, M. 1984. Intensive polyculture of common carp (Cyprinuscarpio) and herbivorous fish (silver carp: Hypophthalmichthysmolitrix and grass carp: Ctenopharyngodon idella). Aqua-culture 38:241–253.

Dionigi, C. P., Champagne, E. T. 1995. Copper-containing aquaticherbicides increase geosmin biosynthesis by Streptomyces ten-dae and Penicillium expansum. Weed Science 43:196–200.

Dodds, W. K., Gudder, D. A. 1992. The ecology of Cladophora.Journal of Phycology 28:415–427.

Doyle, R. D., Smart, R. M. 1998. Competitive reduction of noxiousLyngbya wollei mats by rooted aquatic plants. Aquatic Botany61:17–32.

Edmondson, W. T. 1994. Sixty years of Lake Washington: A curricu-lum vitae. Lake and Reservoir Management 10:75–84.

Edmondson, W. T., Lehman, J. T. 1981. The effect of changes in thenutrient income on the condition of Lake Washington.Limnology and Oceanography 26:1–29.

Effler, S. W., Litten, S., Field, S. D., Tong-Ngork, T., Hale, F., Meyer,M., Quirk, M. 1980. Whole lake response to low level coppersulfate treatment. Water Research 14:1489–1499.

Elakovich, S. D., Wooten, J. W. 1989. Allelopathic potential of six-teen aquatic and wetland plants. Journal of Aquatic PlantManagement 27:78–84.

Elder, J. F., Horne, A. J. 1978. Copper cycles and copper sulphatealgicidal capacity in two California lakes. EnvironmentalManagement 2:17–30.

Elser, J. J., MacKay, N. A. 1989. Experimental evaluations of effectsof zooplankton biomass and size distribution on algal biomassand productivity in 3 nutrient-limited lakes. Archiv furHydrobiologie 114:481–496.

Elser, J. J., Marzolf, E. R., Goldman, C. R. 1990. Phosphorus andnitrogen limitation of phytoplankton growth in freshwaters ofNorth America: A review and critique of experimental enrich-ments. Canadian Journal of Fisheries and Aquatic Sciences47:1468–1477.

EPA. 1990. Lake and reservoir restoration guidance manual, 2nd ed.Office of Water, WH-553, EPA-440/4-90-006, U.S.Environmental Protection Agency, Washington, DC.

EPA. 1998. Clean water action plan: restoring and protectingAmerica’s waters. U.S. Environmental Protection Agency,Washington, DC.

Etnier, C., Guterstam, B., Eds. 1997. Ecological engineering forwastewater treatment. CRC Press, Boca Raton, FL.

Fassett, N. C. 1957. A manual of aquatic plants. University ofWisconsin Press, Madison.

Fitzgerald, G. P. 1959. Bacterial and algicidal properties of some algi-cides for swimming pools. Applied Microbiology 7:205–211.

Fitzgerald, G. P. 1971. Algicides. University of Wisconsin WaterResources Research Center, Eutrophication InformationProgram, Literature No. 2, Madison, WI.

Flores, E., Wolk, C. P. 1986. Production, by filamentous, nitrogen-fixing cyanobacteria, of a bacteriocin and of other antibioticsthat kill related strains. Archives of Microbiology 145:215–219.

Forsberg, C. 1965. Nutritional studies of Chara in axenic cultures.Physiologia Plantarum 18:275–290.

828 Carole A. Lembi

Fowler, M. C., Robson, T. O. 1978. The effects of the food prefer-ences and stocking rates of grass carp (Ctenopharyngodon idellaVal.) on mixed plant communities. Aquatic Botany 5:261–276.

Francke, J. A., Hillebrand, H. 1980. Effects of copper on some fila-mentous Chlorophyta. Aquatic Botany 8:285–289.

Frank, P. A. 1972. Herbicidal residues in aquatic environments, in:Fate of organic pesticides in the aquatic environment. Advancesin Chemistry Series, Oxford University Press, Oxford, UK, Vol.111, pp. 135–148.

Frodge, J. D., Thomas, G. L., Pauley, G. B. 1991. Sediment phospho-rus loading beneath dense canopies of aquatic macrophytes.Lake and Reservoir Management 7:61–71.

Fulton, R. S., III. 1988. Grazing on filamentous algae by herbivorouszooplankton. Freshwater Biology 20:263–271.

Fulton, R. S., Paerl, H. W. 1987. Toxic and inhibitory effects of theblue-green alga Microcystis aeruginosa on herbivorous zoo-plankton. Journal of Plankton Research 9:837–855.

Garrison, P. J., Knauer, D. R. 1984. Long term evaluation of threealum treated lakes, in: Lake and reservoir management. EPA440/5-84-001, pp. 513–517.

Gibbs, G. W. 1973. Cycles of macrophytes and phytoplankton inPukepuke lagoon following a severe drought. Proceedings of theNew Zealand Ecological Society 20:13–20.

Gibson, M. T., Welch, I. M., Barrett, P. R. F., Ridge, I. 1990. Barleystraw as an inhibitor of algal growth II: Laboratory studies.Journal of Applied Phycology 2:241–248.

Gledhill, M., Nimmo, M., Hill, S. J., Brown, M. T. 1997. The toxici-ty of copper(II) species to marine algae with particular referenceto macroalgae. Journal of Phycology 33:2–11.

Gorham, P. R., Carmichael, W. W. 1988. Hazards of freshwaterblue–green algae (cyanobacteria), in: Lembi, C. A., Waaland, J.R., Eds., Algae and human affairs. Cambridge University Press,Cambridge, UK, pp. 403–431.

Gross, E. M., Meyer, H., Schilling, G. 1996. Release and ecologicalimpact of algicidal hydrolysable polyphenols in Myriophyllumspicatum. Phytochemistry 41:133–138.

Gulati, R. D. 1990. Structural and grazing response of zooplanktoncommunity to biomanipulation of some Dutch water bodies.Hydrobiologia 200/201:99–118.

Gunn, G. J., Raferty, A. G., Rafferty, G. C., Cockburn, N., Edwards,C., Beattie, K. A., Codd, G. A. 1992. Fatal canine neurotoxico-sis attributed to blue-green algae (cyanobacteria). VeterinaryRecord 130:301–302.

Halfen, L. N., McCann, M. T. 1975. Behavioral aspects of benthiccommunities of filamentous blue–green algae in lentic habitats.Michigan Botanist 14:49–56.

Hallingse, M. W., Phlips, E. J. 1996. Effects of Cutrine-Plus andCide-Kick II on the growth of algae and cyanobacteria. Journalof Aquatic Plant Management 34:39–40.

Hansen, G. W., Oliver, F. E., Otto, N. E. 1984. Herbicide manual.U.S. Department of the Interior, Bureau of Reclamation,Denver, CO.

Hansen, P. S., Phlips, E. J., Aldridge, F. J. 1997. The effects of sedi-ment resuspension on phosphorus available for algal growth in a shallow subtropical lake, Lake Okeechobee. Lake andReservoir Management 13:154–159.

Hanson, M. J., Stefan, H. G. 1984. Side effects of copper sulfatetreatment of the Fairmont Lakes, Minnesota. Water ResearchBulletin 20:889–900.

Hargeby, A., Andersson, G., Blindow, I., Johansson, S. 1994. Trophicweb structure in a shallow eutrophic lake during a dominanceshift from phytoplankton to submerged macrophytes.Hydrobiologia 280:83–90.

Hauser, W. J., Legner, E. F., Medved, R. A., Platt, S. 1976. Tilapia—A management tool. Fisheries 1:24.

Havens, K. E. 1994. Structural and functional responses of a fresh-water plankton community to acute copper stress.Environmental Pollution 86:259–266.

Havens, K. E., East, T. 1997. In situ responses of Lake Okeechobee(Florida, USA) phytoplankton to nitrogen, phosphorus, andEverglades agricultural area canal water. Lake and ReservoirManagement 13:26–37.

Havens, K. E., James, R. T. 1997. A critical evaluation of phosphorusmanagement goals for Lake Okeechobee, Florida, USA. Lakeand Reservoir Management 13:292–301.

Hawkins, P. R., Griffiths, D. J. 1987. Copper as an algicide in a trop-ical reservoir. Water Research 21:475–480.

Hedtke, S. F. 1984. Structure and function of copper-stressed aquaticmicrocosms. Aquatic Toxicology 5:227–244.

Herbicide Handbook, 7th ed. 1994. Weed Science Society ofAmerica, Lawrence, KS.

Herbst, R. P. 1969. Ecological factors and the distribution ofCladophora glomerata in the Great Lakes. American MidlandNaturalist 82:90–98.

Hinkle, J. 1986. A preliminary literature review on vegetation andfisheries with emphasis on the largemouth bass, bluegill, andhydrilla. Aquatics 8:9–14.

Holm, L. G., Yeo, R. 1981. The biology, control, and utilization ofaquatic weeds. Part III. Weeds Today 12:7–10.

Holz, J. C., Hoagland, K. D. 1999. Effects of phosphorus reductionon water quality: Comparison of alum-treated and untreatedportions of a hypereutrophic lake. Lake and ReservoirManagement 15:70–82.

Horne, A. J., Goldman, C. R. 1974. Suppression of nitrogen fixationby blue-green algae in a eutrophic lake with trace additions ofcopper. Science 183:409–411.

Hoyer, M. V., Canfield, D. E., Jr. 1996a. Lake size, aquatic macro-phytes, and largemouth bass abundance in Florida lakes: Areply. Journal of Aquatic Plant Management 34:48–50.

Hoyer, M. V., Canfield, D. E., Jr. 1996b. Largemouth bass abun-dance and aquatic vegetation in Florida lakes: An empiricalanalysis. Journal of Aquatic Plant Management 34:23–32.

Infante, A., Abella, S. E. B. 1985. Inhibition of Daphnia byOscillatoria in Lake Washington. Limnology and Oceanography30:1046–1052.

Jaag, O. 1972. Oscillatoria rubescens D. C, in: Desikachary, T. V.,Ed., Taxonomy and biology of blue–green algae. University ofMadras, India, pp. 296–299.

Jacoby, J. M., Gibbons, H. L., Stoops, K. B., Bouchard, D. D. 1994.Response of a shallow, polymictic lake to buffered alum treat-ment. Lake and Reservoir Management 10:103–112.

Jochimsen, E. M., Carmichael, W. W., 10 other authors. 1998. Liverfailure and death after exposure to microcystins at a hemodialysiscenter in Brazil. New England Journal of Medicine 338:873–878.

Jones, B. L. 1987. Lake Okeechobee eutrophication research andmanagement. Aquatics 9:21–26.

Jones, J. R., Bachmann, R. W. 1976. Prediction of phosphorus andchlorophyll levels in lakes. Journal of Water Pollution ControlFederation 48:2176–2182.

Jones, R. C., Walti, K., Adams, M. S. 1983. Phytoplankton as a fac-tor in the decline of the submersed macrophyte Myriophyllumspicatum L. in Lake Wingra, Wisconsin. Hydrobiologia107:213–219.

Joyce, J. C. 1993. Practical uses of aquatic weeds, in: Pieterse, A. H.,Murphy, K. J., Eds., Aquatic weeds, the ecology and manage-ment of nuisance aquatic vegetation. Oxford University Press,Oxford, UK, pp. 274–291.

Jupp, B. P., Spence, D. H. N. 1977. Limitations on macrophytes in aeutrophic lake, Lock Leven. I. Effects of phytoplankton. Journalof Ecology 65:175–186.


Jüttner, F., Höfflacher, B., Wurster, K. 1986. Seasonal analysis ofvolatile organic biogenic substances (VOBS) in freshwaterplankton populations dominated by Dinobryon, Microcystis,and Aphanizomenon. Journal of Phycology 22:169–175.

Kallqvist, T., Meadows, B. S. 1978. The toxic effects of copper onalgae and rotifers from a soda lake (Lake Nakaru, East Africa).Water Research 12:771–775.

Kay, S. H. 1997. Barley straw for algae control: A North Carolinaexperience. Aquatics, Spring, pp. 9–10.

Keating, K. I. 1976. Algal metabolite influence on bloom sequence ineutrophied freshwater ponds. U.S. EPA Ecological ResearchService EPA-600/3-76-081.

Keating, K. I. 1977. Allelopathic influence on blue–green bloomsequence in a eutrophic lake. Science 196:885–887.

Keating, K. I. 1978. Blue-green algal inhibition of diatom growth:Transition from mesotrophic to eutrophic community structure.Science 199:971–973.

Khan, M. A. R., Begum, Z. N. T., Rahim, A. T. M. A., Salamatullah,Q. 1996. A comparative study on proximate composition andmineral content of three fresh water green algae. BangladeshJournal of Botany 25:189–196.

Killgore, K. J., Kirk, J. P., Foltz, J. W. 1998. Response of littoral fish-es in upper Lake Marion, South Carolina following hydrillacontrol by triploid grass carp. Journal of Aquatic PlantManagement 36:82–7.

Konopka, A. 1982a. Buoyancy regulation and vertical migration ofOscillatoria rubescens in Crooked Lake, Indiana. BritishPhycological Journal 17:427–442.

Konopka, A. 1982b. Physiological ecology of a metalimneticOscillatoria rubescens population. Limnology andOceanography 27:1154–1161.

Kortmann, R. W., Rich, P. H. 1994. Lake ecosystem energetics: Themissing management link. Lake and Reservoir Management8:77–97.

Kudoh, S., Takahashi, M. 1990. Fungal control of populationchanges of the planktonic diatom Asterionella formosa in ashallow eutrophic lake. Journal of Phycology 26:239–244.

Lackey, J. B. 1968. Ecology of Euglena, in: Buetow, D. E., Ed., Thebiology of Euglena, Vol. I, General biology and ultrastructure.Academic Press, New York, pp. 27–44.

Lage, O. M., Parente, A. M., Vasconcelos, M. T. S. D., Gomes, C. A.R., Salema, R. 1996. Potential tolerance mechanisms ofProrocentrum micans (Dinophyceae) to sublethal levels of cop-per. Journal of Phycology 32:416–423.

Lamon, E. C., III. 1995. A regression model for the prediction ofchlorophyll a in Lake Okeechobee, Florida. Lake and ReservoirManagement 11:283–290.

Lantz, K. E., Davis, J. T., Hughes, J. S., Shafer, H. E. 1964. Waterlevel fluctuation—Its effect on vegetation control and fish popu-lation management. Proceedings of the Southeast AssociatedGame and Fish Commissons 18:483–494.

Lawrence, J. M. 1954. Control of a branched alga, Pithophora, infarm fish ponds. Progress in Fisheries Culture 16:83–86.

Lembi, C. A. 2000. Relative tolerance of mat-forming algae to copper. Journal of Aquatic Plant Management. 38:68–70.

Lembi, C. A., O’Neal, S. W., Spencer, D. F. 1988. Algae as weeds:economic impact, ecology, and management alternatives, in:Lembi, C. A., Waaland, J. R., Eds., Algae and human affairs.Cambridge University Press, Cambridge, UK, pp. 455–481.

Lembi, C. A., Pearlmutter, N. L., Spencer, D. L. 1980. Life cycle,ecology, and management considerations of the green filamen-tous alga, Pithophora. Technical Report No. 130, WaterResources Research Center, Purdue University.

Lembi, C. A., Ritenour, B. G., Iverson, E. M., Forss, E. C. 1978. Theeffect of vegetation removal by grass carp on water chemistry

and phytoplankton in Indiana ponds. Transactions of theAmerican Fisheries Society 107:161–171.

Lembi, C. A., Spencer, D. F., O’Neal, S. W. 1984. Evaluation of non-point pollution control practices and copper treatments on con-trol of filamentous algae. Technical Report No. 170, PurdueUniversity Water Resources Research Center, U. S. GeologicalSurvey, Reston, VA.

Leonardson, L., Ripl, W. 1980. Control of undesirable algae andinduction of algal successions in hypertrophic lake ecostystems,in: Barica, J., Mur, L. R., Eds., Hypertrophic ecosystems. Junk,The Hague, The Netherlands, pp. 57–65.

Letson, M. A., Makarewicz, J. C. 1994. An experimental test of thecrayfish (Orconectes immunis) as a control mechanism for sub-mersed aquatic macrophytes. Lake and Reservoir Management10:127–132.

Lovell, R. T., Lelana, I. Y., Boyd, C. E., Armstrong, M. S. 1986.Geosmin and musty–muddy off-flavors in pond-raised channelcatfish. Transactions of the American Fisheries Society115:485–489.

Maceina, M. J. 1996. Largemouth bass abundance and aquatic vege-tation in Florida lakes: An alternative interpretation. Journal ofAquatic Plant Management 34:43–47.

Maceina, M. J., Cichra, M. F., Betsill, R. K., Bettoli, P. W. 1992.Limnological changes in a large reservoir following vegetationremoval by grass carp. Journal of Freshwater Ecology 7:81–95.

MacKay, N. A., Elser, J. A. 1998. Factors potentially preventingtrophic cascades: Food quality, invertebrate predation, and theirinteraction. Limnology and Oceanography 43:339–347.

Madsen, J. D. 1996. Aquatic plant management evaluation and envi-ronmental assessment, in: Luff, C., Ed., Lake St. Clair,Michigan aquatic plant management investigation. U.S. ArmyCorps of Engineers, Detroit District, April 1996.

Maga, J. A. 1987. Musty/earthy aromas. Food Reviews International3:269–284.

Maloney, T. E., Palmer, C. M. 1956. Toxicity of six chemical com-pounds to thirty cultures of algae. Water and Sewage Works103:509–513.

Marencik, J., Lembi, C. A. 1998. Bioalgicidal potential of reedcanary grass and fresh cut alfalfa. Journal of Phycology34(suppl.):38.

Martin, A. C., Zim, H. S., Nelson, A. L. 1961. American wildlife andplants. A guide to wildlife food habits. Dover, New York.

Martin, E., Benson, R. 1988. Phages of cyanobacteria, in: Calendar,R., Ed., Bacteriophages, Vol. 2. Plenum, New York, pp.607–645.

Martin, E. L., Leach, J. E., Kuo, K. J. 1978. The biological regulationof bloom-causing blue–green algae, in: Louitit, M. W., Miles, J.A. R., Eds., Microbial ecology. Springer-Verlag, Berlin, pp.62–67.

Martin, J. F., Izaguirre, G., Waterstrat, P. 1991. A planktonicOscillatoria species from Mississippi catfish ponds that pro-duces the off-flavor compound 2-methyliso-borneol. WaterResearch 12:1447–1451.

Martin, J. F., McCoy, C. P., Greenleaf, W., Bennett, L. 1988. Analysisof 2- methylisoborneol in water, mud, and channel catfish(Ictalurus punctatus) from commercial culture ponds inMississippi. Canadian Journal of Fisheries and Aquatic Sciences44:909–912.

Mason, C. P., Edwards, K. R., Carlson, R. E., Pignatello, J., Gleason,F. K., Wood, J. M. 1982. Isolation of chlorine-containing antibi-otic from the freshwater cyanobacterium Scytonema hofmanni.Science 215:400–402.

Masuda, K., Boyd, C. E. 1993. Comparative evaluation of the solu-bility and algal toxicity of copper sulfate and chelated copper.Aquaculture 117:287–302.

830 Carole A. Lembi

Mazumder, A., Taylor, W. D., McQueen, D. J., Lean, D. R. S. 1990.Effects of fish and plankton on lake temperature and mixingdepth. Science 247:312–315.

McIntosh, A. W., Kevern, N. R. 1974. Toxicity of copper to zoo-plankton. Journal of Environmental Quality 3:166–171.

McKnight, D. 1981. Chemical and biological processes controllingthe response of a freshwater ecosystem to copper stress: A fieldstudy of the CuSO4 treatment of Mill Pond Reservoir,Burlington, Massachusetts. Limnology and Oceanography26:518–531.

McKnight, D. M., Chisholm, S. W., Harleman, D. R. F. 1983. CuSO4

treatment of nuisance algal blooms in drinking water reservoirs.Environmental Management 7:311–320.

McQueen, D. J. 1990. Manipulating lake community structure:Where do we go from here? Freshwater Biology 23:613–620.

McQueen, D. J., Johannes, M. R. S., Post, J. R., Stewart, T. J., Lean,D. R. S. 1989. Bottom-up and top-down impacts on freshwaterpelagic community structure. Ecological Monographs59:289–309.

Meador, J. P., Taub, F. B., Sibley, T. H. 1993. Copper dynamics andthe mechanism of ecosystem level recovery in a standardizedaquatic microcosm. Ecological Applications 3:139–155.

Melzer, A., Haber, W., Kohler, A. 1977. Foristisch-ökologischeCharakterisierung und Gliederung der Osterseen (Oberbayern)mit Hilfe von submersen Makrophyten. Mitteilungen derFloristisch-Soziologischen Arbeitsgemeinschaft. 19/20:139–151.

Mitchell, D. S., Pieterse, A. H., Murphy, K. J. 1989. Aquatic weedproblems and management in Africa, in: Pieterse, A. H.,Murphy, K. J., Eds., Aquatic weeds. The ecology and manage-ment of nuisance aquatic vegetation. Oxford SciencePublications, pp. 341–354.

Mitsch, W. J., Gosselink, J. G. 1993. Wetlands, 2nd ed. VanNostrand Reinhold, New York.

Monegue, R. L., Phlips, E. J. 1991. The effect of cyanophages on thegrowth and survival of Lyngbya wollei, Anabaena flos-aquae,and Anabaena circinalis. Journal of Aquatic Plant Management29:88–93.

Moore, G. T., Kellerman, K. F. 1905. Copper as an algicide and dis-infectant in water supplies. Bulletin of the Bureau of PlantIndustry USDA. 76:19–55.

Moore, M. V., Winner, R. W. 1989. Relative sensitivies ofCeriodaphnia dubia laboratory tests and pond communities ofzooplankton and benthos to chronic copper stress. AquaticToxicology 15:311–330.

Munawar, M., Munawar, I. F. 1976. A lakewide study of phyto-plankton biomass and its species composition in Lake Erie,April–December, 1970. Journal of the Fisheries Research Boardof Canada 33:581–600.

Murphy, K. J., Barrett, P. R. F. 1993. Chemical control of aquaticweeds, in: Pieterse, A. H., Murphy, K. J., Eds., Aquatic weeds.The ecology and management of nuisance aquatic vegetation.Oxford Science, Oxford, UK, pp. 136–173.

National Academy of Sciences. 1976. Making aquatic weeds useful:Some perspectives for developing countries. Washington, DC.

Neel, J. K., Peterson, S. A., Smith, W. L. 1973. Weed harvest andlake nutrient dynamics. Ecological Research Series, EPA-660/3-73-001, U. S. Environmental Protection Agency, Washington,DC.

Neil, J. H. 1975. Ecology of the Cladophora niche, in: Shear, H.,Konasewich, D. E., Eds., Cladophora in the Great Lakes.International Joint Commission Regional Office, Windsor,Ontario, Canada, pp. 125–127.

Newman, J. R., Barrett, P. R. F. 1993. Control of Microcystis aerugi-nosa by decomposing barley straw. Journal of Aquatic PlantManagement 31:203–206.

Nicholls, K. H. 1996. Barley straw for algae control in ponds. 1996.STB Technical Bulletin No. AqSS-1, Ministry of Environmentand Energy, Ontario, Canada.

Nicholls, K. H., Beaver, J. L., Estabrook, R. H. 1982. Lakewideodours in Ontario and New Hampshire caused by Chrysochro-mulina breviturrita Nich. (Prymnesiophyceae). Hydrobiologia96:91–95.

Nicholls, K. H., Gerrath, J. F. 1985. The taxonomy of Synura(Chrysophyceae) in Ontario with special reference to taste andodour in water supplies. Canadian Journal of Botany63:1482–1493.

Nicholls, K. H., Hopkins, G. J. 1993. Recent changes in Lake Erie(North Shore) phytoplankton: Cumulative impacts of phospho-rus loading reduction and the zebra mussel introduction.Journal of Great Lakes Research 19:637–647.

Nicholls, K. H., Standen, D. W., Hopkins, G. J. 1980. Recentchanges in the near-shore phytoplankton of Lake Erie’s westernbasin at Kingsville, Ontario. Journal of Great Lakes Research6:146–153.

Nicholls, K. H., Standen, D. W., Hopkins, G. J., Carney, E. C. 1977.Declines in the near-shore phytoplankton of Lake Erie’s westernbasin since 1971. Journal of Great Lakes Research 3:72–78.

Nichols, S. A. 1973. The effects of harvesting aquatic macrophyteson algae. Transactions of the Wisconsin Academy of Sciences,Arts and Letters 61:165–172.

Nichols, S. A. 1984. Macrophyte community dynamics in a dredgedWisconsin lake. Water Research Bulletin 20:573–576.

Noonan, T. A. 1998. Como Lake, Minnesota: The long-termresponse of a shallow urban lake to biomanipulation. Lake andReservoir Management 14:92–109.

Ogawa, R. E., Carr, J. F. 1969. The influence of nitrogen on hetero-cyst production in blue-green algae. Limnology andOceanography 14:342–351.

Oliver, R. L. 1994. Floating and sinking in gas-vacuolate cyanobacte-ria. Journal of Phycology 30:161–173.

Olson, R. K., Ed. 1993. Created and natural wetlands for controllingnonpoint source pollution. C. K. Smoley, Boca Raton, FL.

O’Neal, S. W., Lembi, C. A., Spencer, D. L. 1985. Productivity of thefilamentous alga Pithophora oedogonia (Chlorophyta) in SurreyLake, Indiana. Journal of Phycology 21:562–569.

Onodera, H., Satake, M., Oshima, Y., Yasumoto, T., Carmichael, W. W. 1997. New saxitoxin analogues from the freshwater filamentous cyanobacterium Lyngbya wollei. Natural Toxins5:146–151.

Ormerod, G. K. 1970. The relationship between real estate values,algae, and water levels. Report of Lake Erie Task Force,Department of Public Works, Canada.

Ozimek, T., Pieczynska, E., Hankiewicz, A. 1991. Effects of filamen-tous algae on submerged macrophyte growth: A laboratoryexperiment. Aquatic Botany 41:309–315.

Padan, E., Rimon, A., Ginzburg, D., Shilo, M. 1971. A thermosensi-tive cyanophage (LPP-1G) attacking the blue-green algaPlectonema boryanum. Virology 40:773–776.

Paine, R. T. 1980. Food webs: Linkage, interaction strength andcommunity infrastructure. Journal of Animal Ecology49:667–685.

Painter, D. S., Kamaitis, G. 1987. Reduction of Cladophora biomassand tissue phosphorus in Lake Ontario, 1972–1983. CanadianJournal of Fisheries and Aquatic Sciences 44:2212–2215.

Palmer, C. M. 1962. Algae in water supplies. Public Health ServicePublication No. 657. Superintendent of Documents, U.S.Government Printing Office, Washington, DC.

Pearlmutter, N. L., Lembi, C. A. 1986. The effect of copper on thegreen alga Pithophora oedogonia. Weed Science 34:842–849.

Pennak, R. W. 1966. Structure of zooplankton populations in the


littoral macrophyte zone of some Colorado lakes. Transactionsof the American Microscopical Society 85:329–349.

Pennak, R. W. 1973. Some evidence for aquatic macrophytes asrepellents for a limnetic species of Daphnia. International Revueder Gesamten Hydrobiologie 58:569–576.

Perakis, S. S., Welch, E. B., Jacoby, J. M. 1996. Sediment-to-waterblue–green algal recruitment in response to alum and environ-mental factors. Hydrobiologia 318:165–177.

Peterson, H. G., Healey, F. P., Wagemann, R. 1984. Metal toxicity toalgae: A highly pH dependent phenomenon. Canadian Journalof Fisheries and Aquatic Sciences 41:974–979.

Phillips, G. L., Eminson, D., Moss, B. 1978. A mechanism to accountfor macrophyte decline in progressively eutrophicated waters.Aquatic Botany 4:103–126.

Phlips, E. J., Monegue, R. L., Aldridge, F. J. 1990. Cyanophageswhich impact bloom-forming cyanobacteria. Journal of AquaticPlant Management 28:92–97.

Phosphorus Management Strategies Task Force. 1980. Phosphorusmanagement of the Great Lakes. Final Report to the GreatLakes Water Quality Board and the Great Lakes ScienceAdvisory Board. International Joint Commission, Windsor,Ontario, Canada.

Pillinger, J. M., Cooper, J. A., Ridge, I. 1994. Role of phenolic com-pounds in the antialgal activity of barley straw. Journal ofChemical Ecology 20:1557–1569.

Pillinger, J. M., Cooper, J. A., Ridge, I., Barrett, P. R. F. 1992. Barleystraw as an inhibitor of algal growth. III: The role of fungaldecomposition. Journal of Applied Phycology 4:353–355.

Pine, R. T., Anderson, L. W. J. 1991. Plant preferences of triploidgrass carp. Journal of Aquatic Plant Management 29:80–82.

Porter, K. G. 1973. Selective grazing and differential digestion ofalgae by zooplankton. Nature 244:179–180.

Porter, K. G., McDonough, R. 1984. The energetic cost of responseto blue-green algal filaments by cladocerans. Limnology andOceanography 29:365–369.

Power, M. E. 1990. Effects of fish in river food webs. Science250:811–814.

Prepas, E. E., Murphy, R. P., Crosby, J. M., Walty, D. T., Lim, J. T.,Babin, J., Chambers, P. A. 1990. Reduction of phosphorus andchlorophyll a concentrations following CaCO3 and Ca(OH)2

additions to hypereutrophic Figure Eight Lake, Alberta.Environmental Science and Technology 24:1252–1258.

Prepas, E. E., Trew, D. O. 1983. Evaluation of the phosphoruschlorophyll relationship for lakes of the Precambrian Shield inwestern Canada. Canadian Journal of Fisheries and AquaticSciences 40:27–35.

Quade, H. W. 1969. Cladoceran fauna associated with aquaticmacrophytes in some lakes in northwestern Minnesota. Ecology50:170–179.

Quirós, R. 1995. The effects of fish assemblage composition on lakewater quality. Lake and Reservoir Management 11:291–298.

Raman, R. K. 1985. Controlling algae in water supply impound-ments. Journal of the American Water Works Association78:41–43.

Reavie, E. D., Smol, J. P., Carmichael, N. B. 1995. Postsettlementeutrophication histories of six British Columbia (Canada) lakes.Canadian Journal of Fisheries and Aquatic Sciences52:2388–2401.

Redhead, K., Wright, S. J. L. 1978. Isolation and properties of fungithat lyse blue-green algae. Applied and EnvironmentalMicrobiology 35: 962–969.

Reuter, J. E., Rhodes, C. L., Lebo, M. E. , Kotzman, M., Goldman,C. R. 1993. The importance of nitrogen in Pyramid Lake(Nevada, U.S.A.), a saline, desert lake. Hydrobiologia167:179–189.

Rice, E. L. 1984. Allelopathy. Academic Press, Orlando, FL.Riley, E. T., Prepas, E. E. 1984. Role of internal phosphorus loading

in two productive lakes in Alberta, Canada. Canadian Journalof Fisheries and Aquatic Sciences 41:845–855.

Robbins, R. W., Glicker, D. M., Bloem, D. M., Niss, B. M. 1991.Effective watershed management for surface water supplies.American Water Works Association Research Foundation,Denver, CO.

Robinson, P. K., Hawkes, H. A. 1986. Studies on the growth ofCladophora glomerata in laboratory continuous-flow culture.British Phycological Journal 21:437–444.

Rosa, R., Burns, N. M. 1987. Lake Erie central basin oxygen deple-tion changes from 1929–1980. Journal of Great Lakes Research13:684–696.

Rosine, W. N. 1955. The distribution of invertebrates on submergedaquatic plant surfaces in Muskee Lake, Colorado. Ecology36:308–314.

Rothhaupt, K. O. 1991. The influence of toxic and filamentousblue–green algae on feeding and population growth of therotifer Brachionus rubens. International Revue der GesamtenHydrobiologie 76:67–72.

Safferman, R. S., Morris, M. E. 1963. Algal virus: Isolation. Science140:679–680.

Safferman, R. S., Schneider, I. R., Steare, R. L., Morris, M. E.,Diener, T. O. 1969. Phycovirus Sm-1: A virus infecting unicellu-lar blue-green algae. Virology 37:386–395.

Sanchez, I., Lee, G. F. 1978. Environmental chemistry of copper inLake Monona, Wisconsin. Water Research 12:889–903.

Sanders, L., Hoover, J. J., Killgore, K. J. 1991. Triploid grass carp asa biological control of aquatic vegetation. Publication A-91-2,Aquatic Plant Control Research Program, U.S. Army Corps ofEngineers, Waterways Experiment Station, Vicksburg, MS.

Savino, J. E., Stein, R. A. 1982. Predator–prey interactions betweenlargemouth bass and bluegills as influenced by simulated, sub-mersed vegetation. Transactions of the American FisheriesSociety 111:255–266.

Schardt, J. D. 1994. 1994 Florida aquatic plant survey. TechnicalReport No. 972-CGA, Distributed by the Bureau of AquaticPlant Management, Tallahassee, FL.

Schelske, C. L. 1975. Silica and nitrate depletion as related to rate ofeutrophication in Lakes Michigan, Huron, and Superior, in:Hasler, A. D., Ed., Coupling of land and water systems.Springer-Verlag, New York, pp. 277–298.

Schelske, C. L., Stoermer, E. F. 1971. Eutrophication, silica depletion,and predicted changes in algal quality in Lake Michigan.Science 173:423–424.

Schindler, D. W. 1975. Whole lake eutrophication experiments withphosphorus, nitrogen, and carbon. Internationale Vereinigungfür Theoretische und Angewandte Limnologie Verhandlungen19: 3221–3231.

Schindler, D. W. 1977. Evolution of phosphorus limitation in lakes.Science 195:260–262.

Schindler, D. W. 1978. Factors regulating phytoplankton productionand standing crop in the world’s freshwaters. Limnology andOceanography 23:478–486.

Scholze, R. J. 1994. A summary of best management practices fornonpoint source pollution. U.S. Army Corps of Engineers,USACERL Technical Report EP-93/06, National TechnicalInformation Service, Springfield, VA.

Schrader, K. K., Blevins, W. T. 1993. Geosmin-producing species ofStreptomyces and Lyngbya from aquaculture pools. CanadianJournal of Microbiology 39:834–840.

Schrader, K. K., de Regt, M. Q., Tucker, C. S., Duke, S. O. 1997. Arapid bioassay for selective algicides. Weed Technology11:767–74.

832 Carole A. Lembi

Shapiro, J. 1980. The importance of trophic-level interactions to theabundance and species composition of algae in lakes, in: Barica,J., Mur, L. R., Eds., Hypertrophic ecosystems. Junk, TheHague, The Netherlands, pp. 105–116.

Shapiro, J., LaMarra, V., Lynch, M. 1975. Biomanipulation: Anecosystem appropach to lake restoration, in: Brezonik, P. L.,Fox, J. L., Eds., Water quality management through biologicalcontrol. Department of Environmental Engineering Sciences,University of Florida, Gainesville, pp. 85–96.

Shilo, M. 1967. Formation and mode of action of algal toxins.Bacteriological Reviews 31:180–193.

Shireman, J. V., Haller, W. T., Colle, D. E., DuRant, D. F. 1983a.Effects of aquatic macrophytes on native sportfish populationsin Florida, in: Proceedings of the international symposium onaquatic macrophytes, Nijmegen, The Netherlands, pp. 208–214.

Shireman, J. V., Hoyer, M. V., Maceina, M. J., Canfield, D. E. 1985.The water quality and fishing of Lake Baldwin, Florida: 4 yearsafter macrophyte removal by grass carp. Lake and ReservoirManagement 1:201–206.

Shireman, J. V., Rottmann, R. W., Aldridge, F. J. 1983b.Consumption and growth of hybrid grass carp fed four vegeta-tion diets and trout chow in circular tanks. Journal of FishBiology 22:685–693.

Simons, J., Ohm, M., Daalder, R., Boers, P., Rip, W. 1994.Restoration of Botshol (The Netherlands) by reduction of exter-nal nutrients load—Recovery of a characean community, domi-nated by Chara connivens. Hydrobiologia 276:243–253.

Simpson, M. R., MacLeod, B. W. 1991. Comparison of various powdered activated carbons for the removal of geosmin and2–methylisoborneol in selected water conditions, in: 1991Proceedings of the American Water Works Association AnnualConference.

Simpson, P. S., Eaton, J. W. 1986. Comparative studies of the photo-synthesis of the submerged macrophyte Elodea canadensis andthe filamentous algae Cladophora glomerata and Spirogyra sp.Aquatic Botany 24:1–12.

Smeltzer, E. 1990. A successful alum/aluminate treatment of LakeMorey, Vermont. Lake and Reservoir Management 6:9–19.

Smith, C. S., Barko, J. W. 1990. Ecology of Eurasian watermilfoil.Journal of Aquatic Plant Management 28:55–64.

Smith, D. W. 1985. Biological control of excessive phytoplanktongrowth and enhancement of aquacultural production. CanadianJournal of Fisheries and Aquatic Sciences 42:1940–1945.

Smith, V. H. 1983. Low nitrogen to phosphorus ratios favor domi-nance by blue–green algae in lake phytoplankton. Science221:669–671.

Solley, W. B., Merk, C. F., Pierce, R. R. 1988. Estimated use of waterin the U.S. in 1985. U.S. Geological Survey Circular 1004,USGS, Denver, CO.

Sommer, U. 1983. Nutrient competition between phytoplanktonspecies in multispecies chemostat experiments. Archiv fürHydrobiologie 96:399–416.

Sommer, U. 1992. Phosphorus-limited Daphnia: Intraspecific facilita-tion instead of competition. Limnology and Oceanography37:966–973.

Spence, D. H. N. 1976. Light and plant response in freshwater, in:Evans, R. B., Rockham, O., Eds., Light as an ecological factor:II. Blackwell, Oxford, UK, pp. 93–133.

Spencer, C. N., King, D. L. 1984. Role of fish in regulation of plantand animal communities in eutrophic ponds. Canadian Journalof Fisheries and Aquatic Sciences 41:1851–1855.

Spencer, D. L., Lembi, C. A. 1981. Factors regulating the spatial dis-tribution of the filamentous alga Pithophora oedogonia(Chlorophyceae) in an Indiana lake. Journal of Phycology17:168–173.

Speziale, B. J., Dyck, L. A. 1992. Lyngbya infestations: Comparativetaxonomy of Lyngbya wollei comb. nov. (Cyanobacteria).Journal of Phycology 28:693–706.

Stauffer, R. E., Lee, G. F. 1973. The role of themocline migration inregulating algal blooms, in: Middlebrooks, E. J., Falkenbort, D.H., Maloney, T. E., Eds., Modeling the eutrophication process.Utah State University, Water Resources Center, Logan, UT, pp.73–82.

Steinman, A. D. 1996. Effects of grazers on freshwater benthic algae,in: Stevenson, R. J., Bothwell, M. L., Lowe, R. L., Eds., Algalecology. Freshwater benthic systems. Academic Press, SanDiego, pp. 341–373.

Steinman, A. D., Meeker, R. H., Rodusky, A. J., Davis, W. P.,Hwang, S. J. 1997. Ecological properties of charophytes in alarge subtropical lake. Journal of North AmericanBenthological Society 16:781–793.

Steward, K. K. 1993. Aquatic weed problems and management in theeastern United States, in: Pieterse, A. H., Murphy, K. J., Eds.,Aquatic weeds, the ecology and management of nuisance aquaticvegetation. Oxford University Press, Oxford, UK, pp. 391–405.

Stewart, W. D. P., Daft, M. J. 1977. Microbial pathogens ofcyanophycean blooms, in: Droop, M. R., Jannasch, H. W., Eds.,Advances in aquatic microbiology, Vol. 1. Academic Press, NewYork, pp. 177–218.

Stockner, J. G., Shortreed, K. S. 1988. Response of Anabaena andSynechococcus to manipulation of nitrogen: Phosphorus ratiosin a lake fertilization experiment. Limnology and Oceanography33:1348–1361.

Stoermer, E. F. 1988. Algae and the environment: The Great Lakescase, in: Lembi, C. A., Waaland, J. R., Eds., Algae and humanaffairs. Cambridge University Press, Cambridge, UK, pp. 57–83.

Sunderman, F. W. 1978. Carcinogenic effects of metals. FederationProceedings 37:40–46.

Sutton, D. L., Porter, K. M. 1989. Influence of allelochemicals andaqueous plant extracts on growth of duckweed. Journal ofAquatic Plant Management 27:90–95.

Swain, E. B., Monson, B. A., Pillsbury, R. W. 1986. Use of enclosuresto assess the impact of copper sulfate treatments on phyto-plankton. Lake and Reservoir Management 11:303–308.

Taub, F. B., Kindig, A. C., Meador, J. P., Swartzman, G. L. 1989.Effects of seasonal succession and grazing on copper toxicity inaquatic mesocosms. International Vereinigung für Theoretischeund Angewandte Limnologie Verhandlungen 24:2205–2214.

Tilman, D., Kiesling, R. L. 1984. Freshwater algal ecology taxonomictrade-offs in the temperature dependence of nutrient competitiveabilities, in: Klug, M. J., Reddy, C. A., Eds., Current perspec-tives in microbial ecology. American Society for Microbiology,Washington, DC, pp. 314–319.

Tucker, C. S., Boyd, C. E. 1978. Consequences of periodic applica-tions of copper sulfate and simazine for phytoplankton controlin catfish ponds. Transactions of the American Fisheries Society107:316–320.

Tucker, C. S., Busch, R. L., Lloyd, S. W. 1983. Effects of simazinetreatment on channel catfish production and water quality inponds. Journal of Aquatic Plant Management 21:7–11.

Urabe, J. 1993. N and P cycling coupled by grazers’activities: Foodquality and nutrient release by zooplankton. Ecology74:2337–2350.

Urabe, J., Clasen, J., Sterner, R. W. 1997. Phosphorus limitation ofDaphnia growth: Is it real? Limnology and Oceanography42:1436–1443.

van den Berg, M. S., Coops, H., Simons, J., De Keizer, A. 1998a.Competition between Chara aspera and Potamogeton pectina-tus as a function of temperature and light. Aquatic Botany60:241–250.


van den Berg, M. S., Scheffer, M., Coops, H., Simons, J. 1998b. Therole of characean algae in the management of eutrophic shallowlakes. Journal of Phycology 34:750–756.

van der Ploeg, M., Dennis, M. E., de Regt, M. Q. 1995. Biology ofOscillatoria cf. chalybea, a 2-methylisoborneol producing blue-green alga of Mississippi catfish ponds. Water Science andTechnology 31:173–180.

van der Zweerde, W. 1993. Biological control of aquatic weeds bymeans of phytophagous fish, in: Pieterse, A. H., Murphy, K. J.,Eds., Aquatic weeds. The ecology and management of nuisanceaquatic vegetation. Oxford Science, pp. 201–221.

van Donk, E., Grimm, M. P., Gulati, R. D., Kline Breteler, J. P. G.1990. Whole-lake food-web manipulation as a means to studycommunity interactions in a small ecosystem. Hydrobiologia200/201:291–301.

Van Etten, J. L., Lane, L. C., Meints, R. H. 1991. Viruses and virus-like particles of eukaryotic algae. Microbiological Reviews55:586–620.

Vanni, M. J., Findlay, D. L. 1990. Trophic cascades and phytoplank-ton community structure. Ecology 71:921–937.

Vollenweider, R. A. 1969. A manual on methods for measuring pri-mary production in aquatic environments, IBP Handbook 12.Blackwell Science, Oxford, UK.

Wagemann, R., Barica, J. 1979. Speciation and rate of loss of copperfrom lakewater with implications to toxicity. Water Research13:515–523.

Walker, H. L., Higginbotham, L. R. 2000. An aquatic bacterium thatlyses cyanobacteria associated with off-flavor of channel catfish(Ictalurus punctatus). Biological Control 18:71–78.

Walker, W. W., Jr. 1987. Phosphorus removal by urban runoff deten-tion basins. Lake and Reservoir Management 3:314–326.

Walker, W. W., Jr., Westerberg, C. E., Schuler, D. J., Bode, J. A. 1989.Design and evaluation of eutrophication control measures forthe St. Paul water supply. Lake and Reservoir Management5:71–83.

Webster, K. E., Peters, R. H. 1978. Some size-dependent inhibitionsof larger cladoceran filterers in filamentous suspensions.Limnology and Oceanography 23:1238–1244.

Welch, E. B., Cooke, G. D. 1995. Internal phosphorus loading in

shallow lakes: Importance and control. Lake and ReservoirManagement 11:273–281.

Welch, E. B., Cooke, G. D. 1999. Effectiveness and longevity ofphosphorus inactivation with alum. Lake and ReservoirManagement 15:5–27.

Welch, E. B., DeGasperi, C. L., Spyridakis, D. E., Belnick, T. J. 1988.Internal phosphorus loading and alum effectiveness in shallowlakes. Lake and Reservoir Management 4:27–33.

Welch, I. M., Barrett, P. R. F., Gibson, M. T., Ridge, I. 1990. Barleystraw as an inhibitor of algal growth I: Studies in theChesterfield Canal. Journal of Applied Phycology 2:231–239.

Wetzel, R. G. 2001. Limnology, 3rd ed. Academic Press, San Diego.Wetzel, R. G. 1990. Land–water interfaces: Metabolic and limnologi-

cal regulators. International Vereinigung für Theoretische undAngewandte Limnologie Verhandlungen 24:6–24.

Whitaker, J., Barica, J., Kling, H., Buckley, M. 1978. Efficacy of cop-per sulphate in the suppression of Aphanizomenon flos-aquaeblooms in prairie lakes. Environmental Pollution 15:185–194.

Whitton, B. A. 1970. Toxicity of heavy metals to freshwater algae: Areview. Phykos 9:116–125.

Whitton, B. A. 1973. Freshwater plankton, in: Carr, N. G., Whitton,B. A., Eds., The biology of blue–green algae. University ofCalifornia Botanical Monographs, Vol. 1, pp. 353–367.

Wiley, M. J., Gordon, R. W., Waite, S. W., Powless, T. 1984. Therelationship between aquatic macrophytes and sport fish pro-duction in Illinois ponds: A simple model. North AmericanJournal of Fisheries Management 4:111–119

Winner, R. W. 1985. Bioaccumulation and toxicity of copper asaffected by interactions between humic acid and water hard-ness. Water Research 19:449–455.

Winner, R. W., Owen, H. A., Moore, M. V. 1990. Seasonal variabili-ty in the sensitivity of freshwater lentic communities to a chron-ic copper stress. Aquatic Toxicology 17:75–92.

Wium-Andersen, S., Anthoni, U., Christophersen, C., Houen, G.1982. Allelopathic effects on phytoplankton by substances iso-lated from aquatic macrophytes (Charales). Oikos 39:187–190.

Wooten, J. W., Elakovich, S. D. 1991. Comparisons of potentialallelopathy of seven freshwater species of spikerushes(Eleocharis). Journal of Aquatic Plant Management 29:121–125.

834 Carole A. Lembi