sin_lipno global database of methods for cyanobacterial blooms management


Part A

A review of in-lake methods of cyanobacterial blooms control

The correct citation of this material is:
Drabkova M., Marsalek B.: A review of in-lake methods of cyanobacterial blooms control and management.
CyanoData – The Glogal Database of Methods for Cyanobacterial Blooms Management, Centre for Cyanobacteria and their Toxins.
April, 2007


The limiting nutrient for the massive cyanobacterial growth and  development is usually phosphorus (Smith, 1983). Therefore, the first and most important step toward improving lake or reservoir water quality and managing the cyanobacterial blooms is elimination of external nutrient loading from the catchments up stream  and controlling the internal phosphorus turnover (e.g. releasing of P from sediments).  The chance to sufficiently decrease nutrient runoff from watershed is often limited, or this measure may be insufficient due to the internal recycling of nutrients in the lake or reservoir. Many methods are also aimed at further decrease of phosphorus bioavailability in water bodies. The aim of this review is to show the diversity of currently available methods for direct treatment of excessively growing cyanobacteria, The probability of getting water of good quality from hypertrophic lakes is extremely low, therefore, a typical and necessary combination is always to use both, watershed and in-lake methods. The effectiveness of different in-lake methods depends on a number of circumstances and good knowledge of the specific water quality situation. Major differences exist particularly between the possibilities for shallow versus deep lakes.

This review compiles the knowledge of the in-lake methods which may lead to the control of development of harmful cyanobacterial blooms, and provides information about their effectiveness, advantages and limitations.


The main goal of this review is to provide detailed information about in-lake methods and measures. However, very important watershed methods of decreasing nutrient load must not be neglected. Therefore, a brief overview of watershed methods is also added at the beginning of this review.


1.       Remedial measures at the catchments level


In most of the eutrophic lakes affected by cyanobacterial blooms, the primary task should be to decrease nutrient load from watershed. This includes point and non-point nutrient loading.

To non-point sources of phosphorus especially belong agriculture runoff, and erosion from urban and deforested areas. Runoff of nutrients in either dissolved or particulate form is tightly connected to increased runoff of water from landscape. Therefore the preventative measures are in many cases similar to anti-flooding measures. The runoff of water and nutrients can be prevented by revitalisation of regulated and straightened rivers and streams, rehabilitation of riparian zones and especially wetlands restoration and sustainable management. Considerable amounts of nutrients can be also trapped by retention ponds and reservoirs and constructed wetlands, if they are properly managed. In general, the higher is the diversity of the landscape, the higher is its buffering capacity and nutrient fixation. Nutrients runoff from agriculture can be targeted by the best agriculture praxis, especially by a changeover in land-use, methods of fertilization and a proper manure management (Cooke, 2005). 

The most important point sources of phosphorus are municipal wastewaters. Phosphorus load can be therefore substantially decreased by building of new wastewater treatment (WWT) plants, upgrading existing WWT plants by, for example, introducing precipitation and flocculation (tertiary treatment) or adjusting biological treatment to increased phosphorus removal. Very important is also the ban of phosphate detergents. Focusing on problems with cyanobacterial blooms, not only higher concentration of phosphorus in water, but also low N:P ratio supports cyanobacterial growth (Smith, 1983; Stahl-Delbanco et al., 2003). WWT plants without tertiary treatment are usually more efficient in nitrogen removal than in phosphorus, therefore, N:P ratio increases. Moreover nitrates in lakes serve as oxidizing agent and their lack may enhance anaerobic decay of organic sediments, and thus, support phosphorus release from sediments to water (see chapters 2 and 2.6). Therefore, the operation of non-advanced WWT  (e.g. WWT without tertiary treatment)  may support cyanobacterial development as well.

Besides the measures in watershed, there are also few methods which may be employed in the tributary before a particular lake or reservoir.

The high amount of nutrients can be removed in so called pre-reservoirs. These are usually small, shallow reservoirs with short retention time placed closely before the main reservoir in which a high water quality needs to be maintained. Phosphorus in pre-reservoirs is removed by biological production and sedimentation of a particulate matter. Efficiency of pre-reservoirs depends on their proper design and management (especially regular sediment dredging) (Pütz and Benndorf, 1998; Salvia-Castellvi et al., 2001).

Lakes, in which a good water quality is of high importance, can be sometimes protected by the diversion of inflow waters high on nutrients and/or other pollution. These waters are diverted to water body or watershed of lower importance, or to waters with a higher assimilative capacity or a higher volume for dilution. Although there are some positive references (Lake Washington, Edmondson and Lehman, 1981), the possibility of use of this method is limited.

Lakes of high importance can be also protected from excessive input of nutrients by in-stream phosphorus removal. The most effective “phosphorus elimination plant” (PEP) has been proposed for Wahnbach reservoir (Bernhardt, 1980; Clasen and Bernhard, 1987). The method is based on phosphorus precipitation and flocculation by ferric ions and following removal of precipitates by filtration. This method is extremely efficient, capable to decrease P concentration in effluent to 5µg L-1, unfortunately also very expensive.



2.       In-lake sediment treatments and phosphorus     inactivation


Lake bottom sediments accumulate phosphorus over long time periods. Therefore, sediments represent a large internal source of phosphorus, which can be again slowly released into the water. Because of this, a water body may exhibit eutrophic conditions even several years after the external phosphorus load was reduced. Release of phosphorus from sediment to water column can be strongly enhanced when the lake is stratified and the bottom of the lake exhibits anoxic conditions. In anoxic sediments, during degradation of organic matter, sulphate is reduced to hydrogen sulphide.

2CH2O + SO42- + 2 H+       H2S + 2CO2 + 2 H2O

Hydrogen sulphide reacts with hydroxides and phosphates of iron while forming iron sulphide and releasing free phosphates.

2FeO(OH) + 3H2S    2FeS + S + 4H2O

2FePO4 + 3H2S    2FeS + 2PO43- + S + 6H+

Sediments also serve as a long term supply of cyanobacteria, as a place where the cyanobacteria survive adverse conditions and stay alive in huge quantities up to several years.

Also other circumstances, as wind mixing in shallow lakes or turbulence from motor boats and bottom fish, contribute to enhanced release of nutrients from lake bottom sediments to water column, as well as to recruitment of cyanobacteria from sediments. Various procedures are used to decrease the internal phosphorus load – sediment removal, capping and oxidation. Some of these methods also target conditions favourable for cyanobacterial survival in sediments.


2.1.      Sediment removal


Sediment removal might be very effective method for nutrient content decreasing in the lake or reservoir. Removal of upper layers of the lake bottom sediments most rich in phosphorus further reveals the layers with the higher capacity to bind the phosphorus. Together with the removal of upper sediment layer, most of the cyanobacterial inoculum is also removed.

Many case studies of lake restoration by sediment removal has been described, ranged from less than 2 to 1050 ha of lake size and sediment volume from few hundred to over 7 mil m3 (Peterson, 1982; Eiseltová, 1994; Cooke et al., 2005). The cases designed to control internal nutrient cycling show mixed result. The decision, whether the sediments will be removed or better treated in the lake, depends on many circumstances (sediment amount and quality, nutrient content, content of toxic compounds, availability of the disposal area, possibility of their further re-use, costs of the particular techniques, limnological study) which needs to be thoroughly evaluated before final decision about sediment removal is made. Especially the dredged material disposal can be problematic. If the sediment does not contain toxic compounds, it can be used for agricultural purposes as a fertilizer.  In special cases,  the dredged sediment can be applied directly on the fields (Pokorný and Hauser, 2002).

Many different methods of sediment removal is described. The most common method in smaller lakes or ponds is lowering the water level and removal of exposed and dried sediments. However, from various reasons many lakes and reservoirs cannot be emptied or water level lowered mostly because of aquatic life conservation. As the most environmental-friendly technique the suction dredgers are commonly used. This method minimize the undesirable sediment resuspension into the water, however, much bigger volumes need to be transferred (removed sediment containing around 90% of water) and the method is more expensive. Moreover, further problems arise with the following wet sediment transport and disposal. 

Of course the sediment dredging represent big intervention to the lake ecosystem, with possible negative aspects. The most obvious is the destruction of benthic organisms. If the lake basin is dredged completely, 2 to 3 years may be required to re-establish benthic fauna. If portions of bottom are left undredged reestablishment may vary from almost immediate to 1 to 2 years. In any case the effect on benthic fauna is generally acceptable relative to the longer term benefits derived (Cooke et al., 2005). 

Even in the case of high effectiveness and high costs, the sediment removal may not necessarily bring the desired effects, especially if the external nutrient load remains sufficiently high for cyanobacterial mass development. As an example, removal of thick sediment layer from 40 ha Vajgar fishpond (Czech Republic) resulted in negative phosphorus budget (more P was trapped than released by sediments) and absence of cyanobacterial blooms of Microcystis sp. However, due to the unchanged high external nutrient load, this change was temporary, preventing cyanobacterial blooms only for five years after the end of sediment dredging (Pokorný and Hauser, 2002).

To conclude  this frequently discussed topic, the sediment removal is the most expensive  but effective method, which brings a number of further questions (assessing the quality, further use of sediments, removal of benthic fauna, especially  frequently protected molluscofauna etc.). If this method follow the measures  in the catchements  and remove also the cyanobacterial inoculum, can  prevent  cyanobacterial blooms for a number of years. However, the only dredging  of sediments can prevent the cyanobacterial blooms only occasionally.


2.2.      Sediment capping


An alternative and somewhat cheaper technique than the sediment removal is the sediment capping. This technique is used especially for the treatment of sediments polluted by toxic metals or other persistent toxicants, but it can be used for reducing of nutrients or cyanobacterial remobilization to the water column as well. The concept of capping sediments in situ involves the placement of a cover over the sediment to seal it off and minimize the release of contaminants to the water column. The cover material may simply provide a physical barrier over the sediment or may provide an active barrier. 

As the mechanical barrier may serve a „clean“ sediment (without toxic compounds or nutrients), sand or gravel. The layer should be 30 – 40 cm thick to prevent bioturbation of sediments and slightly more coarse than the original sediment to prevent mixing by air and waves (UNEP-IETC, 1999). This method is used only rarely due to difficulties to create uniform and continuous layer under the water.

The active barrier systems are generally pervious geochemicals capable of active demobilizing contaminants/nutrients in the pore water by the adsorption or precipitation processes. Especially various calcite materials have been used to reduce phosphorus release from sediments (Hart et al. 2003). Mixture of aluminium salts and ballast materials is also known to be successfully used for sediment capping (for example in Sweden), aimed to decrease internal phosphorus loading to the lake water. No negative aspects of this method were reported (

Recently a number of active barrier materials have been tested including calcite (CaCO3), zeolites, modified clays and kaolin amorphous derivative, modified humic substances etc. (Jacobs and Forstner, 1999; Hart et al. 2003).


2.3.      Hypolimnetic withdrawal


This method is applicable only to stratified lakes, where the highest phosphorus concentrations are cumulated in the hypolimnion due to the strong release of phosphorus from sediments during anoxic conditions. The method is based on selective discharge of hypolimnetic waters (low in oxygen and rich in phosphate, iron and manganese) from a lake, instead of discharge of low-nutrient upper-layer waters. The decrease of phosphorus and increase of oxygen concentrations might then limit the growth of cyanobacteria particularly in lakes or reservoirs where the internal phosphorus origin dominates. Use of this method is also advisable to accelerate lake restoration after the external phosphorus load has been restricted.

Hypolimnetic waters may be preferentially removed through siphoning, pumping (lakes) or selective discharge (reservoirs). This can be for example achieved by so called Olszewski tube on the principle of siphon, discharging hypolimnetic water to a lower laying place downstream the lake (Olszewski, 1961). During the hypolimnetic withdrawal the destratification should be avoided because it increases transport of hypolimnetic nutrients and anoxic water to the epilimnion. It is applicable only if the discharged amount of water can be replaced by sufficient inflow, to maintain the lake level relatively constant (Cook et al. 2005).

The advantage of this method is relatively low cost. Its use is limited to relatively small and deep lakes and reservoirs. In bigger lakes (>2.5 x 106 m3) the withdrawal might not be sufficient to decrease the anoxy and phosphorus content in the hypolimnion (Nürnberg, 1987). This might be also a case of some reservoirs in the Czech Republic, where this technique is employed inadvertently, when the hypolimnetic water is normally discharged for power generation, however, the decrease of the phosphorus concentration is not significant. Successful cases are described from a range of lakes in the USA, Canada, Finland, Germany and Poland (Nürnberg, 1987; Dunalska et al., 2001; Cooke et al., 2005).

The negative effects might take place downstream due to the discharge of water with a lower temperature, higher nutrients, ammoniac, hydrogen sulphide or other toxic compounds. This may restrict an occurrence of sensitive species and enhance a development of algae and macrophytes. To avoid this negative impact, a mixing with epilimnetic water might be employed (Cook et al. 1993). These adverse effects might be also attenuated by additional chemical phosphorus precipitation in the outlet (Chorus and Mur, 1999).


2.4.      Hypolimnetic aeration and oxygenation


The basic concept of aeration system is to continually maintain oxygen at the bottom of the lake, so that, iron remains in a solid form and phosphorus release from the sediments to water column is reduced. The aeration also supports more rapid degradation of organic sediments by aerobic bacteria.

Most commonly, the aeration is achieved by compressors that introduce air in the bottom of the lake through perforated tubes. The rising bubbles push the anoxic water up to the surface where it is re-aerated with atmospheric oxygen. However, this method can break the stratified conditions in the lake and bring up nutrient-rich water to epilimnion, which may trigger even more intensive algal and cyanobacterial growth. Therefore, specific aeration of hypolimnion is sometimes used. Hypolimnetic aeration is designed to raise an oxygen content in the hypolimnion without destratifying the water column or warming the hypolimnion.

The overview of various designs for hypolimnetic aerators is available in Cooke et al. (2005). The first type is based on a mechanical removal of hypolimnetic water, oxygenation in the air, and return back to the same depth without changing temperature. The disadvantage of this method is poor gas exchange efficiency. Another possibility is an injection of air to the hypolimnion. If the injected air bubbles are small enough (≤ 1 mm radius) and rising plumes weak enough, bubbles should completely dissolve in the hypolimnion (Wüest et al., 1992). Pure O2 is sometimes used instead of air to increase gas transfer efficiency, but this provides less distribution force than air. In some cases small amounts of ozone can be also added to the air to avoid growth of bacteria and fungi inside the aerating tubes. As another type, an injection of air-water mixture or oxygen-rich epilimnetic water to the hypolimnion can be used.

The hypolimnetic aeration may not operate satisfactorily if the water body is too shallow even if stratification exists. The hypolimnetic aeration is not recommended if maximum depth is less than 12 to 15 m (Cooke et al., 2005).  Aerators are usually put on after the spring circulation and run during the whole season until beginning of the autumn circulation. It can be also set off during the winter under the ice cover if necessary. The hypolimnetic aeration should be always designed specifically on conditions in a particular lake. Due to the need of the electric power, the operation costs of this method are relatively high.

Not many adverse effecs are described. The supersaturation of hypolimnetic water with N2 that might lead to a gas bubble disease in fish might be a possible problem in some cases (Kortmann et al., 1994). here may be a possible negative impact of the device transport and installation into the lake. Sometimes oxygenated hypolimnion does not necessarily guarantee that the sediment surface will be oxic enough to sufficiently decrease the P release from the sediments. In some cases the increased diffusion of nutrients to the epilimnion was observed even though the stratification was maintained (Steinberg and Arzet, 1984). The side effects of this method are rather beneficial. The improved oxygen condition improve water quality by decreasing iron, manganese, tastes and odour problems for drinking water supply, decreasing the damage of turbines and other structures by corrosion, and improving downstream water quality (Prepas and Burke, 1997). The aeration also allows zooplankton to swim deeper into the lake where they can hide from predators in the dark bottom during the day (McComas, 2002). Additionally, the expanded aerobic environment should enhance growth and expansion of the coldwater fish and activated mineralization of sediments  can serve the unsatisfactory conditions for the surviving of cyanobacterial  inoculum  in sediments.


2.5.      Phosphorus precipitation and inactivation


This technique focuses on lowering the lake’s P content by removal P from the water column and by retarding release of mobile P from lake sediments. This is achieved by application of so called coagulants. These compounds, when added into the water, precipitate into so called flocks. During the flocks forming, the phosphorus is effectively tied up and converted into a form unavailable for phytoplankton. Some coagulants can also bind small particles including algal and cyanobacterial cells into the flocks. The flocks then settle to the sediments, thus removing phosphate and cyanobacteria from water column. At the bottom of the lake, the coagulum further increases the binding capacity of sediments for phosphorus.

Binding of bioavailable phosphate into the forming flocks is stronger than binding of phosphorus in particulate form (an organic matter, cells etc.). Therefore, it is better to carry out the treatment of lakes with long retention time from late autumn to early spring, when the free phosphate is in maximum, before it is incorporated into intensively growing phytoplankton (Wolter, 1994). The interference with this process occurs in shallow lakes overgrown by macrophytes, and when the external loading exceeds the phosphorus binding capacity of the flocculate (Welch and Cooke, 1999). The effectiveness of this treatment is also low in shallow lakes due to resuspension of phosphorus from sediments by wind/waves. If the main goal of the treatment is the lowering of P concentration, the application might be effective in lakes with long retention time (> one year), in lakes after significant reduction of external P load and/or when the major P input is from sediments (Straškraba and Tundisi, 1999; Cooke et al., 2005). When used in suitable locality, this method may provide long-term effects (Welch and Cooke, 1998).

Besides effects on concentration of P, there is also a possibility to use the coagulants as an alternative to algicidal treatments during the season (see chapter A.5.1).

Coagulants may be applied from the application boat to the whole lake, or the application may be restricted to selected places or particular layers in water column (Sondergaard et al. 2002). Also a continual lakeshore dosing station might be employed in smaller lakes or ponds (McComas, 2002). A mobile plant to withdraw the high nutrient content from the hypolimnion of lakes, called PELICON (Phosphorus ELImination CONtainer), was developed by Keil and Meyer-Jenin (1995). Hypolimnetic water is pumped into the plant, consisting of one or more floating phosphate separators supported by a containerized shore base, supplying coagulant to form sludge that is pumped into a sludge collector (Jørgensen et al., 2005).

There is available a range of compounds that can be used as coagulants - aluminium, iron and calcium salts or their combinations, and some clay materials. Aluminium, iron and calcium have been used for centuries for drinking and wastewater treatments. Lund (1955) appears to be the first who suggested alum addition to streams and lakes to control algal blooms. The above mentioned groups of compounds varies in their effects in aquatic environment, therefore, the aspects of their use are described separately in following sections.



The most commonly used coagulant from aluminium salts is aluminium sulphate (alum, Al2(SO4)3.14H2O). When added in the water, alum quickly forms large, visible, non-toxic precipitates of aluminium hydroxide that grows in size and weight as settles through the water column to the sediments. During the flocks forming and settling, small particles are incorporated. Holz and Hoagland (1998) concluded that alum was extremely effective in controlling sediment phosphorus release rates, improving water clarity, reducing phytoplankton biomass, shifting population species composition from cyanobacteria dominance toward bacillariophytes and chlorophytes, increasing daphnid biomass, and increasing usable fish habitat.

The pH of the solution determines, which aluminium hydrolysis products dominate, and what their solubility will be. At the usual pH 6 to 8, insoluble polymeric Al(OH)3 dominates and P sorption proceeds.

Al3+ + 3H2O        Al(OH)3 + 3H+

The soluble inorganic phosphate binds directly to the aluminium, or adsorbs at the flocks of aluminium hydroxide (Wolter, 1994):

Al(OH)3 + PO43- → AlPO4 + 3 OH-

Al(OH)3 + PO43- → Al(OH)3~PO43-

At pH 4 to 6 various soluble intermediate forms occur (Al(OH)2+, Al(OH)2+...) and at pH less than 4, hydrated and soluble Al3+ dominate. At higher pH levels (> 8.0), as would occur during intense photosynthesis for example, the solubility again increases, the aluminate ion is formed Al(OH)4-, which leads to weaker P sorption (Cooke et al., 2005). Because hydrogen ions are released when an aluminium salt is added to water, in lakes with low or moderate alkalinity (< 30-50 mg CaCO3 L-1) the treatment produces a significant decline in pH, possibly leading to increasing concentration of toxic Al3+. The soluble Al3+ concentration is safe for aquatic organisms up to 50 µg L-1. Also, the pH lower than 6.0 (acidification) itself has adverse effects even without increased Al3+. These pH aspects limits the amount of alum which can be added safely. This problem can be solved by parallel adding a buffer. The sodium hydroxide, calcium hydroxide and sodium carbonate were tested/used for this purpose (Sondergaard et al. 2002). Also sodium aluminate can be used as a buffer with the added benefit of having a high aluminium content (Smeltzer, 1990)

To successfully remove not only dissolved P but also particulate P and to provide sufficient inactivation of sediment P, the goal is to apply as much Al as possible, consistent with environmental safety. Several procedures to estimate proper dose are suggested in Cooke et al. (2005), based on determination of mobile inorganic P in the sediments (Rydin and Welch, 1998, Reitzel et al., 2005), estimated rates of P internal loading from sediments (Kennedy et al., 1987),  or based on lake water alkalinity (Kennedy and Cooke, 1982). The applied concentrations has been reported 5 – 100 g Al m-2  or 5 – 25 g Al m-3 (Welch and Cooke, 1999; Rydin et al. 2000). The inorganic P is removed more effectively than particulate organic P (cells, detritus) (Straškraba and Tundisi, 1999) suggesting that the most effective timing of alum treatment would be in early spring when the content of soluble P is highest. On the other hand, the coagulation is reduced at low temperatures. Therefore the treatment in early summer before cyanobacterial blooms occur is usually considered as the most appropriate. (Cooke et al., 2005). Aluminium is most effectively applied in liquid form. Besides aluminium sulphate or sodium aluminate, also polyaluminium chloride is commonly used for lake treatments. The advantage of using aluminium salts as coagulants for lake treatments is, that low or zero dissolved oxygen concentrations (in lake sediments or commonly present in hypolimnion) do not dissolve the flocks and do not allow P release (Welch et al. 1988).

Aluminium is among the most abundant elements in the Earth’s crust and similarly it is in naturally high concentration in lake sediments. Thus, an alum treatment only slightly increases natural sediment Al content. If applied in reasonable dose not causing persisting lake water acidification, its use is safe. Normally the pH again increases after half to 1 hour after the flocks are formed. Toxicity to aquatic organisms during the flocks formation occurs only in some laboratory experiments or when applied into smaller volumes. Ín lakes or reservoirs, the alum dose is never distributed to the whole lake volume at the same time (treatment of whole lake usually takes several days), therefore, the organisms can escape (Wolter, 1994; Cooke et al., 2005).

Recently, the polyaluminium chloride (trade name PAX18) and aluminium sulphate has been applied in several reservoirs in the Czech Republic to flocculate cyanobacterial cells during the recreational season. In the reservoir Máchovo jezero the PAX18 application (dose 5 mg Al L-1) maintained concentration of cyanobacteria under hygienic limit more than 6 weeks after treatment in June 2005 (unpublished results). Treatments in other lakes in the Czech Republic caused lover or no effects. However, in any cases no adverse effects or damage in aquatic environment has been observed.



Iron is applied usually in the form of FeCl3, but FeCl2 or Fe(SO4)3 may be also used. During application of iron into the water, the flocks of ferric hydroxide are formed, which may transform to mixture of iron oxide and hydroxide:

Fe3+ + 3 H2O     Fe(OH)3  +  3H+

   Fe(OH)3    FeO(OH) + H2O

Similarly as by alum coagulation, the phosphorus can bind directly to the iron or adsorbs at the flocks of ferric hydroxide.

FeO(OH) + H3PO4  → FePO4 + 2 H2O

FeO(OH)  +  PO43-    FeO(OH)~PO43-(aq)

In contrast to alum, the stability of iron flocks is less dependent on pH and iron does not appear in toxic form. Nevertheless, the sorption to Fe(OH)3 is greatest at pH 5 to 7, which is not so common in eutrophic lakes especially if higher phytoplankton densities are present. Phosphorus may be released during periods of high pH (Anderson, 1975). Similarly as during alum treatment, hydrogen ions are released, which may lead to significant decline in pH and toxic effects to fish if pH level decline below 6 (Sondergaard et al. 2002).

Further, the stability of Fe-P compounds is strongly dependent on changes in the redox state.  As the dissolved oxygen in water above sediments drops below 1 mg L-1, iron is used as an alternate electron acceptor. Reduced ferrous ion (Fe2+) is soluble and iron-boud P is released. This change occurs rapidly so that even brief periods of anoxy at the bottom of the lake leads to substantial phosphorus release. To prevent this effect, the aeration usually needs to be employed in parallel to Fe application (Wolter, 1994). Continuous iron application during the summer period has been used with combination with artificial destratification to prevent cyanobaterial blooms in German Bautzen reservoir (Deppe and Benndorf, 2002). Iron application can be also successfully combined with sediment treatment with nitrate as an alternate electron acceptor (see chapter 2.6). 

As another disadvantage, iron binds effectively only inorganic soluble phosphorus. It cannot be usually used to flocculate particulate phosphorus and cells. Therefore, the application during season will not be efficient and lake treatments may possibly take place only in the late autumn or early spring (Sondergaard et al. 2002). Nevertheless, a report about flocculation of cyanobacterial cell is also available (Chow et al. 1998).

No toxic effects of Fe3+ on aquatic organisms are described, at least not in concentrations considered for this treatment. However, in rare occasions the cyanobacterial growth may be supported by iron treatment, when iron is the limiting element in the lake (Chorus and Mur, 1999). In the case of FeCl3 application, concentration of chlorides may reach up to several hundreds mg L-1. However, the concentrations up to 500 mg L-1 should not cause any biological damage. The application may cause temporary brown color of water and the swimming should be restricted during the application (Sondergaard et al., 2002).



Calcium carbonate (calcite, CaCO3) or calcium hydroxide (lime, Ca(OH)2) can be also added to the lake as phosphorus precipitants. Calcite sorbs phosphorus especially when pH exceeds 9.0 and results in significant phosphorus removal from the water column. Phosphate adsorbs at the calcite surface, or binds inside a crystal during the CaCO3 forming when calcium hydroxide is applied (Kleiner, 1988; House, 1990). Various calcite forms have been also reported for their potential use as active barriers for sediment capping to reduce phosphorus release from sediments (Hart et al. 2003).

At high pH, and concentrations of Ca2+ and soluble P, hydroxyapatite is formed.


   10 CaCO3 + 6 HPO42- + 2 H2O -> Ca10(PO4)6(OH)2 + 10 HCO3-


Hydroxyapatite has its lowest solubility at pH >9.5 and bind phosphorus strongly at high pH (Cooke et al, 2005). However, if the pH falls the solubility sharply increases and leads to P release. This occurs especially in zones with intense bacterial respiration near the sediments  (Driscoll et al. 1993).

The described application doses of lime are in a range of 25 – 300 mg Ca L-1 (Sondergaard, 2002). The advantage of lime is its low price and non-toxicity. However adverse effect to aquatic organisms may occur because application of lime increases  the pH of water (Miskimmin et al. 1995; Yee et al. 2000). In soft-water lakes the pH may easily exceeds 11 (Zhang and Prepas, 1996). The lime treatment also temporarily increases turbidity.

As an advantage, lime and calcite may be also used to precipitate cyanobacterial cells from the water column (Zhang and Prepas, 1996).


Clay materials

Also a range of clay materials can be used to bind phosphate from water, as are zeolits, modified clays and kaolins. Specially modified clay PhoslockTM has been reported to succesfully bind phosphorus in the rivers of Canning and Vasse in Australia (Robb et al. 2003). The power of clays to bind phosphorus is also used in constructed wetlands for waste-water treatment (Sakadevan and Bavor, 1998; Drizo et al. 1999).

Clay materials may effectively flocculate and remove cyanobacterial cells. Sengco et al. (2001) in his study demonstrated that the removal efficiency of red- and brown-tide cells were significantly higher in the case of 25 clays than effects of alum or polyaluminium chloride. An application to remove cyanobacterial blooms has been reported from the Swan river in Australia, where a mixture of bentonits and polyaluminium chloride was used (Atkins et al., 2001). However, lake treatments by clay to flocculate cyanobacterial cells in freshwater lakes has not been reported.


2.6.      Chemical methods for sediment oxidation  


The RIPLOX method of sediment oxidation, which focuses on decreasing phosphorus release from sediments, has been widely used in Scandinavia and Germany (Ripl, 1994).  This method  combines surface sediment treatment by calcium nitrate (Ca(NO3)2), ferric chloride (FeCl3) and lime (CaCO3).

Ferric chloride (FeCl3) applied to the sediments low on iron increase the binding capacity for phosphorus. Interstitial phosphorus in sediments is bound by formed Fe(OH)3, or directly to Fe forming FePO4. Ferric sulphate (FeSO4) may be also used instead of FeCl3. A very low pH which is created in the sediment is buffered by added lime to create an optimal pH (7 – 7.5) for denitrification. Subsequently, the calcium nitrate is added to sediments to increase redox potential and stimulate denitrification to deplete organic matter in sediments. In sediments rich in organic matter occurs an intensive bacterial degradation which leads to fast oxygen consumption. As described in the previous section, the lowered redox potential consequently leads to a rapid P release from ferric compounds. During denitrification the organic matter in sediments is degraded into final products – molecular nitrogen, carbon dioxide and water. An advantage of denitrification is, that runs in a range of redox potential, when iron is still in an oxidized ferric state not reduced to soluble ferrous ions. Denitrification normally does not play a big role in the natural lakes, because a sufficient load of nitrogen in oxic state into the lakes is rare (Ripl, 1976; 1994). The increased redox potential also leads to oxidation of iron bound on sulphide and its oxidation to FeOOH, which is also capable to bind phosphorus.

The application is usually performed in late spring. Chemicals may be applied by a direct injection into the upper sediment layer as was used in the treatment of lake Lillesjön (Ripl, 1976). The injection method is most efficient, however, very expensive and applicable only to flat and shallow lake bottoms (Straškraba and Tundisi, 1999). The successful use of Riplox method is also described in restoration project of an former Danube river arm in Vienna, where the application resulted in a significant reduction in nutrient and chlorophyll levels and shift from cyanobacterial dominance to green algae and diatoms. In this case the chemicals were applied into the water column (Donabaum et al. 1999). Application doses of nitrate range from 16 to 140 g N m-2 (Cooke et al., 2005; Donabaum et al. 1999). Ferric chloride and lime addition may be unnecessary in the cases when pH is sufficiently high to promote denitrification and the iron content in sediments is adequate (30-50 mg g-1) for P binding. This brings significant cost savings (Cooke et al., 2005). Another possibility is to apply nitrate to the lake tributaries or to directly bring defosforized wastewaters rich in nitrates (Ripl, 1976, 1994). Successful application of Depox®, a newly developed compound, consisting of Fe(III) and NO3- has been recently reported. This agent has a storage effect for NO3-. NO3- is released slowly, hence the disadvantageous high solubility of NO3- in water can be retarded (Wauer et al., 2005).

No negative effects to aquatic environment have been reported in the cases where the Riplox method was used.


2.7.      Biological treatment for sediment mineralization


Besides the nitrate treatment, the microbial degradation of organic sediments can be also supported by the addition of microorganisms. The mineralization of organic sediments may have two beneficial effects. Lowered content of organic compounds in the sediments decreases the microbial oxygen consumption that occurs during degradation of the organic matter. Therefore, the events of anoxy at the lake bottom followed by high P input into the water column occur less frequently. Moreover, organic sediments are more suitable for overwintering and a long-term reserve of cyanobacterial cells. Thus, the sediment mineralization may also negatively affect cyanobacterial survival in sediments.

Recently lots of biopreparations containing saprophytic microorganisms are commercially available and commonly offered for use in ponds, lakes and reservoirs. These preparations usually consist of selected bacterial strains immobilized on a mineral carrier. Sometimes the preparations are also enriched by bacterial enzymes as biocatalyzators. Some preparations are even enriched by nutrients as growth stimulants at the beginning of growth after addition into the water. However, in most cases, parallel aeration or sediment oxidation by adding another electron acceptor such as nitrate will be necessary to support the growth and activity of added microbes.    

With regard to many bacterial extracelular exudates that have been reported to inhibit growth of cyanobacteria (see chapter A.4.4), also direct effect on cyanobacterial development may be possible. However, no scientific proof of such an effect in the case of application of these commercially available biopreparation is available (Duvall and Anderson, 2001).



3.       Technical and physical in-lake measures



3.1.      Artificial destratification (mixing)


The circulation of water in the lake enhances oxygenation in the entire water column with the main benefits being the reduced P release from oxidized sediments and the enlarging suitable habitat for aerobic animals (see also in chapter 2.4). Besides, the permanent mixing of the water column may have an advantageous direct impact on phytoplankton biomass and composition.

The summer thermal stratification of the lake is favourable for gas-vacuolated cyanobacterial species. In contrast to other phytoplankton species, these are able to regulate their position in the water column (buoyancy) and therefore utilize both, nutrients in nutrient-rich hypolimnion, and sunlight for photosynthesis in epilimnion (Reynolds et al., 1987). The continual mixing of the water column destroys stratified conditions, therefore the advantage of cyanobacterial buoyancy is eliminated on the benefit of faster growing green algae and other non-buoyant algae (Reynolds et al., 1984). The mixing rate velocity should be enough high to exceed flotation velocity of particular buoyant cyanobacterial species present in lake (Visser et al., 1996; Jungo et al., 2001). The cyanobacterial dominance may be shifted to dominance by green algae also in response to decreased pH and associated increase of free CO2, which are conditions more beneficial for green algae (Shapiro,1984, Deppe et al., 1999). The circulation also reduces the whole phytoplankton biomass through light limitation, because the total light received during their brief period in the photic zone is insufficient for net photosynthesis. This would take place most likely in nutrient-rich lakes where the light is more limiting factor than nutrients. This effect of mixing occur mainly in deeper lakes. In some cases it may work also in shallow lakes with higher turbidity and, therefore, higher light extinction. (Cooke et al., 2005).

Continuous mixing is commonly achieved by introducing air bubbles at the bottom of the lake. The rising bubbles push hypolimnetic water up to the surface, which creates a continuous circulation pattern (Verner, 1994). There is also possibility to use jets or pumps. Aerators are usually put on after the spring circulation and run during the whole season until beginning of the autumn circulation. According to Reynolds et al. (1984) sufficient or even better results can be achieved by intermittent mixing (3-weeks periods), rather than continuous. This also lower the operation costs.

There are only few examples reporting negative impacts of artificial destratification. In the case described by Fast and Hulquist (1982), compressed air used to destratify a lake caused supersaturation of the water with dissolved nitrogen and caused downstream fish kills. Mixing can also affect the composition of zooplankton communities (Pastorak et al., 1980). The warming up of the hypolimnion by the mixing with upper and warmer epilimnetic water may have negative impact on some cold water species of fish (McComas, 2002, Cooke et al., 2005). In some cases this method can bring up nutrient rich water to epilimnion, which may trigger even more intensive cyanobacterial growth.

Long-term successful experience with artificial destratification aimed at control of cyanobacterial blooms is known from lake Nieuwe (The Netherands). During the mixing period, the cyanobacterial dominance (80%) was shifted to 5-25% of cyanobacteria, while the proportion of green algae, flagellates and diatoms was elevated. From the cyanobacterial species, especially the representation of Microcystis was reduced most strongly. No water bloom occurred in the years of mixing (Visser et al., 1996).

This method, however, controls the cyanobacterial blooms only to some extent, and might be successful rather in the case of lower phosphorus levels in water and lower cyanobacterial biomass.


3.2       Ultrasound


The most common bloom-forming cyanobacteral species, Microcystis, Anabaena, Planktothrix, Aphanizomenon, and Woronichinia, rank among gas-vacuolated cyanobacteria. The buoyancy allows them to regulate their position in water column, which represent an advantage in competition against other phytoplankton species and therefore supports their excessive growth. Application of the ultrasound (3 seconds) induces the disruption cyanobacterial gas-vesicles and leads to the settling of cyanobacterial cells to the lake bottom. As an advantage, in contrast to algicidal treatment, ultrasonication does not increase release of microcystins from cells (Lee et al., 2001).

However, disrupted gas-vesicles regenerate in a relatively short time after the ultrasonication is interrupted (Ahn et al., 2003a). This accords with results of Walsby (1992) who observed re-synthesis of gas-vesicles in 20h even in the dark. Other aspects related to the effect of the ultrasound have been also studied. After the ultrasonication, impaired photosynthetic activity has been observed (Lee et al., 2001). The ultrasonication also promoted close contact with lysing bacterium Myxobacter, which lead to the destruction of cyanobacterial cells (Lee et al., 2002).

The ultrasound may by applied directly to free water, however, this may have harmful impact on fish in the case of improper parameters setting. Therefore the ultrasound is better to apply in flow-through closed systems. This has been experimentally applied in a Japanese lake, however, with unsatisfactorily results (Nakano et al., 2001). Although the application of the ultrasound still seems to have potential as an effective control method for cyanobacterial blooms, there is a lack of information about the effects of ultrasound in aquatic ecosystems.

Positive effect of ultrasound on the reduction of biomass of  filamentous green algae, or macrophyta should be expected, but  references   are not currently  available.


3.3.      Mechanical removal of cyanobacterial biomass


Under condition of massive cyanobacterial scum, the harvesting  and removal of cyanobacterial biomass  may be used. The units can be dragged by boat and the concentrated scums on surface pumped off. Various types of water filtration can be also employed. Frequently, only small part of cyanobacterial population in the lake can be removed by mechanical removing because of cyanobacterial presence in the entire water column and sediments. This activity must be drived by experts, which can manage the timing of the activity (mostly early morning, when maximum of the scum is located in top layer of water. There is also a consequential problem with a disposal of stinking and often also toxic biomass removed from water. Usually  its combustion or the composting might be taken into account. Flocculation and settling of cyanobacterial biomass to the bottom of the lake  is usually a complementary measures to mechanical removal. A small advantage of the mechanical removal is, that this removes also the nutrients contained in the biomass. Combination of the flocculation and the mechanical removal has been reported from the Swan river in Australia, where a mixture of bentonite clay and polyaluminium chloride was used to flocculate and sink cyanobacterial cells and scums were sucked up using oil spill equipment. Removed cyanobacterial biomass was disposed using sewage treatment facilities (Atkins et al., 2001).


3.4.      Dilution and flushing


On rare occasions the water quality in the lake or reservoir can be improved by the dilution by external water from other sources than from the original reservoir inflow. The concentration of limiting nutrients for cyanobacterial growth is diluted and the water exchange rate is increased, which also leads to a faster loss (flushing) of algae from the lake. The flushing of algae might have an effect even if water high on nutrients is used (Welch, 1981).

The dilution and flushing have worked successfully in several lakes. The significant reduction of cyanobacterial blooms by this method was achieved in Moses Lake in Washington (Welch and Patmond, 1990), after decreasing the water retention time from 10 to 5 days. The cca 50% decrease in nutrients and algae content and increase in water transparency followed after the dilution in the lake Green (Welch et al, 1972). In the lake Veluwe (The Netherlands) the successful rehabilitation was achieved by the flushing with water low on phosphorus but high on nitrate and calcium. The cyanobacterial bloom was replaced by green algae and diatoms (Hosper and Meijer, 1986; Jagtman et al. 1992; Sas, 1989). However, due to high amounts of water needed, this method is rarely applicable.



3.5       Sediment drying


Development of nuisance cyanobacterial blooms is often initiated by recruitment of the overwintering seed population. Poulíčková et al. (1998) reported a reduction in Microcystis blooms for three years after sediment removal as a result of the removal of Microcystis inoculum together with the top layers of sediments. However, dredging projects are generally expensive and require a suitable dumping site for the sediment. Sediment drying may be an easier and a less costly measure than sediment dredging, especially in the case of a shallow reservoir (Tsujimura, 2004). Baker and Bellifemine (2000) demonstrated that desiccation of akinetes of Anabaena circinalis for moderately short periods could significantly impair their capacity to germinate. No reports were found on desiccation of Microcystis sp. However, we can expect lower tolerance to desiccation, because these species do not form akinets, although thick mucilage layer may provide some protection against drying.



4.       Biological control


4.1.      Biomanipulation


Biomanipulation is a term coined by Shapiro et al. (1975) for lake water quality management methods based on food web management and biological interventions. Shapiro included effects of algal biomass from top-down control of zooplanktivores by piscivores and bottom-up effects on algae such as nutrient cycling supported by benthivorous fish. In the contemporary limnology this method is usually referred to as a top-down control. The principle is based on manipulations of a trophic cascade. As a result of the reduction of feeding pressure of fish on zooplankton, large species of zooplankton predominate, which are capable of keeping the phytoplankton levels down. The desired composition of fish populations can be achieved by harvesting non-predatory fish and by introducing predatory fish. The effectiveness of the top-down biomanipulation method is limited. The number of reviews written on this topic shows that if the method is to be successful, specific condition needs to be fulfilled (Gulati et al., 1990; Reynolds, 1994; DeBernardi and Giussani; 1995, Shapiro, 1995; Perrow et al., 1997; Lazzaro, 1997). Biomanipulation is usually not very effective in the case of highly eutrophic lakes and reservoirs where the total phosphorus concentration exceeds 100 µg L-1. Most effective examples of biomanipulation apply to relatively small water bodies because of a great difficulty to continuously manipulate fish populations in large ones. Also, the biomanipulation procedure cannot be considered as a routine method, since it depends on a number of circumstances and can be performed only with the participation of a skilled limnologist. Often, the continuous maintenance is required. (Cooke et al., 2005; Jørgensen et al., 2002).

In the case of Microcystis dominance the zooplankton grazing is limited, mainly because of bigger colonies, which are poorly edible for most of zooplankton species (Boing et al., 1998, Yang et al., 2006). In the case of increased numbers of smaller zooplankton species, the green algae are preferred for grazing, which may even support the cyanobacterial dominance. The toxic effects of cyanobacteria on daphnia species are reported (Thostrup and Christoffersen, 1999, Rohlack et al., 2005), which may also play a role in poor effectiveness if the biomanipulation is focused on cyanobacterial blooms control. However, daphnids in lakes with commonly present cyanobacterial blooms might be already resistant to cyanotoxins (Nandini and Rao, 1998), thus the effective grazing cannot be excluded.

More probably, the biomanipulation may reduce cyanobacterial growth via the bottom-up effect in the case of fish removal. Due to fisheries and economical interests, lakes or ponds are commonly overstocked by benthivorous cyprinids. Benthivorous fishes resuspend large quantities of sediments, which enhances transfer of phosphorus and cyanobacterial cells int the water column. Besides, they uproot aquatic plants and their excretion contributes to phosphorus loads. Therefore the removal of substantial amount of benthivorous fishes is strongly recommended if control of cyanobacterial blooms is the aim of lake-restoration project (Gehrke and Harris, 1994).

More recently, the term biomanipulation has referred to nearly all ecological manipulation to manage algae and aquatic plants. The effects of macrophytes or grazing by herbivorous fish will be mentioned in separate following chapters.


4.2.      Herbivorous fishes


Potential method how to reduce cyanobacterial blooms development is direct grazing by herbivorous fishes. Phytoplankton is the main food especially for the silver carp (Hypophthalmichthys molitrix) and partially also bighead carp (Aristichthys nobilis). However, number of studies reported that metabolic activity of phytoplankton after gut passage remains unaffected or even increases (Miura & Wang 1985, Friedland et al. 2005, Kolmakov and Gladyshev, 2003, Gavel & Marsalek 2004). There are some indications promoted by ecosystem studies describing the effective use of Nile tilapia (Oreochromis niloticus) in cyanobacterial bloom control (Gulati 1990, Figueredo 2005). Nile tilapia is known to have an extremely low pH value in its stomach (Getachew, 1989), therefore enhanced damage of ingested cyanobacteria is probable. More than 95% loss of viability of Microcystis colonies after gut passage has been recently demonstrated (Jančula et al., 2007). Even so, the use of this tropical species to the control growth of cyanobacteria in lakes is limited.

The cyanobacterial growth can be even supported if herbivorous fishes are present, due to increased nutrient release from digested macrophytes, which is the effect called ichtioeutrophication (Opuszynski, 1978). Cyanobacterial dominance may be also supported by preferential grazing of other than cyanobacterial species. Even if substantial grazing of particular cyanobacterial species occurs, it is commonly replaced by blooms of another cyanobacterial species.


4.3       Macrophytes  and periphyton


The presence of macrophytes in lakes and reservoirs is beneficial for many reasons. Macrophyte-dominated lakes are resistant to development of algal or cyanobacterial dominance because rooted plants reduce wind and boat-generated resuspension of sediments, provide a daytime refuge to algae-grazing Daphnia, and shade and therefore cool water in the littoral zones. Macrophytes also remove part of nutrients and act as careers for periphyton, which further removes quantum of dissolved phosphorus (Cooke at al., 2005, McComas, 2002). Some macrophytes also release allelopathic compounds inhibitory to cyanobacteria. Especially, the inhibiting effects of Miriophyllum sp., Chara and Elodea has been reported (Saito et al. 1989; Nakai et al. 1996, Berger and Shagerl, 2003, Erhard and Gross, 2006).

However, there is a nutrient-based stability threshold for a macrophyte dominated state of the lake of about 50-100 µg P L-1. If the nutrient loads remains higher, the phytoplankton or cyanobacterial growth will still overwhelm. Reduction of P loads increases the probability that the cyanobacterial or algal dominated state of the lake can switch to the macrophyte-dominated, clear-water state. However, there is a resistance of the lake to both, the increasing and decreasing nutrient loading. The macrophyte-dominated lake can maintain clear water even in the case of high nutrient load, while water quality in the phytoplankton-dominated lake may not improve even if the nutrients concentrations are substantially reduced (Cooke et al., 2005).

There are several other possible causes that may prevent growth of rooted aquatic plants, which may be partially solved by additional measures - the action of waves (might be solved by temporarily installed wave-brakers), the light limitation (lowering turbidity by precipitation and flocculation), fish uprooting plants (fish stock reduction), ducks, geese, swans or turtles eating the plants (building protective cages), nonactive seeds (can be activated by drought during winter). To accelerate the process of restoring macrophyte dominance, a transplanting plants method can be used, but potential limiting factors must be eliminated before (McComas, 2002). However, the macrophyte-dominated steady state can be achieved only in shallow lakes where the area colonized by macrophyte may potentially close to 100%. Most of deep dimictic lakes have small littoral zones and macrophyte probably have limited effects in maintaining clear water (Cooke et al., 2005).

Macrophytes may act as careers for periphyton, which removes quantum of dissolved phosphorus (McComas, 2002). The removal of phosphorus by periphyton may be also purposely enhanced even in the case of missing macrophytes. Substantial amounts of phosphorus (100 mg of TP per m2) were removed by installing PP-sheets into the lake in the study of Jöbgen et al. (2004).


4.4       Other organisms


Many other aquatic organisms have been considered and studied to limit potentially the cyanobacterial growth. Their use is based on principles of predation, parasitism or release of metabolites suppressing cyanobacterial growth. Studies with viruses, bacteria, algae, fungi, protozoa have been reported in the literature. Especially the parasitism of bacteria and viruses seems to be interesting due to the high specificity only to particular cyanobacterial species and, therefore, no effects to other aquatic organisms. However, the large-scale cultivation of many of these organisms is problematic. It is also common that cyanobacteria develop resistance in the case of the effect based on production of antibiotics (bacteria, fungi). Knowledge in this area is mostly only based on laboratory studies and no successful direct applications in the whole lake scale have been reported. Therefore, this study only brings a brief overview.



Viruses of cyanobacteria (cyanophages) commonly occur in marine and freshwater aquatic environment (Bergh, 1989, Suttle, 1990), where they play an important role in determining the cyanobacteria during the season (Manage et al, 1999, 2001). The first evidence of potential use of viruses to control cyanobacterial blooms were reported by Safferman and Morris (1964). The sudden decline of the cyanobacterial biomass accompanied by the cyanophage occurrence has been reported (Gons et al, 2002). There are few recent studies dealing with the possible control of cyanobacterial development by viruses (Yoshida, 2006, Tucker and Pollard, 2005). However, there are many problems that makes the use of viruses nearly impossible in praxis. The izolation and the cultivation of cyanophage is problematic in the large scale.

Cyanobacteria became resistant to cyanophage so the effect will be temporary (Cannon et al. 1976). Moreover the virus is specific to a particular strain and often does not affect cyanobacteria from another locality even if the species is the same (Waterbury and Valois, 1993). Even after the decline of particular cyanobacteria by viruses, the cyanobacteria can be easily replaced by other cyanobacterial species (van Hannen et al. 1999).   

Viruses and cyanobacteria interaction seems to be a  nice topic for the research, but are problematic for the technological use in the  practical scale of natural ecosystem.



Bacteria can lyse cyanobacteria by the production of extracelular lytic enzymes or by contact lysis. The lytic effects selective to bloom forming cyanobacteria are reported in the case of bacteria Alcaligenes denitrificans (Manage et al. 2000), Bacillus sp. (Reim et al. 1974), Bdellovibrio-like bacteria (Sallal, 1994), Myxococcus sp. (Daft et al. 1985; Daft and Stewart, 1971), Flexibacterium (Gromov et al. 1972), and Pseudomonas sp. or Spingomonas sp. producing lytic agent Argimicin A (Sugiura et al. 1993, Imamura et al. 2000, 2001). Actinomycete Streptomyces exfoliatus caused 50% mortality of Anabaena, Microcystis a Oscillatoria (Sigee et al. 1999). Some cyanobacterial extracellular metabolites also cause inhibition of cyanobacterial growth, photosynthesis or metabolism. Metabolite from Flexibacterium similar to lysozymes inhibited photosynthesis and enzymatic activity of cyanobacterium Oscillatoria williamsi (Sallal, 1994). More recently Ahn et al. (2003b) describes strong selective growth inhibition of cyanobacterium Microcystis aeruginosa and Anabaena affinis  by surfactin produced by Bacillus subtilis C1. Similarly, cyanobacteriolytic substance from Bacillus cereus has been described (Nakamura et al., 2003a, 2003b) and its lytic activity toward Aphanizomenon flos-aquae has been found as well (Shi, 2006). Lysis of Microcystis aeruginosa by  Streptomyces neyagawaensis has been reported (Choi et al., 2005).  

The cultivation and practical application of the certain bacterial strains in the natural ecosystems   has presently a similar limitation like the use of cyanophages, but further research can bring a new advantages.



Besides  filamentous algae, some planktonic algae may also produce allelopathic compounds inhibiting cyanobacterial growth. Wu et al. (1998) demonstrated that extract from dinophyte Peridinium bipes causes changes in membrane permeability in cyanobacterium Microcystis aeruginosa. Data from practical monitoring prove that localities dominated by filamentous algae are usually cyanobacteria free. Further research is needed for better understanding the biotic interactions within algae and cyanobacteria.



Parasitism of a cyanobacterium by the chytridiaceous fungus Rhizophidium planktonicum was shown (Canter and Lund, 1951) however, chytridiaceous fungi were later considered to be of limited use due to difficulties in their large scale culture (Daft et al., 1985). Redhead and Wright (1978) demonstrated specific antagonistic effects of 62 non-chytrid fungi on Anabaena flos-aquae and also other filamentous or unicellular cyanobacteria.



Within aquatic ecosystem, protozoa play an important role in the reduction of phytoplankton populations by grazing and phagocytosis (Canter et al., 1990). The predation on cyanobacteria was described in the case of ciliates Furgasonia (Pajdak-Stos, 2001), Nassula (Brabrand et al. 1983; Canter et al. 1990) and Pseudomicrothorax (Fialkowska and Pajdak-Stos, 2002), amoeba Amoeba (Ho and Alexander, 1974) and flagelate Monas guttula (Sugiura et al. 1990). However, most of bloom-forming cyanobacterial species form colonies, which prevent them from protozoan grazing. Therefore, the use of protozoa as bio-control agents is limited.

4.5       Combined biological treatment

Number of lake restoration projects combines several approaches based on the biological interactions. Bioscaping for example  integrates fish projects  (biomanipultion, zooplankton growth stimulation, removal of roughfish etc.) and  lakescaping (aquatic plants management,  stimulation of fytobentos population growth also called aquascaping etc.).  These approaches require  fine limnological knowledge and can be recommended  only under experts  supervising and only in some localities, which can be selected  by limnological study.

5        Algicides


Algicidal treatment has been one of the most common methods used to control algal and cyanobacterial blooms. In the case of compounds applied to selectively inhibit development of cyanobacteria, the more accurate term cyanocides can be used.  There are many compounds toxic to cyanobacteria, able to kill them (cyanocidal effects), or inhibit their new growth (cyanostatic effects).  Cyanocides or cyanostatics should be considered as a second generation of algicides, which  are characteristic for theit selective toxicity to cyanobacteria (at least one order of magnitude  less toxic to  other non-target aquatic organisms) and are biodegradable in the aquatic environment (e.g. no longer  persistence in aquatic ecosystem than one vegetation season).

 Even in the case when particular compound selectively affects only cyanobacteria, there are still some problematic aspects that limit its use. Killing of already grown cyanobacterial bloom may lead to accidental release of cell contents (including toxins) into water. Besides possible health risks (Lam et al., 1995, Kenefick et al. 1993), this may also lead to accidental anoxia due to bacterial degradation of the huge amount of dying biomass and thus cause fish kills. Therefore, the cyanocidal compounds should be best applied at the beginning of the season, when the cyanobacterial biomass does not reach the state of bloom yet. Moreover, at the beginning of the season, cyanobacteria are more vulnerable due to the higher intake of compounds from surroundings. This, together with a lower density of cyanobacteria in water, also enables to apply cyanocides in lower doses. As another problem, cyanobacteria may also gain a resistance to some algicides and consequently, the increase of application dose is necessary. Moreover, the effects of algicidal treatments are temporary. As soon as the concentration of cyanocide in the lake water decreases (by degradation or dilution), the remaining cyanobacteria will grow again and reach the original density usually in a few weeks or even days.

The broad range of agents with cynocidal or cyanostatic effects encompasses organic or inorganic compounds or materials, chemicals or compounds of natural origin. Their effects, and advantages and limitations of their use are described below subdivided into two chapters – chemicals (A.5.1) and natural compounds (A.5.2).

Despite many disadvantages and risks of algicidal treatments, their use is, unfortunately, in many cases the only one choice to get effects in a  shorter time, or to get any effects at all in the case of limited financial budget, especially in hypereutrophic lakes and reservoirs with a high external P load. Therefore, there is still a need to select or develop new cyanocides with lower negative impact on aquatic environments.


5.1.      Inorganic and organic chemicals



Probably, the most common algicide at all is copper sulphate (CuSO4.5H2O). Toxic effects of copper to algae and cyanobacteria include the inhibitions of photosynthesis, the phosphorus uptake and the nitrogen fixation (Havens, 1994). Although cyanobacteria might be suppressed at concentrations as low as 5-10 µg Cu L-1, in the aquatic environment the effective dose is usually much higher (about 1 mg L-1). In the case of huge cyanobacterial biomass, which is usually accompanied with high pH, even 30-300 mg Cu L-1 may be ineffective (Štěpánek and Červenka, 1974). The effect may be decreased because of the precipitation due to hard water and high pH, adsorption on clay materials, high alkalinity (above 150 mg L-1 of CaCO3), biological uptake (Haughey et al. 2000) and possibly also due to the reduced toxicity of copper by the complexation by cyanobacterial exudates (Štěpánek and Červenka, 1974). On the other hand, the complexation by carrier molecule or compounds such as humic and fulvic acid, or chelating it to non-metal ions keeps copper in solution, which in the contrary increases its effect (Cooke et al., 2005, Tubbing et al., 1994).

The advantages of its use are higher effects on cyanobacteria than on green algae (see table A.2), fast effects and relatively low costs of copper sulphate. However, their use is connected with all negative aspects of algicidal use as described above. It has been proved that cyanobacteria can develop resistance to copper (Shavyrina et al., 2001; Garcia-Villada et al., 2004). Copper is toxic to many other aquatic organisms including fish (Cyrino, 2004). Copper stress also impairs food web functions (Havens, 1994). Copper accumulates in sediments, which may impair benthic invertebrates and cause later problems in sediment removal projects (Hanson and Stefan, 1984, Prepas and Murphy, 1988). Due to the need for higher doses, or more expensive chelated copper preparations, or the need for repeated treatments due to temporary effects the copper treatment are not cost-effective. Although still some successful copper applications are recently reported (Van Hullebusch, 2002) its general use in lakes is not acceptable and should be restricted.


Other inorganic chemicals

There are many other inorganic biocides highly toxic to cyanobacteria as, for example, silver nitrate (Ag NO3, dose 0.04 mg L-1) (Štěpánek and Červenka, 1974), potassium permanganate (KMnO4, dose 1-3 mg L-1) (Lam et al. 1995), and sodium hypochlorite (NaOCl, dose 0.5 – 1.5 mg L-1) (Lam et al. 1995). However, (similarly as in the case of copper sulphate), their application into natural aquatic environment is not conceivable due to their nonselective toxicity to many aquatic organisms.

More recently, few studies exploring effects of other inorganic compounds have been carried out. The present study explores selective effects of hydrogen peroxide H2O2 on cyanobacterial species and photosynthesis, and suggests it as a promising compound for treatment of excessive cyanobacterial growth in lakes and reservoirs. The effective dose of H2O2 vary from 0.3 to 5 mg L-1, depending on particular cyanobacterial species, strains (unicellular x colonial), conditions (laboratory x natural environments) and light intensity. Similarly also hydrogen peroxide in its solid form (sodium carbonate peroxyhydrate, 2Na2CO3.3H2O2) can be used (Schrader et al. 1998b). As an advantage, H2O2 selectively affects cyanobacteria in contrast to fish, aquatic macrophytes or even green algae. Similarly selective effect may be achieved by other types of peroxides, for example organic peroxides, as the result of tests with peracetic acid indicates (table A.2), however, no study comparing various peroxides is available. Its use in lake or reservoir does not lead to the accumulation of toxic residues in the environment and the compound is relatively cheap. There is also a limited chance of gaining a resistance to H2O2. The main limitation of H2O2 application is a short time effect. More details are given in chapter B.1.

The indications about cyanobacterial sensitivity to potassium (K+) have recently been published (Parker et al., 1997, Shukla and Rai, 2006), however, the effective dose for Microcystis species 235 mg L-1 is near the concentration common for sea water, compared to freshwater, where the concentration of potassium is around 10 mg L-1. Therefore, this method can be used only on very special occasions.



In addition to the use of compounds which are in certain way toxic to cyanobacteria, there is also a possibility to use coagulants. These compounds are commonly used in wastewater treatment plants and also directly in lakes to coagulate phosphorus from water (mainly the salts of aluminium and iron). The cells of cyanobacteria are trapped into the flocks, removed from water column and settled at the bottom. As an advantage, in contrast to application of classical algicides, the cyanobacterial cells are not killed/lysed, therefore, toxins are not released into the water (Lam et al., 1995; Chow et al. 1998). Besides the chemical parameters (water pH, hardness, buffering capacity...), the effectiveness of cyanobacterial flocculation depends on density of algae and cyanobacteria in the water, or the size of cyanobacterial colonies. Especially big colonies of Microcystis are poorly precipitated. An effective treatment may provide up to several weeks of clear water until the cyanobacteria will grow up again to original densities.

The flocculation is not selective, also other phytoplankton or bacterioplankton cells or even small zooplankton species are trapped. However, these species are often already suppressed by cyanobacterial biomass and toxins, therefore the removal of dominating cyanobacterial biomass is rather beneficial for them. Moreover, green algae have somewhat faster growing rate.

From the range of coagulants the most capable to flocculate cyanobacterial cells are aluminium salts, calcite and lime, however some studies also describe the use of iron salts and clay materials. More detailed information about possible risks of their use and practical application of these compounds into lakes and reservoirs are described in chapter 2.5.



One of the most recent trends in development of biocides is the use of photocatalysis. The photocatalytic effect is based on application of photosensitizers (photosensitive compounds), which under the irradiation with sunlight produce a range of reactive oxygen species (ROS). Consequently, the ROS have a strong biocidal effect. The most commonly produced ROS are singlet oxygen (1O2), superoxide radical (O2●-), hydrogen peroxide (H2O2), perhydroxy radical (HOO-), and hydroxyl radical (OH). There are many compounds, organic and inorganic, which act as photosenzitizers: free chlorophyll, humic substances, porfyrines, phthalocyanines, titanium dioxide (TiO2) etc. Some of them are used in many applications as are disinfection, water treatment, organic pollutants degradation, self-cleaning surfaces, or the therapy of cancer (Merchat et al., 1996; Linkous et al., 2000; Jori and Brown, 2004; Lonen et al., 2005; , Kühn et al., 2003; Bhatkhande et al., 2001). The employment in the cyanobacterial blooms control has been also explored recently. Except hydrogen peroxide, other ROS are very shortly living, which provides only a very short working distance (few micrometers). Although most of ROS have nonselective biocidal effects, the selectivity to particular species can be, therefore, achieved by a selective binding of photosensitizers to cells.

Titanium dioxide (TiO2) is reported to act mainly by the production of hydroxyl radicals (OH) after UV excitation. Although there are many advantages (a stable, inert and cheap compound, in the dark non-toxic), its use to combat cyanobacterial blooms is limited due to TiO2 insolubility in the water. TiO2 powder, if spread into the water, settles fast, not providing a sufficiently long contact time to cause effects. TiO2 nanoparticles stay in the water column, however, they cause harmful effects in aquatic organisms even without light activation of ROS (Lovern and Klaper, 2006).  If immobilized on some solid surface, the very shortly living hydroxyl radicals provide only a very short working distance (a few micrometers). In spite of that, it was demonstrated that hollow glass beads (2 cm in diameter) coated with TiO2 floating on the water level caused the inhibition of algal growth in enclosed river water. In the laboratory study much lower efficiency was observed for diatom Melosira compared to cyanobacteria Anabaena and Microcystis, which lost their photosynthetic activity in 30-minute reaction. However, the mode of action in the case of such a long working distance remains unexplained (Kim and Lee, 2005). Other ROS are also produced by TiO2 photocatalysis including longer living hydrogen peroxide, which may extend the biocidal effect to a longer distance and also provide a selectivity of its effect to cyanobacteria. However, this mode of action of TiO2 has not been sufficiently explored.

As another example, phthalocyanine derivatives under excitation by light are reported to produce especially singlet oxygen. Their use is known from photodynamic therapy of cancer (Ochsner, 1997) and its effectiveness in killing bacteria, yeast and fungi was also demonstrated (Bertoloni et al., 1992; Minnock et al., 1996, Schaffer and Holtz, 1997). The possible use of phthalocyanines to limit growth of algae and cyanobacteria has been recently explored. Study of a broad range of phthalocyanine derivatives revealed several compounds highly and possibly also selectively toxic to phytoplankton, already in concentrations of 0.06 mg L-1 (original unpublished results).   This research is also a part of the present  projects ongoing in the Centre for Cyanobacteria, thus, can be discussed in more detail with authors of this review.

Among natural compounds as photosensitizers act, for example, humic substances. These compounds normally maintain the low level of reactive oxygen species in freshwaters (Steinberg, 2003). The production of H2O2 by photocatalytic effect of humic substances and other dissolved organic matter has been reported (Cooper and Zika, 1983; Hakkinen et al., 2004). Therefore, a low but continual production of H2O2 in freshwaters with a higher content of humic substances might suppress the development of cyanobacterial species. However, humic substances vary in their properties especially in the proportion of aromatic carbon, which is mostly responsible for photocatalytic effects. The sufficient production of H2O2 to prevent cyanobacterial growth is questionable. The evaluation of various kinds of humic substances  and possible use of  modified humic substances as a photosenzitizer in cyanobacterial blooms control is also a part of the present  projects ongoing in the Centre for Cyanobacteria, thus, can be discussed in more detail with authors of this review.

To develop a photosensitizer effective on cyanobacteria and suitable for use in lakes is a challenge for the future research.


Organic chemicals

The advantage of application of organic compounds can be their biodegradability, although, this by far does not apply to all of them. Usually, the limit of their use is their nonselective toxicity to aquatic organisms and also higher price compared to inorganic agents. Again, usually only a short time effect can be expected, due to continual dilution with inflowing water into the lake, the degradation, and the loss of the compound in aquatic environment. Effects on cyanobacteria have been described for many agents, however, most of them is inapplicable due to the toxicity to other organic organisms. Recently the production and the use of some of them is even prohibited. Still new compounds are studied for their cyanotoxic effects. However, none of organic algicides has been reported to be available and suitable for application in the whole lake scale until now. Therefore, only a brief overview of the compounds follows.

Many pesticides have been found to be highly toxic to cyanobacteria, as are for example Reglone A (diquat,1,1-ethylene-2,2- dipyridium dibromide, dose 2-4 mg L-1) (Lam et al. 1995), simazin (2-chloro-4,6-bis(ethylamino)-s- triazine, dose 0.5 mg L-1) (Lam et al. 1995), Diuron (DCMU; 3,4-Dichlorphenyl 1–1 dimethyl urea; dose 0.1 mg L-1) (Swain et al., 1994), paraquat (N,N'-Dimethyl-4,4'-bipyridinium dichloride, dose 0.026 mg L-1) and others (Schrader et al., 1998b), however, their use is not environmentally and toxicologically safe.

Highly toxic to cyanobacteria and algae are quarternary ammonium compounds. For example, didecyldimethylamonium chloride or n-alkyl-dimethyl-benzyl-ammonium chloride and many other similar compounds are common component in algicides, however, they are not suitable for use in the aquatic ecosystem due to their toxicity to other aquatic organisms. Selectively higher effects to cyanobacteria has not been found in our test (Table A.2).

The active compound of Roundup is glyphosate (citace). Although it is a biocide usually regarded as safe for the environment, it is registered as a herbicide for terrestrial use. The dose effective to algae and cyanobacteria (Table A.2) is too high to consider roundup acceptable for use as a cyanocide.

The use of some antibiotics might provide selective effects on cyanobacteria, but their use at the whole lake scale is not cost-effective (Štěpánek and Červenka, 1974; Robinson et al., 2005, Halling-Sorensen, 2000).

Various quinones have been demonstrated to be highly toxic to cyanobacteria many years ago (Fitzgerald, 1952). A comparison of several quinone compounds is given in Table 1. Similarly, other quinone derivatives have recently been reported to selectively kill cyanobacteria (Schrader, 2003). A commercial biocide SeaClean, a quinone based product (menadione sodium bisulfite), selectively inhibited the growth of cyanobacteria in limnocorral experiments, while green algae and diatoms remained unaffected (Schrader et al., 2004). Derivatives of 9,10 - antraquinone (2- [Methylamino –N- (1′ -methylethyl) ]- 9,10 -anthraquinone) selectively inhibited the growth of cyanobacterium Oscillatoria perornata in laboratory experiments and even in field application in limnocorrals (Schrader et al., 2003). The feasibility of eventual application of some of these compounds and their effects to more target and non-target aquatic species awaits for a future evaluation.

Table 1.  The overview of various quinone compounds and their toxicity to cyanobacterium Microcystis sp. (according to Fitzgerald, 1952).


Concentration (mg L-1)


















Tetrachlor hydroquinone

























Table 2. Toxic effects of various algicidal agents to strains of green algae Pseudokirchneriella subcapitata and Scenedesmus quadricauda, cyanobacteria Microcystis aeruginosa and Synechococcus nidulans in defined growth media, and natural cyanobacterial bloom in lake water. Results are based on 3-day growth inhibition tests in laboratory (authors unpublished results).


Name or



descriptions and composition of tested agents

EC 50

green algae

- laboratory strains


- laboratory strains









Commercially available biocides




mg Cu/L


mg Cu/L


mg Cu/L


mg Cu/L

Blue magic

algicide; acid water solution of Cu ions and hydrotropic anorganic stabilizing agent


 mg Cu/L


mg Cu/L


mg Cu/L


mg Cu/L


algicide; mixture of n-alkyl-dimethyl-benzyl-ammonium chloride, surfactants

0.48 µl/L

0.55  µl/L

2.01  µl/L

2.37  µl/L


algicide; mixture of enzymes, oligobiogenous componets, surfactants

2.4 µl/L

3.26 µl/L

0.8 µl/L

0.7 µl/L


algicide; mixture of dimethylaziridine, aluminium sulphate, natrium chloride, food colourants

73.92  µl/L

81.09  µl/L

80.78  µl/L

118.86  µl/L


herbicide; N-(phosphonomethyl)glycine (glyphosate)


mg glyph./L


mg glyph./L


mg glyph./L


mg glyph./L


antifouling agent; mixture of aliphatic hydrocarbons, with alcohol and amine functionality, in an aqueous emulsion

2.02 µl/L

10.6 µl/L

1.90 µl/L

2.35 µl/L


biocide; solution of didecyldimethylamonium chloride

1.56 µl/L

1.06 µl/L

0.52 µl/L

1.01 µl/L


algicide; active cations of polymeric compounds

0.98 µl/L

2.42 µl/L

0.62 µl/L

1.87 µl/L


algicide; leachate from oak bark, sodium chloride

5.65 ml/L

5.66 ml/L

2.88 ml/L

2.10 ml/L

Oxidative compounds


hydrogen peroxide

5.7 mg/L

5.8 mg/L

0.7 mg/L

0.7 mg/L


peracetic acid

14.0 mg/L

10.6 mg/L

1.8 mg/L

3.7 mg/L


photosensitizer; methylene blue

0.26 mg/L

0.07 mg/L

0.90 mg/L

1.50 mg/L

Natural compounds and materials


food tannins extracted from Quercus robur and Quercus petraea

24.5 mg/L

45.3 mg/L

85.7 mg/L

6.73 mg/L

barley straw

dry natural materials, after 8 weeks of aerobic decay in water





130 mg/L

oak leaves




75 mg/L

oak bark




175 mg/L

pine needles




200 mg/L


Microcystis sp. – natural cyanobacterial bloom



55 mg/L


100 mg/L


72 mg/L

There is an indication about selective effect of L-lysine on cyanobacteria. The cyanobacterium Microcystis was killed by the concentrations of 0.6 - 5 mg L-1 (Hehmann et al., 2002). However, ca. 7 mg L-1 of L-lysine does not prevent growth of natural Microcystis bloom in enclosure experiment, unless applied in combination with 7 mg of malonic acid (Kaya et al., 2005). No report about successful use of lysine in the lake-scale is available.   

Some new algicides that are recently available on the market contain mixture of enzymes and surfactants, but their exact composition of enzymes is not displayed by the distributors. Some of them may exhibit selective effects to cyanobacteria (see Table A.2), but their use in natural aquatic environment is limited due to negative effects on other species and also a higher price.


5.2.      Natural compounds and materials


There is a broad range of natural compounds and materials that exhibit algicidal effects. Some aquatic macrophytes can release allelopathic compounds and thus suppress phytoplankton growth, leaches from various kinds of leaf and plant litter have cyanostatic effects, and many natural compounds that were isolated from various organisms and plant materials show cyanocidal potential. However, most of the natural compounds, has been only studied in laboratory research and never been applied in the lake scale.



Generally, allelopathy refers to the beneficial or harmful effects of compounds released by one species on another. Most often the term allelopathy applies to compounds released by plants to suppress growth of another plants or other photosynthetic organisms (Gross, 2003). This principle can be also employed in the control of cyanobacterial growth. Allelopathic compounds released from aquatic macrophyte Miriophyllum negatively affect growth of cyanobacteria. As the mode of action of Miriophyllum spicatum  was considered a release of hydrolysable polyphenolic compounds (such as ellagic acid, catechin, gallic acid, pyrogallic acid) and tellimagrandin II (syn. eugeniin) (Gross et al., 1996; Nakai et al., 2000). Compounds inhibiting cyanobacterial growth were also extracted from Miriophyllum brasilense. Recently, inhibitory effects of other compounds from Miriophyllum such as fatty acids were also shown to exhibit cyanostatic effects (Nakai et al., 2005). Also many other aquatic macrophytes appear to release inhibitory compounds. The allelopathic potential of exudates from several other aquatic macrophytes has been demonstrated in laboratory studies, e.g. Chara (Wium-Andersen et al., 1982; Mulderij et al., 2003, Berger and Schagerl, 2003), Ceratophyllum (Jasser, 1994; Mjelde and Faafeng, 1997, Körner and Nicklisch, 2002), Stratiotes aloides (Mulderij et al, 2005), Typha lattlifolia (Aliota et al., 1990) and Ellodea canadensis and nutalii (Erhard and Gross, 2006) resulting in changes in phytoplankton biomass, phytoplankton composition or both. Glomski et al. (2002) recently doubted that the exudation of polyphenols by M. spicatum occurs to any significant amount. It is questionable whether the amount of allelopathic compounds released by aquatic macrophytes can be sufficient to reduce cyanobacterial growth in the whole lake. However, such field studies are missed. A further research under in situ conditions is needed to gain more information on the ecological relevance and to assess the effectiveness of the inhibition of cyanobacterial growth.

Some studies report allelopathic interactions between terrestrial and aquatic photoautotrophs. In the last decade, many studies have reported algistatic and cyanostatic effects of “allelopathic” compounds released from barley straw, leaf litter and other plant materials. Especially in small lakes and rivers, leaf litter may strongly influence algal communities (Gross, 2003). However, the use of these materials to prevent cyanobacterial blooms represents an artificial coupling of terrestrial and aquatic organisms and should not be classified as an allelopathic interaction. This use is, therefore, reported in the following sections.


Barley straw

The use of barley straw to inhibit development of cyanobacterial blooms is known especially from Great Britain. The algal-inhibiting properties of decomposing barley straw were first reported by Welch et al. (1990). The barley straw after 4-6 weeks of aerobic decay in water releases compounds with algistatic effects. In the following studies, the effects of the barley straw on cyanobacterial growth have been observed in laboratory tests and also in the reservoir (Newmann and Barrett, 1993; Everall and Lees, 1996). Barley straw was also successfully used in drinking water reservoir with positive results even in the long term study (Barrett et al., 1996, 1999). Inhibitory effect was also demonstrated using barley straw extract. The extract made from decomposed straw containing high lignin content effectively inhibited the growth of Microcystis aeruginosa up to 0.005% dilution (Ball et al. 2001).

In spite of many studies, the mechanism of the barley straw is still not well understood. Several ideas were published. The first idea of the barley straw mode of action was the effect of straw as a result of the antibiotic production of mycoflora or bacteria during the straw decomposition (Pillinger et al., 1992), further, the production of antibiotic agents with bacteria, or the release of phenolic compounds from rotting straw have been discussed. Recently, the release of oxidised polyphenolic and phenolic compounds derived from aerobically decomposed lignin has been proposed as the main algicidal factor of the barley straw (Pillinger, 1994). The production of ROS is thought to be another possible mode of action (Pillinger et al. 1996, Barrett et al. 1996), however, this has not been sufficiently proved. Most probably, the inhibiting effect is not caused by a single compound, but by synergistic effect of different inhibiting components in the system.

The primary requirement for the successful use of the barley straw is to maintain aerobic conditions. Therefore, the straw is usually applied on the water surface filled loosely in the netting with buoys. The inhibition effect begins to act after 4 to 8 weeks of aerobic decomposition in about 10°C. Therefore, the straw should be applied already in early spring to provide effect from the beginning of cyanobacterial development. In higher temperature the decomposition goes faster (1-2 weeks in 20°C). The application dose varies between 6 – 60 g m-3. A higher dose may already cause problems due to rapid loss of oxygen or even the anoxy caused by bacterial degradation. After 4 (up to 6) months the rotted straw needs to be removed andreplaced by new straw, otherwise, the further decomposition would undesirably release nutrients.

According to our experience, the use in the tributary can also be effective. The cages filled with barley straw placed in the tributary of a small shallow recreational lake (1.5 ha, 2 m deep) were re-filled with new straw in regular intervals for 2 vegetation seasons. During 2 years of application the cyanobacterial growth was suppressed sufficiently to achieve hygienic limits, even in the case of hypereutrophic conditions (Maršálek, unpublished results).

The effect of barley straw is rather algistatic than algicidal. It can prevent the new growth, but cannot kill cyanobacteria that are already present in water. The straw does not bring immediate effects, however, it may provide longer-term impact. As an advantage, if properly applied, the barley straw has no negative impact on the aquatic ecosystem. No adverse effects on aquatic macrophytes have been reported, conversely, in shallow eutrophic reservoirs the barley straw application may restore macrophyte dominance (Ridge et al., 1999). The barley straw inhibits growth of some aquatic saprolegniaceous fungi, which is advisable especially in the case of fungal diseases in fishponds (Cooper et al., 1997). There was no negative influence of the barley straw application on 18 invertebrate species and on trout breeding. Barley straw has no effect on Triturus vulgaris or Rana temporaria (Ridge et al., 1999).

Barley straw treatments were in some cases successful and cheap solution mostly in small reservoirs. The considerable disadvantage is the way of application. Especially in larger reservoirs and lakes, which are used for recreational purposes, the manipulating and permanent location of netting with straw is inconvenient. Unfortunately, barley straw is not always effective. According to some reports, the straw application did not bring any significant effects (Kelly and Smith, 1996). With regard to the unsuccessful cases of straw application and the not fully understood mode of action, the effectiveness of barley straw use cannot be guaranteed.

Other plant materials and leaf litter

To explore possible effects similar to barley straw, other plant materials were also studied. The wheat straw did not surprisingly show any significant inhibitory effects (Ball et al. 2001). However, there are several studies reporting a significant reduction of algal or cyanobacterial growth by other plant materials. Although, no reports about their practical application in lakes is available.

Several kinds of leaf litter (13 species) were studied in the laboratory study. The highest effects exhibited extracts from freshly fallen Aescullus hippocastanum, Acer campestre and Quercus robur showing over 85% inhibition of algal growth. The effect of Quercus robur in natural conditions can exhibit inhibitory effects on phytoplankton even without aerobic decomposition due to content of tannins (Ridge et al., 1999). In another study (Park et al., 2006a), 17 various extracts from stem or leaves of nine oak species has been examined. Five extracts exhibited 50% inhibition of Microcystis growth at the concentration of 20 mg L-1 and more than 90% inhibition at 50 mg L-1. The concentration lower than 17 mg L-1 of pure tannic acid did not inhibit cyanobacterial growth. This accords to our results (see Table A.2.).

Coniferous litter also shows potent algicidal effect. Its effect is accompanied with acidification, however, the algicidal effect remains even in buffered solutions  (Ridge et al., 1999).

The study testing brown rotted wood, which also suppressed cyanobacterial growth, revealed the importance of degradation products of lignin (Pillinger et al., 1995).

Rotting barley straw, rice straw, mugwort and chrysanthemum exhibited inhibitory effects in the study of Choe and Jung (2002), where the released phenolic and ester compounds were proposed to be the main inhibitory agents. The effect was stronger in all cases on cyanobacterium Microcystis sp. than on green algae Scenedesmus sp.

The growth of M. aeruginosa was strongly inhibited by rice straw extract concentrations ranging from 0.01 to 10 mg L-1. This activity was proposed to be the synergistic effects of various phenolic compounds in the rice straw (Park et al., 2006b).

Extract from fresh mandarine skin significantly inhibited the growth of Microcystis aeruginosa in a laboratory study (Chen et al., 2004).


Isolated compounds

Hundreds of new compounds isolated from higher plants, algae, cyanobacteria or bacteria have been tested for their potential selective cyanocidal and cyanostatic effects. The results of extensive screening of compounds or extracts isolated from the broad range of organisms are presented in the study of Nagle et al. (2003). Dozens of other natural compounds were tested in several other reports (Schrader et al., 1998a; Schrader, 2003; Schrader and Harries 2001; Schrader et al., 2002). These compounds often belong to groups of alkaloids, phenols and polyphenols, quinones, terpens, organic acids and other groups. Most promising seem to be some quinone derivatives, which are already mentioned in the previous chapter (A.5.1). Compounds for above mentioned studies tests were often chosen according to previous reports of compounds causing algistatic effects of barley straw and other plant litters, or according to the previously described allelopathic effects.

Several other studies describe cyanocidal effects of various metabolites isolated from cyanobacterial species as for example fischerellin, hapalindol A, cyanobacterin, nostocyclamid and others (Smith and Doan, 1999; Gleason and Baxa, 1986; Bagchi et al. 1993). Most recently Blom et al. (2006) describe selective cyanocidal effects of nostocarboline, which is a heterocyclic alkaloid isolated from cyanobacterium Nostoc.

            Other compounds are mentioned above, such as bacterial exudates (chapter A.4.4), organic chemicals (chapter A.5.1), or allelopathic compounds. The advantage of these compounds is their natural origin and degradability. Some natural extracts or compounds showed toxicity to cyanobacteria in very low concentrations and also showed lower toxicity to other aquatic species. Therefore, they have been considered as promising for potential cyanocidal use. However, they have been mostly studied only in laboratory, and none of these compounds has been used for the treatment of cyanobacterial blooms in praxis. Much more studies need to be carried out to evaluate their effectiveness in environmental conditions and their safety for aquatic ecosystem, before some of them could be possibly recommended for the practical application in lakes





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