global database of methods for cyanobacterial blooms management
global database of methods for cyanobacterial blooms management
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.
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.
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.
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 (
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.
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
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).
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
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).
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.
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.
Aluminium
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
Iron
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 (
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
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
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
The RIPLOX method of sediment oxidation, which focuses
on decreasing phosphorus release from sediments, has been widely used in
Scandinavia and
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
No negative effects to aquatic environment have been
reported in the cases where the Riplox method was used.
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).
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
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.
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.
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
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
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.
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.
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
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.
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).
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
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
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.
Algae
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.
Fungi
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.
Protozoa
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.
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.
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.
Copper
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.
Coagulants
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.
Photosensitizers
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).
|
compound |
Concentration (mg L-1) |
Inhibition (%) |
|
Quinone |
0.5 |
100 |
|
Hydroquinone |
0.5 |
100 |
|
Chindhydron |
0.5 |
100 |
|
Catechol |
5.0 |
100 |
|
3,6-Dichlor-2,5-dimetyloxybenzochinon |
16 |
75 |
|
Tetrachlor hydroquinone |
21 |
100 |
|
1,4-naphthoquinone |
1 |
100 |
|
2-Metylnaphthoquinone |
1 |
100 |
|
2-Dimetylamino-3-chlornaphthilquinone |
1 |
100 |
|
Quinonechlorimide |
0.5 |
100 |
|
2,3-dichloronaphthoquinone |
0.002 |
100 |
|
9,10-Phenantroquinone |
0.08 |
100 |
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 abbreviation |
descriptions
and composition of tested agents |
EC
50 |
|||
|
green algae - laboratory strains |
cyanobacteria - laboratory strains |
||||
|
P. subcapitata |
S. quadricauda |
M. aeruginosa |
S. nidulans |
||
|
Commercially available
biocides |
|||||
|
Copper |
CuSO4.5H2O |
1.27
mg Cu/L |
1.15
mg Cu/L |
1.26
mg Cu/L |
1.26
mg Cu/L |
|
Blue magic |
algicide; acid water solution of Cu ions and
hydrotropic anorganic stabilizing agent |
1.01 mg Cu/L |
0.64
mg Cu/L |
0.48
mg Cu/L |
0.57
mg Cu/L |
|
Algiflash |
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 |
|
Algistop |
algicide; mixture of enzymes, oligobiogenous
componets, surfactants |
2.4
µl/L |
3.26
µl/L |
0.8
µl/L |
0.7
µl/L |
|
Algipropre |
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 |
|
Roundup |
herbicide; N-(phosphonomethyl)glycine (glyphosate) |
33.9
mg glyph./L |
14.0
mg glyph./L |
72.7
mg glyph./L |
17.8
mg glyph./L |
|
Mexel |
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 |
|
Dilurit |
biocide; solution of didecyldimethylamonium chloride |
1.56
µl/L |
1.06
µl/L |
0.52
µl/L |
1.01
µl/L |
|
Azurin |
algicide; active cations of polymeric compounds |
0.98
µl/L |
2.42
µl/L |
0.62
µl/L |
1.87
µl/L |
|
Algil |
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 |
|||||
|
H2O2 |
hydrogen peroxide |
5.7 mg/L |
5.8 mg/L |
0.7 mg/L |
0.7 mg/L |
|
PAA |
peracetic
acid |
14.0 mg/L |
10.6 mg/L |
1.8 mg/L |
3.7 mg/L |
|
MB |
photosensitizer;
methylene blue |
0.26 mg/L |
0.07 mg/L |
0.90 mg/L |
1.50 mg/L |
|
Natural compounds and
materials |
|||||
|
QR12 |
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 |
|
|
Coagulants |
Microcystis sp. – natural cyanobacterial bloom |
||||
|
Al2(SO4)3 |
coagulants |
55
mg/L |
|||
|
Fe2(SO4)3 |
100 mg/L |
||||
|
Ca(OH)2 |
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.
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.
Allelopathy
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
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
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