Cylindrospermopsin

(endorsed 2011)

Guideline

Due to the lack of adequate data, no guideline value is set for concentrations of cylindrospermopsin. However given the known toxicity, the relevant health authority or drinking water regulator should be notified immediately if blooms of Cylindrospermopsis raciborskii or other producers of cylindrospermopsin are detected in sources of drinking water.

General description

Cylindrospermopsin is tricyclic guanidine alkaloid cytotoxin with a molecular weight of 415, produced by the freshwater cyanobacteria Cylindrospermopsis raciborskii, Aphanizomenon ovalisporum, Aphanizomenon flos-aquae, Raphidiopsis curvata, and Umezakia natans. There are two structural variants identified in addition to the most common form. It was first characterised and named from an Australian isolate of C. raciborskii (Ohtani et al. 1992). Subsequently cylindrospermopsin has been detected in two other cyanobacteria: Umezakia natans in Japan (Harada et al. 1994, Terao et al. 1994), and Aphanizomenon ovalisporum in Israel (Banker et al. 1997) and Australia (Shaw et al. 1999). In pure form, cylindrospermopsin is predominantly a hepatotoxin, although extracts of C. raciborskii administered to mice induced pathological symptoms in the kidneys, spleen, thymus, heart and eye. Two other structural variants of cylindrospermopsin have been identified (Banker et al. 2000, Norris et al. 1999).

The production of toxins and therefore the presence of toxicity in individual populations of some cyanobacterial species is known to be variable (Chorus and Bartram 1999 Chapter 3). In the case of C. raciborskii, however, the majority of the strains tested so far in Australia appear to produce cylindrospermopsin. It is therefore likely that most blooms of C. raciborskii will have some degree of toxicity. The natural breakdown of cylindrospermopsin in natural waters is influenced by a number of factors including previous occurrence. Degradation occurs within a few weeks in surface water subject to repeated occurrence, but is far slower in waters with no recorded history of occurrence (Chiswell et al. 1999, Smith et al. 2007).

Australian significance

Cylindrospermopsin is believed to have been the causative agent in the Palm Island “mystery disease” poisoning incident in Queensland in 1979, in which 148 people were hospitalised (Byth 1980). It was subsequently shown that water from Solomon Dam on Palm Island contained blooms of toxic C. raciborskii (Hawkins et al. 1985). C. raciborskii has been found in many water supply reservoirs in northern, central and southern Queensland. Although C. raciborskii and A. ovalisporum are both considered to be predominantly tropical/sub-tropical in terms of habitat, with most Australian blooms occurring in Queensland, C. raciborskii also occurs in the Murray-Darling River system (Baker and Humpage 1994). In recent years there has been increasing evidence of detection in the River Murray and C. raciborskii was detected in the major blooms that affected several hundred kilometres of the River Murray on the border between New South Wales and Victoria in 2009 and 2010 (NSW Office of Water 2009, MDBA 2010). C. raciborskii is not a scum-forming organism, but forms dense bands below the water surface in stratified lakes, while A. ovalisporum may form thick brown surface scums (Shaw et al. 1999). Although no reports of human poisoning attributable to cylindrospermopsin have appeared since the Palm Island incident, recent cattle deaths in Queensland are attributed to this toxin (Saker et al. 1999).

Treatment of drinking water

The first line of defence against cyanobacteria is catchment management to minimise nutrient inputs to source waters. Source water management techniques to control cyanobacterial growth include maintaining flow in regulated rivers; water mixing techniques, both to eliminate stratification and reduce nutrient release from sediments in reservoirs; and the use of algicides in dedicated water supply storages. Destratification has been used to attempt to reduce bloom intensities of C. raciborskii in reservoirs in Queensland, however it has not yet been possible to determine the efficacy of this treatment method. Caution is necessary in using algicides if a bloom has developed because these agents will disrupt cells and liberate intracellular cylindrospermopsin that could otherwise be removed by cell removal, as noted below. Once these intracellular toxins are released they are more difficult to manage. The extracellular release of cylindrospermopsin will increase when a developed bloom declines and algal cells lyse, reinforcing the need to prevent blooms as far as possible. Algicide use should be in accordance with local environment and chemical registration regulations. Where multiple intakes are available, withdrawing water selectively from different depths can minimise the intake of high accumulations of cyanobacterial cells at the surface.

The right combination of water treatment processes can be highly effective in removing both cyanobacterial cells and cylindrospermopsin. In contrast to other cyanotoxins, a high proportion of cylindrospermopsin in actively growing C. raciborskii blooms may be found free in the water, i.e. non cell-bound (Chiswell et al. 1999). Only the proportion of cylindrospermopsin that is cell-bound can be removed by coagulation and filtration in a conventional treatment plant (Chorus and Bartram 1999 Chapter 9). It should be noted that using oxidants such as chlorine or ozone to treat water containing cyanobacterial cells, while killing the cells, will also result in the release of free toxin; therefore pre-chlorination or pre-ozonation are not recommended without a subsequent step to remove dissolved toxins.

Cylindrospermopsin is readily oxidised by a range of oxidants including ozone and chlorine. Adequate contact time and pH control are needed to ensure optimum removal of these compounds, and this will be more difficult to achieve in the presence of whole cells (Chorus and Bartram 1999 Chapter 9). Cylindrospermopsin is also adsorbed from solution by both granular activated carbon and powdered activated carbon. Because powdered activated carbon may be a more practical option for intermittent or emergency use, it is important to seek advice and carefully select the most appropriate type for toxin removal, as carbons vary significantly in performance for different compounds. Boiling is not effective for the destruction of cylindrospermopsin. Based on current knowledge, the recommended best-practice treatment scheme for removal of cylindrospermopsin would include conventional treatment (coagulation/filtration) followed by an adsorption or oxidation step.

Method of identification and detection

Animal bioassays (mouse tests) have been used to determine the toxicity of C. raciborskii (Falconer et al. 1999, Seawright et al. 1999). These tests provide a definitive indication of toxicity, although they cannot be used for precise quantification of compounds in water. Instrumental analytical techniques are available for determining the presence of cylindrospermopsin in water, including high performance liquid chromatography (HPLC) with UV detection (Harada et al. 1994) and HPLC-Mass Spectrometry (Eaglesham et al. 1999).

Cyanobacteria are detected by light microscopy, identified using morphological characteristics, and counted per standard volume of water (Hotzel and Croome 1999). Practical keys for the identification are provided in Baker and Fabbro (2002).

Health considerations

The major pathological effects of cylindrospermopsin are damage to the liver, kidneys, lungs, heart, stomach, adrenal glands, the vascular system, and the lymphatic system (Falconer and Humpage 2006). Liver damage is likely to be severe and dose dependent. Cylindrospermopsin is a slow-acting toxin, commonly requiring between 5 and 7 days to produce maximum toxic effect in experimental animals. It has been shown that the $\text{LD}_{50}$ for cylindrospermopsin decreases greatly between 24 hours and 5 days. The 24-hour $\text{LD}_{50}$ for mice (i.p.) is 2 mg/kg, while the 5-6 day i.p. $\text{LD}_{50}$ is 0.2 mg/kg (Ohtani et al. 1992, Terao 1994). The 5-day $\text{LD}_{50}$ for mice by oral administration is approximately 6 mg/kg (Seawright et al. 1999).

A range of sub-chronic oral toxicity studies have demonstrated that the most sensitive responses in mice are in increased liver, kidney, and testis weights, together with a decrease in urine protein content. These studies can be used to derive the maximum no observed adverse effect level (NOAEL) for oral cell extracts of C. raciborskii or purified cylindrospermopsin.

The most detailed sub-chronic oral dosing study was undertaken by Humpage and Falconer (2003). In two trials, mice were exposed to various doses of cylindrospermopsin for 10-11 weeks. Body weights were significantly increased at low doses (30 and 60 μg kg$^{-1}$ d$^{-1}$) and decreased at high doses (432 and 657 μg kg$^{-1}$ d$^{-1}$). Liver and kidney weights were significantly increased at doses of 240 μg kg$^{-1}$ d$^{-1}$ and 60 μg kg$^{-1}$ d$^{-1}$, respectively. Serum bilirubin levels were significantly increased and bile acids significantly decreased at doses of 216 μg kg$^{-1}$ d$^{-1}$ and greater. Serum cholesterol levels were significantly increased at 30 and 60 μg kg$^{-1}$ d$^{-1}$. Urine total protein was significantly decreased at doses above 60 μg kg$^{-1}$ d$^{-1}$. In contrast to previous findings from studies using higher doses and/or shorter exposure times, the kidney rather than the liver appeared to be the more sensitive organ in this trial, although both were clearly affected.

Shaw et al. (2000) calculated a NOAEL of 50 μg kg$^{-1}$ d$^{-1}$ based on fatty infiltration of the liver. Reisner et al. (2004) reported increased serum cholesterol, changes in red blood cell membrane cholesterol, distortion of cell morphology and increased hematocrits in a 21-day oral exposure trial with male mice drinking water containing 600 μg L⁻¹ cylindrospermopsin (estimated daily cylindrospermopsin intake of 66 μg kg⁻¹). Sukenik et al. (2006) reported impacts from giving mice increasing concentrations of cylindrospermopsin in the drinking water over 42 weeks, ranging from initial doses of approximately 10 μg kg$^{-1}$ d$^{-1}$ to 58 μg kg$^{-1}$ d$^{-1}$ Relative kidney weights were significantly increased at 20 weeks whereas liver weights were significantly increased only at 42 weeks. Effects on cholesterol, red cell morphology and hematocrit were observed.

The variations in experimental design of these studies makes an interpretation of dose-response difficult, but overall these findings are in agreement with the 11 week trial described above, both in terms of adverse effects and dose levels producing them.

A NOAEL based on these studies above is estimated to be around 30 μg/kg body weight per day.

Derivation of health alert

The strength of data is insufficient to establish a guideline value. However, an initial health alert can be estimated using the results described above.

$$\text{0.945 μg/L rounded to 1 μg/L} = \dfrac{\text{30 μg/kg bodyweight per day × 70 kg × 0.9}}{\text{2 L/day × 1000}}$$

where:

• 30 mg/kg body weight per day is the No Observed Adverse Effect Level (NOAEL) from the 10 and 11 week ingestion studies with cylindrospermopsin in mice based on liver histopathology, body organ weight and serum enzyme level changes (Humpage and Falconer 2003);

• 70kg is the average weight of an adult;

• 0.9 is the proportion of total daily intake attributed to the consumption of water;

• 2 L/day is the average amount of water consumed by an adult;

• 1000 is the safety factor derived from extrapolation of an animal study to humans (10 for interspecies variability, 10 for intraspecies variability and 10 for limitations in the database, related particularly to the lack of data on chronic toxicity, genotoxicity and carcinogenicity).

In situations where C. raciborskii occurs in drinking water supplies and toxin monitoring data are unavailable, cell numbers may be used to provide a preliminary orientation to the potential hazard to public health. This type of assessment has been used for Microcystis aeruginosa. However, this is slightly more problematic for C. raciborskii, as, at any time, a significant proportion of cylindrospermopsin toxin may be extracellular and free in solution, and this cannot be accounted for in the assessment of cell counts from the raw water.

Nevertheless, in the case of C. raciborskii, local knowledge and experience can allow the development of local thresholds. For example, in Queensland both water and health authorities have extensive monitoring data and experience for a range of populations of toxic C. raciborskii (G McGregor, personal communication). Data from 23 reservoirs indicated that most of the cylindrospermopsin was found in the cell-bound fraction and that concentrations of approximately 1 mg/L were associated with cell concentrations in the range of 15,000-20,000 cells/mL, which is equivalent to a biovolume of 0.6-0.8 $\text{mm}^3\text{/L}$ (based on a mean cell volume of C. raciborskii 42 $\text{mm}^3$). These numbers are indicative only and for health risk assessment, total toxin determination, including both intracellular and extracellular concentrations, is required.

Notification procedure

It is recommended that a notification procedure be developed by water and health authorities. A tiered framework should be considered. Initial notification to health authorities could be provided when numbers of C. raciborskii reach 30% of the density equivalent to 1 μg/L cylindrospermopsin (4,500 cells/mL; biovolume 0.2 $\text{mm}^3\text{/L}$), while an alert could be provided when cell numbers are equivalent to 1 μg/L cylindrospermopsin (15,000 cells/mL; biovolume 0.6 $\text{mm}^3\text{/L}$). For cylindrospermopsin-producing species other than C. raciborskii, notifications and alerts should be based on biovolumes.

In all cases, cell numbers should only be used as preliminary signals and as triggers for toxin testing to enable assessment of potential health risks.

References

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