SEPA
United States Office of Water EPA-820R15103
Environmental Mail Code 4304T June 2015
Protection Agency
Health Effects Support Document
for the Cyanobacterial Toxin
Cylindrospermopsin
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Health Effects Support Document
for the Cyanobacterial Toxin
Cylindrospermopsin
U.S. Environmental Protection Agency
Office of Water (4304T)
Health and Ecological Criteria Division
Washington, DC 20460
EPA Document Number: 820R15103
Date: June 15,2015
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FOREWORD
The Safe Drinking Water Act (SDWA), as amended in 1996, requires the Administrator of the U.S.
Environmental Protection Agency (EPA) to establish a list of unregulated microbiological and chemical
contaminants that are known or anticipated to occur in public water systems and that may need to be
controlled with a national primary drinking water regulation. The SDWA also requires that the Agency
make regulatory determinations on at least five contaminants on the list every five years. For each
contaminant on the Contaminant Candidate List (CCL), the Agency will need to obtain sufficient data to
conduct analyses on the extent of occurrence and the risk posed to populations via drinking water.
Ultimately, this information will assist the Agency in determining the appropriate course of action (e.g.,
develop a regulation, develop guidance or make a decision not to regulate the contaminant in drinking
water).
This document presents information, including occurrence, toxicology and epidemiology data, for the
cyanobacterial toxin cylindrospermopsin to be considered in the development of a Drinking Water Health
Advisory (DWHA). DWHAs serve as the informal technical guidance for unregulated drinking water
contaminants to assist federal, state and local officials, and managers of public or community water
systems in protecting public health as needed. They are not to be construed as legally enforceable federal
standards.
To develop the Health Effects Support Document (HESD) for cylindrospermopsin, a comprehensive
literature search was conducted from January 2013 to May 2014 using Toxicology Literature Online
(TOXLINE), PubMed component and Google Scholar to ensure the most recent published information on
cylindrospermopsin was included. The literature search included the following terms: cylindrospermopsin,
human toxicity, animal toxicity, in vitro toxicity, in vivo toxicity, occurrence, environmental fate, mobility
and persistence. EPA assembled available information on: occurrence; environmental fate; mechanisms of
toxicity; acute, short term, subchronic and chronic toxicity and cancer in humans and animals;
toxicokinetics and exposure.
Additionally, EPA relied on information from the following risk assessments in the development of the
HESD for cylindrospermopsin.
• Health Canada (2012) Toxicity Profile for Cyanobacterial Toxins
• Enzo Funari and Emanuela Testai (2008) Human Health Risk Assessment Related to Cyanotoxins
Exposure
• Tai Nguyen Duy, Paul Lam, Glen Shaw and Des Connell (2000) Toxicology and Risk Assessment
of Freshwater Cyanobacterial (Blue-Green Algal) Toxins in Water
• Cylindrospermopsin [CASRN 143545-90-8] Review of Toxicological Literature (ILS, 2000).
A Reference Dose (RfD) determination assumes that thresholds exist for certain toxic effects, such as
cellular necrosis, significant body or organ weight changes, blood disorders, etc. It is expressed in terms of
milligrams per kilogram per day (mg/kg/day) or micrograms per kilogram per day ((.ig/kg/day). In general,
the RfD is an estimate (with uncertainty spanning perhaps an order of magnitude) of a daily oral exposure
to the human population (including sensitive subgroups) that is likely to be without an appreciable risk of
deleterious effects during a lifetime.
The carcinogenicity assessment includes formal hazard identification and an estimate of tumorigenic
potency if applicable. Hazard identification is a weight-of-evidence judgment of the likelihood that the
agent is a human carcinogen via the oral route and of the conditions under which the carcinogenic effects
may be expressed.
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Development of this hazard identification and dose-response assessment for cylindrospermopsin has
followed the general guidelines for risk assessment as set forth by the National Research Council (1983)
the EPA's (2014b) Framework for Human Health Risk Assessment to Inform Decision Making. EPA
guidelines used in the development of this assessment include the following:
• Guidelines for the Health Risk Assessment of Chemical Mixtures (U.S. EPA, 1986a)
• Guidelines for Mutagenicity Risk Assessment (U.S. EPA, 1986b)
• Recommendations for and Documentation of Biological Values for Use in Risk Assessment
(U.S. EPA, 1988)
• Guidelines for Developmental Toxicity Risk Assessment (U .S. EPA, 1991)
• Interim Policy for Particle Size and Limit Concentration Issues in Inhalation Toxicity Studies
(U.S. EPA, 1994a)
• Methods for Derivation ofInhalation Reference Concentrations and Application of Inhalation
Dosimetry (U.S. EPA, 1994b)
• Use of the Benchmark Dose Approach in Health Risk Assessment (U.S. EPA, 1995)
• Guidelines for Reproductive Toxicity Risk Assessment (U .S. EPA, 1996)
• Guidelines for Neurotoxicity Risk Assessment (U.S. EPA, 1998)
• Science Policy Council Handbook. Peer Review (2nd edition) (U.S. EPA, 2000a)
• Supplemental Guidance for Conducting Health Risk Assessment of Chemical Mixtures (U.S. EPA,
2000c)
• A Review of the Reference Dose and Reference Concentration Processes (U.S. EPA, 2002)
• Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a)
• Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to Carcinogens
(U.S. EPA., 2005b)
• Science Policy Council Handbook: Peer Review (U.S. EPA, 2006a)
• A Framework for Assessing Health Risks of Environmental Exposures to Children (U.S. EPA,
2006b)
• Exposure Factors Handbook 2011 Edition (U.S. EPA, 2011)
• Benchmark Dose Technical Guidance Document (U.S. EPA, 2012)
• Child-Specific Exposure Scenarios Examples (U.S. EPA, 2014a)
• Framework for Human Health Risk Assessment to Inform Decision Making (U.S.EPA, 2014b).
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AUTHORS, CONTRIBUTORS AND REVIEWERS
Authors
Lesley V. D'Anglada, Dr.P.H. (Lead)
Joyce M. Donohue, Ph.D.
Jamie Strong, Ph.D.
Office of Water, Office of Science and Technology
Health and Ecological Criteria Division
U.S. Environmental Protection Agency, Washington D.C.
Belinda Hawkins, Ph.D., DABT
Office of Research and Development, National Center for Environmental Assessment
U.S. Environmental Protection Agency, Cincinnati, OH
The following contactor authors supported the development of this document:
Anthony Q. Armstrong, M.S.
Carol S. Wood, Ph.D., DABT
Oak Ridge National Laboratory, Oak Ridge, TN
The Oak Ridge National Laboratory is managed and operated by UT-Battelle, LLC. for the
U.S. Department of Energy under Contract No. DE-AC05-00OR22725.
The following contractor authors developed earlier unpublished drafts that contributed significantly
to this document:
Carrie Fleming, Ph.D. (former Oak Ridge Institute for Science and Education participant)
Oak Ridge National Laboratory, Oak Ridge, TN
Stephen Bosch, B.S.
Marc Odin, M.S., DABT
David Wohlers, Ph.D.
SRC, Inc., Syracuse, NY
Robyn Blain, Ph.D.
Audrey Ichida, Ph.D.
Kaedra Jones, MPH
William Mendez, Ph.D.
Pamela Ross, MPH
ICF International, Fairfax, VA
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Reviewers
Internal Reviewers
Neil Chernoff, Ph.D.
Armah de la Cruz, Ph.D.
Elizabeth Hilborn, DVM, MPH, DACVPM
Nicole Shao, M.S.
Jody Shoemaker, Ph.D.
External
Lorraine Backer, Ph.D., MPH
Wayne W. Carmichael, Ph.D.
Richard Charron, MSc.
Ian R. Falconer, Ph.D.
Michele Giddings, BSc.
James S. Metcalf , Ph.D.
Brett A. Neilan, Ph.D.
Ian Stewart, Ph.D.
Office of Research and Development, U.S. EPA
Office of Research and Development, U.S. EPA
Office of Research and Development, U.S. EPA
Office of Research and Development, U.S. EPA
Office of Research and Development, U.S. EPA
Centers for Disease Control and Prevention
Wright State University
Water and Air Quality Bureau, Health Canada
University of Adelaide
Water and Air Quality Bureau, Health Canada
Institute for Ethnomedicine
University of New South Wales
South Australian Government's R&D Institute (SARDI)
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TABLE OF CONTENTS
FOREWORD II
AUTHORS, CONTRIBUTORS AND REVIEWERS IV
TABLE OF CONTENTS VI
LIST OF TABLES VIII
LIST OF FIGURES VIII
ABBREVIATIONS AND ACRONYMS IX
EXECUTIVE SUMMARY XII
1.0 IDENTITY: CHEMICAL AND PHYSICAL PROPERTIES 1
2.0 TOXIN SYNTHESIS AND ENVIRONMENTAL FATE 4
2.1 Cyanotoxin Synthesis 4
2.1.1. Environmental Factors that Affect the Fate of Cyanotoxins 4
2.2 Environmental Fate of Cylindrospermopsin 8
2.3 Summary 9
3.0 CYANOTOXIN OCCURRENCE AND EXPOSURE IN WATER 10
3.1 General Occurrence of Cyanobacteria in Water 10
3.2 Cylindrospermopsin Occurrence in Surface Water 11
3.3 Cylindrospermopsin Occurrence in Drinking Water 12
3.4 Summary 12
4.0 OCCURRENCE IN MEDIA OTHER THAN WATER 13
4.1 Occurrence in Soil and Edible Plants 13
4.2 Occurrence in Fish and Shellfish 13
4.3 Occurrence in Dietary Supplements 14
4.4 Summary 14
5.0 TOXICOKINETICS 15
5.1 Absorption 15
5.2 Distribution 15
5.3 Metabolism 16
5.4 Excretion 17
5.5 Pharmacokinetic Considerations 17
6.0 HAZARD IDENTIFICATION 18
6.1 Case Reports and Epidemiology Studies 18
6.2 Animal Studies 19
6.2.1. Acute Toxicity 19
6.2.2. Short Term Studies 22
6.2.3. Subchronic Studies 23
6.2.4. Developmental/Reproductive Toxicity 27
6.2.5. Chronic Toxicity 29
6.3. Carcinogenicity 29
6.4. Other Key Data 30
6.4.1. Mutagenicity and Genotoxicity 30
6.4.2. Immunotoxicity 33
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6.5. Physiological or Mechanistic Studies 34
6.5.1. Noncancer Effects 34
6.5.2. Cancer Effects 37
6.5.3. Interactions with Other Chemicals 37
6.5.4. Structure Activity Relationship 38
7.0 CHARACTERIZATION 01 RISK 40
7.1 Synthesis and Evaluation of Maj or Noncancer Effects 40
7.1.1. Mode of Action for Noncancer Effects 41
7.1.2. Dose Response Characterization for Noncancer Effects 42
7.2 Synthesis and Evaluation of Major Carcinogenic Effects 42
7.2.1. Mode of Action and Implications in Cancer Assessment 43
7.2.2. Weight of Evidence Evaluation for Carcinogenicity 43
7.2.3. Dose Response Characterization for Cancer Effects 43
7.3. Potentially Sensitive Populations 43
7.4. Characterization of Health Risk 44
7.4.1. Choice of Key Study 44
7.4.2. Selection of the Principal Study 44
7.4.3. Selection of the Critical Endpoint 45
7.4.4. RfD Determination 45
8.0 RESEARCH GAPS 47
9.0 REFERENCES 48
APPENDIX A: STUDIES USED IN SUPPORT OF REFERENCE VALUE DERIVATION FOR
CYLINDROSPERMOPSIN 62
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LIST OF TABLES
Table 1-1. Chemical and Physical Properties of Cylindrospermopsin 3
Table 4-1. Bioaccumulation Studies of Cylindrospermopsin in Fish, Shellfish, and Crustaceans 14
Table 6-1. Kidney Weight Data from Oral Toxicity Study of Cylindrospermopsin Administered Daily
over Eleven Weeks (Humpage and Falconer, 2002, 2003) 24
Table 6-2. Selected Clinical Chemistry, Hematology, and Urinalysis Findings (Humpage and Falconer,
2002, 2003) 25
Table 6-3. C. raciborskii Tumor Initiating Results (Falconer and Humpage, 2001) 29
Table 6-4. Genotoxicity of Cylindrospermopsin in vitro 33
LIST OF FIGURES
Figure 1-1. Structure of cylindrospermopsin (de la Cruz et al., 2013) 2
Figure 1-2. Structure of 7-epicylindrospermopsin (de la Cruz et al., 2013) 2
Figure 1-3. Structure of 7-deoxycylindrospermopsin (de la Cruz et al., 2013) 2
Figure 2-1. Environmental factors influencing cyanobacterial blooms (Reproduced from Paerl and Otten,
2013b) 5
Figure 6-1. Structure of cylindrospermopsin and 7-epicylindrospermopsin (de la Cruz et al., 2013) 39
Figure 6-2. Structure of cylindrospermopsin analog AB-MODEL (Runnegar et al., 2002) 39
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ABBREVIATIONS AND ACRONYMS
ADHD Attention Deficit Hyperactivity Disorders
ALT Alanine Aminotransferase
ALP Alkaline Phosphatase
AST Aspartate Aminotransferase
ATP Adenosine Triphosphate
BGAS Bluegreen Algae Supplements
BNCs Binucleated Cells
BSO Buthionine Sulfoximine
BUN Blood Urea Nitrogen
BW Body Weight
CASRN Chemical Abstracts Service Registry Number
CCL Contaminant Candidate List
CBMN Cytokinesis Block Micronucleus Assay
CHO Chinese Hamster Ovary
CI Confidence Interval
CTA Cell Transformation Assay
CYP450 Cytochrome P450
DMSO Dimethylsulfoxide
DNA Deoxyribonucleic Acid
DW Dry Weight
DWHA Drinking Water Health Advisories
ED50 Median Effective Dose
ELISA Enzyme Linked Immunosorbent Assay
EPA United States Environmental Protection Agency
FEL Frank Effect Level
G Gram
GD Gestation Day
GFR Glomerular Filtration Rate
GSH Glutathione
HAB Harmful Algal Bloom
HESD Health Effects Support Document
HPLC High-Performance Liquid Chromatography
HSDB Hazardous Substances Data Bank
IC50 Inhibitory Concentrationso
ILS Integrated Laboratory Systems
I.P. Intraperitoneal
Kg Kilogram
Kow Octanol Water Partition Coefficient
Koc Soil Organic Carbon-Water Partitioning Coefficient
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L
Liter
LCAT
Lecithin-Acyl Cholesterol Transferase
LC/MS/MS
Liquid Chromatography-Tandem Mass Spectrometry
LC50
Median Lethal Concentration
LD50
Median Lethal Dose
LDH
Lactate Dehydrogenase
LOAEL
Lowest-Observed-Adverse-Effect Level
LPS
Lipopolysaccharides
MCH
Mean Corpuscular Hemoglobin
Mg
Microgram
(im
Micromole
MN
Micronuclei
MNBNC
Micronucleated Binucleated Cells
Mg
Milligram
Ml
Milliliter
MN
Mononuclear
MRNA
Messenger RNA
N
Nitrogen
N/A
Not Applicable
NARS
National Aquatic Resource Surveys
ng
Nanogram
NLA
National Lakes Assessment
nmol
Nanomole
NOAEL
No-Observed-Adverse-Effect Level
OECD
Organization for Economic Cooperation and Development
OHEPA
Ohio Environmental Protection Agency
OR
Odds Ratio
P
Phosphorus
PCR
Polymerase Chain Reaction
PMN
Polymorphonuclear
RBC
Red Blood Cell
RfD
Reference Dose
RNA
Ribonucleic Acid
ROS
Reactive Oxygen Species
RT-PCR
Reverse Transcription Polymerase Chain Reaction
SDH
Sorbitol Dehydrogenase
SDWA
Safe Drinking Water Act
SHE
Syrian Hamster Embryo
TEER
Trans-Epithelial Electric Resistance
TOXLINE
Toxicology Literature Online
TPA
O-Tetradecanoylphorbol-13 -Acetate
Ttgase
T ransglutaminase
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UF
Uncertainty Factor
UMP
Uridine Monophosphate
USACOE
United States Army Corps of Engineers
USGS
United States Geological Survey
uv
Ultraviolet
WHO
World Health Organization
WSDE
Washington State Department of Ecology
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EXECUTIVE SUMMARY
Cylindrospermopsin is a toxin produced by a variety of cyanobacteria including: Cylindrospermopsis
raciborskii, Aphanizomenon flos-aquae, Aphanizomenon gracile, Aphanizomenon ovalisporum, Umezakia
natans, Anabaena bergii, Anabaena lapponica, Anabaena planctonica, Lyngbya wollei, Rhaphidiopsis
curvata, and Rhaphidiopsis mediterranea. Under the right environmental conditions, cylindrospermopsin
may be produced and retained within the cell, although it is usually released outside the cell and dissolved
or sorbed to other materials in water. An increase in water column stability, high water temperatures,
elevated concentrations of nutrients, especially nitrogen and low light intensity have been associated with
an increase or dominance of cylindrospermopsin-producing cyanobacteria in surface waters or aquatic
ecosystems.
Cylindrospermopsin is relatively stable to both heat and pH in the dark. In the presence of algal cell
pigments, photochemical degradation can occur rapidly, with reported half-lives of 1.5 hours and
approximately 3 hours. In the absence of pigments, however, there is little decomposition. The
biodegradation of cylindrospermopsin in natural water bodies is a complex process that can be influenced
by many environmental factors, including its concentration, water temperature, sunlight, cell pigments,
and the presence of bacteria. Half-lives of 11 to 15 days and up to 8 weeks have been reported for
cylindrospermopsin in surface waters. Cylindrospermopsin is moderately mobile with low sorbtion to
sediment. Sorption is well correlated with the organic carbon content of soil or sediment.
Cylindrospermopsin-producing cyanobacteria are found in brackish and marine waters, freshwater ponds,
rivers, reservoirs and eutrophic lakes and have been reported in Australia, Asia, Europe, Africa and South,
Central and North America. Cylindrospermopsin has been detected in agricultural soils and edible plants
irrigated with cylindrospermopsin-contaminated water. In the United States, cylindrospermopsin also has
been found in source water and in one case, in finished drinking water.
Human exposure to cyanotoxins can occur by ingestion of toxin contaminated water or food, by
inhalation and dermal contact during bathing or showering, and during recreational activities in
waterbodies containing the toxins. The main source of information on the toxicity of cylindrospermopsin
in humans is from qualitative reports of a hepatoenteritis-like illness attributed to acute or short-term
consumption of drinking water containing Cylindrospermopsis raciborskii. Symptoms reported include
fever, headache, vomiting, bloody diarrhea, hepatomegaly and kidney damage with the loss of water,
electrolytes and protein. No reliable data are available on exposure levels of cylindrospermopsin that
induced these effects.
From limited oral toxicity studies in animals, cylindrospermopsin is likely absorbed from the
gastrointestinal tract. Based on oral and intraperitoneal (i.p.) studies in mice treated with purified
cylindrospermopsin or extracts of Cylindrospermopsis raciborskii cells, the liver and kidneys appear to be
the primary target organs for cylindrospermopsin toxicity. The metabolism and toxicity of
cylindrospermopsin involves the hepatic cytochrome P450 (CYP450) enzyme system. Laboratory studies
have found cylindrospermopsin in the urine, feces, liver, kidney and spleen in mice. Results of in vitro
mutagenic and genotoxic cell assays with cylindrospermopsin are varied with some indication of potential
DNA damage in mouse liver. However, these data are limited and there are no long term bioassays of
purified cylindrospermopsin.
The EPA reference dose (RfD) for cylindrospermopsin is 0.1 (ig/kg/day based on increased relative
kidney weight and decreased urinary protein from a study by Humpage and Falconer (2002, 2003). This
study identified a NOAEL of 30 (ig/kg/day and a LOAEL of 60 (ig/kg/day based on a relative increase in
kidney weight in rats. The composite uncertainty factor includes application of a 10 for intraspecies
variability, 10 for interspecies differences, and a 3 for uncertainties in the database.
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No epidemiological studies of the association of cylindrospermopsin and cancer are available. Also, no
chronic cancer bioassays of purified cylindrospermopsin in animals were identified. Therefore, under the
EPA's (2005) Guidelines for Carcinogen Risk Assessment, there is inadequate information to assess
carcinogenic potential of cylindrospermopsin.
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1.0 IDENTITY: CHEMICAL AND PHYSICAL PROPERTIES
Cyanobacteria, formerly known as blue-green algae (Cyanophyceae), are a group of bacteria containing
chlorophyll-a that can carry out the light and dark phases of photosynthesis (Castenholz and Waterbury,
1989). In addition to chlorophyll-a, other pigments such as carotene, xanthophyll, blue c phycocyanin and
red c phycoerythrin are also present in cyanobacteria (Duy et al., 2000). Most cyanobacteria are aerobic
photoautotrophs, requiring only water, carbon dioxide, inorganic nutrients and light for survival, but
others have heterotrophic properties and can survive long periods in complete darkness (Fay, 1965). Some
species also are capable of nitrogen fixation (i.e., diazotrophy) (Duy et al., 2000) producing inorganic
nitrogen compounds to synthesize nitrogen-containing biomolecules, such as nucleic acids and proteins.
Cyanobacteria can form symbiotic associations with animals and plants, such as fungi, bryophytes,
pteridophytes, gymnosperms and angiosperms, supporting their growth and reproduction (Sarma, 2013;
Hudnell, 2008; Hudnell, 2010; Rai, 1990).
Cyanobacteria can be found in unicellular, colony and multicellular filamentous forms. The unicellular
form occurs when the daughter cells separate after binary fission reproduction. These cells can aggregate
into irregular colonies held together by a slimy matrix secreted during colony growth (WHO, 1999). The
filamentous form occurs when repeated cell divisions happen in a single plane at right angles to the main
axis (WHO, 1999). Reproduction is asexual.
Cyanobacteria are considered gram-negative even though the peptidoglycan layer is thicker than most
gram-negative bacteria. However, studies using electron microscopy show that cyanobacteria possess
properties of both gram-negative and gram-positive bacteria. Compared with heterotrophic bacteria, the
cyanobacterial lipopolysaccharides (LPS) have little or no 2-Keto-3-deoxy-D-manno-octonic acid, lack
phosphate groups, glucosamine and L-glycero-D-mannoheptose, and have long-chain saturated and
unsaturated fatty acids.
Under optimal pH, nutrient availability, light and temperature conditions, cyanobacteria can reproduce
quickly forming a bloom. Studies of the impact of environmental factors on cyanotoxin production are
ongoing, including such factors as nutrient (nitrogen, phosphorus and trace metals) concentrations, light,
temperature, oxidative stressors and interactions with other biota (viruses, bacteria and animal grazers), as
well as the combined effects of these factors (Paerl and Otten 2013a; 2013b). Fulvic and humic acids also
have been reported to encourage cyanobacteria growth (Kosakowska et al., 2007).
Cyanobacteria can produce a wide range of bioactive compounds, some of which have beneficial or
therapeutic effects. These bioactive compounds have been used in pharmacology, as dietary supplements
and as mood enhancers (Jensen et al., 2001). Other cyanobacteria can produce bioactive compounds that
may be harmful, called cyanotoxins. The most commonly recognized bioactive compounds produced by
cyanobacteria fall into four broad groupings: cyclic peptides, alkaloids, amino acids and LPS.
The cyanotoxin cylindrospermopsin is a tricyclic alkaloid with the following molecular formula
C15H21N5O7S (Ohtani et al., 1992) and a molecular weight of 415.43 g/mole. It is zwitterionic (i.e., a
dipolar ion with localized positive and negative charges) (Ohtani et al., 1992) and is believed to be
derived from a polyketide that uses an amino acid starter unit such as glycocyamine or 4-guanidino-3-
oxybutyric acid (Duy et al., 2000). Two naturally occurring congeners of cylindrospermopsin (Figure 1-1)
have been identified including 7-epicylindro-spermopsin (Figure 1-2) and 7-deoxycylindrospermopsin
(Figure. 1-3) (Norris et al., 1999; de la Cruz et al., 2013). Recently, Wimmer et al., (2014) identified two
new analogs, 7-deoxy-desulfo-cylindrospermopsin and 7-deoxy-desulfo-12-acetylcylindrospermopsin,
from the Thai strain of Cylindrospermopsis raciborskii (C. raciborskii). The analogs were identified from
a Thai strain that is very similar to strains isolated from Japan and Australia and in a genetic study by
Chonudomkul et al. (2004) no differences were observed between these geographically separate strains.
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Figure 1-1. Structure of cylindrospermopsin (de la Cruz et al., 2013)
H OH
NH
HN
NH
©
Figure 1-2. Structure of 7-epicylindrospermopsin (de la Cruz et al., 2013)
HO H
NH
HN
NH
©
Figure 1-3. Structure of 7-deoxycylindrospermopsin (de la Cruz et al., 2013)
NH
HN
NH
©
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Cylindrospermopsin is a toxin produced by a variety of cyanobacteria including: Cylindrospermopsis
raciborskii (C. raciborskii), Aphanizomenon flos-aquae, Aphanizomenon gracile, Aphanizomenon
ovalisporum, Umezakia natans, Anabaena bergii, Ancibcienci lapponica, Anabaena planctonica, Lyngbyct
wollei, Rhaphidiopsis curvata, and Rhaphidiopsis mediterrcmea. Table 1-1 provides the chemical and
physical properties of cylindrospermopsin.
Cylindrospermopsin is highly soluble in water (Moore et al., 1998, Chiswell et al., 1999).
Cylindrospermopsin is isolated for commercial use mostly from C. raciborskii with a white powder
appearance. Other physico-chemical properties of cylindrospermopsin in the environment such as vapor
pressure, boiling and melting point, soil (Koc) and living organism's adsorption (Kow) coefficients, and
how it volatize from water and be distributed in the atmosphere (Henry's Law constant) have not been
determined. Limited information is available on the chemical breakdown, biodegradation and distribution
of cylindrospermopsin in the environment (see section 2.2 on Environmental Fate).
Table 1-1. Chemical and Physical Properties of Cylindrospermopsin
Property
Cylindrospermopsin
Chemical Abstracts Registry Number
(CASRN)
143545-90-8
Chemical Formula
C15H21N5O7S
Molecular Weight
415.43 g/mole
Color/Physical State
white powder
Boiling Point
N/A
Melting Point
N/A
Density
2.03g/cm3
Vapor Pressure at 25°C
N/A
Henry's Law Constant
N/A
Kow
N/A
Koc
N/A
Solubility in Water
Highly
Other Solvents
Dimethylsulfoxide (DMSO) and methanol
Sources: Chemical Book, 2012; TOXLINE, 2012
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2.0 TOXIN SYNTHESIS AND ENVIRONMENTAL FATE
2.1 Cyanotoxin Synthesis
Toxin production varies among blooms and within an individual bloom over time (Duy et al., 2000).
Cyanotoxins can be produced by more than one cyanobacterial species and species can produce more than
one toxin at a time, resulting in blooms with different cyanotoxins (Funari and Testai, 2008). The toxicity
of a particular bloom is determined by the mixture of species involved and their strain composition of
toxic and nontoxic genotypes (WHO, 1999). Generally, cyanobacteria toxins are retained within the cell
unless conditions favor cell wall lysis (ILS, 2000). Under the right environmental conditions,
cylindrospermopsin may be produced and retained within the cell, although it is usually released outside
the cell and dissolved or sorbed to other materials in water (Chiswell, et al. 2001). In contrast to other
cyanobacteria, some species of cylindrospermopsin do not form scums (dense accumulations of
cyanobacteria) and the highest cell concentrations can occur below the surface (Falconer 2005).
The synthesis of cyanotoxins is the focus of much research with evidence suggesting that the production
and accumulation of toxin(s) correlates with the cyanobacterial growth rate, with the highest amount
being produced during the late logarithmic growth phase (Funari and Testai, 2008). For example, Sukenik
et al. (1998) found that the concentration of cylindrospermopsin within A. ovalisporum from Lake
Kinneret increased to a plateau during the growth phase and decreased during the stationary phase. The
authors attributed this decrease to cell degradation and the release of the water-soluble toxin into the
medium.
Cylindrospermopsin biosynthesis starts with the production of guanidinoacetate from glycine and
arginine, a natural guanidino donor, followed by successive condensations of five intact acetates, and
subsequently by methylation, ketoreduction, sulfation and cyclizations (Moore et al., 1993; Looper et al.,
2006). Guanidinoacetate is known to be toxic and have been found to accumulate in the
cylindrospermopsin strain, contributing to the total cyanobacteria toxicity and possibly the cause of
increased toxicity in crude extracts in comparison with the purified cyanotoxin (Baron-Sola et al., 2015).
Enzymes encoded by two genes (cyrA- and cyrO) have been detected in cylindrospermopsin-producing
strains of C. raciborskii and are believed to initiate toxin biosynthesis (Schembri et al., 2001; O'Neil et
al., 2012). C-methylation, sulfotransfer and cyclization complete cylindrospermopsin biosynthesis.
Recently, 11 genes involved in cylindrospermopsin biosynthesis in C. raciborskii AWT205 were
identified and the biosynthesis pathway was described (Mihali et al., 2008; Mazmouz et al., 2010).
Little is known about how nitrogen affects cylindrospermopsin production. Saker and Neilan (2001)
observed the highest concentration (on a dry-weight basis) of cylindrospermopsin in cultures of C.
raciborskii in the absence of a fixed nitrogen source (Saker and Neilan, 2001). Some studies have
suggested that increased intracellular cylindrospermopsin content, in the absence of fixed nitrogen, was
due to hyp gene homologs in the C. raciborskii genome associated with the maturation of hydrogenases
(O'Neil et al., 2012). Phosphorus appears to play an important role in cylindrospermopsin production by
C. raciborskii due to the presence of genes that utilize inorganic and organic phosphorus, including those
for high affinity phosphate binding proteins (pstS and sphX), phosphanate transport proteins (phnC,D,E),
and enzymes for metabolism {phnG-M,X, WandphoA).
2.1.1. Environmental Factors that Affect the Fate of Cyanotoxins
Cyanotoxin production is influenced by environmental conditions that promote growth of particular
cyanobacterial species and strains (Fig 2-1). Micronutrient concentrations, temperature, light intensity,
water turbidity, pH, competing bacteria and phytoplankton, turbulence and salinity are all factors that
affect growth and change the dynamics of a cyanobacteria population dynamics as demonstrated in
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Figure 2-1. Although environmental conditions can affect the formation of blooms, the numbers of
cvanobacteria and toxin concentrations produced are not always closely related. Cyanotoxin
concentrations depend on the dominance and diversity of the cyanobacteria strains present within the
bloom, along with environmental and ecosystem influences on bloom dynamics (Hitzfeld et al., 2000;
WHO, 1999).
Figure 2-1. Environmental factors influencing cyanobacteria! blooms (Reproduced from Paerl and
Otten, 2013b)
Positive
1 High P (High N for some)
' Low N (DIN, DON) (only
applies to N2 fixers)
1 Low N:P Ratios
1 Low turbulence
¦ Low water flushing-Long
water residence time
1 High (adequate) light
1 Warm temperatures
1 High dissolved organic
matter
1 Sufficient Fe (+ trace
metals)
¦ Low grazing rates
w
CD
¦*—>
03
DC
Cyanos
CO
O)
¦*—>
CO
CC
Diversity
Modulating factors
Strong biogeochemical gradients (e.g.
persistent stratification, stable benthos)
Heterogeneous and diverse habitats (e.g.
reefs, seagrasses, marshes, sediments,
aggregates)
Selective grazing
"Toxin" production??
Negative
1 High DIN/ total N (only
applies to N2 fixers)
¦ Low P (DIP)
1 High N:P ratios
1 High turbulence & vertical
mixing
1 High water flushing-Short
water residence time
1 Low light (for most taxa)
¦ Cool temperatures
1 Low dissolved organic
matter
1 Low Fe (+ trace metals)
• High grazing rates
1 Viruses (cyanophages)
• Predatory bacteria
Nutrients—Nutrient concentrations are key environmental drivers that influence the proportion of
cyanobacteria in the phytoplankton community, the cyanobacterial biovolume, and the impact that
cyanobactena may have on ecosystem function and water quality. Cyanobacteria production and toxin
concentrations are dependent on nutrient levels (Wang et al., 2002); however, different cyanobacteria
species use organic and inorganic nutrient pools differently. Loading of nitrogen (N) and/or phosphorus
(P) to waterbodies from agricultural, industrial, and urban sources can induce the development of
cyanobacterial blooms and may be related to cyanotoxin production (Paerl et al., 2011).
Smith (1983) first described a strong relationship between the relative amounts of N and P in surface
waters and cyanobacterial blooms. Smith proposed that diazotrophic cyanobacteria should be superior
competitors under conditions of N-limitation because of their unique capacity for N-fixation. The
hypothesis that low N:P ratios favor cyanobactena formation has been intensely debated and challenged
for its poor performance in predicting cyanobacterial dominance (Downing et al., 2001). However, the
dominance of N-fixing cyanobacteria at low N:P ratios has been demonstrated in mesocosm- and
ecosystem-scale experiments in prairie and boreal lakes (Schindler et al., 2008). Eutrophic systems
already subject to bloom events are prone to further expansion of these blooms due to additional N inputs,
especially if sufficient P is available from internal sources. As the trophic state increases, aquatic systems
absorb higher concentrations of N (Paerl and Huisman, 2008; Paerl and Otten, 2013b). Recent surveys of
cyanobacterial and algal productivity in response to nutrient pollution across geographically diverse
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eutrophic lakes, reservoirs, estuarine and coastal waters plus a range of experimental enclosures (<1 L to
over 10,000 L), reveal that greater stimulation is observed in response to both N and P additions,
suggesting that nutrient colimitation is widespread (Elser et al., 2007; Lewis et al., 2011; Paerl et al.,
2011). These results strongly suggest that reductions in both N and P inputs are needed to stem
eutrophication and cyanobacterial bloom expansion.
PreuBel et al. (2014) investigated the influence of nitrogen and phosphorus availability on the production
and the release of cylindrospermopsin in three strains of Aphanizomenon sp. The authors found that
cylindrospermopsin was released from cells under both nitrogen availability and phosphorus limitation.
Under nitrogen-limiting conditions, the authors found a reduction in the release of cylindrospermopsin
from intact cells probably due to changing metabolic activities and the efficiency of resource
consumption.
Light Intensity—Sunlight availability and turbidity influence the predominance of cyanobacteria species
and the depth at which they occur (Falconer et al., 2005; Carey et al., 2012). For example,
Cylindrospermopsis forms dense layers of filaments at depths near the lower bound of the euphotic zone
in deeper rivers, lakes and reservoirs. The relationship of light intensity to toxin production in blooms is
somewhat unclear and continues to be investigated (Duy et al., 2000). While some researchers have found
evidence that toxin production increases with high light intensity (Watanabe and Oishi, 1985), others have
found little variation in toxicity at different levels of light intensity (Codd and Poon, 1988; Codd, 1995).
Deep water mixing and low light have been associated with an increase in the dominance of
C. raciborskii, a toxin producing species (O'Brien et al., 2009).
Kosten et al. (2011) surveyed 143 shallow lakes along a latitudinal gradient (between 5-55°S and
38-68°N) from subarctic Europe to southern South America). Their analyses found a greater proportion of
the total phytoplankton biovolume attributable to cyanobacteria in lakes with high rates of light
absorption. Kosten et al. (2011) could not establish cause and effect from these field data, but other
controlled experiments and field data have demonstrated that light availability can affect the competitive
balance among a large group of shade-tolerant species of cyanobacteria, mainly Oscillatoriales and other
phytoplankton species (Smith, 1986; Scheffer et al., 1997). Overall, results from Kosten et al. (2011)
suggest that higher temperatures interact with nutrient loading and underwater light conditions in
determining the proportion of cyanobacteria in the phytoplankton community in shallow lakes.
Temperature—The increasing body of laboratory and field data (Weyhenmeyer, 2001; Huisman et al.,
2005; Reynolds, 2006; De Senerpont Domis et al., 2007; Jeppesen et al., 2009; Wagner and Adrian, 2009;
Kosten et al., 2011; Carey et al., 2012) suggest that warming may influence cyanobacterial dominance.
Cyanobacteria can benefit more from warming than other phytoplankton groups due to their higher
optimum growth temperatures. The increase in water column stability associated with higher temperatures
also favors cyanobacteria (Wagner and Adrian, 2009; Carey et al., 2012). In their analyses of 143 lakes
along a latitudinal transect from subarctic Europe to southern South America, Kosten et al. (2011)
demonstrated that in shallow lakes the percentage of the total phytoplankton biovolume attributable to
cyanobacteria increased steeply with temperature.
Indirectly, warming also may increase nutrient concentrations by enhancing mineralization (Gudasz et al.,
2010; Kosten et al., 2009 and 2010) and by temperature- or anoxia-mediated sediment phosphorus release
(Jensen and Andersen, 1992; Sondergaard et al., 2003). Thus, temperature may increase cyanobacteria
biomass indirectly through its effect on nutrient concentrations. Others have suggested that warmer
conditions may raise total phytoplankton biomass through an alteration of top-down regulation by grazers
(Jeppesen et al., 2009, 2010; Teixeira-de Mello et al., 2009).
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Rising global temperatures and changing precipitation patterns can stimulate cyanobacteria blooms.
Warmer temperatures favor surface bloom-forming cyanobacterial genera because they are heat-adapted
and their maximal growth rates occur at relatively high temperatures, often in excess of 25°C (Robarts
and Zohary 1987; Reynolds, 2006). At these elevated temperatures, cyanobacteria routinely out-compete
eukaryotic algae (Elliott, 2010; Paerl et al., 2011). Specifically, as the growth rates of the eukaryotic taxa
decline in response to warming, cyanobacterial growth rates reach their optima. Warmer surface waters,
especially in areas of reduced precipitation, are prone to intense vertical stratification. The degree of
vertical stratification depends on the density difference between the warm surface layer and the
underlying cold water which is influenced by amount of precipitation. As temperatures rise due to climate
change, stratification is expected to occur earlier in the spring and persist longer into the fall favoring
cyanobacteria production and release of cylindrospermopsin (Paerl and Otten, 2013b).
Other Environmental Factors—Cyanobacterial blooms have been shown to intensify and persist at pH
levels between six and nine (WHO, 2003). When these blooms are massive or persist for a prolonged
period, they can become harmful. Kosten et al. (2011) noted the impact of pH on cyanobacteria
abundance in lakes along a latitudinal transect from Europe to southern South America. The percentage of
cyanobacteria in the 143 shallow lakes sampled was well correlated with pH, with an increased proportion
of cyanobacteria at higher pH.
Cyanobacteria have a competitive advantage over other phytoplankton species because they are efficient
users of molecular carbon dioxide (Shapiro, 1984; Caraco and Miller, 1998), especially when increasing
pH diminishes the availability of carbon dioxide in the water column. Although this could explain the
positive correlation between pH and the proportion of cyanobacteria, the high proportion of cyanobacteria
at high pH could be the result of an indirect nutrient effect as described previously (see discussion in
Temperature section). As photosynthesis intensifies, pH increases due to carbon dioxide uptake by algae,
resulting in a shift in the carbonic buffer equilibrium and a higher concentration of basic forms of
carbonate. Higher pH in the water column can be a reflection of higher photosynthetic rates, which can be
linked with high nutrient concentrations (Duy et al., 2000) that stimulate phytoplankton growth and
bloom formation.
Most phytoplankton-cyanobacteria blooms occur in late summer and early fall and the phytoplankton
community can become vertically stratified. The vertical phytoplankton biomass structure and cyanotoxin
production can be influenced by seasonal changes as well as weather conditions (e.g., wind, rainfall), and
also by runoff. At times, the bottom layer can have more biomass and display different population
dynamics than the upper water column. Conversely, seasonal influences with increases in temperature and
changes in wind patterns may favorably influence the upper water column cyanobacterial community to
become dominant. This vertical variability is common and attributed to four causes, each of which may
occur at different times, including: (a) sinking of dead/dying cells; (b) density stratification of the water
column, especially nutrient concentrations and light, which affects all aspects of the cyanobacteria
growth; (c) nutrient supply from organic-rich bottom sediment (even when the water body is not density-
stratified), encouraging growth at or near the sediment; and, (d) species-specific factors (Drake et al.,
2010). In addition, there are microbial interactions within blooms, such as competition and adaptation
between toxic and nontoxic cyanobacterial strains, as well as attacks of cyanobacteria by viruses. Each of
these factors can cause fluctuations in bloom development and composition. When the composition of the
cyanobacterial bloom changes, the toxins present and their concentrations may change as well (Honjo et
al., 2006; Paerl and Otten, 2013b). The concentration of cyanotoxins observed in the water column when
a bloom collapses, such as from cell aging or from algaecide treatment, depends on dilution of the toxin in
the impacted water due to water column mixing, the degree of adsorption to sediment or particulates, and
the rate of toxin biodegradation (Funari and Testai, 2008).
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In summary, there is a complex interplay of environmental factors that dictates the spatial and temporal
pattern in the concentration of cyanobacteria cells and their toxins with respect to the dominant species as
illustrated in Figure 2-1 (Paerl and Otten, 2013b). Factors such as the N:P ratio, organic matter
availability, temperature, and light attenuation, as well as other water and physico-chemical processes,
can play a role in determining harmful algal bloom (HAB) composition and toxin production (Paerl and
Huisman, 2008; Paerl and Otten, 2013b). Dynamics of microflora competition as blooms develop and
collapse can also impact cyanotoxin concentrations in surface waters. In addition, impacts of climate
change including potential warming of surface waters on ecosystem dynamics that lead to more frequent
formation of cyanobacteria blooms and their associated toxins (Paerl and Huisman, 2008; Paerl et al.,
2011; Paerl and Otten, 2013b).
2.2 Environmental Fate of Cylindrospermopsin
Hydrolysis—Cylindrospermopsin is relatively stable to heat and pH in the dark (Moore et al., 1998).
Studies have found cylindrospermopsin is stable at temperatures from 4°C to 50°C for up to five weeks in
the dark (ILS, 2000).
Photolysis—Chiswell et al., 1999 reported that cylindrospermopsin in an algal extract solution
decomposes rapidly (half-life of 1.5 h) when exposed to sunlight; however, no decomposition was
recorded in pure cylindrospermopsin and Milli-Q water solutions. They further observed that
cylindrospermopsin remains a potent toxin even after boiling for 15 minutes. Pure cylindrospermopsin is
relatively stable in sunlight, but in the presence of cell pigments, photochemical degradation can occur
rapidly. Researchers have noted that degradation rates are concentration-dependent. When
cylindrospermopsin (1 mg/L aqueous media) was exposed to normal sunlight, 54% remained after
3 hours; at 4 mg/L, cylindrospermopsin degraded more rapidly, with 29% of the original concentration
remaining after 3 hours (ILS 2000). When cell pigments are present, photolysis has been shown to
degrade more than 90% of cylindrospermopsin within 2 to 3 days (Chiswell et al., 1999).
Metabolism—Toxins released from cyanobacteria into lakes are decomposed by bacteria (Falconer,
1998). A half-life of 11 to 15 days has been reported for cylindrospermopsin in surface waters (Funari and
Testai, 2008). However, at pH 4, 7 and 10, cylindrospermopsin can remain stable for a period of up to
eight weeks (ILS, 2000). Smith et al. (2008) concluded that the biodegradation of cylindrospermopsin in
natural water bodies is a complex process that can be influenced by many environmental factors,
including: concentration, temperature and the presence of copper-based algaecides. Studies by Klitzke
and Fastner (2012) found that degradation of cylindrospermopsin in sediment was completely inhibited or
retarded under anoxic conditions (T1/2 oxic =2.4 days; T1/2 anoxic =23.6 days). A decrease in temperature
from 20 °C to 10 °C slowed down degradation rates by a factor of 10. Smith et al. (2008) reported an
optimum degradation rate between 25 °C and 30°C. Mohamed and Alamri (2012) reported that Bacillus
strain (AMRI-03) isolated from cyanobacterial blooms degraded cylindrospermopsin in laboratory
studies. Cylindrospermopsin degradation occurred rapidly, with a complete degradation based on the
initial concentration of cylindrospermopsin. Degradation occurred after 6 days at the highest tested
concentration (300 |ig/L) compared to seven and eight days at lower concentrations (10 and 100 |ig/L.
respectively) and depended on temperature (25 and 30°C) and pH (7 and 8).
Transport—Klitzke et al. (2011) reported low sorption of cylindrospermopsin to sediments and moderate
mobility. Sorption was non-linear and results were best fit using a Langmuir model. Organic carbon
proved to be the main parameter in sediment that determines sorption of cylindrospermopsin, with little
sorption observed on sandy and silt sediments. Cation exchange played only a minor role in comparison
to sorption to organic carbon. Sorption of cylindrospermopsin to sediment increased at low pH (Klitzke et
al., 2011). The authors suggested that the low sorption of cylindrospermopsin to sediment could be due to
its high polarity and tendency to remain in solution.
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2.3 Summary
Cylindrospermopsin is produced by a variety of cyanobacteria. Environmental conditions such as
nutrients, pH, light intensity and temperature can influence the growth of cyanobacteria and encourage
toxin production. Some species of cyanobacteria do not form scums; high cell concentrations occur below
the water surface because cyanobacteria have gas vacuoles to regulate their position in the water column.
Cylindrospermopsin may be retained within the cell, but is usually found dissolved or attached to other
materials in water. Cylindrospermopsin is relatively stable in the dark and remains potent even after
boiling for 15 minutes. In sunlight, photochemical degradation of cylindrospermopsin in water can occur
rapidly, within 2 to 3 days, especially when cell pigments are present. The biodegradation of
cylindrospermopsin in natural water bodies is pH and temperature dependent. The optimum degradation
rate has been reported between 25°C and 30°C. Its half-life in surface water ranges from 11 to 15 days,
but, cylindrospermopsin can remain stable for a period of up to eight weeks (at pH 4, 7 and 10).
Cylindrospermopsin adsorbs onto sediment and is moderately mobile. Organic carbon content is a key
sediment parameter determining sorption, with little sorption observed on sandy and silt sediment.
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3.0 CYANOTOXIN OCCURRENCE AND EXPOSURE IN WATER
The presence of detectable concentrations of cyanotoxins in the environment is closely associated with
blooms of cyanobacteria. Cyanobacteria flourish in various natural environments including salty, brackish
or fresh water, cold and hot springs and in environments where no other microalgae can exist, including
desert sand, volcanic ash and rocks (Jaag, 1945; Dor and Danin, 1996). Cyanobacteria also form
symbiotic associations with aquatic animals and plants, and cyanotoxins are known to bioaccumulate in
common aquatic vertebrates and invertebrates (Ettoumi et al., 2011).
Currently, there is no national database recording freshwater harmful algal blooms (HAB) events. Instead,
states and local governments document HAB occurrences in various ways depending on the monitoring
methods used and the availability of laboratories capable of conducting algal toxin analyses.
Human exposure to cyanotoxins, including cylindrospermopsin, may occur by direct ingestion of toxin-
contaminated water or food, and by inhalation and dermal contact during bathing, showering or during
recreational activities in waterbodies contaminated with the toxins. Cylindrospermopsin may be retained
within the cell, but most of the time (50/50 ratio) it is found in the water (extracellular) or attached to
particulates present in the water (Chiswell et al., 2011). Exposure through drinking water can occur if
there are toxins in the water source and the existing water treatment technologies were not designed for
removal of cyanotoxins. Because children consume more water per unit body weight than do adults,
children potentially may receive a higher dose than adults. Exposures are usually not chronic; however,
they can be repeated in regions where cyanobacterial blooms are more extensive or persistent. As
described above, cylindrospermopsin is not considered persistent in natural waters, thus exposure from
ambient surface waters is more likely to be acute or subacute. People, particularly children, recreating
close to lakes and beach shores also can be at potential risk from exposure to nearshore blooms.
Livestock and pets are potentially exposed to higher concentrations of cyanobacterial toxins than humans
because they are more likely to consume scum and mats when drinking cyanobacteria-contaminated water
(Backer et al., 2013). Dogs are particularly at risk as they may lick cyanobacteria from their fur after
swimming in a water body with an ongoing bloom.
3.1 General Occurrence of Cyanobacteria in Water
Species of cyanobacteria are predominantly found in eutrophic (nutrient-rich) water bodies in freshwater
and marine environments (ILS, 2000), including salt marshes. Most marine cyanobacteria of known
public health concern grow along the shore in benthic vegetation between the low- and high-tidewater
marks but can grow as free-floating water blooms (Walsh et al, 2008). The marine planktonic forms have
a global distribution. They also can be found in hot springs (Castenholz, 1973; Mohamed, 2008),
mountain streams (Kann, 1988), Arctic and Antarctic lakes (Skulberg, 1996) and in snow and ice
(Laamanen, 1996).
Gas vacuoles of A. ovalisporum and C. raciborskii can regulate the position of the cyanobacteria in the
water column. These species of cyanobacteria do not form a floating scum, but concentrate (with densities
up to 100,000 cells/mL) several meters below the surface. Because the cells remain suspended in the
water column, potentially toxin-producing blooms of these cyanobacteria may not be readily observable.
In older blooms, some cyanotoxins (including cylindrospermopsin) may be found at higher concentrations
dissolved in the water column (Rucker et al., 2007).
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3.2 Cylindrospermopsin Occurrence in Surface Water
C. raciborskii occurs in freshwater ponds, rivers, reservoirs and eutrophic lakes and has been found in
Australia, Asia, Europe, Africa and South, Central and North America (Fuentes et al., 2010).
Cylindrospermopsin-producing cyanobacteria occur in tropical or subtropical regions, but also have been
detected in warmer temperate regions. Surveys conducted in Florida, the Great Lakes and the Midwest,
and monitoring efforts in Ohio and Washington indicate that freshwater cyanotoxins are prevalent in the
U.S., mostly during warm seasons (Hudnell, 2010; Graham et al., 2010).
According to a survey conducted in Florida in 1999 from June to November, the most frequently
observed toxigenic cyanobacteria were Microcystis (43.1%), Cylindrospermopsis (39.5%), and Anabaena
spp (28.7%) (Burns, 2008). Of 167 surface water samples taken from 75 waterbodies, 88 samples were
positive for cyanotoxins. The actual cylindrospermopsin concentrations in ambient water were not
reported.
Concentrations of cylindrospermopsin have been reported at concentrations between 0.05 and 0.2 mg/L in
Florida since 1999 by The Harmful Algal Bloom Task Force (Pelaez et al., 2010). C. raciborskii have
also been detected, in some cases at more than 100,000 cells per mL. Additional data collected from the
Florida Department of Health found consistent cylindrospermopsin production in specific lakes at
concentrations ranging from 0.5 to 1.6 mg/L during the months of July through October. Samples
collected in the St. Johns River in 2008 around the same months (June through October) found
cylindrospermopsin consistently present ranging from 0.05 to 0.44 mg/L.
Samples collected from 2000 to 2004 in Lake Erie and analyzed by protein phosphatase inhibition assay
(PPIA) detected cylindrospermopsin in 3% of the samples at concentrations greater than 0.01 mg/L
(Pelaez et al., 2010).
Between 2000 and 2004, water samples were collected for cyanotoxin analysis from 81 different New
York lakes during June to October (Boyer et al., 2008). Cylindrospermopsin was measured by high
performance liquid chromatography (HPLC) and detected in 8 of the 366 samples with concentrations
less than 0.25(ig/L.
In Oklahoma during 2005, the U.S. Army Corps of Engineers (USACE) detected cylindrospermopsin at a
maximum concentration of 1.6 (ig/L (Lynch and Clyde, 2009). During the same year in Wisconsin, sixty-
five samples were taken in Castle Rock and Petenwell lakes for blue-green algae and toxin identification
(Evans, 2011). Cylindrospermopsis, which is not commonly found in Wisconsin, was present in only 6%
of the samples.
In 2005, Washington State Department of Ecology developed the Ecology Freshwater Algae Program,
focuses on the monitoring and management of cyanobacteria in Washington lakes, ponds, and streams
(WSDE, 2012). Data have been summarized in a series of reports for the Washington State Legislature
(Hamel, 2009; 2012). Cylindrospermopsin was below the state recreational guidance level of 1 |ig/L in
41 lakes tested in 2010, and was not detected in 46 lakes sampled in 2011.
In Florida, C. raciborskii was found to be the dominant cyanobacteria species in one lake all year round
(Burns, 2008). Cylindrospermopsin was also detected from Aphanizomenon ovalisporum in levels ranging
from 7.39 to 9.33 |ig/mg freeze-dried cells (Yilmaz et al., 2008). This finding supports the potential of
cylindrospermopsin to be produced by other cyanotoxin-producing species.
In 2006, C. raciborskii was detected in lakes in southern Louisiana (Fuentes et al., 2010). Conditions
promoting its growth were identified as shallow, warm surface water (over 30°C) and low light
intensities. The highest concentrations of C. raciborskii were observed from June through August with
densities ranging from 37,000 cells/mL to more than 160,000 cells/mL. In a study of two lakes directly
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connected to Lake Michigan, Hong et al., (2006) found low concentrations only in the late summer and
these were associated with elevated bottom water temperatures and phosphorus concentrations.
In 2006, the U.S. Geological Survey (USGS) conducted a study of 23 Midwestern lakes in which
cyanobacterial blooms were sampled and analyzed by enzyme-linked immunosorbent assays (ELISA) and
by direct-inject multianalyte liquid chromatography/tandem (LC/MS/MS) to determine the co-occurrence
of toxins and taste-and-odor compounds in cyanobacterial blooms (Graham et al., 2010). Microcystin was
detected in all the blooms, anatoxin-a was detected in 30% of the blooms, and cylindrospermopsin was
detected in 9% of the blooms sampled. The low concentrations of cylindrospermopsin (0.12 to 0.14 (ig/L)
detected in these studies were associated with algal communities dominated by Aphanizomenon or
Anabaena and/or Microcystis, but not in those dominated by Cylindrospermopsis. The authors attributed
the low concentration of cylindrospermopsin to either the lack of toxin production by Cylindrospermopsis
strains in the U.S as compared to elsewhere in the world, or to the lack of favorable environmental
conditions for the toxic strains and/or toxin production in the lakes sampled.
EPA's National Aquatic Resource Surveys (NARS) generate national estimates of pollutant occurrence
every 5 years. In 2007, the National Lakes Assessment (NLA) conducted the first-ever national
probability-based survey of algal toxins in the nation's lakes. A total of 1,028 lakes were sampled for the
NLA during summer 2007, representing the condition of about 50,000 lakes nationwide. The NLA looked
at actual cyanobacterial cell counts and chlorophyll-a concentrations as indicators of the potential for the
presence of algal toxins including microcystin and cylindrospermopsin. However, concentrations of
cylindrospermopsin were not reported. The USGS subsequently analyzed the stored samples collected
during the NLA and reported the presence, but not actual concentrations of cylindrospermopsin, in 5% of
the samples collected (Loftin and Graham, 2014). Future NARS plan to include other algal toxins,
including cylindrospermopsin.
Since 2007, Ohio EPA (OHEPA, 2012) has been monitoring inland lakes for cyanotoxins. In 2010,
OHEPA sampled Grand Lake St. Marys for anatoxin-a, cylindrospermopsin, microcystin, and saxitoxin.
Cylindrospermopsin concentrations ranged from below the detection limit (<0.15) to 9 (ig/L.
3.3 Cylindrospermopsin Occurrence in Drinking Water
The occurrence of cyanotoxins in finished drinking water depends on their levels in the raw source water
and the effectiveness of the treatment methods used for removing cyanobacteria and cyanotoxins.
Currently, there is no federal or state program in place that requires monitoring for cyanotoxins at U.S.
drinking water treatment plants. Therefore, data on the presence or absence of cyanotoxins in finished
drinking water are limited.
A survey conducted in 2000 in Florida (Burns, 2008) found cylindrospermopsin in raw drinking water
and in nine finished drinking water samples at concentrations ranging from 8 (ig/L to 97 |_ig/L.
3.4 Summary
Cylindrospermopsin-producing cyanobacteria occur in freshwater systems in tropical or subtropical
regions, but also can occur in warmer temperate regions. No national database on the occurrence of
freshwater cylindrospermopsin is available, and no federal or state program is in place to monitor for
cyanotoxins at U.S. drinking water treatment plants.
Exposure to cylindrospermopsin from contaminated drinking water could occur via oral exposure
(e.g. ingestion of contaminated drinking), dermal exposure (contact of exposed parts of the body with
water containing toxins) and inhalation exposure. Exposure to cylindrospermopsin during recreational
activities could occur through direct contact, inhalation and/or ingestion. Exposures usually are not
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chronic with the exception of regions with extensive and persistent cyanobacterial blooms. Since
cylindrospermopsin is not expected to be persistent in surface waters, exposure will depend on the
formation and persistence of the blooms and the related toxin concentration.
4.0 OCCURRENCE IN MEDIA OTHER THAN WATER
4.1 Occurrence in Soil and Edible Plants
Cyanobacteria are highly adaptable and have been found to colonize infertile substrates, such as volcanic
ash and desert sand (Jaag, 1945; Dor and Danin, 1996; Metcalf et al., 2012). They also have been found in
soil, at the surface or several centimeters below the surface, where they play a functional role in nutrient
cycling. Cyanobacteria are known to survive on rocks or tree trunks, and in snow and ice (Adhikary,
1996). They have been reported in deeper soil layers likely transported by percolating water or burrowing
animals. Some freshwater species are halotolerant (salt tolerant) and have been found in saline
environments such as salt works or salt marshes (WHO, 1999). Cyanobacterial cells can bioaccumulate in
zooplankton (Watanabe et al., 1992). As a result of higher trophic level grazing, the damaged or residual
cyanobacterial cells may settle out of the water column and accumulate in sediment where breakdown by
sediment bacteria and protozoa can release their toxins (Watanabe et al., 1992).
Cyanobacterial cells and toxins can contaminate spray irrigation water and subsequently be taken up by
crop plants after spray irrigation (Corbel et al., 2014). Water contaminated with toxins produced by
cyanobacterial cells that is then used for spray irrigation may produce food chain contamination since low
levels of cyanotoxins could be absorbed by roots, migrate to shoots, and then translocated to grains and or
fruits. Cyanotoxins can be accumulated in plant leaves. Kittler et al. (2012) found that crop plants
irrigated with cylindrospermopsin-contaminated water showed significant cylindrospermopsin uptake in
the leaves at 10% to 21% of the cylindrospermopsin concentration applied to the roots. Water
contaminated with cyanotoxins used for spray irrigation of crop plants inhibited plant growth and induced
visible effects such as the appearance of brown leaves (Funari and Testai, 2008). Therefore, according to
the authors, affected plants and crops will most likely not be used for eating purposes. Further
investigation is needed to understand the uptake and fate of cylindrospermopsin and other cyanobacterial
toxins by food plants.
4.2 Occurrence in Fish and Shellfish
Cyanotoxins can bioaccumulate in common aquatic vertebrates and invertebrates, including fish, snails
(Carbis et al., 1997; Beattie et al., 1998; Berry et al., 2012) and mussels (Eriksson et al., 1989; Falconer et
al., 1992; Prepas et al., 1997; Watanabe et al., 1997; Funari and Testai, 2008). Human exposure to
cyanotoxins may occur if fish are consumed from reservoirs with existing blooms of toxin-producing
cyanobacteria (Magalhaes et al., 2001).
The health risk from consumption depends on the bioaccumulation of cyanotoxins in edible fish tissue
compared to organs such as the liver. Levels of cylindrospermopsin found in tissues of aquatic species
potentially consumed by humans are shown in Table 4-1. One study (Saker and Eaglesham, 1999)
determined the concentration of cylindrospermopsin in redclaw crayfish and rainbow fish from
aquaculture ponds. Cylindrospermopsin concentrations were 0.9 and 4.3 jj.g/g freeze-dried tissue in
crayfish muscle and hepatopancreas, respectively, and 1.2 jj.g/g freeze-dried tissue in the viscera of
rainbow fish. This study also demonstrated that bioaccumulation can occur in fish that are exposed for
longer periods of time to a cyanobacterial bloom. Recent reviews also included levels of
cylindrospermopsin in freshwater mussels and prawns (Kinnear, 2010; Funari and Testai, 2008; Ibelings
and Chorus, 2007). No cases of toxicity in humans following ingestion of fish or shellfish exposed to
cylindrospermopsin have been documented.
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Table 4-1. Bioaccumulation Studies of Cylindrospermopsin in Fish, Shellfish, and Crustaceans.
Species/tissue
Concentration
Conditions
Reference
Fish
Rainbow fish - viscera
1.2 |jg/g freeze dried
tissue
Aquaculture pond during
bloom; 589 |jg/L
cylindrospermopsin
Saker and
Eaglesham,
1999
Shellfish
Alathyria pertexta
0.13-0.56 |jg/g fresh
tissue
Experimental exposure to
reservoir water; <0.8 |jg/L
cylindrospermopsin
Kinnear, 2010
Swan mussel
Hemolymph
Viscera
Whole body
61.5 |jg/g dry tissue
5.9 |jg/g dry tissue
2.9 |jg/g dry tissue
Experimental; 14-90 |jg/L
cylindrospermopsin
Kinnear, 2010
Mussel
Whole body
Viscera
0.247 |jg/g wet wt.
1.099 |jg/g wet wt.
Experimental exposure
concentration not given;
secondary citation
Saker et al.,
2004
Crustaceans
Crayfish muscle tissue
hepatopancreas
0.9 |jg/g freeze dried
tissue
4.3 |jg/g freeze dried
tissue
Aquaculture pond during
bloom; 589 |jg/L
cylindrospermopsin
Saker and
Eaglesham,
1999
Prawns - flesh
0.205 |jg/g wet wt.
Survey;
cylindrospermopsin
concentrations not given
Ibelings and
Chorus, 2007
4.3 Occurrence in Dietary Supplements
Extracts from Arlhrospira (Spirulina spp.) and Aphcmizomenon flos-aquae (AFA) have been used as
dietary bluegreen algae supplements (BGAS) (Funari and Testai, 2008). These supplements are reported
to have beneficial health effects including supporting weight loss, and increasing alertness, energy and
mood elevation for people suffering from depression (Jensen et al., 2001). In children, they have been
used as an alternative, natural therapy to treat attention deficit hyperactivity disorders (ADHD).
Heussner et al. (2012) analyzed 18 commercially available BGAS for the presence of toxins. Neither
anatoxin-a nor cylindrospermopsin were found in any of the supplements.
4.4 Summary
Cylindrospermopsin could be detected in aquatic animals, field soils and edible plants. Bioaccumulation
occurs mostly in the viscera of fish, shellfish and crustaceans, but cylindrospermopsin has also been
detected in fish tissue. No cases of toxicity in humans following ingestion of fish or shellfish exposed to
cyanotoxins have been documented.
Cylindrospermopsin has not been found in any of the tested commercially-available blue-green algal
supplements. Exposure to cylindrospermopsin for the general population is most likely through the
ingestion of drinking water and incidental ingestion when recreating in a water source contaminated with
cylindrospermopsin.
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5.0 TOXICOKINETICS
The available toxicokinetic data for cylindrospermopsin are from studies that do not reflect environmental
exposure conditions. All studies identified for this assessment were generated using intraperitoneal (i.p)
exposures to mice or in vitro assays rather than by the oral, dermal and/or inhalation routes applicable to
humans and domestic animals.
5.1 Absorption
Data on human and animal absorption of cylindrospermopsin after inhalation or dermal exposure were not
located. In two oral animal studies (Humpage and Falconer, 2002, 2003; Shaw et al., 2000, 2001), mice
were exposed to pure cylindrospermopsin for 14 days and 11 weeks, respectively. Systemic effects
observed in these studies following oral administration of cylindrospermopsin suggest absorption from
the gastrointestinal tract. The structural and conformational properties of the cylindrospermopsin
molecule suggest that uptake by the intestines and other tissues likely involves facilitated transport. No
data were identified relative to potential membrane receptors with properties compatible with the
properties of the cylindrospermopsin ion.
5.2 Distribution
Total tissue distribution of cylindrospermopsin following oral, inhalation or dermal exposure is unknown.
A series of three studies were done in six-week old male Quackenbush mice exposed to sublethal and
lethal doses of 14C-cylindrospermopsin (>95% pure) in normal saline by intraperitoneal (i.p.)
administration (Norris et al., 2001). At 48 hours, analysis of kidney, liver, and spleen after a 0.1 mg/kg
dose demonstrated 13.1% 14C recovery of the dose in the liver and <1% in the rest of the tissues. In each
of the four mice tested the total recovery of radiolabel cylindrospermopsin from tissues and excreta was
85-90% of the administered dose; 68% of the dose was found in urine and 15.5% in the feces.
In the second study, Norris et al. (2001) administered a single dose of 0.2 mg/kg dose of
14C-cylindrospermopsin by i.p. to 12 mice. After 12 and 24 hours, urine and feces in all animals had
detectable levels of 14C content. Five mice euthanized after 5-6 days (due to unspecified toxicity), had 14C
content in the liver, kidney and spleen. The remaining 7 mice, also had 14C content after 7 days with no
signs of toxicity. After 5 to 7 days, the overall mean (and standard deviation) recoveries of 14C were 2.1 ±
2.1 in the liver, 0.15 ± 0.14 in the kidneys and <0.1% (no standard deviation provided) of the dose in the
spleen. The broad standard deviations are indicative of considerable inter-individual differences in
response. There was no clear relationship between the signs of toxicity and the observed tissue
distribution. However, Norris et al. (2001) proposed that the lack of toxicity could be explained by a
tendency toward decreased liver retention in surviving mice.
In the third experiment, Norris et al. (2001) evaluated the excretion and tissue distribution in four mice
after the administration of a 0.2 mg/kg i.p. dose of 14C-cylindrospermopsin. After 6 hours, liver, kidney,
heart, lung, spleen, blood and bile were examined for 14C content. Detection of 14C was observed in all
tissues, however, mean 14C content was higher in the liver (20.6% (range 14.6 to 27.9), and 4.3% (range
3.7 to 4.7) of the dose in the kidneys. After a week, around 2% of the 14C content was detected in the
liver.
A slow, progressive, non-energy dependent uptake of purified cylindrospermopsin was detected in a
cultured African green monkey kidney cell line (Vero cells) (Froscio et al., 2009). Although, conducted in
vitro, these results suggest facilitated transport as a mechanism for uptake by the kidney.
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Studies on the distribution of cylindrospermopsin in fish using immunohistochemical (IHC) techniques
have found immunopositive results in the liver, followed by the kidney, intestines, and gills (Guzman-
Guillen, et al., 2014). IHC techniques were used in fish (Oreochromis niloticus) to determine the
distribution of 200 (.ig pure CYN/Kg body weight (bw) administered by i.p. or by gavage and evaluated
after 5 days of exposure. In addition, fish were also exposed to CYL by immersion to either 10 or
100 (ig/L of lyophilized A. ovalisporum cells for 7 or 14 days. Results were similar in both experimental
methods. Immunolabeling intensified with increasing time in both experiments, and with increasing dose,
with the highest immunolabeling at the highest concentration (100 (ig/L). and at the longest time of
exposure (14 days). These results suggest a delay in the toxicity of cylindrospermopsin.
5.3 Metabolism
Metabolism and toxicity of cylindrospermopsin appear to be related to the hepatic CYP450 enzyme
system. In a study done by Froscio et al., 2003, hepatocytes were pretreated with known inhibitors of
CYP450 (50 (.iM proadifen or ketoconazole). A reduction in the in vitro cytotoxicity of
cylindrospermopsin was observed. Norris et al., (2002), demonstrated that in male Quackenbush mice
pretreated with the CYP450 inhibitor, piperonyl butoxide, protection against the acute lethality of
cylindrospermopsin occurred. Shaw et al (2000, 2001) also noted the involvement of the CYP450s and
demonstrated that cylindrospermopsin targets periacinar region of the liver, an area where xenobiotic
metabolism mediated by CYP450 occurs.
In a series of studies done by Norris et al. (2001), the distribution and metabolism of
14C-cylindrospermopsin (>95% pure) also was tested. A single i.p. dose of 0.1 mg/kg was administered to
4 male Quackenbush mice, and 0.2 mg/kg was given to 12 mice (Norris et al., 2001). After 12 hours of
dosing, body weights were taken and urine and fecal samples were collected. The group of mice receiving
the lower dose (0.1 mg/kg) were sacrificed after 48 hours of dosing and samples of urine, feces, plus liver
and kidney tissues were treated with methanol to precipitate proteins. The protein precipitates were not
fractionated to identify 14C radiolabel. HPLC was used to fraction the 14C in the methanol supernatant and
to detect metabolites in urine and feces. The HPLC of the urine reveal one major, one moderate and one
minor peak. The minor peak was not present in all samples. It eluted early appearing to be more
hydrophilic than cylindrospermopsin. The major peak appeared to be cylindrospermopsin.
Approximately 23.5% of the urinary 14C was detected in the protein precipitates, indicating the presence
of a protein-bound metabolite (Norris et al., 2001). Results did not indicate whether the levels of proteins
found in the urine were normal or increased. An aqueous extract of the fecal matter from one mouse
indicated that a compound that elutes at the retention time for cylindrospermopsin accounted for 93% of
the administered radiolabel cylindrospermopsin.
Liver tissue analysis of both the protein precipitate and the aqueous supernatant showed the presence of
14C (Norris et al., 2001). After 14C was fractioned by HPLC, the liver supernatant showed the same
elution characteristics as the urine methanol supernatant, indicating the presence of cylindrospermopsin
plus what appeared to be the minor metabolite from urine based on elution time. There were differences
across the samples evaluated, with two animals showing high levels of the minor metabolite. All four
animals had cylindrospermopsin present in the supernatant; for three of the animals the
cylindrospermopsin accounted for less than 50% of the radiolabel present. In the case of the kidney
supernatant, cylindrospermopsin accounted for about 90% of the radiolabel in the supernatant for two
mice evaluated.
Evidence from Runnegar et al. (1995) and Shaw et al. (2000) studies suggests the extractable 14C might be
a cylindrospermopsin metabolite. Runnegar and Shaw also provided evidence of the need for the
activation of cylindrospermopsin for toxicity to occur, suggesting the presence of one or more
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metabolites. Although, no identification of metabolites was performed, results indicate the metabolite is
either more polar than cylindrospermopsin, or that cylindrospermopsin is fragmented during metabolism.
5.4 Excretion
The excretion of cylindrospermopsin following oral, inhalation or dermal exposure has not been reported.
Norris et al. (2001) reported the excretion of 14C-cylindrospermopsin (>95% pure) after the i.p.
administration of sublethal and lethal doses in male Quackenbush mice.
In the first study, 0.1 mg/kg was administered by i.p. to four mice and urine and feces samples were
collected at 12 hour intervals for 48 hours (Norris et al., 2001). After 12 hours, the mean cumulative
excretion of 14C in the urine was 62.8 ± 25.3% (of the 0.1 mg/kg dose), and 15.5 ± 26.9% in the feces.
One of the animals excreted a total of 15.5% of 14C content in the feces (nearly 60% of the dose in this
one mouse compared to less than 5% in the other mice), indicating the possibility that this single high
value occurred because of injection into the upper gastrointestinal tract. However, the authors discounted
this possibility due to the injection technique used. After 24 hours, little additional excretion of 14C in
either the urine or feces was observed. In each of the four mice, the total mean recovery of the 14C in the
urine, feces, liver, kidney and spleen was 85-90%.
In the second part of the study, Norris et al. (2001) administered by i.p. 0.2 mg/kg of 14C-
cylindrospermopsin to 12 mice and collected the urine and feces after 12 and 24 hours. In this study,
continued 14C excretion in urine and feces was observed over 24 hours. After 12 hours, the mean
cumulative excretion of 14C in the urine was 66.0 ±27.1% and in the feces was 5.7 ± 5.6% of the dose.
After 24 hours, 68.4 ± 26.7% was detected in the urine and 8.5 ± 8.1% in the feces, with a mean total
recovery of 76.9% of the administered dose. There was no clear relationship between the signs of toxicity
and the excretion patterns among the mice with signs of toxicity or those with no signs of toxicity.
However, there was a trend in survivors towards increased urinary and decreased fecal excretion and liver
retention.
In the third study, Norris et al. (2001) administered a 0.2 mg/kg i.p. dose of 14C-cylindrospermopsin and
collected the urine and feces after 6 hours. The mean cumulative excretion in the urine was 48.2 ± 29.3%
and in the feces was 11.9 ± 21.4% of the administered dose. The authors reported that 40% of the 14C
dose was excreted in the feces of one of the four mice.
5.5 Pharmacokinetic Considerations
No data on half-life or other quantitative pharmacokinetic data applicable to cylindrospermopsin were
identified. Gastrointestinal uptake of cylindrospermopsin is assumed based on the adverse effects
observed in mice following dosing with both extract and pure cylindrospermopsin. Studies using i.p.
administration of labeled compound demonstrate distribution to the liver, kidney, lung, spleen and heart
in descending order. Some of the label in the liver is bound to protein. There is evidence for hepatic
oxidation by the CYP450 system generating oxidized metabolites that are more toxic than the parent
compound. Pretreatment with CYP450 inhibitors decreased manifestations of toxicity. The presence of
labeled cylindrospermopsin in urine demonstrates the kidney is the principal excretory organ for absorbed
cylindrospermopsin. In mice, a portion of labeled cylindrospermopsin in urine was bound to protein.
Detection of the labeled compound in the feces after i.p. dosing likely reflects some biliary excretion.
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6.0 HAZARD IDENTIFICATION
6.1 Case Reports and Epidemiology Studies
Oral Exposure—In 1979, 148 residents of aboriginal descent in Palm Island in Queensland, Australia
were affected by a hepatoenteritis-like illness (Byth, 1980 and Griffiths and Saker, 2003). Although the
total number of people exposed was not determined, 148 cases were reported. Of those, 138 cases were
children between the ages of 2-16 years (41% boys and 59% girls), and 10 were adults (no sex or age was
reported). Most of the cases required hospitalization and presented symptoms of vomiting, headache,
fever and profuse, bloody diarrhea. Hepatomegaly and renal damage (represented by the presence of
substances in urine such as proteinuria (89%), glycosuria (74%), ketonuria (53%), hematuria (20%), and
urobilinogenuria (8%), were observed. Many (69%) of the patients received intravenous therapy for fluids
and electrolyte imbalance; 12% received intravenous plasma proteins for hypovolemia (decreased volume
of circulating blood) and acidosis. The prevalence of illness in children compared to adults may be due to
the fact that children ingest larger amounts of tainted water compared to adults. Eighty two percent of the
children developed hypokalemia (deficiency of potassium in the blood) and acidosis (Byth, 1980).
Solomon Dam reservoir, the major drinking water supply for Palm Island, was treated a few days prior to
the outbreak with unreported levels of copper sulfate to control a dense algal bloom in the reservoir
(Griffiths and Saker, 2003). Only people in those households connected to the reservoir were affected by
the outbreak. C. raciborskii was identified by retrospective analyses, including epidemiological and
ecological assessments, as the predominant cyanobacterial species in the reservoir and the likely source of
the illness (Griffiths and Saker, 2003; Hawkins et al., 1985). Ohtani et al. (1992) later identified
cylindrospermopsin as the toxin in the reservoir. Some of the reported symptoms (headache, nausea,
vomiting and diarrhea) are effects that are associated with acute oral exposure to concentrations of copper
as sulfate at doses > 3mg/L; with a no effect level (NOAEL) of 1 mg/L (Pizzaro et al., 1999). Although
the copper sulfate treatment could have accounted for reports of nausea, vomiting, headache and diarrhea,
the cyanotoxins in the drinking water are the most likely cause of the observed adverse health effects in
the ill people, assuming the copper sulfate was applied at the recommended 1 mg/L level. Had excess
copper sulfate been added to the water or if concentrations were not uniformly distributed in the water
body, copper could have contributed to the symptoms observed. No other case reports or epidemiological
studies were identified for oral exposure to cylindrospermopsin.
Dermal Exposure—Skin-patch testing in humans was done by Pilotto et al., (2004) to test the potential of
cylindrospermopsin to irritate the skin. Laboratory-grown C. raciborskii cells, both whole and lysed, were
applied using adhesive patches at concentrations ranging from <5,000 to 200,000 cells/mL, to the skin of
50 adult volunteers. The cell concentrations (densities) used were similar to those that could be found in
C. raciborskii-contaminated water bodies used for recreational activities. The patch itself and the culture
media were used as the negative controls, and 1 and 5% solutions of sodium lauryl sulfate were used as
the positive control. After 24 hours, patches were removed and evaluation of the erythematous reactions
were graded (by a dermatologist who was not provided identifying information on the patch concentration
used) using a scale of from 0 to 4: 0 = no reaction or erythema; 1= minimal or very weak spotty erythema;
2= mild diffuse erythema; 3= moderate diffuse erythema; and 4= severe diffuse erythema with edema.
Logistic regression modeling and odds ratios (OR) evaluation was used to determine the distribution of
clinical responses relative to patch concentration.
Analysis of volunteer reactions to patches treated with whole cells showed an OR of 2.13 and a 95%
Confidence Interval (CI) of 1.79-4.21 (p<0.001). Lysed cells patch analysis showed an OR of 3.41 and a
95% CI of 2.00-5.84 (p<0.001). No statistically significant increase or dose-response between skin
reactions and increasing cell concentrations for either patches (whole or lysed) was observed. Subjects
had skin reactions to the cylindrospermopsin and positive control patches more frequently than to the
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negative control patches. The mean percentage of subjects with a reaction was 20% (95% CI 15-31%).
For subjects reacting to negative controls (39), the mean percentage was 11% (95% CI 6-18%).
Evaluation of erythematous reactions showed that mild irritations (grade 2) were resolved in all cases
within 24 to 72 hours. The difference in reaction rates between the whole and lysed cells was minimal and
no evidence for a threshold effect (i.e., a particular concentration above which there were frequent or
strong reactions) was observed.
Stewart et al. (2006) also conducted skin patch testing on 19 human volunteers using lyophilized
C. raciborskii. Up to 160 ng of cyanotoxin was applied to filter paper discs adhered to the back of each
volunteer; patches were removed after 48 hours and the exposed skin was scored after 48 and 96 hours.
No individual developed a clinically detectable skin reaction.
Other Routes of Exposures—In February 1996, there was an outbreak of acute liver failure in
hemodialysis patients at a clinic in Caruaru, Brazil (Carmichael et al., 2001). One hundred and sixteen of
131 patients who received their routine hemodialysis treatment at that time, experienced headache, eye
pain, blurred vision, nausea and vomiting. Of the affected patients, 100 developed acute liver failure and
76 of these patients died. Analysis of the carbon, sand and cation/anion exchange resins from in-house
water treatment filters from the clinic demonstrated the presence of both microcystins and
cylindrospermopsin. Microcystins, but not cylindrospermopsin, were found in blood, sera and liver
samples from the patients. Analysis of liver samples for cylindrospermopsin by HPLC-MS/MS did not
reveal the toxin. However, the method used to detect the more polar alkaloid cylindrospermopsin may
have been inadequate. Based on comparisons between liver pathology data from animal studies of
microcystins and cylindrospermopsin and the symptoms observed in the outbreak, intravenous exposure
to microcystins, and possibly cylindrospermopsin was most likely the cause of death of the dialysis
patients.
6.2 Animal Studies
6.2.1. Acute Toxicity
Oral Exposure—Acute toxicity to cylindrospermopsin-equivalent of freeze-dried C. raciborskii cells
(strains PHAWT/M or PHAWT/1) was tested in male MF1 mice by gavage (Seawright et al., 1999).
Twelve mice were administered a single dose of 4.4, 5.3, 5.7 (to only two mice), 5.8, 6.2, 6.5, 6.7, 6.8,
6.9, 8.0 and 8.3 mg/kg by gavage and observed after 8 days. Of the 12 mice, 8 died two to six days after
treatment. The lowest lethal dose was 4.4 mg/kg, and the highest non-lethal dose was 6.9 mg/kg. An
average lethal dose was approximately 6 mg/kg. Histological examinations showed fatty liver effects with
periacinar coagulative necrosis, acute renal tubular necrosis and atrophy of the lymphoid tissue of the
spleen and thymus. Subepicardial and myocardial hemorrhages in the heart and ulceration of the
esophageal section of the gastric mucosa also were observed. The authors reported thrombohemorrhagic
lesions in one or both eye orbits in some of the animals.
Falconer et al. (1999) administered a single gavage dose of 1,400 mg extract/kg of a cell-free extract of
freeze-dried and sonicated C. raciborskii Woloszynska (AWT 205) cells to an unreported number of male
Swiss mice. Although not specified in this experiment, concurrent i.p. experiments stated the content of
cylindrospermopsin in the extract ranged from 1.3 to 5.4 mg/g extract. This indicates that the
cylindrospermopsin-equivalent gavage dose likely ranged from 1.8 to 7.6 mg/kg. Although not fatal, the
authors observed severe liver and kidney pathology at this dose. No other information on the design and
results of the oral study were provided (Falconer et al., 1999). In a subsequent gavage study Falconer and
Humpage (2001) reported that 2,500 mg extract/kg was the minimum oral lethal dose of freeze-dried
C. raciborskii cells (strain AWT 205) in male Swiss albino mice.
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Shaw et al. (2001) administered a single gavage dose of 0, 1, 2, 4, 6 or 8 mg cylindrospermopsin/kg of
cell-free extract of freeze-dried and sonicated C. raciborskii cells (strain AWT 205) in water to groups of
four Quackenbush mice. After 7 days, all animals were evaluated for gross pathological and histological
(liver, kidney, spleen, heart, lungs and thymus) changes. Different hepatic effects were observed at
different doses as follows:
• 1 and 2 mg/kg showed foamy hepatocellular cytoplasmic changes;
• 4 mg/kg resulted in lipid infiltration with some hepatocyte necrosis in the periacinar region;
• 6 mg/kg resulted in uniformly pale and mottled livers with lipid infiltration throughout and cell
necrosis mainly in the periacinar region;
• 6 mg/kg caused the death of two of four mice within 5 days; and
• 8 mg/kg caused the death of all of the mice within 24 to 48 hours.
In the second part of a genotoxicity study (described in Section 6.4.1), Bazin et al., (2012), administered
1, 2, and 4 mg/kg of cylindrospermopsin (98% purity) by gavage to mice (three per dose). Clinical signs
and tissue sample evaluations were done 24 hours after treatment as well as histological examination. One
mouse in the 2 mg/kg dose group died, and one of the three mice treated with the highest dose (4 mg/kg)
was moribund. Histological evaluation found a dark red liver and intestinal hemorrhage. Another mouse
manifested intestinal bleeding and liquid stools at the same dose. Apoptosis was observed in the liver and
the kidneys at 2 and 4 mg/kg, involving up to 5% of hepatocytes within some sections and apoptosis of
lymphocytes within the Peyer's patches in some mice within these two dose groups. Authors concluded
that the liver and kidneys are target organs, but the kidneys appeared to be the most sensitive organ
following gavage of cylindrospermopsin (Bazin et al., 2012)
Other Routes of Exposure—Acute i.p. lethality values have been determined for cylindrospermopsin
purified from extracts of cultured C. raciborskii or if. natans cells (Ohtani et al., 1992; Shaw et al., 2000,
2001; Terao et al., 1994). In male CH3 mice, 24-hour and 5-to 6-day LD50 values of 2.1 and 0.2 mg/kg
body weight (bw), respectively, were reported for a single i.p. dose of purified cylindrospermopsin
(percent purity not reported) (Ohtani et al., 1992). Another study found that a single 0.2 mg/kg i.p. dose
of purified cylindrospermopsin (percent purity not reported) caused 50% moribundity after 31 hours of
exposure in Quackenbush mice (Shaw et al., 2001). The main pathological findings in the moribund
animals were lipid infiltration and cell necrosis of the liver.
The results of acute i.p. studies of extracts of freeze-dried and sonicated C. raciborskii cells are generally
similar to those of the i.p. studies of purified cylindrospermopsin. A single 0.2 mg/kg
cylindrospermopsin-equivalent dose caused 50% moribundity in Quackenbush mice after 98 hours (Shaw
et al., 2000, 2001). Other single-dose LD50 values, expressed as cylindrospermopsin-equivalent doses
included 24-hour and 7-day values of 0.29 and 0.18 mg/kg, respectively, in male Swiss mice (Hawkins et
al., 1997). A 24-hour LD50 from exposure to extract containing cylindrospermopsin was lower than the
24-hour i.p. LD50 of 2.1 mg/kg for purified cylindrospermopsin in mice, leading the authors to suggest
that the extract contained more than one toxin (Ohtani et al., 1992). Although the liver was the main
target organ in the extract studies, lesions also occurred in the kidney, adrenal gland, lung and intestine
(Hawkins et al., 1985, 1997; Shaw et al., 2000, 2001).
A single dose i.p. LD50 value of 64 mg freeze-dried culture/kg was determined in mice observed for
24 hours (Hawkins et al., 1985). The principal tissue injury was severe centrilobular hepatic necrosis.
Evidence for histological damage also was observed in the kidney, adrenal glands, lungs and intestines.
Falconer et al. (1999) assessed the acute lethality and liver and kidney effects of four different
preparations of cell-free extracts of sonicated freeze-dried C. raciborskii cells in male Swiss albino mice
treated by single i.p. injection. Reported 24-hour and 7-day LD50 values for the four preparations were 50
to 110 and 20 to 65 mg extract/kg, respectively. The cylindrospermopsin content in the four preparations
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varied from 1.3 to 5.4 mg/g extract, indicating that the cylindrospermopsin-equivalent LD50 values were
0.07-0.6 mg/kg (24-hour) and 0.03-0.4 mg/kg (7-day). Centrilobular liver damage was characterized by
cellular vacuolation, intercellular spaces and dark nuclear and cytoplasmic staining. In the kidney, there
was a reduction in the number of erythrocytes in the glomerulus, an increase in the space around the
glomerulus, proximal tubule epithelial necrosis and the presence of proteinaceous material in the distal
tubules. Transmission electron microscopy suggested that the material in the distal tubules was cell debris
from necrosis. The nature, location and time course of the histological damage were similar for oral and
1.p. administration, with maximum damage observed from 2 to 3 days after treatment. There was no clear
correlation between cylindrospermopsin preparation concentration and the LD50 values or the severity of
liver or kidney lesions, leading the study authors to conclude that more than one toxin was present in the
extract.
Terao et al. (1994) examined toxicity in 24 male ICR mice administered a single 0.2 mg/kg i.p. dose of
purified cylindrospermopsin (percent purity not reported). Ultrastructural examination of the organs was
conducted using electron microcopy following sacrifice of 3 mice at each of 8 time points (16 hr. to
100 hr.) after exposure. The liver cells were isolated and found to be the main toxicity target. Ribosomes
were detached from the endoplasmic reticulum and there was an increase in the smooth endoplasmic
reticulum plus Golgi apparatus. Nucleoli became dense and reduced in size. Severe necrosis was present
in the centrilobular region. There was a dramatic increase in intracellular fat vacuoles impacting the
orientation of the microorganelles. After 100 hours, the lobular hepatocytes were destroyed. Histological
changes in the kidney included proliferation of the endoplasmic reticulum and fat droplet accumulation in
cells along the brush borders of the tubules plus limited single cell necrosis. The thymus also was
impacted as indicated by massive necrosis of lymphocytes in the cortex. Occasional single cell necrosis
was identified in the heart.
Inhalation Exposure—Oliveira et al. (2012) evaluated the effects in the lung of cylindrospermopsin after
intratracheal instillation of a lethal dose in BALB/c mice. Semi-purified extract of cylindrospermopsin
was instilled at 70 (ig/kg-bw into 52 mice. Control group (12 mice) received a single intratracheal
instillation of 50 |iL of saline solution. Animals were analyzed 2, 8, 24, and 96 hours after instillation for
the presence of cylindrospermopsin in the lungs and liver. Pulmonary mechanics to measure airflow, lung
volume, lung resistive, static elastance, viscoelastic/inhomogeneous pressures and viscoelastic component
of elastance were performed. The fraction of collapsed and normal alveoli areas were determined
microscopically and expressed relative to the total area examined. Polymorpho- (PMN) and mono-nuclear
(MN) cells, and pulmonary tissue were determined by histological analysis. Biochemical analyses
determined the total protein content, inflammation changes (myeloperoxidase activity), oxidative stress
analyses, and to determine the presence of cylindrospermopsin in the liver and lungs (Oliveira et al.,
2012).
No deaths occurred during the experiment by Oliveira et al. (2012). After 24 hours, the authors detected a
higher concentration of cylindrospermopsin in the lung, and after 96 hours, concentration in the liver
increased significantly (p<0.05). Histological evaluation revealed that after 24 hours, static elastance,
PMN influx into lung parenchyma and myeloperoxidase activity increased. Alveolar collapse (indicative
of a partial collapse of lung tissues) was apparent at 8 h and increased after 24 hours. However, the
authors did not observe intra-alveolar hemorrhage. According to the authors, the increase in alveolar
collapse may have occurred as a result of the production of the reactive oxygen species (ROS) from
cylindrospermopsin and/or its metabolites, or by activated defense cells involved with the inflammatory
process (Oliveira et al., 2012).
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6.2.2. Short Term Studies
Oral Exposure—Four Quackenbush mice were administered either cell-free extract of C. raciborskii
(strain AWT 205) or purified toxin (Shaw et al., 2001). Doses for the extract ranged from 0 to
0.3 mg/kg/day. The no-observed-adverse-effect-level (NOAEL) for cell-free extract was <0.005
mg/kg/day for lymphophagocytosis in the spleen following 14-day gavage administration. When purified
cylindrospermopsin was given at doses of 0 to 0.3 mg/kg/day for 14 days, the low dose (0.05 mg/kg/day)
was the lowest-observed-adverse-effect-level (LOAEL) for fatty infiltration of the liver;
lymphophagocytosis did not occur at any dose (Shaw et al., 2001). One animal had a retro-orbital
hematoma of one eye. The observation that lymphophagocytosis was present only in the animals that
received the extract suggests the presence of an additional toxin in the extract. The authors suggested the
impurity might be a lipopolysaccharide.
Shaw et al. (2001) administered drinking water containing 800 (ig/L cylindrospermopsin to six
Quackenbush mice and two Wistar rats for 21 days. The drinking water was "sourced" from a dammed
impoundment containing cylindrospermopsin. Based on water consumption, the reported approximate
daily dose for both species was 0.2 mg cylindrospermopsin/kg/day. No effects were observed in gross
pathological and histological examinations of the liver, kidney, spleen, heart, lungs and thymus indicating
a NOAEL of 0.2 mg/kg/day in rats and mice. No additional information on the experimental design and
results was reported by the authors (Shaw et al., 2001).
Significant increases in hematocrit, acanthocytes (abnormal form of a red-blood cell that has a spiked,
thorn-like cell membrane), and liver and testes weights were observed in a study of purified
cylindrospermopsin from Aphanizomenon ovalisporum (isolated from Lake Kinneret during a 1994
bloom) by Reisner et al. (2004). Groups of eight, 4 week-old ICR mice were exposed to drinking water
containing 0.6 mg/L cylindrospermopsin for 3 weeks. The dose was estimated by the authors as
66 (ig/kg/day. The study was designed in order to test a hypothesis introduced by Banker et al. (2000) that
the toxicity of cylindrospermopsin could be a reflection of its inhibition of one of the enzymes involved in
synthesis of uridine monophosphate (UMP).
Blood was collected once per week for determination of hematocrit, red blood cells (RBC) counts and
plasma cholesterol (Reisner et al., 2004). At the end of the exposure period, the animals were sacrificed;
the liver, kidney, and spleen were removed, and weighed. The liver was then homogenized. A sample of
the homogenate was analyzed for total cholesterol and the crude protein extract was frozen for later
analysis of uracil monophosphate.
Body weight increased across the duration of the study for both the controls and treated animals and did
not differ significantly between groups at 21 days (Reisner et al., 2004). Significant (p<0.05) increases in
relative liver and testes weights were noted when compared to controls; relative kidney weight also
increased, but was not statistically significant. At the end of three weeks, urinary orotic acid (a pyrimidine
precursor) concentration and hematocrit were significantly increased (p<0.05) in the treated animals
compared to the controls. There was a decrease in the urine excretion rate for both the controls and treated
animals over the three week exposure period with the decrease in the treated animals being significantly
greater (p < 0.05) in those exposed than that for controls at the end of the three week period. Acanthocyte-
like RBCs were observed in Numansky light micrographs of the blood samples collected from the treated
animals at the end of each exposure week. The cholesterol content of the RBC membranes and plasma
were significantly (p<0.05) greater than the levels in controls after the three week exposure and the liver
levels were significantly lower than controls (Reisner et al., 2004).
The authors attributed the acanthocyte (abnormal RBC) formation to the increase in RBC membrane
cholesterol (Reisner et al., 2004). An increase in the ratio of RBC membrane cholesterol to phospholipids
is believed to be a factor responsible for acanthocyte formation. The authors hypothesized that this change
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is the consequence of decreased activity of plasma lecithin-acyl cholesterol transferase (LCAT), an
enzyme associated with high density lipoproteins that regulates the formation of cholesterol esters
(Garrett and Crisham, 1999). Effects on the cholesterol content of the RBC membrane can occur with
inhibition of the enzyme increasing membrane fluidity and mean corpuscular volume. Removal of the
abnormal blood cells by the spleen increases both spleen weight and serum bilirubin stimulating
hematopoiesis. Additional research is needed to examine the LCAT enzyme inhibition hypothesis to
confirm whether it accounts for the effects on the RBCs following cylindrospermopsin exposure. The
authors proposed that there is a relationship between the cylindrospermopsin-induced liver and/or kidney
damage and the decreased LCAT activity.
As stated above, the original goal of the Reisner et al. (2004) study was to investigate the role of the
uracyl moiety of cylindrospermopsin as an inhibitor of uridine synthesis. Although the study revealed that
the toxin was a noncompetitive inhibitor of the UMP synthase complex, there were minimal in vitro
consequences of inhibition at the cylindrospermopsin dose evaluated (66 (ig/kg/day) or in vivo evidence
of orotic aciduria, the expected consequence from UMP synthase inhibition.
Other Routes of Exposure—In addition to the oral exposure studies discussed above, Shaw et al. (2001)
also studied the effects of i.p. exposures. Four Quackenbush mice were dosed by i.p. with either cell free
extract of C. raciborskii (strain AWT 205) or purified toxin for 14 days. The doses for the extract ranged
from 0 to 0.05 mg/kg/day. The LOAEL was <0.005 mg/kg/day for slight foamy cytoplasmic changes in
the liver and for lymphophagocytosis in the spleen. No NOAEL was identified. When doses of 0 to
0.025 mg/kg/day of purified cylindrospermopsin were given for 14 days, the low dose (0.005 mg/kg/day)
was a LOAEL for foamy hepatocellular cytoplasm, but lymphophagocytosis did not occur at any dose
(Shaw et al., 2001).
6.2.3. Subchronic Studies
Oral Exposure—Doses of 0, 30, 60, 120 or 240 (ig/kg/day purified cylindrospermopsin in water was
administered by gavage to groups of male Swiss albino mice (10 mice per dose for all but the highest
dose group which included 6 mice) for 11 weeks (Humpage and Falconer, 2002, 2003). The
cylindrospermopsin was from an extract of freeze-dried C. raciborskii cells (strain AWT 205) purified
using Sephadex size-exclusion gel (G-10). The individual sephadex fractions were assayed using HPLC
and concentrated to a sample that was 47% cylindrospermopsin by dry weight and 53% phenylalanine.
Food and water consumption and body weight were examined throughout the study. After 9 weeks of
exposure, a clinical examination consisting of physiological and behavioral signs of toxicity was
conducted, the study authors did not report specific tests. Hematology evaluations (4 to 5 per dose group,
except the high dose) was done. Serum chemistry (4 to 6 per dose group), and urinalysis (6 or 10 per dose
group) were also conducted. All the evaluations were conducted either near or at the end of the treatment
period.
Postmortem examinations were done on the following organ weights: liver, spleen, kidneys, adrenal
glands, heart, testis, epididymis and brain. Comprehensive histological evaluations were conducted in
accordance with the recommendations from the Organization for Economic Cooperation and
Development (OECD).
No deaths or clinical signs of toxicity were reported in mice exposed to purified cylindrospermopsin
under the study conditions. The mean final body weight was 7-15% higher in all dose groups compared to
controls, but not dose-related and only statistically significant at 30 and 60 |ig/kg/day (Humpage and
Falconer, 2003). No significant changes were observed in food consumption. In all dose groups, the water
intake was significantly reduced.
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Relative kidney weight was significantly increased in a dose-related manner at >60 (ig/kg/day (12-23%
greater than controls; see Table 6-1), and only at the highest dose (240 (ig/kg/day) relative liver weight
was significantly increased (13% greater than controls). Relative spleen, adrenal and testes weights were
increased for doses >60 (ig/kg/day, but the differences from control were not statistically significant
(Humpage and Falconer, 2002).
Selected serum chemistry (n= 4-6), hematology (n=4-5) and urinalysis (n=6-10) results are shown in
Table 6-2. The hematology and serum chemistry evaluations showed no dose-related, statistically-
significant changes, although serum albumin, total bilirubin, and cholesterol were increased compared to
controls at all doses (Humpage and Falconer, 2002). The increases in cholesterol were significant for the
30 and 60 (ig/kg/day groups, but not at the higher doses. The serum urea concentration was slightly
decreased at the two highest doses. A non-significant increase in red cell polychromasia (high number of
RBCs), was indicated for all doses, but quantitative data were not presented. Packed red cell volume was
slightly increased and mean corpuscular hemoglobin was slightly decreased (Table 6-2) when compared
to controls, although the changes were not dose related. When combined with the bilirubin results and the
increased relative spleen weight, the hematological data suggest a possibility for minor red blood cell
effects. One of the limitations of the serum chemistry and hematology data, is the small number of
samples evaluated, a factor that impacts the determination of statistical significance (Humpage and
Falconer, 2002).
Table 6-1. Kidney Weight Data from Oral Toxicity Study of Cylindrospermopsin Administered Daily
over Eleven Weeks (Humpage and Falconer, 2002, 2003)
Dose
(pg/kg/day)
Relative Kidney Weight
%
Change
Number
Control
g/100g bw
Exposed
g/100g bw
Significance
30
10
1.48
1.57
+6
Not significant
60
9
1.48
1.66
+12
p <0.001
120
9
1.48
1.82
+23
p <0.001
240
6
1.48
1.78
+20
P <0.001
There was a significant decrease in the urine protein-creatinine ratio (g/mmol creatinine) at 120 and
240 (ig/kg/day compared to that of controls (51% and 37% of controls, respectively; both p<0.001)
(Humpage and Falconer, 2002). Also, a significant decrease in urine specific gravity normalized for
creatinine was seen at 240 (ig/kg/day compared to the control (p<0.001). The renal glomerular filtration
rate (GFR) was decreased compared to controls at all doses, but the differences were not dose-dependent
or statistically significantly different from controls. The renal failure index1 was decreased slightly at
> 120 (ig/kg/day; the differences from control were not statistically significant (Humpage and Falconer,
2002). Tubular retention of the low molecular weight urinary proteins could account for the decreased
urinary protein and possibly the increased kidney weight. Although effects on kidney weight and urine
protein levels were observed in male mice, the biological relevance of the latter effect and whether it
would also occur in in female mice needs further investigation. Mice are known to excrete a group of
highly polymorphic, low-molecular-weight urinary proteins that play important roles in social recognition
1 Renal failure index= (urinary sodium concentration x plasma creatinine concentration) / urinary creatinine
concentration
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and mate assessment (Cheetham et al., 2009). The relevance of the urinary protein findings in mice to
humans is unknown.
Table 6-2. Selected Clinical Chemistry, Hematology, and Urinalysis Findings (Humpage and
Falconer, 2002, 2003)
Endpoint
N
Dose (|jg/kg/day)
0
30
60
120
240
Clinical Chemistry
Urea (mmol/L)
4-6
9.24
9.22
8.55
7.51
7.92
Albumin (g/L)
4-6
23.8
26.6
26.0
26.0
25.8
Cholesterol (mmol/L)
4-6
3.26
4.60**
4.65**
3.68
4.08
Bilirubin (mmol/L)
4-6
2.62
2.72
2.88
3.06
3.07
Hematology
Packed Cell volume (L/L)
4-5
0.38
0.39
0.39
0.39
ND
Mean Corpuscular
Hemoglobin (MCH, pg/L)
4-5
16.8
15.7
16.4
16.4
ND
Urinalysis
Volume (mL)
6-10
9.85
11.18
10.38
11.74
6.74
Creatinine (mmol/L)
6-10
0.57
0.49
0.54
0.51
0.72**
Specific gravity/creatinine
6-10
1.79
2.04
1.91
1.99
1.44*
Protein/creatinine
(g/mmol)
6-10
4.3
3.6
3.3
2 2**
1.6**
Renal Failure Index
(mmol/L)
4-6
4.3
4.3
4.5
3.6
3.6
ND = not determined
Significantly different from control: *p<0.05; **p<0.01.
Although cylindrospermopsin appeared to inhibit protein synthesis in the liver, based on the histological
evidence of ribosomal detachment from the endoplasmic reticulum after i.p. exposure to a 0.2 mg/kg dose
(see previous discussion of Terao et al., 1994), serum albumin and total serum protein were not decreased
in Humpage and Falconer studies (2002, 2003). The most sensitive effects observed by Humpage and
Falconer (2002, 2003) were dose-related decreases in the urinary protein/creatinine ratio at >120
(ig/kg/day and increased relative kidney weight at >60 (ig/kg/day. The noted decrease in urinary protein
excretion could reflect an impact on excretion of mouse urinary proteins given the fact that total serum
protein was not significantly increased compared to controls for all dose groups. Mouse urinary proteins
are synthesized in the liver (Clissold and Bishop, 1982) and transported to the kidney for excretion. If the
cylindrospermopsin were to reduce liver protein synthesis a decrease in total serum protein would be
expected. However, this was not the case suggesting a lack of an effect on synthesis of the urinary
proteins in the liver.
The Humpage and Falconer (2002, 2003) postmortem tissue examinations showed histopathological
damage to the liver based on scores assigned for necrosis, inflammatory foci and bile duct changes at
>120 (ig/kg/day. The percent of animals with liver lesions in the 120 and 240 (ig/kg/day dose groups was
60% and 90% when compared to 10%, 10%, and 20% for the 0, 30 and 60 (ig/kg/day dose groups,
respectively. Severity scores were not given and the liver lesions were not further described. There was
proximal renal tubular damage in kidney sections from two mice in the 240 (ig/kg/day dose group
(Humpage and Falconer, 2002, 2003). A NOAEL and LOAEL of 30 and 60 (ig/kg/day, respectively, was
identified based on the dose related and statistically significant increase in relative kidney weight.
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Humpage and Falconer (2002, 2003) also conducted a study of the crude cylindro-spermopsin extract that
was purified for the studies described above. Exposure occurred over a 10 week period in groups of 10
male Swiss albino mice given doses of 0, 216, 432 or 657 (ig/kg/day cylindrospermopsin in drinking
water. Significantly decreased body weight was observed at the two highest doses. Liver and kidney
weights were increased in a dose-related manner at 216, 432, and 657 g/kg/day (p<0.0001). Serum ALP
was significantly increased (p<0.05) for the 432 (ig/kg/day dose group. There was a dose-related,
significant increase in total serum bilirubin (p<0.05, p<0.001) and a decrease in serum bile acids for the
two low dose groups (p<0.001) data were not presented for the high dose group. The urinary
protein/creatine ratio was significantly decreased (p<0.001) compared to controls for the two high dose
groups; it also was decreased for the low dose but the difference from controls was not statistically
significant. The renal failure index was decreased significantly at 432 |_ig/kg (p<0.01); no data were
presented for the 657 j^ig/kg high dose group. Glomerular filtration was increased (142% of control) for
the 432 mg/kg/day dose group but the difference from controls was not significant. Glomerular filtration
was increased for the 432 mg/kg/day dose group. The lowest dose tested (216 mg/kg/day) was a LOAEL.
In a follow-on study to Reisner et al. (2004), the potential effects of cylindrospermopsin were investigated
in a 42-week mouse oral dose step-up protocol2 (Sukenik et al., 2006). Food and water ad libitum were
administered to four-week old weaned male and female ICR mice (initial body weight 24-28 g). Animals
were divided into two groups of 20 males and 20 females in each group. The mice in the control group
received freshly prepared cyanobacterial growth medium as their drinking water, whereas mice in the
experimental group received spent medium that contained several different concentrations of
cylindrospermopsin obtained from the spent medium on which cultures of Aphanizomenon ovalisporum
had been grown. Cylindrospermopsin concentrations in the spent medium were quantified by HPLC using
a UV detector set at 263 nm to identify the cylindrospermopsin peak. No other toxins were quantified in
the spent medium although it is known that cyanobacterium spent medium contains an array of secondary
metabolites and other compounds (A. Sukenik, personal communication, 2014).
The concentration of cylindrospermopsin in drinking water was increased gradually from 100 to 550 (ig/L
(Sukenik et al., 2006). The daily intake of the toxin by animals in the experimental group was -10 j^ig/kg
for weeks 0-8; -15 j^ig/kg for males and -17 |_ig/kg for females for weeks 8-16; -30 |_ig/kg for males and
-34 (ig/kg for females for weeks 16-24; and -48 j^ig/kg for males and -55 |_ig/kg for females for weeks
24-42 (these data were presented graphically). Body weight was measured weekly. Water consumption
and urine excretion rates were estimated using metabolic cages every two weeks. Blood samples were
obtained every 4 weeks to determine hematocrit. Ten animals were sacrificed after 20 weeks and ten
animals were sacrificed after 42 weeks of treatment. At sacrifice, liver, spleen, kidney and testes were
weighed and examined grossly for pathological symptoms. Cholesterol levels were measured in the liver.
There were no significant changes in body weight while relative kidney weights were significantly
increased (p<0.05) in males and females at 20 weeks and 42 weeks. Relative liver weight was increased
(p<0.05) in males and females only at 42 weeks. Relative testes weights were increased in males at
42 weeks. Absolute organ weight data were not given.
Hematocrit levels were significantly (p <0.05) elevated compared to controls in both male and female
mice from 16 to 32 weeks of exposure to cylindrospermopsin in drinking water, but returned to control
levels by 36 weeks (Sukenik et al., 2006). The observed changes in the hematocrit level were
accompanied by increased numbers of acanthocytes in the blood as observed by light microscopy. At 20
weeks "many" RBCs were present as acanthocytes (abnormal RBCs), and at 42 weeks very "few normal"
2 In a step-up dose approach the same animals are sequentially exposed to a consecutive series of increasing doses.
Each dose is given for a specified period of time. In Sukenik et al. (2006) doses were increased on weeks 8, 16, and
24. At week 20 half the animals were sacrificed for examination.
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cells were present in the collected blood samples. The number of normal versus acanthocyte cells was not
quantified.
As explained above, RBC conversion to acanthocytes appears to be related to increased cholesterol in the
RBC membrane. The authors measured the cholesterol in the RBC membrane, plasma and liver for 8
males and 8 females at 20 and 42 weeks. Cholesterol was significantly increased in the RBC membrane
and decreased in the liver for both males and females at 42 weeks. At 20 weeks, there was a significant
decrease in liver cholesterol in males, but not in females. Plasma cholesterol increased slightly at 42
weeks and the difference was significant only for the females (Sukenik et al., 2006).
Based on changes in hematocrit at 16 weeks, the authors proposed a dose of 20 (ig/kg/day (equivalent to
200 (ig/L) as the maximal daily intake of cylindrospermopsin during the first 16 weeks that resulted in
adverse effects and the proposed lowest-observed adverse-effect level for both male and female mice
(Sukenik et al., 2006).
In a 90-day study, Quackenbush mice were administered a cell-free extract of freeze-dried and sonicated
C. raciborskii cells (strain AWT 205) in drinking water (Shaw et al., 2001). Gross pathological and
histological (liver, kidney, spleen, heart, lungs and thymus) examinations showed no effects at dose levels
as high as 0.15 mg/kg/day. Animals were examined for mortality and other clinical signs of toxicity.
Neither the number of animals per dose group nor other details of both the experimental design and
results were reported.
Neurotoxicity—The published literature does not provide sufficient data to determine if
cylindrospermopsin elicits neurotoxicity. Humpage and Falconer (2002) reported that they examined
brain, spinal cord and peripheral nerve histopathology, but no results are given in their published report.
6.2.4. Developmental/Reproductive Toxicity
Pregnant rats (10 per dose group and control) were exposed by gavage to 0 (control group), 0.03, 0.3 and
3 mg/kg/bw purified cylindrospermopsin (purity not specified) solutions (Sibaldo de Almeida et al.,
2013). The rats were exposed daily from GD 1-20 and water intake, food consumption and organ and
body weight were recorded during the treatment. Histopathological evaluations were conducted to tissue
portions from liver and kidney. The authors used half of the fetuses from each litter to study visceral
malformations (teratogenic action), and the other half to study skeleton malformations. The authors did
not find significant differences (no statistical significance was provided) between the control and treated
rats in body weight gain, water and food intake or in the histopathological analysis of tissue. No visceral
or skeletal malformations in the fetus were observed (Sibaldo de Almeida et al., 2013).
A series of studies was conducted with cylindrospermopsin in pregnant CD-I mice to investigate
developmental toxicity (Rogers et al., 2007) as well as characterize maternal toxicity and recovery post-
partum (Chernoff et al., 2011). Purified cylindrospermopsin (>98%) was supplied by the Australian
Water Quality Centre and administered in distilled water by i.p. injection for all experiments. All controls
were given distilled water. Dosages were calculated based on maternal body weight on gestation day
(GD) 6 and remained constant. The first set of studies included a standard developmental toxicity study
evaluating dose-response effects in dams and fetuses; this was followed by evaluation of post-natal
growth and development after maternal exposure to a single dose level (Rogers et al., 2007). Dams whose
offspring were used for post-natal evaluation were subsequently used to further characterize maternal
effects and recovery from the single dose level during gestation (Chernoff, et al., 2011).
In a standard developmental toxicity study, groups of 20 to 25 pregnant females were administered 0, 8,
16, 32, 64, 96 or 128 |_ig cylindrospermopsin/kg/day on GDs 8-12 (Rogers et al., 2007). Animals were
sacrificed on GD 17 and the uterine contents examined. Increased maternal mortality occurred in mice
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given 32 (ig/kg/day (4/20) or higher (17-19/20). Average time to death ranged from 6.5 days at
32 (ig/kg/day to 4.4 days at 128 (ig/kg/day. A significant (p <0.01 or 0.05) dose-related increase in liver-
to-body-weight ratio was observed in the dams from the 8, 16 and 32 (ig/kg/day dose groups compared
with control (+13, +15, +30%, respectively). Maternal body and absolute liver weights were not given.
Fetal body weight and numbers of live and dead fetuses per litter were not significantly affected by
maternal treatment. No treatment-related external, skeletal or visceral anomalies were observed. The
LOAEL was 8 (ig/kg/day based on increased relative liver weight; the 32 (ig/kg/day was a frank effect
level (FEL) based on maternal mortality.
The post-natal evaluation experiments from Rogers et al. (2007) were conducted using two groups per
exposure period. The dams (23 to 51 mice per group) were dosed i.p. with 50 (ig/kg/day on either GD
8-12 or GD 13-17 and allowed to litter. At birth, litters were examined and pups from control and treated
dams were combined (total of 10/litter) and cross-fostered with control dams until post-natal day 5-6. A
subset of male pups from dams treated on GDs 13-17 was weaned and their growth monitored for
15 months. Dosing on GDs 8-12 resulted in marked maternal toxicity including: vaginal bleeding,
reduced activity, blood in the tail tips, and combined mortality of 49/79 animals. In contrast, dams treated
on GDs 13-17 showed very low incidences of bleeding around the eyes and vaginal bleeding and only
1/71 animals died. A slight decrease in gestation length for dams treated on GDs 13-17 was noted by the
authors and described as unusual. The dosing during the earlier period of gestation (GD 8-12) resulted in
greater manifestations of maternal toxicity that the later dosing period.
Significant (p <0.01 or 0.05) reductions in litter size at birth were observed in both the GD 8-12 and GD
13-17 treated groups (Rogers et al., 2007). No evidence of late fetal or early postnatal deaths was found;
numbers of implantations were not assessed. The litters born to the dams in the GD 8-12 group were
fewer than those for the controls but the differences were not statistically significant. Pup body weight
and survival through post-natal day 6 were not affected by maternal treatment on GDs 8-12. In contrast,
maternal treatment on GDs 13-17 resulted in significantly (p <0.01) decreased survival as well as reduced
pup body weight at birth and on lactation days 1 and 5-6. Necropsy of pups that died revealed blood-filled
intestines. Lower body weight persisted in male pups from dams exposed on GDs 13-17 throughout the
15-month post-weaning interval.
To further characterize effects on the adult animal, 3-5 dams/group, whose litters were evaluated for post-
natal growth (i.p., 50(ig cylindrospermopsin/kg/day; GDs 8-12 [n=42] or GDs 13-17 [n=42]), were
sacrificed the day following the last dose and on post-treatment days 7 and 14 for both exposures, days 28
and 42 for the GDs 8-12 exposure and on days 35 and 49 for the GD 13-17 exposures (Chernoff et al.,
2011). Blood, liver and kidney samples were obtained at each time point for further analyses. Endpoints
measured included maternal weight and clinical signs of toxicity, serum chemistries indicative of hepatic
and/or renal function and general homeostasis, histopathology of liver and kidney tissues, and hepatic
gene expression after the dosing period.
Dosing on GDs 8-12 resulted in maternal toxicity and death as described above (Chernoff et al., 2011).
Maternal body weight gain was reduced (p<0.05) throughout treatment on GDs 8-12 resulting in
significantly lower body weight (p<0.01) at termination one day after the last dose. Treatment on GDs
13-17 caused a reduced weight gain only after the second dose. Mice sacrificed the day after the last dose
from either regimen had decreased albumin and numerous elevated serum enzymes, including alanine
aminotransferase (ALT), aspartate aminotransferase (AST), alpha-1-antitrypsin, sorbitol dehydrogenase
(SDH) and lactate dehydrogenase (only 2-5 mice per assay). Blood urea nitrogen (BUN) and creatinine
(indicators of kidney damage) were also significantly increased the day after exposure ended. All clinical
chemistry endpoints had returned to control levels 7 days after exposure. No significant differences in
relative kidney or liver weights were observed at any time. The day after the last dose, histopathology
revealed hepatocyte necrosis in 7/19 of GD 8-12 treated animals and 4/19 of GD 13-17 treated animals
compared with 1/19 of both control groups. Moderate nephrosis and/or renal inflammation was found in
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5/19 animals treated on GDs 8-12, but in none of the other treated and control mice. Microscopic lesions
had resolved by one week post-dosing. Analysis of gene expression in liver tissue showed alterations in
expression of genes involved in ribosomal biogenesis, xenobiotic and lipid metabolism, inflammatory
response, and oxidative stress. The response was similar between both exposure groups, persisted for
2 weeks after treatment ended and returned to normal by 4 weeks (Chernoff et al., 2011).
6.2.5. Chronic Toxicity
No information regarding the chronic toxicity of cylindrospermopsin was located.
6.3. Carcinogenicity
In vivo Studies—Falconer and Humpage (2001) tested the tumor initiating activity of
cylindrospermopsin in male Swiss mice using O-tetradecanoylphorbol 13-acetate (TPA) as the promoter.
Saline extract of freeze-dried C. raciborskii cells (strain AWT 205) of 500 or 1500 mg/kg doses were
given to those treated with cylindrospermopsin, and control mice were administered saline. Three oral
doses separated by a two-week recovery period between each dose were given to each control and treated
group. The number of animals initially assigned to each group was not reported. However, of those that
received oral doses of 1500 mg/kg, 70% died within one week of the second dose. Surviving animals
were not dosed again. The cylindrospermopsin-equivalent doses in the 500 extract/kg group was
2.75 mg/kg, and in the 1500 mg extract/kg group was 8.25 mg/kg, based on the reported
cylindrospermopsin content of 5.5 mg/g extract. Two weeks after the final dose, the saline and 500 mg
extract/kg groups were fed liquid food containing TPA dissolved in DMSO, or food containing DMSO
alone, for 24 hours two times per week for 30 weeks and divided into subgroups of 13 to 18 mice. All of
the surviving mice in the 1500 mg/kg groups were similarly exposed to TPA-containing liquid food only
and were not exposed to food containing DMSO alone.
At the end of the 30-week promotion period, histological examinations of the liver, kidneys, spleen and
grossly abnormal organs were performed on all groups. No neoplastic changes were found in any of the
27 control mice. There were three tumors and two areas of dysplastic foci in 5 cylindrospermopsin-treated
mice. No clear pattern in the neoplastic changes was observed because they occurred in different animals,
target organs and treatment groups (Table 6-3). The results of the study do not indicate that the
cyanobacterial extract was a tumor initiator. However, the study is limited by the number of animals
tested, design of the dosing regimen, and by the 30 week observation period.
Table 6-3. C. raciborskii Tumor Initiating Results (Falconer and Humpage, 2001)
Oral Treatment
(mg extract/kg)
Number of Mice
Histological Findings*
1 x 1500/TPA
14
2 hepatocellular dysplastic foci
1 fibroblastic osteosarcoma
2x1500/TPA
5
No neoplasia observed
3 x 500/DMSO
18
1 hepatocellular carcinoma, 1 lymphoma
3 x 500/TPA
16
No neoplasia observed
Saline/DMSO
14
No neoplasia observed
Saline/TPA
13
No neoplasia observed
* Findings in different animals.
* DMSO (dimethylsulphoxide); TPA (tetradecanoly phorbol acetate)
In vitro Studies—The carcinogenic potential of cylindrospermopsin was assessed in vitro via the cell
transformation assay (CTA) on Syrian hamster embryo (SHE) cells (Marie et al., 2010). This assay is
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recommended by OECD Guidelines (2007) as an alternative to in vivo long term experiments for
carcinogenic potential of chemicals because SHE cells are genetically-stable, normal diploid cells that
are capable of metabolic activation. Purified cylindrospermopsin, supplied by the Australian Water
Quality Centre (>98% purity; Adelaide, Australia) was dissolved in water and applied to SHE cells at
cylindrospermopsin concentrations of 1 x 10~5 to 1 x 10"1 ng/mL for the evaluation of cytotoxicity and
1 x 10"7 to 1 x 10"3 ng/mL for the evaluation of cell transformation for seven days. Relative cloning
efficiency was used as an indicator of cytotoxicity. Transformation frequency was determined
microscopically based on the cell morphology (spindle shaped cells, the nucleoplasm to cytoplasm ratio
and basophilic staining properties). Benzo(a)pyrene was used as the positive control and
dimethylsulfoxide as the negative control.
Cylindrospermopsin exhibited transformation at concentrations lower than those causing cytotoxicity
(Marie et al., 2010). There was no change in cloning efficiency at any concentration. However, cloning
efficiency was significantly decreased at a 1 x 10 2 ng/mL concentration in a range-finding study
conducted prior to the main experiment. Transformation frequency was significantly increased over the
positive control at concentrations from 1 x 10"2 to lxlO"7 ng/mL but not at the 1 or lxlO"1 ng/mL. The lack
of a positive response for the lxlO1 and 1 ng/mL concentrations may reflect the fact that only very few
colonies (3 and 4 colonies/concentration) were transformed at those concentrations compared to the
colonies with the elevated transformation frequencies (34-111 transformed colonies).
6.4. Other Key Data
6.4.1. Mutagenicity and Genotoxicity
Studies investigating the in vivo and in vitro genotoxicity (evaluation of DNA damage) from exposure to
cylindrospermopsin are few in number and are discussed below.
In vivo Studies—Shen et al (2002) injected BALB/c mice i.p. with 0.2 mg/kg cylindrospermopsin. The
animals were sacrificed after 6, 12, 24, 48 and 72 hours. The livers were removed and the DNA examined
for strand breaks using alkaline gel electrophoresis. DNA strand breaks were characterized based on the
median molecular lengths of the fragments. The fragment lengths were significantly shorter than those for
the controls at all time points except 72 hours, when the differences in length were not statistically
significant.
Covalent binding of cylindrospermopsin or a metabolite to DNA was detected in the liver of
Quackenbush mice given a single i.p. injection of a cell-free extract of C. raciborskii (dose levels not
reported). DNA was isolated from the liver using a phenol-chloroform purification technique, hydrolyzed
and labeled with 32P. Individual nucleotides were separated using two-dimensional thin layer
chromatography and adducted nucleotides visualized by autoradiography (Shaw et al., 2000). A single
adduct spot was found in each case. The authors concluded that either cylindrospermopsin or a metabolite
was bound to one of the DNA nucleotides.
Based on structural characteristics (the nucleoside structure and potentially reactive guanidine group) of
cylindrospermopsin, it has been speculated that cylindrospermopsin may exert its toxic effects via
pathways that could include reactions with DNA and/or RNA (see Humpage et al., 2000).
Ames MPF microplate format mutagenicity assay was used to assess the mutagenic potential of
cyanobacterial extracts (with different proportions of cyanobacteria) and pure microcystin-LR, (+)-
anatoxin fumarate, and cylindrospermopsin (Sieroslawska, 2013). Pure toxins (purity not reported) were
tested at concentrations of 0.312, 0.625, 1.25, 2.5, 5 and 10 mg/ml with four strains of S. typhimurium and
three strains of E.coli. Cylindrospermopsin was detected at low concentrations (no statistical significance
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reported) in only 2 of the 10 extracts. In one extract (E6), composed of Aphanizomenon flos-aquae,
P. agardhii, and D. planctonicum, cylindrospermopsin was detected at 0.5 1 (ig/L; in the other extract
(E10) composed of D. flos-aquae, D. planctonicum, Aphanizomenon flos-aquae, P. agardhii, and
M. aeruginosa was detected at 0.89(.ig/L. Of all the tested extracts, four (E3, E6, E8 and E10) were
mutagenic, suggesting the presence of other substances able to induce mutations and maybe synergistic
interactions with cyanotoxins.
In vitro Studies—The genotoxicity of cylindrospermopsin was assessed in vitro with two human cell
lines (HepaRG and Caco-2) that represent known target organs of cylindrospermopsin (Bazin et al.,
2010). The objective of this study was to investigate how changes in phenotype associated with cell
differentiation affect toxic response to cylindrospermopsin exposure. In their differentiated state, HepaRG
cells express metabolic enzymes at levels comparable to those found in cultured primary human
hepatocytes. Therefore, HepaRG are metabolically competent cells derived from a human hepatoma that
represent a suitable model to study the genotoxicity of protoxicants in the human liver. However, as the
major route of human exposure to cylindrospermopsin is likely to be ingestion of contaminated water
(i.e., during recreational activities or from drinking), cylindrospermopsin genotoxicity also was
investigated in a human colon adenocarcinoma cell line, Caco-2. After differentiation, Caco-2 cells
display morphological and biochemical characteristics of human enterocytes. Cylindrospermopsin
genotoxicity was assessed using the cytokinesis-block micronucleus assay to assess various cytotoxic and
genotoxic outcomes in these cells. In addition, the involvement of CYP metabolism in the cytotoxicity
and genotoxicity of cylindrospermopsin was determined by the addition of the CYP3A4 inhibitor
ketoconazole.
Cylindrospermopsin (>98% purity, from the Australian Water Quality Center in Adelaide, Australia) was
dissolved in physiological saline. Caco-2 cells in both differentiated and undifferentiated states and
undifferentiated HepaRG cells were exposed to cylindro-spermopsin at concentrations ranging from
0.5 to 2 (ig/mL while differentiated HepaRG cells were exposed to 0.04 to 0.4 (ig/mL for 24 hours (Bazin
et al., 2010). Exposure to 0.5-1.5 (ig/mL cylindrospermopsin resulted in a significant increase in
micronucleated binucleate cells (MNBNC) by approximately three-fold above controls in both
differentiated and undifferentiated Caco-2 cells. Above this concentration, the MNBNC frequency
reached a plateau. Similarly, in differentiated HepaRG cells, MNBNC increased to a maximum of
1.8-fold over controls at 0.06 (ig/mL and leveled-off above this concentration. No change in MNBNC
frequency was seen in undifferentiated HepaRG cells exposed to cylindrospermopsin. The plateau in the
genotoxicity results likely reflects the increase in cytotoxicity as the exposure concentrations increase.
Addition of ketoconazole reduced both cytotoxicity and genotoxicity suggesting that activation by
CYP450 is necessary for both cytotoxicity and genotoxicity.
Lankoff et al. (2007) examined the carcinogenic potential of cylindrospermopsin in vitro through the
formation of chromosomal aberrations in Chinese hamster ovary (CHO)-Kl cells. Cylindrospermopsin
isolated from two cultures of C. raciborskii in AWT 205 (Australian Water Technology Center) and Thai
(from a fish pond in Thailand), was prepared in solution. CHO-K1 cells were exposed to 0, 0.05, 0.1, 0.2,
0.5, 1 and 2 (ig/mL with and without metabolic activation (S9) for 3, 16 and 21 hours. No significant
influence on the frequency of chromosome aberrations in cells treated with cylindrospermopsin with or
without S9 compared to control groups was found. The study showed that neither cylindrospermopsin nor
the S9 fraction-induced metabolites were clastogenic in CHO-K1 cells. However, significant (p <0.05)
decreases in the frequency of mitotic indices were observed after various exposure durations at
concentrations of 0.1 |_ig/mL and above. Furthermore, significant (p <0.05) increases in the frequency of
apoptotic cells (1 (ig/mL and above) and necrotic cells (0.5 (ig/mL and above) after 21 hours were
observed compared to the controls in a dose and time-dependent manner. The presence of metabolic
activation influences susceptibility to necrotic cell death, but not apoptosis.
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To confirm that cylindrospermopsin metabolism is necessary for the manifestation of genotoxicity and to
characterize CYP450 involvement in activation, the micronucleus assay also was conducted with a
CYP450 inhibitor (Lankoff et al., 2007). CYP3A4 is the major CYP450 form in the human small
intestine, responsible for metabolizing a large number of xenobiotics (Pelkonen et al., 2008). Cells were
treated with ketoconazole, widely known to inhibit CYP3A4. Results indicate ketoconazole protects
undifferentiated Caco-2 cells from the induction of (micronuclei) MN induced by cylindrospermopsin.
This further suggests that a CYP450-mediated metabolite is involved in the genotoxic effect at
noncytotoxic concentrations in the Caco-2 cell model. This finding is in agreement with Humpage et al.
(2005) who demonstrated that omeprazole, a CYP3A4 inhibitor less specific than ketoconazole, was
effective in protecting mouse primary hepatocytes from cylindrospermopsin-induced genotoxicity. These
results are also in accordance with Fessard and Bernard (2003) and Lankoff et al. (2007) who observed
that cylindrospermopsin does not react directly with DNA in metabolically-incompetent CHO K1 cells
(Table 6-2).
Humpage et al. (2000) reported that purified cylindrospermopsin caused an increase in the frequency of
micronuclei in the human lymphoblastoid cell line, WIL2-NS. WIL2-NS cells were exposed to 1-
10 (ig/mL cylindrospermopsin for 24 hours to evaluate micronucleus frequency and cellular ploidy.
Cylindrospermopsin caused a dose-dependent increase in the incidence of MN in binucleated cells
(BNCs) at >3 (ig/mL. There was an 8 fold increase in MN/lOOOBNCs over the control.
Cylindrospermopsin also produced "multimicronucleated" cells indicating chromosomal damage,
although the underlying mechanism was unclear. An increase in centromeres was observed in MNBNCs
suggesting cylindrospermopsin could be a spindle poison causing changes in the centromere/kinetochore
function. Two mechanisms were suggested as the cause of cytogenetic damage: the first one leading to
strand breaks at the DNA level, and the other, at the level of kinetochore/spindle function, which induces
loss of whole chromosomes (Humpage et al., 2000).
Fessard and Bernard (2003) examined the genotoxic potential of cylindrospermopsin in (CHO) K1 using
the comet assay. Doses of 0.5 and 1 (ig/mL of purified cylindrospermopsin caused cell growth inhibition
and altered cell morphology linked to effects on the cytoskeleton. No apoptosis or DNA strand breaks
were observed after 24 h of treatment with cylindrospermopsin. Cell mitosis was decreased at
cylindrospermopsin concentrations between 0.33 and 1 (ig/mL.
Humpage et al. (2005) examined the integrity of hepatocyte DNA using a comet assay following exposure
to concentrations of 0.05 to 0.5 (.iM purified cylindrospermopsin (98% pure). Clofibrate was used as the
positive control. After exposure of cultured cells to the toxin, the cells were lysed and the DNA isolated,
and denatured using an alkaline pH to generate double strand breaks. The treated DNA was stained and
visualized for scoring of the comet tail moment. Cylindrospermopsin produced significant DNA
fragmentation at concentrations as low as 0.05 (.iM. The addition of CYP450 inhibitors (omeprazole and
SKF525A) to the culture medium reduced the number of DNA strand breaks. The ability of
cylindrospermopsin to induce DNA damage in isolated human peripheral blood lymphocytes was
investigated by Zegura et al. (2011). Whole blood samples were treated with cylindrospermopsin
concentrations (0, 0.05, 0.1, and 0.5 |ig/mL) for the comet assay and the cytokinesis-block micronucleus
(CBMN) assay at 4 and 24- hours of exposure. The number of cells containing micronuclei increased
significantly following 0.5 |ig/mL treatment at 4 hours incubation and after a 24-hour incubation at a
concentration of 0.1 (ig/mL. Nuclear buds were observed in binucleated human peripheral blood
lymphocytes at 0.05 and 0.1 |ig/mL after 4 hours and at 0.1 |ig/mL after 24-hour exposures. This was
accompanied by a significant decrease in the nuclear division index after 24 hours of exposure to the
0.1 and 0.5 |ig/mL concentrations. Exposure to cylindrospermopsin was associated with a slight but
significant increase in strand breaks at 24 hours. Increases in in nuclear bridges were not significant
(Zegura et al., 2011).
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The genotoxicity of cylindrospermopsin in a human hepatoma cell line (HepG2) was studied by Straser et
al., (2011) using an alkaline comet assay and CBMN with different cylindrospermopsin concentrations
(0, 0.005, 0.01, 0.05, 0.1, 0.5, 1 and 5 (ig/mL) for 4, 12 and 24-hours incubation. Cell viability was
significantly decreased at concentrations of 1 and 5 |ig/mL. After exposure to 0.5 |ig/mL
cylindrospermopsin for 24 hours, a significant decrease in the nuclear division index was observed in
HepG2 cells (Straser et al., 2011). The frequency of cells with micronuclei and nuclear buds increased
significantly at 0.05 and 0.5 |ig/mL cylindrospermopsin. Nuclear bridges increased at both concentrations,
but were only statistically significant in cells exposed to 0.05 |ig/mL. These results demonstrate the
occurrence of complex genomic changes including gene amplification (nuclear buds) and chromosomal
rearrangements. The results of the in vitro genotoxicity are summarized in Table 6-4 below.
Table 6-4. Genotoxicity of Cylindrospermopsin in vitro
Species
(test system)
End-point
Results
Reference
Human cell lines
(HepaRG and
Caco-2)
DNA damage
Significant increase in MNBNC in both HepaRG
at 0.04-0.06 |jg/ml_ and Caco-2 cells
at 0.5-1.5 |jg/ml_
Bazin et al.,
2010
Human
lymphoblastoid
WIL2-NS cells
DNA damage
Exposure to 3, 6 and 10 |jg/ml_ increased
frequency of MN in WIL2-NS cells
Humpage et
al., 2000
Chinese Hamster
Ovary-K1 cells
DNA damage
Comet assay showed altered cell growth and
morphology but no interaction with DNA at
0.5 and 1.0 |jg/ml_
Fessard and
Bernard, 2003
Chinese Hamster
Ovary-K1 cells
DNA damage
Chromosome aberration not observed in CHO-
K1 cells; apoptotic cells (1 |jg/ml_ and above) and
necrotic cells (0.5 |jg/ml_ and above) observed
Lankoff et al.,
2007
Hepatocytes from
Male Albino Swiss
Mouse
DNA damage
Comet assay showed concentration dependent
increase in comet tail length, area, and moment
in cells at 0.05 |jM - 0.5 |jM
Humpage et
al., 2005
Human hepatoma
cell lines
(HepG2)
DNA damage
Significant increases in micronuclei and nuclear
buds at 0.05 and 0.5 |jg/ml_ (statistically
significant) and a decrease in nuclear division at
0.5 |jg/ml_ after 24 hours
Straser et al.,
2011
Human peripheral
blood lymphocytes
DNA damage
Increases in micronuclei at 0.5 |jg/ml_ at 4 hours
incubation and after a 24-hour incubation at 0.05
|jg/ml_. Nuclear buds were observed at 0.05 and
0.1 |jg/ml_ after 4 hours and at 0.1 |jg/ml_ after
24-hours and a decrease in nuclear division
index after 24 hours at 0.1 |jg/ml_.
Zegura et al.,
2011
6.4.2. Immunotoxicity
Data on the effects of cylindrospermopsin on immune function were not located. However, in single and
short-term studies of high-level exposures, immune related effects were observed. A single 0.2 mg/kg
dose of cylindrospermopsin purified (percent purity not reported) from cultured U. natans cells was
administered i.p. to male ICR mice. (Terao et al., 1994).
While there was massive necrosis of lymphocytes in the cortical layer of the thymus, large lymphocytes
in the medulla survived.
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A single gavage dose of a suspension of freeze-dried C. raciborskii cells, in the lethal dose range of 4.4 to
8.3 mg/kg, was given to MF1 mice (Seawright et al., 1999). Effects observed included atrophy in the
thymus (degeneration and necrosis of cortical lymphocytes) and at the lymphoid tissue of the spleen
(follicular lymphocyte loss due to lymphophagocytosis). Shaw et al. (2001) administered a nonlethal dose
of 0.05 mg/kg/day of a cell-free extract of freeze-dried and sonicated C. raciborskii cells by gavage to
Quackenbush mice for 14 days. Lymphophagocytosis was observed in the spleen of exposed mice. A
similar effect did not occur with purified cylindrospermopsin from the same source (Shaw et al., 2001).
In a study done by Poniedzialek et al. (2014b), cylindrospermopsin significantly inhibited (p<0.001) cell
proliferation of cultured human T-lymphocytes. After exposing T-cell phytohaemagglutinin-L (PHA-L)
to as much as 1 (.ig/m L of purified cylindrospermopsin (>95%) isolated from C. raciborskii, cell viability
and lymphocyte proliferation were measured after 72 hours of lymphocyte culture. The authors reported
inhibition of human T-cell lymphocytes after 6 hours (91.0%), 24 hour (81.1%) and 48 hours (69.0%) of
the 72 hour culture period (19.0%). At lower concentrations (0, 0.01 and 0.1 (.ig/niL). cylindrospermopsin
did not induce significant differences (p>0.05) in T-lymphocyte proliferation when compared to controls,
regardless of the time the toxin was added to cell culture. At the highest dose (1 (.ig/niL). the authors also
observed a decrease in the viability of human T-lymphocytes in a time-dependent manner. A statistically
significant (p<0.001) decrease in live cells was observed at 6 hours (6.8%), 24 hours (10.7%), 48 hours
(11.5%) and 72 hours (2.8%) along with an increase in necrotic cells. Lower concentrations did not
induce significant changes in lymphocyte viability (p >0.05); however, at 0.1 (ig/mL of
cylindrospermopsin, a statistically significant (p<0.05) increase in necrotic cells was observed at the
beginning (1.0%) and after 6 hours of cell culture (0.9%). The greatest alterations were observed at
1 |_ig/m L after 24 h of culturing.
Poniedzialek et al. (2014a) also studied the effect of cylindrospermopsin on the oxidative burst capacity
of human neutrophils. A decline in the production of ROS in stimulated and unstimulated neutrophils
from healthy donors was observed after 1 hour of exposure to purified (95% purity) cylindrospermopsin.
Generation of ROS is an important step in pathogen inactivation by neutrophils. The concentrations
evaluated were 0 (control), 0.01, 0.1 and 1.0 (ig/mL. The decrease in ROS levels was statistically
significant (p<0.01) for all the concentrations evaluated. Cylindrospermopsin had no effect on the
neutrophils numbers in whole blood based on a stable number of apoptotic or necrotic cells in the exposed
samples. There was no impact on the proportion of phagocytic neutrophils in the blood samples or in their
ability to engulf bacteria at all cylindrospermopsin concentrations. However, exposure to
cylindrospermopsin reduced the ability of the human neutrophils to disable the pathogens because of the
decrease in their ROS production (Poniedzialek et al., 2014a).
6.5. Physiological or Mechanistic Studies
6.5.1. Noncancer Effects
Mechanistic studies have mostly assessed hepatic endpoints because the liver has been historically been
regarded as the primary target of cylindrospermopsin toxicity. Although not clearly understood, the
specific mechanism for liver toxicity may involve more than one mode of action. Terao et al. (1994)
concluded that inhibition of protein synthesis following i.p. administration of 0.2 mg/kg could occur in
various tissues because electron microscopy of liver cells revealed ribosome detachment from the
endoplasmic reticulum. Hepatocyte cytotoxicity, as evidenced by lactic dehydrogenase leakage in
cultured cells, co-occurred with protein synthesis inhibition, but by a mechanism that is independent of
the inhibition of protein synthesis (Froscio et al., 2003).
To examine the inhibition of protein synthesis hypothesis, Terao et al. (1994) isolated liver microsomes
from 6 mice (4 controls and two treated) and measured protein levels colorimetrically (Lowrey method).
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Levels were lower in the treated mice than in the control, but the differences were not statistically
significant (16.6 ± 1.3 mg/g liver for controls vs. 11.0± 1.2 mg/g liver for the treated mice). Using a
rabbit cell-free reticulocyte system as a platform for globulin synthesis, there was a concentration-
dependent decrease in leucine incorporation at concentrations up to 48 ng/mL cylindrospermopsin.
Cylindrospermopsin induced concentration- and time-dependent toxicity and inhibition of protein
synthesis in cultured hepatocytes isolated from male Swiss mice using radiolabeled leucine uptake as a
measure of protein synthesis (Froscio et al., 2003). Diminished leucine incorporation was apparent for
concentrations > 0.5|_iM over a 20-hour period, but not for a 0.1 (.iM concentration. The authors also
looked at cell leakage of lactate dehydrogenase as an indicator of cytotoxicity. A significant increase in
LDH leakage at 18 hours occurred at concentrations > 1 (.iM. The broad-spectrum CYP450 inhibitors
proadifen (SKF525A) and ketoconazole diminished cytotoxicity, but did not diminish the inhibition of
protein synthesis. These findings suggest that the cytotoxic effects of cylindrospermopsin may be more
related to oxidized metabolites than the inhibition of protein synthesis, presumably by the parent
compound. The protein synthesis inhibition was not reversed by removal of the toxin or washing of the
hepatocytes.
Froscio et al. (2008) extended their studies of the impact of cylindrospermopsin on protein synthesis
using extracts from plant (wheat germ) and mammalian tissues (rabbit reticulocytes) as protein synthesis
platforms. The template for protein synthesis was the mRNA for luciferase. They measured luminescence
to determine the amount of luciferase formed. Cylindrospermopsin was able to inhibit protein synthesis
with similar potency for both wheat germ and mammalian tissues. Radio labeled cylindrospermopsin
binding to rabbit reticulocyte ribosomes increased as cylindrospermopsin concentration increased and was
associated with the inhibition of protein synthesis. Unlabeled cylindrospermopsin displaced the
radiolabeled cylindrospermopsin from ribosomes suggesting noncovalent binding. A toxin to ribosome
ratio of 0.02:1 completely inhibited protein synthesis in samples with 300nM labeled toxin (Froscio et al.,
2008).
The authors also examined the extent of radiolabeled cylindrospermopsin binding to detached 80S
ribosomes. They concluded that the ribosome was not the target of the cylindrospermopsin because of the
ease with which it could be detached from the ribosome during elution and size exclusion filtration.
Eluted toxin was associated with a >100kD elution fraction, leading Froscio et al. (2008) to hypothesize
that the cylindrospermopsin was bound to the elongation or initiation factors necessary for protein
synthesis.
Cylindrospermopsin-induced effects on cellular protein synthesis in Vero cells (originally isolated from
the kidney epithelial cells of African Green monkeys) were studied by Froscio et al. (2009). The cells
were cultured to express a green fluorescent protein. Effects of cylindrospermopsin on protein synthesis
were examined in vitro using a rabbit reticulocyte lysate as a cell-free platform and inhibition was
evaluated as the decrease in green protein fluorescence overtime. Fluorescence decreased after 4-hours.
(IC50 = 5.9 (.iM) and 24- hour exposures to the toxin. There was a concentration-related decrease in cell
viability that roughly paralleled the decrease in protein fluorescence. When the medium containing toxin
was replaced with medium free of toxin, fluorescence continued to decline. The decrease was significant
for the cells originally exposed to the 100 and 300 (.iM concentrations but not for the 30 (.iM
concentration. The strong interaction of the toxin with its targets indicated that cylindrospermopsin
remained in the intracellular environment for an extended period.
Cylindrospermopsin-induced depletion of Quackenbush mouse hepatic glutathione was demonstrated in
vivo by Norris et al. (2002) although the study authors did not consider the effect to be of sufficient
magnitude to represent the primary mechanism of cylindrospermopsin toxicity. After the mice were
treated with a dose of 0.2 mg/kg toxin following pretreatment with glutathione (GSH) depleting agents
(butathione and diethylmalate) the 7 day survival rate was 5/13 (38%) compared to 9/14 (64%) for the
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controls yet the difference between GSH levels between the exposed and control animals was small
(quantitative measures not provided). The results after treatment with piperonyl butoxide, a CYP450
inhibitor, were protective with 100% survival in for the exposed and control mice.
Cylindrospermopsin caused decreased glutathione levels, as well as decreased synthesis of glutathione
and protein, in cultured rat hepatocytes (Runnegar et al., 1994, 1995, 2002). In the Runnegar et al. (1994)
study, pretreatment with the CYP450 inhibitor, a-naphthoflavone, partially protected against cytotoxicity
and cellular glutathione depletion, indicating involvement of the CYP450 enzyme system in
cylindrospermopsin metabolism. Inhibition of glutathione synthesis was the predominant mechanism for
the observed reduction in glutathione; other mechanisms, including increased utilization of glutathione,
increased formation of oxidized glutathione, increased glutathione efflux, decreased glutathione precursor
availability and decreased cellular adenosine triphosphate (ATP) were effectively ruled out.
Runnegar et al. (1995) investigated the decrease in cellular GSH and its role in the metabolism and
toxicity of cylindrospermopsin in primary cultures of rat hepatocytes. To ascertain whether the reduction
in GSH was due to decreased GSH synthesis or increased GSH consumption, total GSH was measured
after treatment with 5 mM buthionine sulfoximine (BSO, an irreversible inhibitor of GSH synthesis). The
rate of fall in total GSH (nmol/106cells/hr.) was 8.2 ±2.5, 6.0 ±1.7, and 5.9 ±1.3 for control, 2.5 (iM and
5 (.iM cylindrospermopsin pretreated cells, respectively. This suggests that the toxin-induced decrease in
GSH induced was due to the inhibition of GSH synthesis rather than increased consumption, because, in
the latter case, the rate of decrease in GSH would have been accelerated by toxin pretreatment.
Furthermore, excess GSH precursor (20 mM N-acetylcysteine), which supported GSH synthesis in
control cells, did not prevent the decrease in GSH or toxicity induced by cylindrospermopsin. Addition of
CYP450 inhibitors a-naphthoflavone, SKF525A and cimetidine partially prevented the decrease in cell
GSH induced by cylindrospermopsin. Results suggest that an oxidized and/or glutathione-conjugated
derivative of cylindrospermopsin is formed and could be a more potent inhibitor of GSH synthesis than
the parent cylindrospermopsin.
Humpage et al. (2005) used inhibitors of specific CYP450 isoforms, furafylline (CYP1A2) and
omeprazole (CYP3A4 and CYP2C19) to determine if they would protect against cylindrospermopsin
cytotoxicity in an in vitro mouse hepatocyte system. However, inhibitors of CYP450s 2A6, 2D6 and 2E1
(reduced glutathione (GSSG) reductase (GSSG-rd.) inhibitor l,3-bis(chloroethyl)-l-nitrosourea (BCNU))
were not cytoprotective (Humpage et al., 2005). There was no indication that reductions in glutathione
levels by cylindrospermopsin increased levels of ROS. The authors concluded that CYP450 derived
metabolites were responsible for the cytotoxicity and genotoxicity of cylindrospermopsin and that ROS
were not involved. In another study, the addition of the CYP3A4 inhibitor ketoconazole to cultured
HepaRG cells reduced both cytotoxicity and genotoxicity of cylindrospermopsin (Bazin et al., 2010).
The CYP450 inhibitors omeprazole (100 \\M) and SKF525A (50 \xM) completely inhibited the
genotoxicity induced by CYN. The toxin also inhibits production of glutathione (GSH), a finding
confirmed in this study. This could potentiate cytotoxicity, and by implication genotoxicity, via reduced
reactive oxygen species (ROS) quenching. The lipid peroxidation marker, malondialdehyde (MDA) was
quantified in CYN-treated cells, and the effect of the reduced glutathione (GSSG) reductase (GSSG-rd.)
inhibitor l,3-bis(chloroethyl)-l-nitrosourea (BCNU) on both MDA production and lactate dehydrogenase
(LDH) leakage was examined. MDA levels were not elevated by CYN treatment, and block of GSH
regeneration by BCNU did not affect lipid peroxidation or cytotoxicity. It therefore seems likely that
CYP450-derived metabolites are responsible for both the acute cytotoxicity and genotoxicity induced by
CYN.
Shen et al. (2003) found that cylindrospermopsin induced up-regulation of the tissue transglutaminase
(tTGase) gene in the liver in Balb/c mice following i.p. injection of a single 100 |ig/kg dose. Tissue
TGase catalyzes the post-translational modification of proteins via Ca2+-dependent cross-linking reactions
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as well as hydrolysis of GTP and functions as a protein kinase. It also takes part in cell adhesion processes
and stabilization of the extracellular matrix (Onyekacji and Coussons, 2014). Up-regulation oftTGase can
lead to cell injury and apoptosis. Using semi quantitative, real-time PCR and primer sets for mouse
tTGase with enzyme separation by gel electrophoresis, Shen et al. (2003) found that toxin-exposed cells
had higher levels of the enzyme at 6, 72 and 96 hours than unexposed control cells.
Zegura et al. (2011) measured gene expression in human peripheral blood lymphocytes after incubation
with 0.5 (ig/mL cylindrospermopsin for 24 hours. Genes for metabolism (CYP1A1 and CYP1A2), DNA
damage response (P53 and downstream regulated genes), apoptosis (BCL-2 and MAX) and oxidative stress
response (GPX1, SOD1, GSR and GCLC) were up-regulated.
In an in vitro study by Fernandez et al. (2014), cylindrospermopsin was analyzed for its ability to cross
the intestinal epithelium and enter systemic circulation. To determine the effect of cylindrospermopsin on
the Caco-2 monolayer integrity, Caco-2 monolayer, trans-epithelial electric resistance (TEER) was
measured after 3,10 and 24 hours of incubation with 1 mM, 5 mM and 10 mM of the toxin. TEER values
after exposure to cylindrospermopsin did not show any significant difference when compared with
controls (0), indicating that the monolayer was not disrupted or altered by cylindrospermopsin at the
concentrations tested.
To test the ability of cylindrospermopsin to cross cell membranes, Fernandez et al. (2014), incubated
differentiated Caco-2 cells and arat Clone 9 hepatic cell line with 0.1, 0.25, 0.5,1, 1.5, 2.5, 5 or 10 mM
cylindrospermopsin. The molecular mass and hydrophilic nature of cylindrospermopsin make it a poor
candidate for simple diffusion across a cell membrane without the aid of a transport channel. After first
determining that the cylindrospermopsin did not increase cytotoxicity and weaken the cell membrane, the
apical side of the cell monolayer was exposed to concentrations of 1, 5, or 10 (.iM cylindrospermopsin for
3, 10, or 24 hours. Even after 24 hours, relatively small percentages of the applied cylindrospermopsin
had crossed the basolateral membrane (16.7 to 20.5%). Caco-2 cells are often used as a surrogate for the
intestinal membrane. The results from this study indicate that uptake from the intestines after oral
exposure is limited and is in agreement with the differences observed in LD50 values for oral versus i.p.
exposures (Terao et al., 1994; Ohtani et al 1992).
When the Clone 9 hepatic cell monolayers were exposed to varying concentrations of cylindrospermopsin
there was a time and concentration dependent decrease in viability when compared to the controls. Clone
9 cell viability decreased by more than 40% at 24 hours and more than 65% at 48 hours of exposure to
5 mM of cylindrospermopsin. Observations with a phase-contrast microscope determined that after
48 hours of exposure to 5 mM of cylindrospermopsin, Clone 9 monolayer cells showed morphological
signs of cellular damage, detachment from their substrate, and decreased viability. Levels of GSH
increased over time, especially 48 hours after exposure to 1 and 5 mM, probably as a means of
minimizing cell damage caused by the toxin. Based on the GSH increase and increases in the proteins
(3-tumulin and actin, Fernandez et al. (2014) concluded that the toxicity in the Cone 9 hepatocyte cultures
was not related to GSH depletion or impaired protein synthesis.
6.5.2. Cancer Effects
There are no long term bioassay studies that examined the tumorigenicity of cylindrospermopsin. Thus,
mechanistic data for this endpoint are lacking.
6.5.3. Interactions with Other Chemicals
No studies of mixtures of cylindrospermopsin with other specific chemicals (except those identified as
being present in the growth media) were identified. Although, extracts can contain chemicals other than
cylindrospermopsin, in no case were those chemicals identified other than the presence of the essential
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amino acid phenylalanine in the purified cylindrospermopsin isolated by Humpage and Falconer (2002).
The Caruaru outbreak (Section 6.1) involved exposure of patients at a renal dialysis clinic in Caruaru,
Brazil to a mixture of microcystin and cylindrospermopsin. However, the data do not reveal any
quantitative information on the toxicity of the mixture compared to its individual components.
An aquatic invertebrate study using brine shrimp (Artemia salina, Daphnia magna and Daphnia galeata)
to determine the toxicity of microcystin and cylindrospermopsin in combination with cyanobacterial LPS
found that pre-exposure to LPS increased the lethal concentration (LC50) of cylindrospermopsin 8-fold
(Lyndsay et al., 2006). The authors concluded that the decrease in susceptibility to cylindrospermopsin
was due to the effects of LPS on detoxification enzyme pathways; LPS decreased toxic metabolites of
cylindrospermopsin by suppressing the invertebrate cytochrome P450 system, thus decreasing toxicity.
6.5.4. Structure Activity Relationship
There is some evidence that the most toxic form of cylindrospermopsin is an unidentified metabolite
produced by hepatic CYP450. Pretreatment of hepatocytes with known inhibitors of CYP450 diminished
the in vitro cytotoxicity of cylindrospermopsin (Froscio et al., 2003). Similarly, Norris et al. (2002) found
that pretreatment of mice with a CYP450 inhibitor protected against the acute lethality of
cylindrospermopsin in male Quackenbush mice.
According to Banker et al. (2000), the uracil portion of cylindrospermopsin may play an important role in
cylindrospermopsin toxicity. Banker et al. (2000) found that chlorination eliminated the acute lethality of
cylindrospermopsin in mice resulting in the formation of 5-chlorocylindrospermopsin or cleavage of the
pyrimidine ring to form cylindrospermic acid. This was shown by a 5-day i.p. LD50 value of 0.2 mg/kg for
cylindrospermopsin, and 10-day i.p. LD50 values of >10 mg/kg for 5-chlorocylindrospermopsin, and
>10 mg/kg for cylindrospermic acid.
Norris et al. (1999) treated male white Quackenbush mice with a 0.8 mg/kg i.p. dose of
7-deoxycylindrospermopsin, an analogue of cylindrospermopsin isolated and purified from C. raciborskii.
After 5 days of administration, 7-deoxycylindrospermopsin did not appear to be toxic.
Runnegar et al. (2002) conducted an analysis of natural cylindrospermopsin, synthetic (racemic)
cylindrospermopsin and selected synthetically-produced cylindrospermopsin structural analogues to
determine the effects on protein synthesis in the rabbit reticulocyte lysate system and in cultured rat
hepatocytes. Orientation of the hydroxyl group at C7 in the carbon bridge was not important because the
C7 epimer of cylindrospermopsin (Figure 6-1) and its corresponding diol showed similar inhibition of
protein synthesis compared to that of synthetic (racemic) cylindrospermopsin. An analogue with a
hydroxyl functional group in place of the sulfate substituent on C12 had a comparable impact on protein
synthesis. Another analogue (Figure 6-2), lacking the 5-membered heterocyclic ring but retaining cationic
nitrogen and without the methyl and sulfate substituents on the six membered ring (AB-MODEL), was
500 to 1000-fold less effective in the inhibition of in vitro protein synthesis using a rabbit reticulocyte
system.
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Figure 6-1. Structure of cylindrospermopsin and 7-epicylindrospermopsin (de la Cruz et al., 2013)
H OH HO H
© ©
Cylindrospermopsin 7-epicylindrospermopsin
Figure 6-2. Structure of cylindrospermopsin analog AB-MODEL (Runnegar et al., 2002)
H
>Uli
.0
N N N^.NH
V h Y
NH," OH
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7.0 CHARACTERIZATION OF RISK
7.1 Synthesis and Evaluation of Major Noncancer Effects
Information on the human health effects of cylindrospermopsin is limited to the observations from the
Australian Palm Island outbreak involving acute and/or short-term drinking water exposure to
C. raciborskii (Byth, 1980; Griffiths and Saker, 2003). The clinical picture of the illness is we 11-defined
and includes fever, headache, vomiting, bloody diarrhea, hepatomegaly and kidney damage with renal
loss of water, electrolytes and protein, but no data are available on the exposure levels of
cylindrospermopsin that induced these effects. Furthermore, the concentration of copper sulfate used to
treat the lake (a source of drinking water) to control harmful algal blooms before illness was observed is
not known; thus, the presence of elevated copper concentrations in the drinking water could have been a
contributing factor for some of the symptoms observed.
Most of the information in animals on the non-cancer effects of cylindrospermopsin is from oral and i.p.
administration studies in mice exposed to purified compound or extracts of C. raciborskii cells. Studies
conducted with purified toxin (Riesner et al., 2004 and one component of Humpage and Falconer 2002)
are preferred over other extracts which could contain other toxins or compounds with similar chemical
physical properties that co-elute with the toxin. Based on available studies, effects on the liver and
kidney, increases in the hematocrit level in serum and deformation of RBCs were the most sensitive
endpoints observed in mice exposed to cylindrospermopsin (Humpage and Falconer, 2003; Reisner at al.,
2004; Sukenik et al., 2006).
Increased relative kidney weight, observed in several studies (Humpage and Falconer, 2002, 2003;
Sukenik et al., 2006; Reisner et al., 2004), was the most sensitive endpoint, co-occurring with decreased
urinary protein excretion (Humpage and Falconer, 2002, 2003), and RBC changes (Humpage and
Falconer, 2002, 2003; Sukenik et al., 2006; Reisner et al., 2004). The mode of action for the decreased
urinary protein levels observed in male mice in both drinking water and gavage studies need further
investigation but are consistent with either an impact on hepatic mouse urinary protein synthesis and/or
retention of mouse urinary proteins by the kidney. The relevance of kidney effects in mice to humans is
supported by the results obtained from the Palm Island water poisoning incident. The most severe impacts
observed in the exposed population reflected impairment of kidney function, with decreased serum
electrolytes, glycosuria, proteinurea, ketonurea and hematourea that led to hospitalization. Treatment
included intravenous electrolyte and replacement of serum proteins. Without this support for kidney
malfunction, several would have most likely died. Children were particularly sensitive to hypokalemia
and acidosis as evidenced by the fact that 82% displayed these conditions when hospitalized. It is
important to note that in mice the effect observed was diminished excretion of protein. In mice, decreased
protein excretion is as much as a reflection of altered kidney function as is increased protein loss in
humans because urinary proteins in mice have distinct physiological functions.
The long-term extract study by Sukenik et al. (2006) and the shorter duration study of purified
cylindrospermopsin by Reisner et al. (2004) showed structural changes in the RBC wherein the cells
acquired a spiked external membrane (acanthocytes) rather than their normal appearance. The
acanthocytes (abnormal RBCs) were associated with an increased hematocrit value as an indicator for
adverse changes along with increased kidney and liver weights. Humpage and Falconer (2002, 2003) did
not find significant changes in RBC membrane parameters in their 11-week mouse subchronic study.
They did observed a trend towards increased serum bilirubin and spleen weight across the 30 (ig/kg/day
to 120 (ig/kg/day range plus non-significant increases in polychromasia. These observations are consistent
with removal of abnormal or hemolysized RBCs by the spleen. Sample sizes were small (4-5 per dose
group) which is a limiting factor for determining statistical significance.
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An increase in the ratio of RBC membrane cholesterol to phospholipids could be responsible for
acanthocyte formation according to Reisner et al. (2004). Studies of the effects of cylindrospermopsin on
LCAT structure and/or function in altering the cholesterol content of the red cell membrane could shed
light on the mode of action for this effect. There are no data from human epidemiology studies of
cylindrospermopsin that have examined RBC morphology.
No information was located regarding the chronic toxicity or neurotoxicity of cylindrospermopsin. Effects
on the liver/body weight ratio were seen in maternal CD-I mice exposed to 8 (ig/kg/day and a
32 |_ig/kg/day dose was identified as a FEL (Rogers et al., 2007). There was no significant difference in
the number of live fetuses/per litter at either dose. In separate studies by the same authors, there was some
evidence for an impact on postnatal growth and survival after a maternal i.p. dose of 50 (ig/kg/day on
GDs 13-17, but not on GDs 8-12 (Rogers et al., 2007). Sibaldo de Almeida et al. (2013) did not find any
visceral or skeletal malformations in the offspring of pregnant rats receiving an oral dose of 3 mg/kg/day
purified cylindrospermopsin during gestation (GD 1-20).
7.1.1. Mode of Action for Noncancer Effects
Liver—The occurrence of toxicity in the liver suggests a protein-synthesis inhibition mechanism of action
for cylindrospermopsin. In vitro and in vivo studies have been conducted to demonstrate the ability of
cylindrospermopsin to inhibit hepatic protein synthesis, which could impact mouse urinary protein
production leading to decreased urinary excretion of these proteins (Froscio et al., 2008, 2009; Terao et
al., 1994). Available evidence indicates that protein synthesis inhibition is not decreased by broad-
spectrum CYP450 inhibitors, but they do reduce cytotoxicity (Froscio et al., 2003; Bazin et al., 2010).
Hepatotoxicity appears to be CYP450-dependent which indicates a possible involvement of oxidized and
or fragmented metabolites and mechanisms other than protein synthesis inhibition (Froscio et al., 2003;
Humpage et al., 2005; Norris et al., 2001, 2002). Despite the number of studies that have been published,
the mechanism for liver and kidney toxicity by cylindrospermopsin are not completely characterized.
Red Blood Cells—There was evidence of effects on RBCs in the Reisner et al. (2004) and Humpage and
Falconer (2002) studies of purified cylindrospermopsin. In the Reisner et al. (2004) report, microscopic
examination of blood samples showed the presence of RBCs with spiked surfaces rather than their normal
biconcave-disc shape. The authors attributed the acanthocyte formation to an increase in the RBC
membrane cholesterol to phospholipid ratio. Phospholipids constitute the matrix material of cell
membranes. The authors hypothesized that this change was the consequence of decreased activity of
plasma lecithin-acyl cholesterol transferase (LCAT), an enzyme associated with high density lipoproteins
and the esterification of plasma cholesterol. Effects on the cholesterol content of the RBC membrane can
occur with inhibition of the enzyme increasing membrane fluidity and mean corpuscular
volume. Associated effects were observed in the Reisner et al. (2004) and Humpage and Falconer (2002)
studies. Removal of the abnormal blood cells by the spleen increases both spleen weight and serum
bilirubin plus stimulates hematopoiesis. Additional research is needed to examine the LCAT enzyme
inhibition hypothesis in order to confirm whether it accounts for the effects on the RBC as a result of
cylindrospermopsin exposure.
Kidney—No information on the mode of action for the kidney effects observed in the studies of
cylindrospermopsin was provided by the researchers. Since all of the studies were conducted using mice,
a species that excretes low molecular weight proteins in urine, there is a need to conduct a study of
cylindrospermopsin in a laboratory species that does not excrete protein in the urine in order to determine
whether there are comparable effects on kidney weight, protein excretion and renal cellular damage.
Kidney necrosis and the decreased renal failure index at the high cylindrospermopsin doses provides
support for the effects on the kidney. Numerous signs of renal damage including proteinuria, glycosuria,
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and hematuria were observed after the Palm Island incident, all of which are associated with impaired
kidney function (Byth, 1980).
7.1.2. Dose Response Characterization for Noncancer Effects
Human Data—The limited information on the toxicity of cylindrospermopsin in humans is from
qualitative reports of a hepatoenteritis-like illness attributed to the acute or short-term consumption of
drinking water containing C. raciborskii (Byth, 1980; Griffiths and Saker, 2003). The clinical picture of
the illness includes fever, headache, vomiting, bloody diarrhea, hepatomegaly and kidney damage with
loss of serum electrolytes, proteinuria, glycosuria, and hematuria, but no data are available on exposure
levels of cylindrospermopsin that can induce these effects. Thus, human data on the oral toxicity of
cylindrospermopsin are limited by lack of quantitative information and by potential co-exposures to other
cyanobacterial toxins and microorganisms.
Animal Data—The information on non-cancer effects of cylindrospermopsin in animals is available from
oral and i.p. administration studies in mice exposed to purified compound or extracts of C. raciborskii
cells. Based on available studies, the liver, kidneys and RBCs appear to be the main targets of
cylindrospermopsin toxicity.
Studies conducted with purified toxin are preferred over those of extracts, which may contain other toxins
or compounds with similar chemical physical properties that co-elute with the toxin. Effects on the liver
and kidney, including changes in organ weights and histopathological lesions, along with increases in the
hematocrit level in serum and deformation of RBC are observed following short term and subchronic oral
exposure to cylindrospermopsin (Humpage and Falconer, 2002, 2003; Reisner at al., 2004; Sukenik et al.,
2006). Oral and i.p. acute toxicity studies in mice also report histopathological effects in both liver and
kidney.
No oral reproductive or developmental and chronic toxicity studies are available for cylindrospermopsin.
Developmental toxicity studies following i.p. administration of cylindrospermopsin provide some
evidence for maternal toxicity and decreased post-natal pup survival and body weight (Rogers et al.,
2007; Chernoff et al., 2011).
7.2 Synthesis and Evaluation of Major Carcinogenic Effects
Studies investigating the in vitro and in vivo genotoxicity (evaluation of DNA damage) from exposure to
cylindrospermopsin are relatively few in number. In vitro mutagenic and genotoxic cell assays with
cylindrospermopsin show varied results with some indications of potential damage to DNA. The human
hepatocytic and enterocytic models for HepaRG and Caco-2 cells showed increased MNBNC (Bazin et
al., 2010). Micronucleated cells were observed in a study with human lymphoblastoid WIL2-NS cells
(Humpage et al., 2002). DNA breaks have been observed in primary hepatocytes by the comet assay
indicating that DNA strand breakage could be a mechanism for cylindrospermopsin-induced cytogenetic
damage (Humpage et al., 2005). Following i.p. exposure, DNA strand breakage was observed in the liver
of Balb/c mice (Shen et al., 2002) and covalent binding between DNA and cylindrospermopsin, or a
metabolite, occurred in a study on Quackenbush mouse liver (Shaw et al., 2000). However, these data are
limited and there has been no long term bioassay of purified cylindrospermopsin. The study by Falconer
and Humpage (2001) on initiation with TPA promotion did not support classification of
cylindrospermopsin as a tumor initiator.
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7.2.1. Mode of Action and Implications in Cancer Assessment
There is minimal information available to inform a cancer mode of action hypothesis for
cylindrospermopsin. The study by Falconer and Humpage (2001) noted only one tumor and two areas of
dysplastic foci in a study of two doses of C. raciborskii extracts or three doses of freeze-dried cells
combined in treatment with TPA as a promoter. Genotoxicity studies indicate a potential for
cylindrospermopsin to cause DNA strand breaks in cell lines with activated CYP450s as well as provide
evidence for possible gene amplification and chromosomal alterations form in vitro (Straser et al., 2011;
Zegura et al., 2011) and in vivo studies (Shen et al., 2002). The data from the Marie et al. (2010) study in
SHE cells, the clastogenicity seen in the comet assays (Humpage et al. 2000, 2005), and micronuclei
observed by Bazin et al. (2010) and Lankoff et al. (2007) support the need for additional research.
7.2.2. Weight of Evidence Evaluation for Carcinogenicity
There are no studies in humans evaluating cancer and no chronic cancer bioassays in animals available for
cylindrospermopsin. In accordance with the Guidelines for Carcinogen Risk Assessment (U.S. EPA,
2005a), there is inadequate information to assess carcinogenic potential of cylindrospermopsin.
7.2.3. Dose Response Characterization for Cancer Effects
Dose-response data regarding the carcinogenicity of cylindrospermopsin are not available.
7.3. Potentially Sensitive Populations
A review of the available animal data does not support a definitive difference in the response of males
versus females following oral exposure to cylindrospermopsin. Based on animal study results, individuals
with liver and/or kidney disease might be more susceptible than the general population because of
compromised detoxification mechanisms in the liver and impaired excretory mechanisms in the kidney.
Results of an episode in a dialysis clinic in Caruaru, Brazil, where microcystins (and possibly
cylindrospermopsin) were not removed by treatment of dialysis water, suggest that dialysis patients are a
population of potential concern in cases where the drinking water source is contaminated with
cyanotoxins.
The data on RBC acanthocytes (abnormal RBCs) identifies individuals that suffer from anemia
(e.g. hemolytic or iron-deficiency) as a potentially sensitive population. Several rare genetic defects such
as abetalipoproteinemia and hypobetalipoproteinemia are associated with RBCs acanthocytes and appears
to result from a defect in expression of hepatic apoprotein B-100, a component of serum low density
lipoprotein complexes (Kane and Havel, 1989). Individuals with either condition might be sensitive to
cylindrospermopsin.
Based on the currently available science, evidence is lacking to assess differences in susceptibility
between infants and children and adults. However, for cyanotoxins including cylindrospermopsin,
drinking water contributes the highest risk of the total cyanotoxin intake for infants to one year old fed
exclusively with powdered formula prepared with tap water containing cyanotoxins. Based on average
drinking water intake rates for children <12 months (0.15 L/kg-day), the exposure of children is 5 times
higher than those of adults >21 years old on a body weight basis (0.03 L/kg-day).
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7.4. Characterization of Health Risk
7.4.1. Choice of Key Study
Human data on the toxicity of cylindrospermopsin are limited by the lack of quantitative information and
by potential co-exposures to other cyanobacterial toxins and microorganisms. The limited information on
the toxicity of cylindrospermopsin in humans is from qualitative reports of a hepatoenteritis-like illness
that is attributed to the acute or short-term consumption of drinking water containing C. raciborskii (Byth,
1980; Griffiths and Saker, 2003). Although clinical symptoms of the illness from the Australian Palm
Island poisoning incident are well-defined and documented, no data are available on the exposure levels
for cylindrospermopsin that induced these effects.
Observed health effects following single, oral exposures included mortality and toxicity in the liver and
kidneys (previously described in Section 6.2.1). Although these effects were observed, they are
inadequate to support the derivation of an RfD due to limitations such as: testing inadequate numbers of
animals, incomplete reporting, failure to measure clinical and pathological endpoints and exposure to a
single dose. Additionally, these studies include exposure to C. raciborskii cells or cell extracts (not
purified cylindrospermopsin).
Several short term and developmental toxicity studies that evaluated the effects of cylindrospermopsin are
also available (Shaw et al., 2001, Rogers et al., 2007, Chernoff et al, 2011). These studies were not
selected for the derivation of the RfD due to limitations including use of extract (Shaw et al, 2001), i.p.
route of administration (Rogers et al., 2007, Chernoff et al., 2011), lack of adequate numbers of animals
and monitored endpoints (Shaw et al., 2001), and the limited number of doses tested (Shaw et al., 2001).
The oral data for purified extract from Shaw et al. (2001) identified fatty liver as an adverse effect in mice
following a 14 day gavage exposure to 0.05 mg/kg/day. The only effects mentioned in the published
paper are the liver effects and an absence of lymphophagocytosis in the spleen.
The critical study selected for the derivation of the RfD is Humpage and Falconer (2002, 2003). Humpage
and Falconer (2002, 2003) is a comprehensive toxicity study in which mice were exposed by gavage to
purified cylindrospermopsin from cell extract for 1 lweeks. The study authors also utilized four dose
groups, adequate numbers of animals per dose group and evaluated a variety of endpoints. Statistically
significant, dose-related effects on the kidney, liver and serum chemistry were observed. The kidney was
the most sensitive target of toxicity. The Humpage and Falconer (2002) data are supported by the Riesner
et al. (2004) results showing increased kidney weights and hematological effects (acanthocytes) following
a 30 day exposure.
7.4.2. Selection of the Principal Study
The subchronic study by Humpage and Falconer (2002, 2003) and the short term studies by Sukenik et al.
(2006) and Reisner et al. (2004) all identified increases in kidney weight and hematological effects as the
result of exposure to cylindrospermopsin. Humpage and Falconer (2002, 2003) found signs indicative of
hemolysis (increased bilirubin, spleen weight and polychromasia), while Reisner et al. (2004) and
Sukenik et al. (2006) found acanthocytes and increased hematocrit. Increases in kidney weight were
significant for Humpage and Falconer (2002, 2003) and Sukenik et al. (2006), but not significant for
Reisner et al. (2004). Humpage and Falconer (2002, 2003) and Reisner et al. (2004) used purified
cylindrospermopsin, while Sukenik et al. (2006) used an extract in spent medium.
Sukenik et al., (2006) was a step up dose study. The authors used an extract that was not purified and
identified 20 (ig/L as the LOAEL at about 20 weeks. Because it was an extract study using spent medium
dissolved in water as the exposure vehicle the study was not selected to derive the RfD.
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The results from the Reisner et al. (2004) single dose study demonstrated that hematological and kidney
weight effects were present in young (4 week) male ICR mice (8 mice) after a three week exposure with a
LOAEL of 66 (ig/kg/day. Humpage and Falconer (2002, 2003) used 10 male Swiss mice per dose group
and evaluated 5 doses. Because of the similarity in the LOAELs from Humpage and Falconer (2002,
2003) and Reisner et al. (2004) and the type of effects observed, the selection of Humpage and Falconer
(2002, 2003) was determined to be the most appropriate study for derivation of the RfD, despite its
subchronic duration of exposure.
7.4.3. Selection of the Critical Endpoint
Upon considering all effects observed by Humpage and Falconer (2002, 2003), increased relative kidney
weight was considered the most appropriate basis for quantitation. Adverse effects on the kidney were
manifested by decreased urinary protein concentration and increased relative kidney weight. The study
authors reported significant increased relative kidney weight at >60 (ig/kg/day, decreased urinary protein
and liver lesions at >120 (ig/kg/day and renal tubular lesions at 240 (ig/kg/day (Humpage and Falconer,
2002, 2003). Relative kidney weight was increased in a significant, dose-related manner beginning at
60 (ig/kg/day (12-23% greater than controls) and relative liver weight was significantly increased at the
high dose of 240 (ig/kg/day (13% greater than controls). Relative spleen, adrenal and testes weights were
increased for doses >60 (ig/kg/day, but the differences from control, although dose-related, were not
statistically significant. Humpage and Falconer (2002, 2003) identified the LOAEL as 60 (ig/kg/day and a
NOAEL of 30 (ig/kg/day based on the dose related and statistically significant increase in relative kidney
weight. These adverse effects are potential indicators of suppressed hepatic protein synthesis and/or
increased retention of low molecular weight of mouse urinary proteins by the kidney because of damage
to the renal tubules.
7.4.4. RfD Determination
The NOAEL from the Humpage and Falconer studies (2002, 2003) was the 30 (ig/kg/day dose based on
increased relative kidney weight. The RfD is calculated as follows:
30ng/kg/day =
300
where:
30 (ig/kg/day = The NOAEL for kidney effects in mice exposed to
cylindrospermopsin in water for 11 weeks (Humpage and Falconer,
2002, 2003).
300 = Composite uncertainty factor including a 10 for intraspecies
variability (UFh), a 10 for interspecies differences (UFa), and 3 for
uncertainties in database (UFd).
Uncertainty Factor Application:
• UFh. A Ten-fold value is applied to account for variability in the human population. No
information was available to characterize inter-individual and age-related variability in the
toxicokinetics or toxicodynamics among humans. Individuals with a low-red-cell count as a result
of genetic or nutritional factors could be more sensitive to cylindrospermopsin exposures than the
general population. Individuals with pre-existing kidney problems may also be more sensitive.
• UFa. A Ten-fold value is applied to account for uncertainty in extrapolating from laboratory
animals to humans (i.e., interspecies variability). Information to quantitatively assess
toxicokinetic or toxicodynamic differences between animals and humans is unavailable for
Health Effects Support Document for Cylindrospermopsin - June, 2015
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cylindrospermopsin. Information to quantitatively assess toxicokinetic or toxicodynamic
differences between animals and humans is unavailable for microcystin. Allometric scaling is not
applied in the development of the Ten-Day HA values for microcystin. The allometric scaling
approach is derived from the relationship between body surface area and basal metabolic rate in
adults (U.S. EPA, 2011). For infants and children, surface area and basal metabolic rates are very
different than adults.
• UFd. An uncertainty factor of 3 (10°5 = 3.16) is selected to account for deficiencies in the
database for cylindrospermopsin. The database for cylindrospermopsin includes limited human
studies. The database for studies in laboratory animals includes oral exposure acute, short-term
and subchronic studies, but many of them lacked a comprehensive evaluation of a wide spectrum
of effects. The database lacks chronic toxicity and multi-generation reproductive and
developmental toxicity studies using the oral route of exposure. There is a lack of data on
neurological and immunological endpoints. The RBC parameters evaluated differed between the
Humpage and Falconer (2002, 2003) and Reisner et al. (2004) studies.
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8.0 RESEARCH GAPS
The deficiencies in the toxicological database for cylindrospermopsin are many. The nature of the
problem limits research in humans to outbreak reports and case studies. Both are retrospective scenarios
with confounding variables related to the composition of the toxins in the water source, the timing of
exposure and the dose. Controlled animal systemic studies have been conducted only in male mice
making it difficult to determine whether the critical effects (increase in kidney weight and decreased
protein excretion) are relevant to species that do not normally excrete urinary protein for functional
reasons. Mode of action information is lacking for the liver, kidney and hematological.
Research is needed to improve the quantitative assessment for human health consequences from exposure
to cylindrospermopsin in drinking water. Key research gaps were identified during the development of
this document and are not intended to be an exhaustive list. Additional research efforts are needed on:
• Quantification for the absorption, distribution, and elimination of cylindrospermopsin in humans
or animals following oral, inhalation and/or dermal exposure.
• The clinical significance in humans for biological changes observed in experimental animals such
as increased kidney weight, decreased urinary protein levels, decrease in renal failure index, and
the formation of acanthocytes.
• Health risks posed by repeated, low-level exposures to cylindrospermopsin in a second species.
• The chronic toxicity of cylindrospermopsin. Whole-lifetime toxicity studies showing cumulative
detrimental effects.
• The immunotoxic, neurotoxic and developmental/reproductive toxicity of cylindrospermopsin
following oral exposure.
• The in vivo genotoxicity of cylindrospermopsin exposure.
• The carcinogenic potential of cylindrospermopsin, including lifetime carcinogenicity studies.
• Health risks from exposure to mixtures of cylindrospermopsin with other cyanotoxins, bioactives,
and chemical stressors present in ambient and or drinking water supplies.
• Populations that might be sensitive to cylindrospermopsin exposure via the oral, dermal, and/or
inhalation routes.
Health Effects Support Document for Cylindrospermopsin - June, 2015
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Appendix A: Studies Used in Support of Reference Value Derivation for
Cylindrospermopsin
Humpage and Falconer (2002)
Reisner et al.,
2004
Sukenik et al., 2006
Study 1
Study 2
Test Substance
Extract
Compound
Compound
Spent medium with:
95% CYL
5% 7-epiCYL
Species
Male Swiss
Albino mice
Male Swiss
Albino mice
Male ICR mice
ICR mice
Number
10
10 (4-5 for
hematology)
8
20/sex
Route
Drinking Water
Gavage
Drinking Water
Drinking Water
Duration
10 weeks
11 weeks
3 weeks
42 weeks (step up dose
design
Doses
0.216, 423, 657
jjg/kg/day
1, 30, 60, 120,
240 jjg/kg/day
0, 66 jjg/kg/day
10-50 jjg/kg/day
Histopathology
Liver injury (432)
Liver injury
(>120)
ND
ND
Liver weight*
T
ND
T
| (42 weeks male)
Kidney weight*
T
T
T
| 20 and 42 weeks
Testes weight*
ND
T
T
| (42 weeks)
Spleen weight*
| trend
<—>
<—>
Hematocrit or
paced cell volume
ND
| PCV trend
(N=5)
1 with duration
1 (32 weeks) J, 42 weeks
Acanthocytes
ND
ND
Detected
Detected
Polychromasia
T
Serum bilirubin
T
| trend (N=5)
ND
ND
Cholesterol
ND
|30 and 60
jjg/kg/day
| higher doses
but <60 (N=5)
T
T
Bile acids
ND
ND
Urinary protein
4
4
ND
ND
NOAEL jjg/kg/day
-
30
-
-
LOAEL jjg/kg/day
216
60
66
20 (based on 20 week
data)
ND = Not Determined; * Relative organ weight
-Shaded cell entries signify that the study provides numeric values for the effects with or without accompanying
graphic representation.
-From Reisner et al. (2004) and Sukenik et al., (2006) the data on statistical significance represent a relationship
between exposure duration and response, not a relationship between dose and response.
-The term trend does not denote a statistical test for trend, it indicates a uniform direction for the change as reported in
the Humpage and Falconer, 2002.
-Reisner et al. (2004) hypothesis for acanthocyte formation: Studies in humans and rats indicate that acanthocytes
form due to alterations in RBC membrane lipoproteins that increase the ratio of cholesterol to phospholipids.
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