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Edited by Leon P. Zann
Great Barrier Reef Marine Park Authority, Townsville Queensland
Ocean Rescue 2000 Program
Department of the Environment, Sport and Territories, Canberra, 1995
ISBN 0 642 17406 7
Nicholas J. Ashbolt
AWT Science & Environment
West Ryde, NSW 2114
The purpose of this review is to summarise present understanding of the micro-organisms of public health concern in the Australian marine environment. Human health aspects relating to chemicals released into the Australian marine environment are covered elsewhere in the SOMER documentation (Philip 1994). Guidelines for exposure to micro-organisms, as with other health-related contaminants, are categorised by water use. This review covers waters used for primary contact recreation (e.g. swimming/immersion in water), secondary contact recreation (e.g. boating) and fisheries.
From a human health perspective, most of Australia's coastal marine environment is distant from the focus of concern, that is regions of urban development. While little is known about the disease-causing micro-organisms (pathogens) in most of our marine environment, overseas experience points to two microbial groups of significance: 1) those introduced from animal and human wastes and 2) indigenous pathogens.
Both groups contain opportunistic pathogens, which only present a health hazard under certain conditions. For example, if an individuals' normal defence mechanisms are compromised, such as by a break in the skin or by the presence of immunosuppressive agents, or if pathogen(s) grow to high densities in the presence of increased organic wastes. The important routes of pathogen uptake are ingestion, aerosol/liquid inhalation and through breaks in the skin. Doses required to cause infection vary widely with micro-organism and status of individuals. However, in general, 10-100 viruses, 1-10 000 parasites and 1000-1 000 000 bacteria are required for infection.
Introduced pathogens (from faecal material) make their way to the marine environment from a number of sources including sewage, septic seepage, boats, stormwater and diffuse run-off via estuaries and major rivers/harbours. These protozoan, helminth, bacterial and viral pathogens may arrive freely suspended, but they are more likely to be associated with particulate matter. Considerable protection from die-off/predation may be conveyed to pathogens which associate with particulate material, and this may present problems in enumeration. Hence, caution is needed in interpreting and comparing quantitative results presented in the literature. Nonetheless, while predicting die-off is a complex problem involving many factors, dilution has the major impact.
In Australia, faecal pollution in seawater is inferred from the presence of certain indicator bacteria, primarily faecal coliforms and/or faecal streptococci (includes the enterococci). However, epidemiological studies of waterborne illness indicate that the common aetiological agents are more likely to be viruses and parasitic protozoa than bacteria (Moore et al. 1994; Seyfried et al. 1985; Cabelli et al. 1982). Furthermore, the recent literature illustrates the poor correlations between waterborne human viruses and faecal coliforms in marine waters (Deuter et al. 1991). This lack of a relationship relates in part to the sporadic presence of pathogens in sewage, which reflects the incidence of illness in the population. In addition, there are problems in sampling (Fleisher 1990), and many pathogens survive longer than the faecal indicator bacteria determined by culturing methods (Evison 1988), with viruses present when indicator bacteria are absent (Hughes et al. 1991). A further complication is the occurrence of nonculturable but viable indicator bacteria which are not enumerated by standard methods and result in further underestimation of pathogen presence (Byrd et al. 1991; Green et al. 1991).
Therefore, this review covers recent literature on the prevalence of a range of key micro-organisms and points to research and management's responsibilities. Secondary public health risks, such as the transfer of antibiotic resistance genes from sewage bacteria to the sediment microbiota and shellfishes (Belliveau et al. 1991; Avilés et al. 1993) are not examined. Also, diseases such as dengue fever and malaria, which are transmitted by insects associated with water, are not considered in this review.
In Australia over 80% of the population reside in large coastal cities with aging sewerage systems. Sewage discharged from sewerage treatment works to estuarine and coastal waters receives at least, primary treatment. As it is currently considered impractical to monitor for the range of pathogens possibly present in seawater, indicator organisms have been chosen (ANZECC 1992).
There is, however, no single suitable group of indicator micro-organisms available (Elliot & Colwell 1985), although the bacterial groups of faecal coliforms and faecal streptococci correlate reasonably well with some of the bacterial pathogens, such as salmonellae (Moriñingo et al. 1992). Furthermore, extension of indicator behaviour, as determined in extensive temperate water studies, to tropical waters must be questioned (McNeill 1992).
Factors affecting faecal coliform decay have been applied to complex models (Canale et al. 1993) in an attempt to predict health risk. Such an approach is doomed to failure, as a range of pathogens in sewage effluent survive for considerably longer periods in marine waters than the indicator bacteria (Tables 1 & 2). In addition, the ratio of faecal coliforms to streptococci is of no use in identifying pollution sources (Pourcher et al. 1991), although the use of strain specific molecular probes may prove useful in the not too distant future.
It is important to note when reviewing the literature that though 99.9% reductions in pathogens may at first appear satisfactory, this is often not the case. For example, infective doses of viruses or protozoa (possibly as low as 1-50 particles) will still be present after 99.9% removal from raw sewage, viz. secondary treated sewage could represent a significant health risk if further disinfection/dilution is not guaranteed.
Once an enteric pathogen of human or animal origin enters a marine or estuarine environment, there are a number of factors affecting its fate. These include sedimentation, predation (by copepods and protozoa), parasitism, inactivation by sunlight, temperature, osmotic stress, or toxic chemicals (Kueh et al. 1991). While dilution is clearly a major factor reducing the likelihood of infection, there is a range of interacting factors, such as presence of nutrients and suspended solids, that influence the survival of pathogens in seawater (McNeill 1992).
An indication of mortality rates, in terms of T90 or T99.9 values (times for 1 and 3 log reduction in numbers, respectively), for selected indicators and pathogens is given in Table 2. Cooler waters and lack of sunlight increase survival. For some bacteria there appears to be a linear relationship between temperature and log T90, and between light intensity and T90 (Evison 1988). Hence, release of subsurface sewage plume material not only substantially reduces harmful light effects but also allows for adaptation to higher salt concentrations with less cell stress, pushing T90's out to several tens of hours for bacteria (Pommepuy et al. 1992) and many months for enteric viruses (Gerba & Goyal 1988).
There is a paucity of data on viral illness associated with direct contact (swimming) in polluted waters. This is a result of the fact that of the over 120 classified enteric viruses, most are very difficult to isolate from the aquatic environment and many are simply nonculturable (USEPA 1985). Nevertheless, enteroviruses (polioviruses, coxsackie viruses, echoviruses) hepatitis A & E, adenoviruses, rotaviruses, and caliciviruses (including Norwalk and small round structured viruses) have all been associated with swimming-related illness (Dufour 1986).
Table 1: Key faecal micro-organisms in sewage and typical % removals by various treatment processes
|Source||Escherichia coli||Salmonella/Campylobacter||Enteric Viruses||Giardia Cysts|
|Raw Sewage (L-1)||108-109||40 000||100-15 000||5200-22 700|
|% Removal by:|
|Primary treatment||50-90, 27-96||50-90, 15||0-30||55|
|Secondary treatment||91-99||96-99||30-75, 76-99||99|
Table 2: Major potential pathogens/indicators in the marine environment
|Group of Organism||Source(s)||Symptom(s)||Marine Survival|
|Adenovirus||animal/human faeces||C Co F G||H R 50 d|
|Calicivirus (inc.Norwalk)||human faeces||G||Unknown|
|Coxsackie A&B||human faeces||B C D E-M F H R S||2 d - 46 wk|
|Echovirus||human faeces||C E-M F G R P S||2 d - 46 wk|
|Hepatitis A||human faeces||H||> 24 d|
|Poliovirus||human faeces||C F E-M P R||2-130 d|
|Reovirus||animal/human faeces||None known||> 4 d|
|Rotovirus||animal/human faeces||G||2-34 d|
|Aeromonas spp.||animal/human faeces||G S W||'indigenous'|
|Campylobacter jejuni||animal/human faeces||G-F||Poor|
|Enterotoxigenic Escherichia coli||animal/human faeces||G 5||h-2 d|
|Faecal coliforms||animal/human faeces||indicator organism||2 h -2 d|
|Faecal streptococci||animal/human faeces||indicator organism||2 h - 12 d|
|Mycobacterium marinum||seawater||S W||'indigenous'|
|Salmonella spp.||animal/human faeces||G-F||12h - 5d|
|Shigella spp.||animal/human faeces||Bloody diarrhoea||<15->70 d|
|Vibrio spp.||seawater, faeces||G W||'indigenous'/<6d|
|Yersinia enterocolitica||animal/human faeces||Appendicitis-like G||days-weeks|
|Cryptosporidium parvum||animal/human faeces||Watery diarrhoea F||Unknown|
|Entamoeba histolytica histolytica||faeces||G/dysentery||Unknown|
|Giardia intestinalis||animal/human faeces||Diarrhoea/bloating||Unknown|
|Ascaris spp.||animal/human faeces||Roundworm||Unknown|
|Taenia spp.||animal/human faeces||Tapeworm||Unknown|
|Gambierdiscus toxicus||seawater||ciguatera shellfish poisoning||'indigenous'|
Source: Chung & Sobsey 1993; Hallegraeff 1992; Evison 1988; McNeill 1985; Akin et al. 1976; Blawat et al. 1976. Key: C-carditis, Co-conjunctivitis, F-fever, D-diabetes, E-M-encephalitis-meningitis, G-gastroenteritis, G-F- gastro+fever, H-hepatitis, P-paralysis, PSP-paralytic shellfish poisoning, R-respiratory infection, S-skin infection, W-wound infection. T90 or T99.9 : times for 1 or 3 log reduction in numbers respectively at 10-25°C.
Seasonality plays a role in the presence of enteric viruses in the community and sewage effluent. Therefore, it may be necessary to detect a range of viruses to determine viral risk in receiving waters. Identification of many of these nonculturable viruses requires molecular methods, which in the case of enteroviruses appear to largely identify infective viruses (Enriquez et al. 1993).
Viruses which specifically infect bacteria (bacteriophages) are useful surrogates for human enteric viruses in survival studies (Elliot & Colwell 1985). Bacteriophages are readily enumerated by plaque assays on bacterial host cells and can serve as useful tracers of sewage effluent, as demonstrated off Geelong, Victoria (Richardson et al. 1993). However, it is important to note that coliphage presence in sewage is related to initial coliform concentration, not the concentration/presence of enteric viruses (Grimes et al. 1986).
To date, only the enterovirus group is monitored in beach waters (in 10 L volumes) under a European directive (Council of the European Communities 1976) and proposed guidelines in some American states for recreational waters (Bitton 1980). As a consequence, most information for enteric viruses relates to the enterovirus group. A problem with this approach is that the majority of culturable enteric viruses in marine waters are not enteroviruses, but generally reoviruses (Grabow et al. 1989; Grohmann et al. 1993). Hence, as discussed in the Sydney region studies (Grohmann et al. 1993), it is important to focus on a range of viral groups, else risk in considerably underestimating enteric virus risk.
The majority of shellfish-associated illness reported in the US is of unknown aetiology, with the largest group identified being due to nonculturable Hepatitis A (20% of illnesses) and a small percentage due to nonculturable Norwalk viruses (Craun 1986). In contrast, Norwalk-like viruses have been the main cause of viral food poisoning from sewage contaminated oysters in Australia (Grohmann et al. 1980). Based on a virus prevalence of 19% in shellfish, Rose and Sobsey (1993) estimated that there was a one in a hundred chance of becoming infected with an enteric virus by consuming raw shellfish. While such an analysis lacks sufficient viral prevalence data, reliance on bacteriological standards appears to be quite deficient in protecting human health. For example, a multi-state outbreak of hepatitis A resulted from eating raw oysters in waters meeting bacteriological standards (Desenclos et al. 1991).
Viruses in waters generally adsorb to solids which protect them from inactivation by biological, chemical and physical factors (USEPA 1985). Marine sediments, for example can adsorb more than 99% of a poliovirus suspension (containing 108 plaque forming units per mL) and may contain 10-10 000 times the concentration of viruses in overlying water (Schaiberger et al. 1982; LaBelle & Gerba 1979). Hence, surface water sampling alone may not give a true indication of the potential viral hazard (Rao et al. 1984). Such sediment-bound viruses can also be taken up by shellfish, thus allowing their bio accumulation in marine life near sewage outfalls (Lewis et al. 1986).
Virus survival in this adsorbed state is of particular interest, as it is well known that virus removal from shellfish by seawater flushing (depuration) occurs at a significantly slower rate than for bacteria (Lewis et al. 1986). Furthermore, sediment-associated viruses are known to maintain their infectivity (Taylor et al. 1980).
Due to considerable variability in enumerating enteric viruses, results are generally discussed on a presence/absence basis. For example, based on the European directive of no enterovirus in 10 L, studies on selected beach waters indicated 27-35% of southern Welsh, 100% of Yorkshire, 29% of English and 46% of Northern Ireland beaches failed the enterovirus directive over various periods in the last seven years (Deuter et al. 1991; Hughes et al. 1992). In Sydney, tests on 100 L samples indicated enteric viruses (entero-, adeno- and reo-viruses) were present in about 20% of city beach waters (Grohmann et al. 1993) prior to commissioning three deepwater ocean outfalls (discharging primary treated sewage effluent). In the two years since commissioning the deepwater outfalls, only two beach water samples (of over 300) cultured positive for an enteric virus (Grohmann, unpublished).
In most western countries, including Australia, gastrointestinal illnesses due to Campylobacter species outnumber those due to any other identified pathogen. However, illness from these bacteria will almost solely be due to food or person-to-person contact (Cohen & Gangarosa 1978). Of the introduced bacterial pathogens (Table 2) salmonellae and shigellas probably survive the longest in marine waters. Nevertheless, the most recent outbreak of swimming-associated Salmonella illness world-wide was typhoid fever (Salmonella typhi) which occurred in Western Australia in 1958 (Anon. 1961). Shigella gastroenteritis has only been implicated from swimming in fresh waters (Herwaldt et al. 1991).
The causative agent of cholera, Vibrio cholerae 01 (and recently type 0139 in Asia) is endemic in various countries, but not Australia. Nevertheless, pandemic strains of cholera could be transported to Australian marine waters in ships ballast water. Whereas non01 strains are not uncommon in warm Australian fresh waters, 01 types have only rarely been isolated from fresh waters (Desmarchelier 1989). This contrasts to the other pathogenic vibrios, V. parahaemolyticus and V. vulnificus which are endemic to temperate Australian seawaters and are two of the most common causes of bacterial food poisoning from shellfish. Furthermore, these endemic vibrios (V. alginolyticus, V. parahaemolyticus, V. vulnificus or V. mimicus) have been reported in Australia and overseas as the cause of swimmer's ear following contact with warm seawater, whereas Pseudomonas aeruginosa is the causative agent following freshwater contact (Dufour 1986). However, the significance of P. aeruginosa in seawater, as a causative agent of swimmer illness, is still being debated and some recommend monitoring due to its possible presence in the absence of faecal coliforms (Mates 1992). Similarly, Staphylococcus aureus has been shown to be a useful indicator for ear, respiratory and total illness developed by bathers at densely populated bathing beaches (Cheung et al. 1991). In contrast to these overseas findings, slight to heavily polluted bathing beaches in Sydney contained very low densities of P. aeruginosa and S. aureus (Ashbolt et al. 1993).
Indigenous marine vibrios are also the major aetiological agent in wound infections observed worldwide (Dufour 1986). It is not known if there is a relationship between pollution and occurrence of pathogenic vibrios or other wound bacteria, such as Mycobacterium marinum. The latter is also indigenous to seawater, and infection is so common that it is often considered an occupational hazard among commercial fishers (Dufour 1986).
Pneumonia, resulting from the inhalation of seawater containing Pseudomonas putrefaciens, Staphylococcus aureus or Aeromonas hydrophila have been associated with near-drowning accidents (Dufour 1986). Whereas the low densities of S. aureus are unlikely to be a problem in Sydney waters, Aeromonas and Vibrio spp. are present at concentrations which could cause wound infections or septicemia under certain conditions (Kueh et al. 1992).
Examination of survival characteristics of some bacteria in the laboratory has shown that traditional cultural methods do not detect all viable bacteria present. The portion of the viable population which do not grow in culture have been termed 'viable but nonculturable' or somnicells (Byrd et al. 1991). Somnicells have been reported for a range of pathogens, and such cells may retain their ability to cause disease (references cited in Turpin et al. ).
Lack of correlation between pathogens and indicator bacteria may in part be due to light stress induced somnicell formation in Escherichia coli and Enterococcus faecalis (Byrd et al. 1991; Barcina et al. 1990). Furthermore, Lewis et al. (1991) suggest that E. coli in seawater (in the dark) may be underestimated by up to 100-fold by relying on culturing methods.
For the general population, cysts of Giardia, and oocysts of Cryptosporidium are probably the most important parasites in Australian sewage. In contrast there is an absence of data on the health significance of helminths in marine waters.
Though no relevant survival data for Giardia or Cryptosporidium was available for this review, artificially added cysts may remain viable for weeks in freshwater (De Regnier et al. 1989), and high numbers of cysts (10-100/L) have been detected in marine waters off San Juan, Puerto Rico (Correa et al. 1990); in Victoria Harbour, Hong Kong (Hutton, unpublished); and oocysts (up to 400/L) have been found in the Georges River, Sydney (Ferguson et al. 1994). Given that only a few parasite cysts cause infection and that both have been contracted during swimming (Sorvillo et al. 1992), the prevalence of viable parasites in environmental waters requires investigation.
Cases of illness transmitted through inhalation of aerosols from contaminated waste water should not be ruled out, particularly near a surface sewage/contaminated stormwater plume. Aerosols produced in surf can contain a 200-fold increase in viruses per ml than that present in the seawater (Baylor et al. 1977). Also, bacterial aerosols have been reported to cause skin sensitivity in bathers (Gruft et al. 1975).
Harmful algal blooms that have occurred in Australia were reviewed by Hallegraeff (1992) and are summarised below. Many Australian red tide species (primarily Trichodesmium erythraeum), although possibly more common in recent years in waters north from Sydney or Perth, have not been reported as containing toxic compounds. On the other hand, toxic dinoflagellate (Alexandrium minutum) red tides have occurred annually during spring since 1986 near metropolitan Adelaide. Other paralytic shellfish poisoning (PSP) dinoflagellates, such as Alexandrium catenella and Gymnodinium catenatum, have also become a problem during the 1980s in the southern states. For example, blooms of the latter have resulted in shellfish farm closures in Tasmania for up to 6 months. The sudden occurrence of toxic species and their detection in ship ballast water has led to speculation that they have recently been introduced to Australia rather than indicating a rise in eutrophication.
There are however, toxic dinoflagellates endemic to tropical Australian waters: they include Gambierdiscus toxicus, the cause of ciguatera poisoning from fish (Hallegraeff 1992). There is also recent fossil record of the tropical PSP species Pyrodinium bahamense as far south as Newcastle, but unless ocean warming eventuates, Hallegraeff suggests the latter is unlikely to return to Australian waters.
A range of prospective epidemiological studies have been undertaken worldwide over the last decade (Corbett et al. 1993; Harrington et al. 1993; Cheung et al. 1991). Most, however, rely on insufficient water samples (of variable volumes) and poor recovery of stressed cells resulting in an error of at least 50% in estimating the likely concentration of bacteria to which an individual swimmer is exposed to (Fleisher 1990). As a consequence, different index bacteria are proposed with varying correlations to illnesses.
Sewage release from surface ocean outfalls has definitively caused gastrointestinal illness in swimmers at nearby bathing beaches (Zagorski et al. 1984) and children are particularly vulnerable to such illnesses (Alexander et al. 1992). However, the majority of illness reported in prospective epidemiological studies, including two major studies undertaken in Sydney, have been respiratory and unrelated to standard faecal indicator bacteria presence (Corbett et al. 1993; Harrington et al. 1993; Cheung et al. 1991). Nevertheless, unidentified agents from sewage may cause some of these respiratory infections via aerosols, and several children apparently acquired a Norwalk-like gastrointestinal virus while swimming near a stormwater outlet in Sydney Harbour (Ferson et al. 1993). Except for the 1958 typhoid outbreak, no epidemiological study has been reported for Australian marine waters outside the Sydney region.
On the other hand, in Harrington et al. study, there was tenuous correlation between densities of the faecal bacterium Clostridium perfringens over 50 cfu/100 mL and increased respiratory illness (unpublished). Furthermore, staphylococci have strongly correlated with swimmer illness (respiratory and ear) at beaches with very high swimmer densities, suggesting staphylococci indicate person-to-person transmission rather than sewage effluent contamination (Cheung et al. 1991).
The microbiological guidelines for recreational waters in Australia are expressed in terms of concentrations of colony forming units (cfu) of faecal coliforms and enterococci (NH&MRC 1990). These and the new draft standards (ANZECC 1992) are similar to those reported around the world for faecal indicator bacteria. However, what is clearly missing are standards for pathogens, including viruses, parasitic protozoa and the indigenous bacterial and algal pathogens. With the advent of molecular detection methods for such pathogens (Dorsch et al. 1994; Vesey et al. 1994; Enriquez et al. 1993; Tsai et al. 1993), it is timely to reassess microbiological standards.
The epidemiological evidence for accepting the enterococci standard proposed by Cabelli for marine waters (Cabelli et al. 1979) has been put in doubt (Fleisher 1991). Furthermore, the longer survival of enterococci in marine waters compared to faecal coliforms is of little consequence when in their absence, enteric viruses are present in either seawater or shellfish (Schaiberger et al. 1982).
On the other hand, the sewage indicator bacterium, Clostridium perfringens is significantly elevated at sewage/sludge disposal sites (Ashbolt et al. 1993; Hill et al. 1993). This bacterium may indicate the presence of long lived faecal micro-organisms as illustrated by elevated levels hundreds of kilometres down current of a sewage sludge dump site (Hill et al. 1993). Currently no standards exist for Clostridium perfringens, although a Hawaiian (Fujioka pers. comm.) and the Sydney (Harrington et al. 1993) epidemiological studies are evaluating its usefulness as a pathogen index organism.
Coastal water microbiology has been monitored in the USA since 1955, mounting to a cost of $US130 million in 1989 (Ludwig et al. 1992). Most of this data had not been condensed into useful information (i.e. no computerised database). In addition, the data exist in pools of intensive samplings around discharge points, with very little overall study of the coastal waters. Hence, normal background levels are generally not known and monitoring programs are not coupled to the decision making and resource allocation processes.
A similar situation occurs in Australia. Government agencies require beach waters to be monitored, but streams, lagoons or stormwater drains nearby generally are not. For example, despite the use of submarine outfalls in cities like Sydney or Rio de Janeiro, unmonitored local stormwater discharges/leaky sewers still result in unacceptably high coliform counts at ocean bathing beaches during storm events (Kueh et al. 1991; Jordäo & Leitäo 1990).
If indeed the current faecal indicator bacteria are poor predictors of health risk in temperate and/or tropical Australian waters, where should we be headed? Clearly the presence of fresh sewage material is a health risk; hence a coordinated national measure by applying ANZECC guidelines could be a move in the right direction. On the other hand, a health risk may be present in the absence of these short-lived indicators, or they may be unsatisfactory due to their growth in warmer waters or nonculturability.
Such problems can only be ironed out by a coordinated national program specifically designed to answer the various aspects mentioned above. Taking primary recreational waters as an example, a nationwide database of faecal indicator bacteria for bathing waters could lead onto a better understanding of their distribution (seasonal trend and the impacts of rainfall events) and relevance in protecting public health. However, additional data, such as time of sampling, state of tide during sampling and preceding rainfall is also required along with financial resources to reduce such data into useful information.
Furthermore, a simultaneous step should be to coordinated epidemiological data gathered by health departments and major pathology laboratories to identify likely water or fisheries exposure routes. A reasonable attempt to collect such data has been made for example, in the United States (Moore et al. 1994). Next, specific epidemiological studies and statistical evaluation of risk exposure from measured concentrations of key pathogens and indicators is required to assess what organisms to use and most importantly, what level of risk should be targeted by the Australian Standards. Additional focused monitoring and quantitative risk assessment studies would also be required to understand the role of the nonfaecal pathogens.
While the above example was for primary contact waters, secondary contact standards could rationally be deduced from the primary standards for each major water type (tropical, temperate, ocean, estuarine). Standards for fish/shellfish and fisheries waters should be generated by an analogous program focusing on these foods and source waters and applying quantitative risk assessment methods (Rose & Sobsey 1993) to aid in the establishment of Australian standards.
Until such a program is undertaken, little progress can be made on rationally setting national standards or using appropriate indicators/pathogens for environmental reporting. Key data are available from overseas on likely target micro-organisms, doses required to cause infection and models to apply to estimate risk exposure. However, no progress has been made on establishing health-based microbiological standards for primary or secondary contact waters since the US-EPA program in the late seventies (Cabelli et al. 1979). Therefore, to develop standards suitable for Australian waters we need to use the new methods available in Australia for monitoring target micro-organisms and collate the health data available from the larger pathology laboratories. In summary, it appears that a national priority and appropriately coordinated financing are all that are required to produce health-based standards to protect our diverse range of marine waters.
Acknowledgments: I would like to acknowledge the constructive comments by Anne McNeill and Gary Grohmann on the previous draft of this manuscript. Views expressed in this review are not necessarily those of AWT Science & Environment.
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