Cranfield Centre for EcoChemistry Cranfield University, Silsoe, Beds MK45 4DT, UK Tel: 01525 863000 Fax 01525 863253 E-mail: [email protected]Web: http://www.cranfield.ac.uk/ecochemistry Higher Tier Laboratory Aquatic Toxicity Testing A Boxall, C Brown & K Barrett Final report March 2001 DETR Contract Reference No. EPG 1/5/131 Cranfield Centre for EcoChemistry Contract No. JA4317E
70
Embed
Higher Tier Laboratory Aquatic Toxicity Testing · Cranfield Centre for EcoChemistry 2 FOREWORD This report reviews the role of higher tier laboratory aquatic toxicity studies in
This document is posted to help you gain knowledge. Please leave a comment to let me know what you think about it! Share it to your friends and learn new things together.
Transcript
Cranfield Centre for EcoChemistryCranfield University, Silsoe, Beds MK45 4DT, UK
TABLE OF CONTENTS .....................................................................................................................................3
2.2 STUDIES USING ADDITIONAL SPECIES ......................................................................................................232.2.1 Selection of additional test species ...............................................................................................232.2.2 Analysis of data.............................................................................................................................272.2.3 Current limitations to the approach..............................................................................................292.2.4 Conclusions and recommendations ..............................................................................................29
2.3 ORGANISM RECOVERY ............................................................................................................................322.4 SYSTEM RECOVERY .................................................................................................................................332.5 SENSITIVE LIFE STAGE STUDIES ...............................................................................................................362.6 POPULATION LEVEL STUDIES...................................................................................................................382.7 INDOOR MULTI-SPECIES STUDIES .............................................................................................................41
3. APPLICATION OF HIGHER TIER LABORATORY STUDIES IN ENVIRONMENTAL RISKASSESSMENT....................................................................................................................................................43
3.1 USE OF FORMULATED PRODUCT OR ACTIVE INGREDIENT IN HIGHER TIER STUDIES ..................................483.2 IN-LIFE MONITORING OF CHEMICAL FATE ................................................................................................493.3 REPLICATION...........................................................................................................................................503.4 USE OF STATISTICS ..................................................................................................................................503.5 QUALITY CONTROL .................................................................................................................................52
4. RECOMMENDATIONS FOR FURTHER WORK ..............................................................................52
6.22) and fenvalerate (log Kow 5.0) have demonstrated that the presence of soil or sediment
in the test system results in a significant reduction in the observed effect of a pesticide (Table
4). This observation can be explained by the fact that very little pesticide (i.e. <1% for lamda-
cyhalothrin) is available in the water column due to partitioning of the compound from the
aqueous phase to sediment (Hamer et al., 1999). The addition of sediment had less effect on
the hydrophilic compounds isoproturon (log Kow 2.5) and pirimicarb (log Kow 1.7).
Cranfield Centre for EcoChemistry
16
Table 4 Changes in observed toxicity caused by a range of modifications to water-only exposure studies (Clark et al., 1989; Maund et al., 1998; Shillabeer
et al., 2000; Schulz and Liess, in press)
Reduction factor in toxicity or BCF relative to water-only study
Compound/test species Simulation ofspray drift
Simulation of spraydrift and suspended
sediment
Application to soilprior to addition of
water
Simulation ofsuspendedsediment
Simulation of bedsediment (not
stirred)
Simulation ofsediment (stirred)
Simulation ofeffects of soil
lambda cyhalothrin
Daphnia magna 3 40 175 - 120-140 81-280 -
glyphosate
Daphnia magna - - - 3.3 - - -
esfenvalerate
Daphnia magna
Lepomis macrochirus
- - - - 3.0
5.0
- -
fenvalerate
Limephilus lunatus
grass shrimp
- - - 10-100
-
- -
9,700
-
pirimicarb
Daphnia magna (BCF) - - - - 0.7 - -
isoproturon
Scenedesmus subspicatus - - - - - 2 -
herbicide 1
Selenastrum capricornutum
herbicide 2
Navicula pelliculosa
-
-
-
-
-
-
-
-
>17
>900
-
-
-
-
Cranfield Centre for EcoChemistry
17
Water/sediment systems have also been used to assess the impact of herbicides on
macrophytes. For example, Forney and Davis (1981) investigated the effects of herbicides on
aquatic plants. Studies with atrazine in soil/water systems demonstrated that atrazine
associated with the soil was not toxic because submerged plants lack a transpiration stream
whereas atrazine in the water column exerted a toxic effect.
The studies using the hydrophilic carbamate pesticide, pirimicarb (log Kow 1.7) have
demonstrated that whilst the presence of sediment had no effect on concentrations of the study
compound in the water column, the bioconcentration factor for Daphnia magna was reduced
(Kusk, 1996). This observation was possibly due to the effects of dissolved organic carbon
released from the sediment.
In the natural environment, water contains dissolved or colloidal organic matter (DOM).
These substances have been shown to interact with organic compounds resulting in effects on
bioavailability because only freely dissolved compounds are generally assumed to be
accumulated by organisms. Previous studies using pesticides, surfactants and polycyclic
aromatic hydrocarbons have demonstrated that the presence of dissolved organic carbon
reduces accumulation and toxicity (e.g. Landrum, et al., 1985; McCarthy et al., 1985; Black
and McCarthy 1988; Kukkonen and Oikari, 1991; Kukkonen and Pellinen, 1994; Oris et al.,
1990; Weinstein and Oris, 1999). However, at low concentrations of DOM (i.e. less than 10
mg/l) bioaccumulation may be enhanced (e.g. Haitzer et al., 1998).
In order to simulate the effects of natural concentrations of dissolved organic matter on
ecotoxicity, studies have also been performed using water samples collected in the field (e.g.
Isnard et al., 2000; S. Marshall, personal communication). Studies on a range of compounds
have demonstrated that generally the apparent toxicity of chemicals is only slightly reduced in
tests using natural waters (Isnard et al., 2000; Table 5). This suggests that the use of
laboratory water does not necessarily introduce a significant source of error. The use of this
type of study as a higher tier test may not therefore be particularly useful.
Cranfield Centre for EcoChemistry
18
Table 5 Mean ratios of EC50s obtained using tests on 15 natural waters to EC50s obtained using
ISO standard media (taken from Isnard et al., 2000)
Log Kow Raphidocelis
subcapitata
microplates
Raphidocelis
subcapitata
erlens
Brachionus
calyciflorus
Daphnia
magna
zinc sulphate na 5.7 3.0 2.1 1.2
4-nonylphenol 5.92 0.6 0.7 1.2 0.9
phosalone 4.30 0.7 1.2 1.1 1.7
pentachlorophenol 5.12* 2.7 2.1 1.2 3.1
2,4,5-trichloroaniline 3.45 1.9 2.2 2.1 1.0
Mean 2.3 1.8 1.5 1.6
Standard ecotoxicity studies usually require that the test concentration at the end of the study
is 80% of the starting concentration. For compounds that are volatile or which are degraded
abiotically (i.e. via hydrolysis) this can mean that the test system is covered and that a flow-
through test system is required or that the test solution is replaced regularly. Thus the impact
of abiotic degradation processes on toxicity could readily be assessed using a static exposure
system with no replacement of test solution.
The test matrix used in dissipation studies (e.g. sediment or natural water) should be selected
based on available data from fate studies and modelling work and should represent a realistic
‘worst case’. Two types of test matrix could be used, namely an artificial matrix (e.g. artificial
sediment) or a natural test matrix. Ideally, test sediment should be uncontaminated by
chemicals in toxic amounts, free of biota, and possess characteristics similar to the sediment
type of interest (Suedel et al., 1996). Formulated sediments have therefore been developed to
simulate field-collected sediments (e.g. Suedel and Rogers, 1994). Studies with copper
(Suedel et al., 1996) indicate that formulated sediments can be used as reference sediments.
The selection of the test matrix type will be dependent on the objectives of the higher tier test.
Natural materials will mimic actual conditions in the field more closely, whereas artificial
matrices are standardised and thus facilitate comparison of results with those from similar
studies. At present, higher tier regulatory studies seek to increase the realism of the test
system and the use of natural water and sediment materials might be preferred on this basis.
Single test systems should be based on sediment with realistic worst-case organic carbon
contents (1-2%), whereas studies with two sediment types (as for fate water-sediment
experiments) allow assessment of effects under a broader range of conditions. Artificial
sediments are continually improving and studies with such matrices are routinely accepted.
Cranfield Centre for EcoChemistry
19
If a natural test matrix is to be used, then it should be obtained from sites that are
uncontaminated and that have not previously been exposed to the test substance. It must also
be able to support the survival of the test organisms. The sampling methodology for both
water and sediment should be considered and involve minimum disturbance. If possible, the
test matrix should be initially characterised on-site in terms of dissolved oxygen and pH. The
matrix should be used as soon as practicable after collection. On return to the laboratory,
samples should be fully characterised and the characterisation should also reflect the objective
of the study. For example, if the effects of humic acids in the water column are being
investigated, then the dissolved organic carbon concentration of the test solution should be
determined.
2.1.4.1 Time-varying toxicokinetic models
The results of modified exposure studies can be used in conjunction with toxicokinetic models
to assess toxicant effects under non-steady state conditions. A range of approaches is
available (Table 6), including: compartment based (these describe the movement of toxicants
between compartments); physiologically based (describe accumulation, elimination and
distribution); and bioenergetic based (describe accumulation and loss in terms of the
organism’s energy requirements). To apply the models, information is required on body
residue levels and toxic effects. This is a neglected area in ecotoxicology and further work is
required before it can be incorporated into the risk assessment process.
Table 6 Toxicokinetic models that can be used in the assessment of results from modified
exposure studies
Model Model type Major inputs Reference
FGETS Bioenergetic-
based
MW, MP, Log Kow, fish
species, fish weight, water
temperature
Barber et al., 1988
PULSETOX Compartment
based
Kow, MW, HLC, organism
volume, lipid fraction, BCF,
uptake and depuration rates,
test concentration, water and
toxicant flow rate
Landrum et al., 1992;
Hickie et al., 1995
DEBtox Bioenergetics
based
Concentration of compound
in solution; uptake and
depuration rates
Kooijman and
Bedaux, 1996
Cranfield Centre for EcoChemistry
20
2.1.5 Conclusions and recommendations
A wide range of approaches is available for assessing the impact of fate processes on the
toxicity of a substance. These include:
1. analysis of core data to assess the time to effect – this approach is most appropriate for
chemicals that enter the environment in pulses or which dissipate rapidly;
2. standard tests without maintenance of test substance concentration – this approach is
appropriate for volatile compounds or compounds that readily hydrolyse;
3. pulsed exposure studies – these are appropriate for chemicals that enter the environment
in pulses;
4. simulation of spray drift;
5. simulation of the effects of sediment or suspended sediment – for compounds that readily
partition to solids and/or which are readily biodegradable; and
6. tests using natural water – appropriate for highly hydrophobic substances that may
associate with dissolved organic carbon.
Differences between the results of standard and modified exposure studies can be used to
demonstrate the influence of a particular fate process or dissipation of a compound. Generally,
the results should be expressed as an initial toxicant concentration and then compared to an
initial PEC. Care should be taken to ensure that the process studied has not been incorporated
into the PEC calculation. The uncertainty factors associated with the preliminary risk
characterisation should be used.
Cranfield Centre for EcoChemistry
21
Table 7 Modified exposure studies used for the assessment of chemicals that were identified in the survey
Test title Species Culture source Organism age Brief methodology Rationale for test QA Procedures +statistics
References
Modified exposure Daphnia magna Laboratory < 24 h old Acute 48 h in awater/sedimentsystem
Incorporatesadsorptionprocesses
GLP with EC50
determinationConfidential report
Modified exposure Daphnia magna Laboratory <24 h old 21 d reproductionin a water/sediment system
Incorporatesadsorptionprocesses
GLP with EC50 andNOECdetermination
Confidential report
Modified exposure Daphnia magna Laboratory Day 6 gravid 21 d reproductionin a water/sediment system
Incorporatesadsorption process
GLP with EC50 andNOECdetermination
Confidential report
Modified exposure Oncorhynchusmykiss
Laboratory Juvenile Acute 96 h in awater/sedimentsystem
Incorporatesadsorptionprocesses
GLP with LC50
determinationConfidential report
Modified exposure Lemna minor Laboratory Mature plant 14 d exposure andrecovery in awater/sedimentsystem
growth inhibition in lightgrowth inhibition in dark
protozoa Tetrahymena pyriformis growth inhibition
rotifera Brachionus calyciflorus mortalityswimming activityfeeding activitypopulation growthfull life cyclepartial life cyclemultigeneration life cycle
Field Juvenile Acute 48 h Establishesspecies sensitivity
GLP with EC50
determinationConfidential report
Cranfield Centre for EcoChemistry
32
2.3 Organism recovery
A number of approaches have been used to assess the potential for an organism to recover
after exposure to a toxicant (e.g. Table 12). Assessment of recovery is a standard component
of the lower tier toxicity tests for algae and the higher plant, Lemna. An aliquot of previously
exposed cells or fronds is placed into clean media to establish whether cell division is
reversibly or irreversibly retarded. However, the principle can be extended to aquatic plants,
invertebrates and to sub-lethal effects in fish (e.g. swimming abnormalities, colouring,
lethargy or impaired feeding and growth). A few studies have been reported for pesticides,
including organophosphorous and carbamate pesticides and for organisms including daphnids,
chironomids and blackfly (e.g. Kallander et al., 1997; Sanchez et al., 1999).
The potential for recovery will be determined by the duration before further exposure and the
toxic mode of action of the compound being considered. For example, studies into the effects
of carbamate pesticides on chironomids indicated that toxic effects were reduced if animals
were allowed to recover for more than 6 hours (Kallander et al., 1997). This observation was
probably due to the reactivation of acetylcholinesterase during the recovery period –
carbamate pesticides have been shown to be reversible inhibitors of AChE. No such recovery
was observed for organisms exposed to organophosphorus insecticides which are considered
irreversible inhibitors (Kallander et al., 1997). RADAR is a useful tool to evaluate duration
of pesticide exposure and time for recovery which has been proposed by ECOFRAM
(Hendley & Giddings, 1999).
The results of organism recovery studies could be used to refine the toxicity value used in the
risk assessment process. For example, if a compound is shown in standard toxicity studies to
affect a test organism at a particular concentration but is able to recover when subjected to a
more realistic exposure scenario, the effect value used in the assessment could be raised. One
problem with this approach is in demonstrating the relevance of recovery in the laboratory for
a whole community and this could be particularly so where species replacement occurs.
Cranfield Centre for EcoChemistry
33
2.4 System recovery
Once a system has been impacted by a contaminant, its rate of recovery will be determined by
the persistence of the contaminant and the ecology, physiology and biochemistry of organisms
in the system and in proximity to the system.
One approach to predict the likely recovery of a system is based on the dissipation half-life of
the compound, the initial exposure concentration and the hazardous concentration for 5% of
species (Van Straalen et al., 1992). If recovery is required within a year after a single
application, then the chemical’s half-life should meet the condition:
<
5
1/2
ln
2ln
HC
CT
o
where T1/2 is measured in years, C0 is the initial concentration and HC5 is the hazardous
concentration for 5% of species. Although degradation will be necessary for complete
recovery, under some conditions it may not be sufficient and ecological recovery may lag
behind the disappearance of the chemical. This lag will depend on a range of factors
including: accessibility of shelters; survival of resistant life stages; and presence of untreated
areas from which recolonisation can occur. In addition, ability to recolonise will depend on
habitat selection and life history
A number of experimental studies have been proposed for determining whether an impacted
system is likely to be re-colonised. This is particularly relevant for mobile organisms such as
fish which may be resident in a section of a water body for a short time and may be able to
move to leave a toxic system. The study design consists of adding the test organism at the
start of the study and at regular intervals thereafter. A similar approach can also be used with
those invertebrate species with a high potential for re-colonisation, with the re-introduction of
individuals into a previously treated system, and monitoring of their subsequent survival and
performance. For example, Crane et al. (1999) investigated the toxicity of pirimiphos-methyl
to Gammarus pulex. Beakers were spiked with the study compound at day 0 and toxicity to
Gammarus pulex was assessed over 24 h at 1, 4, 8 and 12 d after application. Generally
mortality decreased with increasing time from spiking, probably as a result of pesticide
degradation.
Cranfield Centre for EcoChemistry
34
The factors that influence the recovery of a population after a significant perturbation are
complex. One important factor affecting recovery is the life history of the individual species
(Sherratt et al., 1999). Aquatic organisms can vary widely in their reproductive and dispersal
characteristics. For example some species may reproduce continuously whereas others may
reproduce at discrete times of the year. Generation times can vary from days (e.g. daphnids)
to years (e.g. odonates). The ability for recolonisation can also vary - many aquatic insects
have adult life stages with wings whereas other taxa (e.g. crustaceans and molluscs) will rely
on recolonisation by more passive forms of dispersal (wind, flooding, transportation by birds).
The method for inclusion of system recovery in the risk assessment process is likely to be
dependent on the nature of the organisms most at risk. For example, if a particularly sensitive
group has a high potential for recolonisation, then the refinement of the effect concentration
used in the risk assessment process may be justified. If there is low potential for
recolonisation, then the original effect concentration should be used. A wide range of factors
will affect recolonisation potential and many of these are not fully understood. Significant
work is therefore required in this area before useful guidelines on application to the risk
assessment process can be developed.
Cranfield Centre for EcoChemistry
35
Table 12. Higher tier laboratory studies identified in the survey that are performed to assess the potential for organisms to recover following exposure to toxicants
Test title Species Culture source Organism age Brief methodology Rationale for test QA Procedures +statistics
References
Recovery Selenastrumcapricornutum
Laboratory Exponential growth Cell densitydetermined during4 d exposure toseveral testconcentrations.Aliquot of culture isthen taken fromselected testconcentrations anddiluted at leat 100xin fresh media andalgal density isthen meaured overseveral days
Determines if algalgrowth resumesonce testconcentration fallsbelow a NOEC
3 replicates,calculate areaunder growth curveand growth rate c.f.untreated control
Internal industryreport
Recovery Lemna minor Laboratory Mature plant 7 and 14 dexposure andrecovery
Incorporatesrecovery potential
GLP with EC50 andNOECdetermination
Confidential report
Recovery Lemna gibba orLemna minor
Laboratory Rapidly dividing Frond numberdetermined over 4-14 d exposure toselected testconcentrations.Equal number offronds/plantsremoved to freshmedia and frondnumberdetermined forseveral days
Determine whetherincrease in frondnumber resumesafter testsubstanceremoved
144 or 400 cm2 with 5cm deep sediment layerand sewater with 1.6L/min/tube flow.
chlorpyrifos Flemer et al., 1997
Communityrepresentative ofepilimnon of amesotrophictemperate lake
3 l; 3 species of algaemix of lake bacterialspecies
atrazine Genoni, 1992
Alagae + benthicinvertebrates
120 l re-circulatingsystem; field collectedinvertebrates andperiphyton
atrazine Gruessner and Watzin,1996
Algae + daphniamagna
100 ml of watercontaining 1.2 x 105
algal cells/ml + 1daphnid
chromium Gorbi and Corradi,1993
Cranfield Centre for EcoChemistry
43
3. APPLICATION OF HIGHER TIER LABORATORY STUDIES IN
ENVIRONMENTAL RISK ASSESSMENT
Generally, higher tier studies should be carried out in response to indications of adverse risk
at lower tiers. It is assumed that a full and critical evaluation of the baseline fate and effects
data package precedes any move to higher tier studies. Higher tier laboratory methods can be
used to:
• reduce the magnitude of uncertainty factors in hazard/risk assessment;
• provide a more realistic assessment of effects on a particular species;
• set dose rates, duration and identity of key species of concern for multi-species testing;
• provide input data to ecological models;
• assess compounds with a high Koc and/or affinity for binding to sediment/humic acids;
• provide information on the potential duration of effects;
• assess whether effects are reversible.
The exact nature of the test to be performed will be dependent on a number of factors
including: the results of lower tier studies; the physicochemical properties of the compound of
interest; and the degradability of the study compound. There are currently no guidelines
describing the incorporation of higher tier laboratory studies into the risk assessment process.
However, it would seem appropriate to use a stepped system, involving the initial
interrogation of core data, followed by the identification of sensitive species and finally a
combination of modified exposure studies and population studies on the identified species. A
summary of the major higher tier studies identified is provided in Table 16. The combination
of a number of the approaches (e.g. use of population studies coupled to pulsed exposure and
dissipation studies) may also be appropriate.
Higher tier laboratory approaches have a number of advantages over the use of standard
ecotoxicity tests and/or mesocosm/field studies. These include:
• the endpoints are more pertinent to actual environmental exposure;
• they provide scope for the integration of ‘realistic’ measures of fate and exposure into the
risk assessment process;
• they should provide more confidence when predicting actual effects in the environment;
Cranfield Centre for EcoChemistry
44
• they can be designed to determine effects on specific endpoints, e.g. mortality, growth,
behaviour, population level;
• they can be used to identify more clearly those species that may be at risk and assist in the
targeting of multi-species tests;
• the results may help explain observed effects (e.g. mode/specificity of action);
• recovery can be assessed for both individuals and populations;
• there are none of the complications that are associated with the interpretation of data from
multi-species studies;
• relative to mesocosm studies, they can be performed with less regard to the season, the
test organisms are usually readily available and the system is more controlled;
• sublethal effects can be determined.
However the approaches also have limitations:
• they cannot provide information on species interactions;
• results can be more difficult to interpret and compare than those from standard single
species tests;
• the tests may be more costly than standard single species studies;
• the test methods are undefined;
• the quality of available test species cannot always be guaranteed.
Cranfield Centre for EcoChemistry
45
Table 16. Higher tier toxicity methods
Test type Rationale for test Design of test Data analysis Output Advantages/Disadvantages
Testing ofadditonalspecies
To reduce uncertaintyfactors and/or developspecies sensitivitydistributions
Test a minimum of 8 additionalspecies: –
From a range of groups
or
For compounds whose mode ofaction indicates that a particulargroup will be sensitive, testsshould be performed on speciesfrom this group
or
If fish are likely to be sensitive,then only 5 additional speciesshould be tested
Lower uncertainty factorsby up to an order ofmagnitude or use statisticalapproaches to generatespecies sensitivitydistributions
Revised TER or HCx • identification of most sensitivespecies
• more relevance than standardspecies
• refinement of TERs• allows probabilistic risk
assessment
• no guidance available onuncertainty factors to use
• no guidance available on whatlevel of population protection isacceptable
• lack of standard test methods foradditional species
• may need to obtain organismsfrom the field
Time-to-eventanalysis
To assess the toxiceffects of exposure topesticides overdurations shorter thanthe standard testduration
Use data recorded over time forstandard ecotoxicity tests
Compare time-to-eventdata with expectedexposure duration
If duration < time toeffect then reduceeffect estimate
• more realistic assessment ofecotoxicity
• resolution of observations instandard tests may not beappropriate
• no guidance on how effectsestimate can be reduced
Cranfield Centre for EcoChemistry
46
Test type Rationale for test Design of test Data analysis Output Advantages/Disadvantages
Short-termexposurestudies
To assess effects ofshort exposuredurations on pesticidetoxicity
Perform test over exposureduration predicted for the field
Use standard toxicity teststatistical procedures
Revised effectmeasurement for usein risk characerisation
• more realistic assessment ofexposure
• does not consider delayed effects
Pulsed-exposurestudies
To simulate pulsedexposure conditionsthat are likely in the field
Static renewal or flow-throughsystem, depending on detailrequired
Use standard toxicity teststatistical proceduresand/or toxicokinetic models
Revised effectmeasurement for usein risk characerisation
• more realistic assessment ofexposure
• includes potential for organismsto recover
• results affected by a number ofvariables (including pulseduration, time between pulsed,organism recovery) so difficult tointerpret
Dissipationstudies
To assess the impact ofdissipation processeson toxicity
Generally performed usingsediment/water systems.Studies could also beperformed to assess effects ofphotodegradation (byperforming studies in the light)or volatilisation.
Use standard toxicity teststatistical procedures
Revised effectmeasurement for usein risk characerisation
• more realistic assessment ofexposure
• can include effects of metabolites• guidelines not available on
selection of test matrices
• not suitable for all species (e.g.algae)
• methods not available forassessing photodegradation andvolatilisation
Cranfield Centre for EcoChemistry
47
Test type Rationale for test Design of test Data analysis Output Advantages/Disadvantages
Studies usingnatural matrices
Account for differencesbetween bioavailabilityin natural and laboratorysystems
Perform toxicity studies usingwater samples collected fromthe field
Use standard toxicity teststatistical procedures
Revised effectmeasurement for usein risk characerisation
• assessment of bioavailabilityunder natural conditions
• no guidance available onselection of test matrix
Effects onsensitive lifestages
To determine whether aparticular life stage issensitive or not
A range of approaches areavailable for a number ofspecies.
Use standard toxicity teststatistical procedures
Toxicity to sensitivelife stages – possiblereduction inuncertainty factor
• reduction of uncertainty factors
Populationstudies
Perform studies on apopulation of aparticular species overa prolonged time period
Experimental and modellingapproaches are available
Use population modellingapproaches to interpretresults
More ecologicallyrelevant assessmentof effects – possiblereduction inuncertainty factor
• includes effects of recovery• includes selection of tolerant
organisms
• may be more ecologicallyrepresentative
• difficult to perform on fish
Recoverystudies
To determine whetherpopulations can recoverafter exposure to apesticide
Expose organisms and transferto clean media to determinetime to recovery
Assessment of timerequired fororganisms to recover– within 2 monthsmay be acceptable(CLASSIC, in prep.)
Cranfield Centre for EcoChemistry
48
When designing higher tier tests for use in the risk assessment of chemicals, a number of
factors should be considered. These include: whether formulated or technical product should
be use; in-life monitoring of chemical concentrations in the test; replication and statistics; and
quality control. The approaches used to address each of these issues will vary according to the
nature of the test.
3.1 Use of formulated product or active ingredient in higher tier studies
A number of studies have investigated the effects of co-formulants on pesticide toxicity (e.g.
Zitko et al., 1979; Bradbury et al., 1985; Holdway et al., 1994). The results indicate that
formulations can be either more or less toxic than the active substance. For example, technical
grade esfenvalerate and fenvalerate were 1.7 and 2.3 times more toxic than emulsified
formulations containing these active substances. The results were probably due to a slower
rate of uptake of the pesticides from the formulated product (Holdway et al., 1994). In
contrast, studies on one emulsified pesticide indicated that this was more toxic than the
technical grade pesticide, possibly due to a longer half-life as a result of the emulsifier. If the
formulated product is more toxic, it should be used in the higher tier studies.
Generally, it is expected that formulated material will be used for higher tier studies as it will
better mimic drift inputs to surface waters. Dosing to nominal concentrations above the limit
of solubility makes interpretation of test results difficult and may result in physical effects (e.g
formulated material in excess of solubility may form a film on the water surface); such doses
are only appropriate in support of a particular application of concern. Formulated product is
also preferred where the formulation has been shown to have a significant impact on exposure
and/or effects. For example, it may affect the stability and residence time of the compound in
the water column and thus increase bioavailability. If formulated product is used, testing of a
formulation blank may be appropriate where levels of toxicity are observed that were not
anticipated. However, use of such a blank will normally be triggered at lower tiers of the
testing procedure.
The technical product may be easier to work with in some instances. A co-solvent can be
used to aid dispersion. In addition, the results of studies on technical product can be
extrapolated to different formulations of the same product. Technical product is likely to be
preferred for species sensitivity tests as this will allow for a direct comparison with the
available Tier I data, and again for extrapolation to a range of formulations. Where toxicity
tests are mimicking an input of chemical to surface water via overland flow or drainflow, it is
recommended that technical product should be used. Although it is not clear from fate studies
Cranfield Centre for EcoChemistry
49
how long a chemical and its formulation are likely to remain associated after introduction into
the environment, the current assumption is that the two are instantaneously decoupled upon
reaching the soil and that the formulation plays no further part in the behaviour of the active
substance.
3.2 In-life monitoring of chemical fate
Monitoring of chemical fate over the course of a particular higher tier laboratory toxicity test
is generally required as at lower tiers. This is true for tests both with maintained
concentrations (e.g. life cycle tests, species sensitivity tests) and those with a degradation or
removal process. Monitoring will show the effect of any modified design on the behaviour of
a compound and also allow comparisons between results of laboratory toxicity tests at lower
and higher tiers.
Frequency of in-life monitoring of chemical fate needs to be appropriate to the test system.
For example, more frequent monitoring will be required if chemical is added in a series of
pulses (e.g. to simulate drift from multiple applications) than if a single dose of chemical is
added. In certain cases (e.g. the introduction of a sediment layer), the higher tier toxicity test
will closely match a fate and behaviour study (e.g. a water-sediment study) and in-life
monitoring of fate may be reduced to the minimum necessary to show similar behaviour of
the chemical. This is only the case if the system designs are closely related.
Where tests examine effects of a chemical at a particular concentration (e.g. a PEC), it would
be appropriate to define a target threshold whereby aqueous concentrations should be within a
certain percentage range of the target value. Use of radiolabelled compound may prove an
effective way of monitoring fate and distribution at low levels where sufficiently sensitive and
specific methods of analysis may not be available. It is important to remember that for many
systems the size will not allow for removal of adequate sample for analysis. This may mean
that analysis has to be limited to the start and end of the experiment, or alternatively that
additional units should be set up and taken for destructive analysis at timed intervals.
Cranfield Centre for EcoChemistry
50
3.3 Replication
As a general principle, the replication specified in existing standardised guidelines for lower
tiers of testing should be adhered to, wherever the study designs are comparable. However,
there may be practical issues which prevent this – for example, the availability of some field
collected animals (e.g. dragon fly larvae) may prohibit testing on this species with the same
replication as prescribed for a standard invertebrate test such as Daphnia. In addition, their
carnivorous nature may dictate a different holding regime of individually held animals.
Generally, as the test system becomes more natural and complex, the response is likely to
become more variable and the level of replication required is likely to increase. A decision on
replication required should be made in consultation with a statistician.
An important, related point is that acceptance criteria in the form of mortality thresholds
should not be prescriptive for higher tier tests because of additional difficulties in handling.
Decisions on the acceptability and validity of a study should be based on comparison with an
appropriate control system run in parallel. As indicated earlier, the levels of background
mortality experienced with these non-standard higher tier tests is likely to be greater and this
should be taken into consideration. Statistical aspects of replication are briefly considered in
Section 3.4 below.
3.4 Use of statistics
Two statistical activities are involved in ecotoxicity testing, namely: experimental design and
data analysis. Of these, experimental design is the most important and the following should
be considered when higher tier studies are being designed (Chapman et al., 1996):
Randomisation – treatments should be allocated to experimental units in a random fashion as
the experimental variation in higher tier studies may be large (due for example to the use of
natural populations or test matrices). Moreover, all handling of experimental units after the
start of a study should also be performed in a random fashion.
Replication – replication allows the power or precision of a test to be controlled. Larger
numbers of replicates will lead to more precise estimates or tests with greater power. If
ANOVA is the likely form of data analysis, then the higher tier test procedure should specify
the required precision and power so that the most appropriate number of replicates can be
chosen. If a dose response model is to be fitted, then the precision of the numbers of
Cranfield Centre for EcoChemistry
51
concentrations tested will also affect an estimate of effective concentrations (e.g. EC50,
LC50).
Number of organisms per unit - Having more than one organism per experimental unit will
usually improve the power and precision of the test. The extent of the improvement will be
dependent on the size of the within-unit variation. An assessment of the within-unit variation
associated with a range of higher tier approaches would therefore be useful in order to identify
the most approproate number of test organisms to be used in a particular test.
Number and spacing of concentrations – the choice of concentrations, both their number and
their value, will affect the precision of the LC and EC estimates. Standard test guidelines
currently require four or five geometrically spaced concentrations. Five test concentrations are
also likely to be appropriate for higher tier studies.
Optimum times for taking measurements – standard guidelines often require measurements to
be made at more than one time point. The timing and frequency of measurements in a higher
tier study may well be more critical. For example a pulsed exposure may require many
observations.
Blind assessing – the endpoints used in certain higher tier studies may be quite subjective.
For example, evaluation of effects such as lethargy or swimming abnormalities in sublethal
tests. Under these circumstances, blind assessment of test vessels is recommended.
Use of linear regression to generate a dose response is generally preferred to an ANOVA
approach to experimental design, although both options are available in the OECD guideline
for mesocosm studies. ANOVA requires a larger number of replicates and the value of
NOECs has recently been questioned due to the different design requirements of dose-
Estimation of an effective concentration (EC) using regression analysis has a number of
advantages over the derivation of NOECs: the EC is not restricted to being one of the test
concentrations; its value does not depend on the precision of the experiment; its precision can
be estimated; its interpretation is straightforward; and regression modelling uses data from all
concentrations (Chapman et al., 1996).
Cranfield Centre for EcoChemistry
52
3.5 Quality control
Higher tier studies should always be conducted to GLP. The standard 10% and 20% mortality
thresholds are likely to be exceeded in some non-standard studies. However, acceptable
thresholds cannot be established without ring tests to investigate inter- and intra-laboratory
variation. This is considered impractical because of the inherent flexibility and wide range of
available study designs at higher tiers. In the absence of appropriate mortality thresholds for
higher tier tests, expert judgement should be used in deciding the acceptability of a given
study relative to an appropriate control system. Results should be adequate to provide a valid
statistical evaluation.
A potential tool in quality control which is little used at present is the parallel testing of a
reference toxicant. In attempting to make higher-tier tests more like the real world, there is a
clear need to evaluate the success of a particular system. Reference toxicants offer the
opportunity to compare behaviour in the test system with known behaviour in the wider
environment and thus to establish the validity of the higher tier study. A suite of reference
toxicants would be required covering a range of physico-chemical properties and degradation
rates so that the chemical with properties most similar to the test compound could be selected.
4. RECOMMENDATIONS FOR FURTHER WORK
Higher tier laboratory tests provide a useful intermediate between standard toxicity studies
and semi-field / field tests. The review presented here has shown that there are a wide range
of higher tier studies available and that a significant body of work has been generated on
most, but not all, of these methods. Three main priorities are identified to take this area of
work forward and improve the regulatory acceptability of higher tier laboratory tests:
1. There is currently no guidance available to support the use of higher tier laboratory tests
and a priority is to put in place agreed documentation to do this. Aspects which would
need to be covered within the guidance include: a) a scheme of how and when to use each
particular test; b) major constraints and methodological pointers for each test type; c) how
to interpret the results; d) acceptability of the data; and e) how the results should be
incorporated into the risk assessment process, including indications of appropriate safety
factors to apply. Flexibility is an intrinsic component of higher-tier testing and it is
Cranfield Centre for EcoChemistry
53
envisaged that the documentation would set broad constraints around the use and
interpretation of tests rather than a rigid set of guidelines. Given the complexity of the
issues involved and the range of tests available, it is likely that most benefit will be gained
through a joint initiative between scientists from regulatory agencies, industry and
research organisations. SETAC or a similar body may offer a useful, common forum for
discussion as with the HARAP workshop.
2. A number of higher tier tests are available and there is a need to validate these methods in
a consistent and systematic way. To do this, comparisons should be undertaken between
data and risk assessments generated from standard laboratory testing, higher-tier
laboratory testing and mesocosms. This would allow evaluation of the advantages and
limitations of laboratory higher tier testing compared to standard studies and mesocosm
studies. Costs involved in carrying out validation could be minimised if case studies were
performed on well studied compounds where some laboratory higher tier data are already
available (e.g. atrazine, lamda-cyhalothrin, chlorpyrifos). Supplementary data could then
be generated to fill any data gaps before the risk of the compound was characterised.
3. It would be greatly beneficial if a means could be found to pool expertise and knowledge
on working with higher tier tests and with particular aspects such as studying non-
standard species. It is not realistic to expect ring-tested methods to be available for higher
tier tests even in the medium term. However, informal meetings could be held in a
manner similar to the Joint Initiative on Beneficial Insects which led to the production of
a series of guidance notes. The aim would be to produce guidance documents to improve
the quality of data produced and harmonise aspects of procedure where appropriate. One
possibility for the UK would be to re-start the UK Aquatic ad-hoc group to address this
issue. Pesticides Safety Directorate (PSD) and DETR could play an important role in
driving the process, but the main sources of practical experience will lie within industry
and contract laboratories.
A number of secondary requirements for further work are set out below:
1. For assessments of species sensitivity, the number of additional species for testing has
been identified. However, the majority of available test methods (Appendix B) have
generally been developed in North America. The suitability of these tests relative to UK
aquatic systems needs to be assessed and recommendations should be provided on the
most appropriate test species to use for additional studies. Test methods for these
additional species may need developing. The development of geographical information
Cranfield Centre for EcoChemistry
54
systems on the ecology of water bodies in agricultural areas should also be considered.
These systems would enable sensitive species to be readily identified for testing and assist
in setting priorities for method development.
2. Statistical analyses of species sensitivity datasets result in a hazard concentration for a
fixed percentage of species. There is some conflict in the literature with published studies
indicating that either a 5% or a 10% level is appropriate to protect aquatic systems. A
small study should be undertaken to support the selection of a common value and to
demonstrate its potential to protect the environment. A large volume of data (e.g. the
AQUIRE database, the EAT database, industry held data) is available that could enable
this work to be performed.
3. Datasets on the ecotoxicity of chemicals to species should be further analysed based on
information on the mode of action of a compound. The development of lists of sensitive
species associated with a particular mode of action would be highly beneficial when
selecting organisms for testing.
4. The results of modified exposure studies are likely to be highly dependent on the test
matrix. For sediment/water studies, research should focus on the effects of variability in
sediment characteristics on test results so that guidelines to standardise properties can be
put in place.
5. When a large dataset is available on the ecotoxicity of a compound, statistical approaches
should be used to assess effects levels for use in the risk assessment process. If few data
are available, then the assessment factor approach should be used. This latter approach
has a number of limitations including a lack of stability and no improved precision when
more species are tested. Work is required to develop a better approach for analysing
small datasets. One possibility would be a method based on assessment factors derived
from results of statistical approaches (Roman et al., 1999).
6. Current methods for assessing effects on communities from distributions of species
sensitivity assume that all species are equally important. This is probably not the case and
consideration should be given to identifying key species in aquatic ecosystems. In part,
this may be an output from a current MAFF/PSD project on “Aquatic ecosystems in the
UK agricultural landscape”. Approaches should be developed for ensuring that key
species are protected by risk assessments.
Cranfield Centre for EcoChemistry
55
7. A range of modelling approaches has been identified to support/complement higher tier
toxicity studies. There is a need to validate models in order to allow better predictions of
effects and extrapolations to other scenarios in the future.
8. A number of multispecies studies have been identified . It may be possible to refine these
methods based on a knowledge of major interactions observed in the field or large
mesocosm studies. The use of food web models to assess the likely impact of a compound
should also be considered.
9. In situ toxicity tests are outside the scope of this review. A number of recent publications
suggest that such studies may provide a useful supplement for higher tier laboratory tests.
ACKNOWLEDGEMENTS
The authors would like to thank Dr Mike Collins and Dr Steve Norman for useful dialogue
during the course of the study and Professor Allen Burton, Dr Andy Girling, Dr David Pascoe
and Dr Martin Streloke for their constructive comments on the draft final manuscript for this
project. This study was funded by the Department of Environment, Transport and the
Regions.
Cranfield Centre for EcoChemistry
56
REFERENCES
1. ABEL, P.D. (1980) A new method for assessing the lethal impact of short-term high-level discharges ofpollutants on aquatic animals. Prog Water Technol. 13: 347-352.
2. ACEVEDO, M.F., WALLER, W.T., SMITH, D.P., POAGE, D.W., MCINTYRE, P.B. (1995) Modellingcladoceran population to stress with particular reference to sexual reproduction. Nonlinear World 2: 97-129.
3. ALABASTER, J.S., ABRAM, F.S.H. (1965) Development and use of a direct method of evaluating toxicityto fish. In The Proceedings of the second international water pollution research conference, Tokyo, 1964.Pergamon Press, NY.
4. ALDENBERG, T., SLOB, W. (1993) Confidence limits for hazardous concentrations based on logisticallydistributed NOEC toxicity data. Ecotoxicology and Environmental Safety 25(1): 48-63.
5. ALLISON, P.D. (1995) Survival analysis using the SAS system. A practical guide. SAS Institute, Cary,NC.
6. AMERICAN SOCIETY FOR TESTING MATERIALS (1992) Standard guide for conducting acute toxicitytests with fishes, macroinvertebrates and amphibians. Designation E 729-88a Annual book of ASTMmethods, Vol. 11.04:403-422.
7. AQUATIC DIALOGUE GROUP (1994) Pesticide risk assessment and mitigation. SETAC Press,Pensacola, FL.
8. AXELSEN, J.A., HOLST, N., HAMERS, T., KROGH, P.H. (1997) Simulations of the presator-preyinteractions in a two species ecotoxicological test system. Ecological Modelling 101: 15-25.
9. BARBER, M.C., SUAREZ, L.A., LASSITER, R.R. (1988) Bioconcentration of nonpolar organics by fish.Environ. Toxicol. Chem. 7: 545-558.
11. BARRY, M.J., LOGAN, D.C. (1998) The use of temporary pond microcosms for aquatic toxicity testing:direct and indirect effects of endosulfan on community structure. Aquatic Toxicology 41: 101-124.
12. BAVECO, J.M., DEROOS, A.M. (1996) Assessig the impact of pesticides on lumbricid populations: anindividual-based modelling approach. Journal of Applied Ecology 33: 1451-1468.
13. BEATTIE, J.H., PASCOE, D. (1978) Cadmium uptake by rainbow trout, Salmo gairdneri eggs and alevins.J. Fish Biol 13: 631-637.
14. BELANGER, S.E. (1992) Use of mesocosms in predicting ecosystem risk from cationic surfactantexposure. In: Cairns, J., Neiderlehner, B.R., Orvos, D.R. (Eds) Predicting ecosystem risk. PricetonScientific, Princeton, NJ.
15. BELANGER, S.E., GUCKERT, J.B., BOWLING, J.W., BEGLEY, W.M., DAVIDSON, D.H., LEBLANC,E.M., LEE, D.M. (2000) Responses of aquatic communities to 25-6 alcohol ethoxylate in model streamecosystems. Aquatic Toxicology 48: 135-150.
16. BERRILL, M., BERTRAM, S., PAULI, B., COULSON, D., KOLOHON, M., OSTRANDER, D. (1995)Comparative sensitivity of amphibian tadpoles to single and pulsed exposures of the forest-useinsecticide fenitrothion. Environmental Toxicology and Chemistry. 14(6): 1011-1018.
17. BERRILL, M., COULSON, D., MCGILLIVRAY, L., PAULI, B. (1998) Toxicity of endosulfan to aquaticstages of anuran amphibians. Environ. Toxicol. Chem. 17(9): 1738-1744.
18. BLACK, M.C., MCCARTHY, J.F. (1988) Dissolved organic macromolecules reduce the uptake ofhydrophobic organic contaminants by the gills of rainbow trout (Salmo gardneri). EnvironmentalToxicology and Chemistry 7: 593-600.
19. BLANCK, H. (1984) Species dependent variation among aquatic organisms in their sensitivity tochemicals. Ecol. Bull. 36: 107-119.
21. BLOCKWELL, S.J., MAUND, S.J., PASCOE, D. (1999) Effects of the organochlorine insecticide lindane(γ-C6H6Cl6) on the population responses of the freshwater amphipod Hyalella azteca. EnvironmentalToxicology and Chemistry 18(6): 1264-1269.
22. BRADBURY, S.P., COATS, J.R., MCKIM, J.M. (1985) Differential toxicity and uptake of two fenvalerateformulations in fathead minnows. Environmental Toxicology and Chemistry 4: 533-541.
23. BRECK, J.E. (1988) Relationships among models for acute toxicity effects: applications to fluctuatingconcentrations. Environmental Toxicology and Chemistry 7: 775-778.
25. BROWN, V.M., JORDAN, D.H.M., TILLER, B.A. (1969) The acute toxicity to rainbow trout offluctuating concentrations and mixtures of ammonia, phenol and zinc. J. Fish. Biol. 1: 1-9.
26. BURNISON, B.K., HODSON, P.V., NUTTLEY, D.J., EFLER, S. (1996) A bleached-kraft mill effluentfraction causing induction of a fish mixed function oxygenase enzyme. Environ. Toxicol. Chem. 15:1524-1531.
27. CAIRNS, J. (1986) The myth of the most sensitive species. BioScience 36(10): 670-672.
28. CALOW, P., SIBLY, R.M., FORBES, V.E. (1997) Risk assessment on the basis of simplified life historyscenarios. Environ. Toxicol. Chem. 16: 1983-1989.
29. CAMPBELL, P.J., ARNOLD, D.J.S, BROCK, T.C.M., GRANDY, N.J., HEGER, W., HEIMBACH, F.,MAUND, S.J. & STRELOKE, M. (1999). Guidance document on higher-tier aquatic risk assessment forpesticides (HARAP). SETAC-Europe, Brussels, 179p.
30. CHANDLER, G.T., GREEN, A.S. (1995) A 14-day harpacticoid copepod reproduction bioassay forlaboratory and field contaminated muddy sediments. In: Ostrander, G. (ed) New techniques in aquatictoxicology. Lewis Publishers, Boca Raton.
31. CHAPMAN, P.F., CRANE, M., WILES, J., NOPPERT, F., MCINDOE, E. (1996) Improving the quality ofstatistics in regulatory ecotoxicity tests. Ecotoxicology 5(3): 169-186.
32. CHAPMAN, P.F., CRANE, M., WILES, J.A., NOPPERT, F., MCINDOE, E.C. (1996) Asking the rightquestions: ecotoxicology and statistics. SETAC-Europe, Brussels.
33. CLARK, J.R., GOODMAN, L.R., BOTHWICK, P.W., PATRICK, J.R., CRIPE, G.M., MOODY, P.M.,MOORE, J.C., LORES, E.M. (1989) Toxicity of pyrethroids to marine invertebrates and fish: a literaturereview and test results with sediment-sorbed chemicals. Environ Toxicol Chem 8: 393-401.
34. CORRELL, D. L., WU, T. L. (1982) Atrazine toxicity to submerged vascular plants in simulated esturinemicrocosms. Aquat. Bot. 12: 151-158.
35. COX, D.R., OAKES, D. (1994) Analysis of survival data. Wiley and Sons, New York.
36. CRANE, M. (1997) Research needs for predictive multispecies tests in aquatic toxicology. Hydrobiologia346: 149-155.
37. CRANE, M., ATTWOOD, C., SHEAHAN, D., MORRIS, S. (1999) Toxicity and bioavailability of theorganophosphorous insecticide pirimiphos methyl to the freshwater amphipod Gammarus pulex L. inlaboratory and mesocosm systems. Environmental Toxicology and Chemistry 18(7): 1456-1461.
38. CUPPEN, J.G.M., VAN DEN BRINK, P.J., CAMPS, E., UIL, K.F., BROCK, T.C.M. (2000) Impact of thefungicide carbendazim in freshwater microcosms.1. water quality, breakdown of particulate organicmatter and responses of macroinvertebrates. Aquatic Toxicology 48: 233-250.
39. CURTIS, L.R., SEIM, W.K., SIDDENS, L.K., MEAGER, D.A., CARCHMAN, R.A., CARTER, W.H.,CHAPMAN, G.A. (1989) Role of exposure duration in hydrogen ion toxicity to brook (Salvelinusfontinalis) and rainbow trout (Salmo gardneiri). Can J. Fish Aquat. Sci. 46: 33-40.
40. DAY, K., KAUSHIK, N.K. (1987) The adsorption of fenvaerate to laboratory glassware and the algaChlamydomonas reinhardii, and its effects on uptake of the pesticide by Daphnia galeata mendotae.Aquatic Toxicology 10: 131-142.
Cranfield Centre for EcoChemistry
58
41. DE ANGELIS, D., GODBOUT, L.L., SHUTER, B.J. (1991) An individual-based approach to predictingdensity-dependent compensation in smallmouth bass populations. Ecological Modelling 57: 91-115.
42. DEBUS, R., FLIEDNER, A., SCHAFERS, C. (1996) An artificial stream mesocosm to simulate fate andeffects of chemicals: technical data and initial experience with biocenosis. Chemosphere 32(9): 1813-1822.
43. DYER, S.D., BELANGER, S.E. (1999) Determination of the sensitivity of macroinvertebrates in streammesocosms through field derived-assessments. Environmental Toxicology and Chemistry, 18(12), 2903-2907.
45. FAIRCHILD, J.F., LAPOINT, T.W., ZAJICEK, J.L., NELSON, M.K., DWYER, F.J., LOVELY, P. (1992)Population, communiy and ecosystem level responses of aquatic mesocosms to pulsed doses of apyrethroid insecticide. Environmental Toxicology and Chemistry 11: 115-129.
46. FAIRCHILD, J.F., DWYER, F.J., LAPOINT, T.W., BURCH, S.A., INGEROLL, C.G. (1993) Evaluationof laboratory-generated NOEC for linear alkylbenzene sulfonate in outdoor experimental streams.Environmental Toxcology and Chemistry, 12: 1763-1776.
47. FAIRCHILD, J.F., RUESSLER, D.S., CARLSON, A.R. (1999) Comparative sensitivity of five species ofmacrophytes and six species of algae to atrazine, metribuzin, alachlor and metolachlor. EnvironmentalToxicology and Chemistry 17(9) 1830-1834.
48. FARMER, D., HILL, I.R., MAUND, S.J. (1995) A compaarison of the fate and effects of two pyrethroidinsecticides (lamda-cyhalothrin and cypermethrin) in pond mesocosms. Ecotoxicology 4: 219-244.
49. FLIEDNER, A., REMDE, A., NIEMANN, R., SCHAFERS, C. (1997) Effects of the organotin pesticideazocyclotin in aquatic microcosms. Chemosphere 35: 209-222.
50. FLEMER, D.A., RUTH, B.A., BUNDRICK, C.M.,MOORE, J.C. (1997) Laboratory effects of microcosmsize and the pesticide chlorpyrifos on benthic macroinvertebrate colonization of soft estuarine sediments.Marine Environmental Research 43(4): 243-263.
51. FORNEY, D. R., DAVIES, D. E. (1981) Effects of low concentrations of herbicides on submersed plants.Weed Sci. 29: 677-685.
52. GENONI, G.P. (1992) Short-term effect of a toxicant on scope for change in ascendency in a microcosmcommunity. Ecotoxicology and Environmental Safety 24: 179-191.
53. GIDDINGS, J.M., BIEVER, R.C., ANNUNZIATO, M.F., HOSMER, A.J. (1996) Effects of diazinon on largeoutdoor pond microcosms. Environmental Toxicology and Chemistry 15(5): 618-629.
54. GIDDINGS, J.M. (1997) Aquatic mesocosm studies and field studies with pyrethroids: observed effects andtheir ecological significance. Springborn Laboratories Report 97-6-7014.
55. GIDDINGS, J.M., BIEVER, R.C., RACKE, K.D. (1997) Fate of chlorpyrifos in outdoor pond microcosmsand effects on growth and survival on bluegill sunfish. Environmental Toxicology and Chemistry 16(11):2353-2362.
56. GIRLING, A.E., PASCOE, D., JANSSEN, C.R., PEITHER, A., WENZEL, A., SCHAFER, H.,NEUMEIER, B., MITCHELL, G.C., TAYLOR, E.J., MAUND, S.J., LAY, J.P., JUTTNER, I.,CROSSLAND, N.O., STEPHENSON, R.R., PERSONNE, G. (2000) Development of methods forevaluating toxicity to freshwater ecosystems. Ecotoxicology and Environmental Safety, 45(2): 148-176.
57. GIRLING, A.E., TATTERSFIELD, L., MITCHELL, G.C., CROSSLAND, N.O., PASCOE, D.,BLOCKWELL, S.J., MAUND, S.J., TAYLOR, E.J., WENZEL, A., JANSSEN, C.R., JUTTNER, I.(2000) Derivation of predicted no-effect concentrations for lindane, 3,4-dichloroaniline, atrazine andcopper. Ecotoxicology and Environmental Safety 46: 148-162.
58. GOMEZ, A., CECCINE, G., SNELL, T.W. (1997) Effects of pentachlorophenol on predator-preyinteractions of two rotifers. Aquatic Toxicology 37: 271-282.
59. GORBI, G., CORRADI, M.G. (1993) Chromium toxicity on two linked trophic levels. Ecotoxicology andEnvironmental Safety 25: 64-71.
60. GRADE, R., GONZALEZ-VALERO, J., HOCHT, P., PFEIFLE, V. (2000) A higher tier flow-throughtoxicity test with the green alga Selenastrum capricornutum. Sci. Total Environ. 247: 355-361.
Cranfield Centre for EcoChemistry
59
61. GREEN, D.W.J., WILLIAMS, K.A., PASCOE, D. (1986) The acute and chronic toxicity of cadmium todifferent life history stages of the freshwater crustacean Asellus aquaticus (L). Archives EnvironmentalConatmination and Toxicology 15: 465-471.
62. GRUESSNER, B., WATZIN, M.C. (1996) Response of aquatic communities from a Vermont stream toenvironmentally realistic atrazine exposure in laboratory microcosms. Environmental Toxicology andChemistry 15(4), 410-419.
63. GURNEY, W.S.C., MCCAULEYE., NISBET, R.M., MURDOCH, W.W. (1990) The physiological ecologyof Daphnia: a dynamic model of growth and reproduction. Ecology 71: 716-732.
64. HAITZER, M., HOSS, S., TRAUNSPURGER, W., STEINBERG, C. (1998) Effects of dissolved organicmatter (DOM) on the bioconcentration of organic chemicals in aquatic organisms: a review. Chemosphere37(7): 1335-1362.
65. HALLAM, T.G., LASSITER, R.R. (1994) Individual-based mathematical modelling approaches inecotoxicology: a promising direction for aquatic population and community ecological risk assessment.In Kendall, R.J., Lacher, T.E. (eds) Wildlife toxicology and population modelling. Lewis Publishers,Boca Raton.
66. HAM, L., QUINN, R., PASCOE, D. (1995) Effects of cadmium on the predator-prey interaction betweenthe turbellarian Dendrocoelum lacteum (Muller, 1774) and the isopod crustacean Asellus aquaticus (L).Archives of Environmental Contamination and Toxicology 29: 358-365.
67. HAMER, M.J., MAUND, S.J., HILL, I.R. (1992) Laboratory methods for evaluating the impact ofpesticides on water/sediment organisms. Proceedings of the British Crop Protection Council Conference(Pests and diseases), Brighton, UK.
68. HAMER, M.J., GOGGIN, U.M., MULLER, K., MAUND, S.J. (1999) Bioavailability of lamda-cyhalothrinto Chironomus riparius in sediment-water and water-only systems. Aquatic Ecosystem Health andManagement 2: 403-412.
69. HAMERS, T., KROGH, P.H. (1997) Predator-prey relationships in a two-species toxicity test system.Ecotox. Environ. Safety 37: 202-212.
70. HAMILTON, P.B., JACKSON, G.S., KAUSHIK, N.K., SOLOMON, K.R., STEPHENSON, G.L. (1988)The impact of two applications of atrazine on the plankton communities of in situ enclosures. AquaticToxicology 13: 123-140.
71. HANAZATO, T. (1998) Response of a zooplankton community to insecticide application in experimentalponds: a review and the implications of the effects of chemicals on the structure and functioning offreshwater communities.
72. HANDY, R.D. (1994) Intermittent exposure to aquatic pollutants: assessment, toxicity and sublethalresponses in fish and invertebrates. Mini review: Comp Biochem Physiol C 107: 171-184.
73. HANSEN, C.R., KAWATSKI, J.A. (1976) Application of the 24 h post-exposure observation to acutetoxicity studies with invertebrates. J Fish Res Board Can 33: 1198-1201.
74. HENDLEY, P., GIDDINGS, J. (1999) Draft report of the Aquatic Workgroups of ECOFRAM (EcologicalCommittee on FIFRA Risk Assessment) - Aqex_ecofram_Peer01_may499.doc.
75. HICKIE, B.E., MCCARTY, L.S., DIXON, D.G. (1995) A residue-based toxicokinetic model for pulseexposure toxicity in aquatic systems. Environ. Toxicol. Chem. 14: 2187-2197.
76. HINMAN, M. L., KLAINE, S. J. (1992) Uptake and translocation of selected organic pesticides by therooted aquatic plant Hydrilla verticillata Royle. Environ. Toxicol.Chem. 26: 609-613.
77. HOLDWAY, D.A., BARRY, M.J., LOGAN, D.C., ROBERTSON, D., YOUNG, V., AHOKAS, J.T. (1994)Toxicity of pulse-exposed fenvalerate and esfenvalerate to larval australian crimson-spotted rainbow fish(Melanotaenia fluviatilis). Aquatic Toxicology 28(3-4): 169-187.
78. HOSMER, A.J., WARREN, L.W., WARD, T.J. (1998) Chronic toxicity of pulsed-dosed fenoxycarb toDaphnia magna exposed to environmentally realistic concentrations. Environ. Toxicol. Chem. 17: 1860-1866.
79. HUTCHINSON, T.H., SOLBE, J., KLOEPPER-SAMS, P.J. (1998) Analysis of the ECETOC aquatic toxicity(EAT) database.III- Comparative toxicity of chemical substances to different life stages of aquaticorganisms. Chemosphere 36:129-142.
Cranfield Centre for EcoChemistry
60
80. ISNARD, P., VASSEUR, P., GRAFF, L., NARBONNE, J.F., CLERANDEAU, C., BUDZINSKI, H.,AUGAGNEUR, S., BASTIDE, J., CAMBON, J.P., CELLIER, P., ROMAN, G. (2000) Comparing theecotoxicity of chemicals in standard media and natural waters. Poster presented at the 3rd SETAC WorldCongress, 21-25 May, 2000, Brighton, UK.
81. JOHNSON, P.C., KENNEDY, J.H., MORRIS, R.G., HAMBLETON, F.E. (1994) Fate and effects ofcyfluthrin (pyrethroid insecticide) in pond mesocosms and concrete microcosms. In Graney, R.L., Kennedy,J.H., Rodgers, J.H. (eds) Aquatic mesocosm studies in ecological risk assessment. Lewis Publishers,Michigan.
82. KALLANDER, D.B., FISHER, S.W., LYDY, M.J. (1997) Recovery following pulsed exposure toorganophosphorous and carbamate insecticides in the midge, Chironomus riparius. Archives ofEnvironmental Contamination and Toxicology 33(1): 29-33.
83. KAY, S. H. HALLER, W. T., GARRARD, L. A. (1984) Effects of heavy metals on water hyacinths(Eichhornia crassipes (Mart.)Solms). Aquat. Toxicol. 5: 117-128.
84. KERSTING, K. (1991) Microecosystems state and its response to the introduction of a pesticide. Verh. Int.Verein. Limnol. 23: 1641-1646.
85. KERSTING, K., VAN WIJNGAARDEN, P.A. (1999) Effects of a pulsed treatment with the herbicide afalon(active ingredient linuron) on macrophyte-dominated mesocosms. I. responses of ecosystem metabolism.Environmental Toxicology and Chemistry. 18(12): 2859-2865.
86. KEVAN, S.H., DIXON, D.G. (1996) Effects of age and coion (K+ and Na+) on the toxicity of thiocyanate torainbow trout (Oncorhynchus mykiss) during pulse or continuous exposure. Ecotoxicology andEnvironmental Safety 35: 288-193.
87. KLUTTGEN, B., KUNTZ, N., RATTE, H.T. (1996) Combined effects of 3,4-dichloroaniline and foodconcentration on life table data of two related clacocerans, Daphnia magna and Ceriodaphnia quadrangula.Chemosphere 32: 2015-2028.
88. KOOIJMAN, S.A.L.M. (1981) Parametric analyses of mortality rates in bioassays. Water Research 17:747-759.
89. KOOIJMAN, S.A.L.M. (1987) A safety factor for LC50 values allowing for differences in sensitivityamongst species. Water Research 21: 269-276.
90. KOOIJMAN, S.A.L.M., METZ, J.A.J. (1984) On the dynamics of chemically stressed populations: thededuction of population consequences from effects on individuals. Ecotoxicol. Environ. Safety 8: 254-274.
91. KOOIJMAN, S.A.L.M., BEDAUX, J.J.M. (1996) The analysis of aquatic toxicity data (includes DEBtox,vers. 1.0) VU University Press, Amsterdam.
92. KOSTEL, J.A., WANG, H., AMAND, A.L.S., GRAY, K.A. (1999) Use of a novel laboratory streamsystem to study the ecological impact of PCB exposure in a periphytic biolayer. Water Research 33(18):3735-3748.
93. KREIGER, K.A., BAKER, D.B., KRAMER, J.W. (1988) Effects of herbicides on stream aufwuchsproductivity and nutrient uptake. Archives of Environmental Contamination and Toxicology 17: 299-306.
94. KUKKONEN, J., OIKARI, A. (1991) Bioavailability of organic pollutants in boreal waters with varyinglevels of dissolved organic material. Water Research 25: 455-463.
95. KUKKONEN, J., PELLINEN, J. (1994) Binding of organic xenobiotics to dissolved organicmacromolecules:comparison of analystical methods. Science of the Total Environment 152: 19-29.
96. KUSK, K.O. (1996) Bioavailability and effect of pirimicarb on Daphnia magna in a laboratoryfreshwater/sediment system. Archives of Environmental Contamination and Toxicology 31(2):252-255.
97. LANDRUM, P.F., REINHOLD, M.D., NIHART, S.R., EADIE, B.J. (1985) Predicting the bioavailability oforganic xenobiotics to Pontoporeia hoyi in the presence of humic and fulvic materials and naturaldissolved organic matter. Environ. Toxicol. Chem. 4, 459-467.
98. LANDRUM, P.F., LEE, H., LYDY, M.J. (1992) Toxicokinetics in aquatic systems: model comparisons anduse in hazard assessment. Environ. Toxicol. Chem. 11: 1709-1725.
Cranfield Centre for EcoChemistry
61
99. LEEUWANGH, P., BROCK, T.C.M., KERSTING, K. (1994) An evaluation of four types of freshwatermodel ecosystem for assessing the hazard of pesticides. Human and Experimental Toxicology, 13: 888-899.
101. MACEK, K.J., SLEIGHT, B.H. (1977) Utility of toxicity tests with embryos and fry of fish in evaluatinghazards associated with the chronic toxicity of substances to fishes. In Mayer, F.L. and Hamelink, J.L.(eds) Aquatic Toxicology and Hazard Evaluation. ASTM STP, Philadelphia.
102. MANCINI, J.L. (1983) A method for calculating effects on aquatic organisms of time-varying exposureconcentrations. Water Researh 17: 1355-1362.
103. MARUBINI, E., VALSECCHI, M.G. (1995) Analysing survival data from clinical trials and observationalstudies. John Wiley and Sons, NY.
104. MAUND, S.J., TAYLOR, E.J., PASCOE, D. (1992) Population responses of the freshwater amphipodcrustacean Gammarus pulex (L) to copper. Freshwater Biol. 28:29-36.
105. MAUND, S.J., SHERRATT, T.N., STICKLAND, T., BIGGS, J., WILLIAMS, N., SHILLABEER, N.,JEPSON, P. (1997) Ecological considerations in risk assessment for pesticides in aquatic ecosystems.Pesticide Science 49: 185-190.
106. MAUND, S.J., HAMER, M.J., WARINTON, J.S. AND KEDWARDS, T.J. (1998) Aquatic ecotoxicologyof the pyrethroid insecticide lamda-cyhalothrin: considerations for higher-tier aquatic risk assessment.Pesticide Science 54(4): 408-417.
107. MCCAHON, C.P., PASCOE, D. (1988) Cadmium toxicity to the freshwater amphipod Gammarus pulex(L.) during the molt cycle. Freshwater Biol 19: 197-203.
108. MCCARTHY, J.F., JIMINEZ, B.D., BARBEE, T. (1985) Effect of dissolved humic material onaccumulation of polycyclic aromatic hydrocarbons: structure-activity relationships. Aquatic Toxicology,7:15-24.
109. MCKIM, J.M. (1977) Evaluation of tests with early life stages of fish for predicting long-term toxicity.Journal of the Fish Research Board of Canada 34: 1148-1154.
110. MCKIM, J.M. (1985) Early life stage toxicity tests. In Rand, G.M. and Petrocelli, S.R. (Eds.) Fundamentalsof aquatic toxicology, First edition. Hemisphere Publishing Corporation, New York.
111. MEYER, J.S., GULLEY, D.D., GOODRICH, M.S., SZMANIA, D.C., BROOKS, A.S. (1995) Modelingtoxicity due to intermittent exposure of rainbow trout and common shiners to monochloroamine.Environmental Toxicology and Chemistry 14(1): 165-175.
112. MILLER, R.G. (1981) Survival Analysis. Wiley and Sons, NY.
113. NADDY, R.B., JOHNSON, K.A., KLAINE, S.J. (2000) Response of daphnia magna to pulsed exposures ofchlorpyrifos. Environ. Toxicol. Chem. 19(2): 423-431.
114. NEUGEBAUR, K., ZIERIS, F.J., HUBER, W. (1990) Ecological effects of atrazine on two outdoorartificial freshwater ecosystems. Wasser Abwasser Forsch 23: 11-17.
115. NEWMAN, M.C., McCLOSKEY, J. (1996) Time-to-event analysis of ecotoxicity data. Ecotoxicology 5:187-196.
116. NEWMAN, M., OWNBY, D.R., MEZIN, L.C.A., POWELL, D.C., CHRISTENSSEN, T.R.L., LERBERG,S.B., ANDERSON, B.A. (2000) Applying species-sensitivity distributions in ecological risk assessment:assumptions of distribution type and sufficient numbers of species. Environ. Toxicol. Chem. 19(2): 508-515.
117. OKKERMAN, P.C., PLASSCHE, E.J.V.D., EMANS, H.J.B., CANTON, J.H. (1993) Validation of someextrapolation methods with toxicity data derived from multiple species experiments. Ecotoxicology andEnvironmental Safety 25: 341-359.
118. ORIS, J. T., HALL, A. T. & TYLKA, J. D. (1990) Humic acids reduce the photo induced toxicity ofanthracene to fish and Daphnia. Environmental Toxicology and Chemistry 9: 575-584.
119. PACK, S. (1993) A review of statistical data analysis and experimental design in OECD aquatic toxicologytest guidelines. Report to the OECD, Paris.
121. PARSONS, J.T., STURGEONER, G.A. (1991) Acute toxicity of permethrin, fenitrothion, carbaryl andcarbofuran to mosquito larvae during single or mutliple-pulse exposures. Environmental Toxicology andChemistry 10: 1229-1233.
122. PASCOE, D., SHAZILI, N.A.M. (1986) Episodic pollution – a comparison of brief and continuousexposure of rainbow trout to cadmium. Ecotoxicol Environ Saf 12: 189-198.
123. PASCOE, D., WENZEL, A., JANSSEN, C., GIRLING, A.E., JUTTNER, I., FLIEDNER, A.,BLOCKWELL, S.J., MAUND, S.J., TAYLOR, E.J., DIEDRICH, M., PERSOONE, G., VERHELST, P.,STEPHENSON, R.R., CROSSLAND, N.O., MITCHELL, G.C., PEARSON, N., TATTERSFIELD, L.,LAY, J-P., PEITHER, A., NEUMEIR, B., VELLETTI, A-R. (2000) The development of toxicity testsfor freshwater pollutants and their validation in stream and pond mesocosms. Water Research 34(8):2323-2329.
124. PAULI, B.D., COULSON, D.R., BERRILL, M. (1999) Sensitivity of amphibian embryos and tadpoles tomimic 240 lv insecticide following single or double exposures. Environ. Toxicol. Chem. 18(11) 2538-2544.
125. PEITHER, A., JUTTNER, I., KETTRUP, A., LAY, J-P. (1996) A pond mesocosm study to determine thedirect and indirect effects of lindane on a natural zooplankton community. Environmental Pollution93(1): 49-56.
126. PERSOONE G. & JANSSEN C.R. (1993) Freshwater Invertebrate Tests. In Handbook ofEcotoxicologyVolume 1. Ed. P. Calow. Pages51-65, Blackwell Science Ltd, London.
127. PUSEY, B.J., ARTHRINGTON, A.H., MCCLEAN, J. (1994) The effect of a pulsed application ofchlorpyrifos on macroinvertebrate communities in outdoor artitificial stream system. Ecotoxicology andEnvironmental Safety 27: 221-250.
128. ROEX, E.W.M., VANGESTEL, C.A.M., VAN WEZEL, A.P., VANSTRAALEN, N.M. (2000) Ratiosbeween acute aquatic toxicity and effects on population growth rates in relation to toxicant mode ofaction. Environmental Toxicology and Chemistry 19(3): 685-693.
129. ROMAN, G., ISNARD, P., JOUANY, J.M. (1999) Critical analysis of methods for assessment of predictedno-effect concentration. Ecotoxicology and Environmental Safety, 43(2): 117-125.
130. RUSSOM, C.L., BRADBURY, S.P., BRODERIUS, S.J., HAMMERMEISTER, D.E., DRUMMOND, R.A.(1997) Predicting the modes of toxic action from chemical structure: acute toxicity in the fatheadminnow (Pimephales promelas). Environmental Toxicology and Chemistry 16(5): 948-967.
131. SANCHEZ, M., FERRANDO, M.D., SANCHO, E., ANDREU, E. (1999) Assessment of the toxicity of apesticide with a two generation reproduction test using Daphnia magna. Comparative Biochemistry andPhysiology Part C 124: 247-252.
132. SCHULZ, R., LIESS, M. (in press) Toxicity of aqueous-phase and suspended-particle-associatedfenvalerate: chronic effects following pulse-dosed exposure of Limnephilus lunatus (trichoptera).Environ. Toxicol Chem.
133. SEIM, W.K., CURTIS, L.R., GLEN, S.W., CHAPMAN, G.A. (1984) Growth and survival of developingsteelhead trout (Salmo gairdneri) continuously or intermittently exposed to copper. Can J. Fish. Aquat.Sci. 41: 433-438.
134. SHAW, J.L., MAUND, S.J., HILL, I.R. (1995) Fathead minnow (Pimephales promelas Rafinesque)reproduction in outdoor microcosms: an assessment of the ecological effects of fish density. EnvironToxicol Chem 14: 1763-1772.
135. SHAZILI, N.A.M., PASCOE, D. (1986) Variable sensitivity of rainbow trout (Salmo gairdneri) eggs andalevins to heavy metals. Bulletin Environmental Contamination and Toxicology 36: 468-474.
136. SHERRATT, T.N., ROBERTS, G., WILLIAMS, P., WHITLFIELD, M., BIGGS, J., SHILLABEER, N.,MAUND, S.J. (1999) A life-history approach to predicting the recovery of aquatic invertebratepopulations after exposure to xenobiotic chemicals. Environmental Toxicology and Chemistry 18(11):2512-2518.
Cranfield Centre for EcoChemistry
63
137. SHILLABEER, N., SMYTH, D.V., TATTERSFIELD, L. (2000) Higher tier risk assessment ofagrochemicals, incorporating sediment into algal test systems. Proceedings of th BCPC Conference –Pests and Diseases, Brighton, 2000. pp 359-364
138. SIBLY, R.M. (1996) Effects of pollutants on individual life histories and population growth rates. InNewman, M.C., Jagoe, C.H. (eds) Ecotoxicology: a hierarchical approach. Lewis Publishers, BocaRaton, Fl.
139. SIDDENS, L.K., SEIM, W.K., CURTIS, L.R., CHAPMAN, G.A. (1986) Comparison of continuous andepisodic exposure to acidic aluminium-contaminated waters of brook trout (Salvelinus fontinalis). Can. J.Fish. Aquat. Sci. 43: 2036-2040.
140. SKALSKI, J.R. (1981) Statistical incosistencies in the use of no-observed-effect levels in toxicity testing.In: Branson, D.R. and Dickson, K.L. (eds) Aquatic Toxicology and Hazard Assessment: FourthConference ASTM STP737. ASTM, Philadelphia.
141. SLOOFF, W., CANTON, J.H. (1983) Comparison of the susceptibility of 11 freshwater species to 8chemical compounds. 2. (semi)chronic toxicity tests. Aquatic Toxicology 4(3): 271-282.
142. SLOOFF, W., VAN OERS, J.A.M., DEZWART, D. (1986) Margins of uncertainty in ecotoxicologicalhazard assessment. Environ Toxicol Chem. 5: 841-852.
143. SOLOMON, K.R. (1996) Overview of recent developments in ecotoxicological risk assessment. RiskAnalysis 16(5): 627-633.
144. SOLOMON, K.R., BAKER, D., RICHARD, P.R., DIXON, K.D., KLAINE, S.J., LA POINT, T.W.,KENDALL, R.J., WEISSKOPF, C.P., GIDDINGS, J.M., GIESY, J.P., HALL, L.W., WILLIAMS, W.M.(1996) Ecological risk assessment of atrazine in north american surface waters. Environ. Toxicol. Chem.,15(1) 31-76.
145. STARK, J.D., BANKEN, J.A.O. (1999) Importance of population structure at the time of toxicantexposure. Ecotoxicology and Environmental Safety 42: 282-287.
146. STEPHAN, C.E., MOUNT, D.I., HANSEN, D.J., GENTILE, J.H., CHAPMAN, G.A., O’NEILL, R. (1985)Guidelines for deriving numerical national water quality criteria for the protection of aquatic organismsand their uses. USEPA, PB85-227049.
147. STUIJFZAND, S.C., POORT, L., GREVE, G.D., VANDERGEEST, H.G., KRAAK, M.H.S. (2000)Variables determining the impact of diazinon on aquatic insects: taxon, developmental stage, andexposure time. Environmental Toxicology and Chemistry 19(3): 582-587.
148. SUEDEL, B.C., ROGERS, J.H. (1994) Development of a formulated reference sediments for freshwaterand estuarine sediment toxicity testing. Environmental Toxicology and Chemistry 13: 1163-1175.
149. SUEDEL, B.C., DEAVER, E., RODGERS, J.H. (1996) Formulated sediment as a reference and dilutionsediment in definitive toxicity tests. Archives of Environmental Toxicology and Chemistry 30(1): 47-52.
150. SUTER, G.W., BARNTHOUSE, L.W., BRECK, J.E., GARDNER, R.H. AND O’NEILL, R.V. (1985)Extrapolation from the laboratory to the field: how uncertain are you? In: Cardwell, R.D., Bahner, R.C.(Eds) Aquatic Toxicology and Hazard Assessment. American Society for Testing of Materials,Philadelphia, Pa.
151. TAUB, F.B. (1969) A biological model of a freshwater community: a gnotobiotic ecosystem. Limnol.Oceanogr. 14: 136-142.
152. TAYLOR, E.J., BLOCKWELL, S.J., MAUND, S.J., PASCOE, D. (1992) Effects of lindane on the lifecycle of a freshwater invertebrate Chironomus riparius Meigen (Insecta: Diptera). Arch. Environ.Contam. Toxicol. 24:145-150
153. TAYLOR, E.J., MORRISON, J.E., BLOCKWELL, S.J., TARR, A., PASCOE, D. (1995) Effects of lindaneon the predator-prey interaction between Hydra oligactis Pallas and Daphnia magna Strauss. ArchivesEnvironmental Contamination and Toxicology 29: 291-296.
154. THURSTON, R.V., CHAKOUMATOS, C., RUSSO, R.C. (1981) Effect of fluctuating exposures on theacute toxicity of ammonia to rainbow trout (Salmo gardneiri) and cut throat trout (S. clarki). WaterResearch 15: 911-917.
Cranfield Centre for EcoChemistry
64
155. TRAUNSPURGER, W., SCHAFER, H., REMDE, A. (1996) Comparative investigation on the effect of aherbicide on aquatic organisms in single species tests and aquatic microcosms. Chemosphere 33(6): 1129-1141.
156. USEPA (1984). Estimating concern levels for concentrations of chemical substances in the environment.Environmental Effects Branch, Health and Environmental Review Division, USEPA.
158. USEPA (2000) Technical report of the implementation plan for probabilitic ecological assessments: aquaticsystems. USEPA, Duluth, MN.
159. VAAL, M.A. VANLEEUWEN, C.J., HOEKSTRA, J.A., HERMENS, J.L.M. (2000) Variation onsensitivity of aquatic species to toxicants: Practical consequences for effect assessment of chemicalsubstances. Environmental Management 25(4): 415-423
160. VAN DEN BRINK, P.J., HARTGERS, E.M., FELTWEIS, U., CRUM, S.J.H, VAN DONK, E., BROCK,T.C.M. (1997) Sensitivity of macrophyte-domiated freshwater microcosms to chronic levels of theherbicide linuron. Ecotoxicology and Environmental Safety 38(1): 13-24.
161. VAN DEN BRINK, P.J.,HATTINK, J., BRANSEN, F., VAN DONK, E., BROCK, T.C.M. (2000) Impactof the fungicide carbendazim in freshwater microcosms. II. zooplankton, primary producers and finalconclusions. Aquatic Toxicology 48, 251-264.
162. VAN DEN BRINK, P.J., PSTHUMA, L., BROCK, T.C.M. (2000) Verification of the SSD-concept: fieldrelevance and implications for ecological risk assessment. Poster presented at the SETAC WorldCongress, Brighton, May 2000.
163. VAN GEEST, G.J., ZWAARDEMAKER, N.G., VAN WIJNGAARDEN, R.P.A., CUPPEN, J.G.M. (1999)Effects of a pulsed treatment with the herbicide afalon (active ingredient linuron) on macrophyte-dominated mesocosms. II. structural responses. Environmental Toxicology and Chemistry 18(12): 2866-2874.
164. VAN STRAALEN, N.M., DENNEMAN, G..A.J. (1989) Ecotoxicological evaluation of soil quality criteria.Ecotoxicol Environ Safety 18: 241-251.
165. VAN STRAALEN, N.M., SCHOBBEN, J.H.M., TRAAS, T.P. (1992) The use of ecotoxicological riskassessment in deriving maximum acceptable half-lifes of pesticides. Pesticide Science 34: 227-231.
166. VERHAAR, H.J.M., VAN LEEUWEN, C.J., HERMENS, J.L.M. (1992) Classifying environmentalpollutants 1: structure-activity relationships for prediction of aquatic toxicity. Chemosphere 25: 471-491.
167. WAGNER, C., LOKKE, H. (1991) Estimation of ecotoxicological protection levels from NOEC toxicitydata. Water Research 25(10): 1237-1242.
168. WANG, M.P., HANSON, S.A. (1985) The acute toxicity of chlorine on freshwater organisms: time-concentration relationships of constant and intermittent exposures. In: R.C. Bahner and D.J. Hansen(Eds) Aquatic toxicology and hazard assessment: eigth symposium. STP891. ASTM, Philadelphia, PA.
169. WANG, W., WILLIAMS, J. M. (1990) The use of phytotoxicity tests (common duckweed, cabbage andmillet) for determining effluent toxicity. Envionment Monitor Assess 14: 45-58.
170. WANG, H., KOSTEL, J., AMAND, A.L.S., GRAY, K.A. (1999) The response of a laboratory streamsystem to PCB exposure: study of periphytic and sediment accumulation patterns. Water Research33(18): 3749-3761.
171. WARD, S., ARTHINGTON, A.H., PUSEY, B.J. (1995) The effects of chronic application of chlorpyrifoson the macroinvertebrate fauna in an outdoor artificial stream system. species responses. Ecotoxicologyand Environmental Safety 30(1): 2-23.
172. WEBBER, E.C., DEUTSCH, W.G., BAYNE, D.R., SEESOCK, W.C. (1992) Ecosystem-level testing of asynthetic pyrethroid insecticide in aquatic mesocosms. Environmental Toxicology and Chemistry 11: 87-105.
173. WEINSTEIN, J.E., ORIS, J.T. (1999) Humic acids reduce the bioaccumulation and photoinduced toxicityof fluoranthene to fish. Environmental Toxicology and Chemistry 18(9): 2087-2094.
Cranfield Centre for EcoChemistry
65
174. WILLIAMS, K.A., GREEN, D.W.J., PASCOE, D., GOWER, D.E. (1986) The acute toxicity of cadmiumto different larval stages of Chironomus riparius (diptera: chironomidae) and its ecological significancefor pollution regulation. Oecologia 70: 362-366
175. WRIGHT, A. (1976) The use of recovery as a criterion for toxicity. Bull Environ Contam Toxicol 15: 747-749.
176. ZITKO,V., MCCLEESE, D.W., METCALFE, C.D., CARLSON, W.G. (1979) Toxicity of permethrin,decamethrin and related pyrethroids to salmon (Salmo salar) and lobster (Homerus americanus). Bull.Environ. Contam. Toxicol. 21: 338-343.
Cranfield Centre for EcoChemistry
66
APPENDIX A – ORGANISATIONS CONTACTED DURING THE
STUDY*
Alterra Green World Research American Cyanamid
AstraZeneca Ltd Aventis Crop Science
BASF Aktiengesselschaft Bayer-AG
BBA CEFAS
Covance Inc Crop Protection Association
DETR Dow AgroSciences
DuPont Crop Protection Environment Agency
Fraunhofer Institut Health & Safety Executive
Huntingdon Life Sciences INIA
Inveresk Research International Monsanto Company
Novartis Crop Protection AG Pesticides Safety Directorate
RIVM SafePharm Laboratories Ltd
Shell Chemicals Stirling University
Unilever University of London
University of Sheffield USEPA
Veterinary Medicines Directorate Water Research Centre
Wright State University Zeneca Agrochemicals
* Company names at time questionnaire issued
Cranfield Centre for EcoChemistry
67
APPENDIX B – AVAILABLE SINGLE SPECIES ECOTOXICITY
METHODS
Table A1: Invertebrate species commonly used for fresh water toxicity tests