Toxicology

CHAPTER

Toxicology Chris D Metcalfe

-t

Environmental and Resource Studies, Trent University, Peterborough, Ontario, Canada

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Introduction The discipline of toxicology involves studying the nature and mechanisms of toxic lesions, and evaluating in a quantitative manner the spectrum of biological changes produced by exposure to chemicals. In this chapter, we will mainly concern ourselves with the methods and techniques used to quantitatively evaluate toxicological responses in fish. It is important to realize that every chemical can be toxic to fish under certain exposure conditions. For every chemical there should be an exposure condition (i.e. dose or concentration) that is 'safe' and an exposure condition that is 'toxic' to fish (Sprague, 1971). As illustrated in Table 37.1, the range of concentrations or doses that are toxic to fish may span several orders of magnitude. Therefore, toxicology is primarily concerned with establishing relationships between 'dose (concentration) and effect' for single test organisms or 'dose (concentration) and response' for a population of test organisms. It is also important to determine toxic 'thresholds'; that is, concentrations or doses above which toxicity occurs and below which, it does not. Until relatively recently, toxicological studies with fish focused almost exclusively on very toxic substances which produce 'acutely lethal' responses; that is, mortalities in fish exposed to chemicals for only Copyright 9 2000AcademicPress

short periods. Recently, we have become concerned with substances that may produce 'sublethal' responses in fish after 'chronic' exposure. Much of the terminology associated with toxicology is relatively specialized, so it is useful to start this chapter with a glossary of terms commonly used in fish toxicology.

0

H'I X m

m

z TABLE 37.1: Approximate toxic concentrations of chemicals in water for early life stages of fish

I--

0 r"-

Nitriloacetic acid (NTA)

I00 000

Fluoride

I0 000

Zinc

I 000

Cadmium

I00

DDT

I0

Methyl mercury

I

PCB congener 126

0.1

2,3,7,8-tetrachlorodibenzo-

0.01

p-dioxin

Glossary of terms used in fish toxicity studies

0 .J 0

_v x 0

@ _.1

a o ..J

z

I.U

U.l X

9 Dose Describes quantitatively the amount of chemical to which the fish has been exposed. Experimentally, exposure is usually done by dosing the food (per os), force-feeding solid material (gavage) or liquid through a tube (intubation), or by injecting the chemical into an egg (in ovo injection) or into the fish intraperitoneally (i.p.), intramuscularily (i.m.) or sub-cutaneously (s.c.). Dose is usually calculated as the weight of chemical (ng, ~lg, mg) administered per gram or kilogram of body weight. 1 I~gkg - l = 1 ngg -1 = 1 ppb; 1 mgkg -1 = 1 lagg -1 = 1 ppm. 9 Concentration Fish are often exposed to chemicals dissolved in water, and the exposure conditions are described according to the weight of the chemical dissolved in a volume of water. 1 lag L - 1 = 1 ngmL -1 = 1 ppb;1 mgL -1 = 1 lagmL -1 = 1 ppm. 9 Toxicity The capacity of a chemical to cause injury to a living organism. Lethality is a specific term refering to the capacity of a chemical to cause death to a living organism. Toxicity or lethality cannot be defined for fish without providing information on: - The species and life stage of the fish. - The dose or concentration of the chemical to which the fish were exposed. - The distribution of exposure in time (i.e. single, repeated or continuous dose). - The type and severity of injury (i.e. lethal response, sublethal response). - The time needed to produce toxicity (i.e. acute, chronic). 9 EDso (ECs0) Toxicity is often assessed quantita, r . tively as a dose-response or concentration-

response'. The EDs0 and ECs0 are the median effective dose and concentration, respectively; that is, the dose or concentration of a chemical at which 50% of the test population demonstrates a toxic effect. A specific response involving death of the organism is the median lethal concentration or dose (i.e. LCs0 or LDs0). The duration of exposure must be specified (e.g. 96-h LCs0). LOED (LOEC) The'lowest-observed-effect'dose or concentration is the lowest dose or concentration of a chemical that causes injury to a fish. N O E D (NOEC) The 'no-observed-effect' dose or concentration is the highest dose or concentration of a chemical that does not cause injury to a fish. M A T C The 'maximum acceptable toxicant concentration' is the hypothetical toxic threshold lying in a range bounded by the LOEC and the NOEC.

Factors affecting toxicity to fish There are several chemical, physical and biological factors that influence the toxicity of chemicals to fish, including the properties of the chemical in water, the water quality conditions, the route of exposure, and the species and life stage of the fish being tested. Therefore, there are several parameters that must be taken into account or controlled when designing fish toxicity tests if we are to eliminate or reduce variability in the toxic responses due to these factors (Table 37.2). In this chapter, we will examine the chemical, physiological and biological bases of why these factors influence toxicity.

TABLE 37.2: Parameters that must be controlled or standardized in toxicity tests with fish

Water temperature

Fish species and source

Photoperiod

Life stage

Dissolved oxygen concentration

Sex

pH Alkalinity

Size and condition

Reproductive status

Hardness

Health

Salinity

Acclimation period

Dissolved organic carbon

Nutrition

Solubility in water The solubility of a chemical in water governs the extent to which the chemical mixes with the aqueous phase to form a homogeneous solution9 Since fish live in water, the extent to which fish are exposed to a chemical is dependent on aqueous solubility9 The solubility of non-ionic chemicals, such as organic compounds and elemental forms of toxic metals (e.g. Hg~ is influenced by the polarity of the compound. The intermolecular forces that occur between nonionic molecules are dipole-dipole interactions for polar compounds, and van der Waals interactions for non-polar molecules; the former being much stronger than the latter (Connell, 1997)9 When a chemical is added to water, the solute will not dissolve in water to any appreciable extent unless it is sufficiently polar to overcome the dipole-dipole interactions of adjacent water molecules9 Thus, polar solutes such as methanol readily dissolve in water9 Non-polar solutes such as carbon tetrachloride do not dissolve readily in water, but tend to partition into other environmental media, such as lipids and sediments. This is of interest in fish toxicology because biotic membranes are composed primarily of lipid9 Thus, the tendency for a non-ionic chemical to partition between water and fish is governed by the polarity of the compound (Gobas and Russell, 1992). Thus, non-polar chemicals are termed, 'hydrophobic' or 'lipophilic', whereas polar chemicals are 'hydrophilic' or 'lipophobic'. A partition coefficient which is used as an indicator of lipophilicity is the 'octanol-water partition coefficient' (Kow). This coefficient describes the distribution of a chemical between water and a nonpolar substance, octanol. Chemicals that tend to dissolve in the octanol rather than water have high Kow and are classified as hydrophobic. The importance of hydrophobicity in determining the toxicity of organic compounds in fish is illustrated in studies where the lethal toxicity of organic compounds to fish has been shown to increase with log Kow (Yoshioka et al., 1986) or decrease with water solubility (Neely, 1984)9 In fact, theoretical relationships have been developed which predict toxicity based upon estimates of how readily chemicals bioconcentrate in fish and reach 'critical body residues' (Barron et al., 1997). The solubility of ionic chemicals, which include most salts of toxic metals and some ionic organic compounds, is usually much higher than that of non-ionic compounds. Ions tend to dissociate freely in aqueous

solution, as polar water molecules can overcome the forces between ions. Ions are not actually 'free' within an aqueous matrix, but are associated with water molecules as an 'aqua ion complex'9 In many cases, the toxic form of ionic chemicals is the 'free ion'; that is, the cation (Me+), anion (Me-), or oxyanion (MEOW-). Therefore, the amount of free ion in solution is an important factor in the toxicity of the element (Campbell, 1995)9 The degree of dissociation into the free ionic species is pH dependent. For instance, under acidic conditions, free cations (e.g. Zn 2+, A13+) should predominate 9 Therefore, the process of 'speciation' of ionic chemicals in water is an important factor governing toxicity (Wang, 1987).

Water quality conditions As mentioned above, ions may be dissolved in water in non-toxic forms. For instance, ions may form complexes with inorganic and organic 'ligands'. Inorganic ligands for cations in fresh water include carbonate ( C O 32 - -) * , sulfate (SO24-), and fluoride (F-) ions, and C1- is an important ligand in saline water (Morel and Hering, 1993). Complexes between cations and inorganic ligands tend to be fairly'labile' or reversible, depending on the concentration of the ion and the ligand, and the pH. However, complexes with organic ligands, such as humic acids, tend to be relatively r 9 9 , 9 non-labile. Alkahmty, which is primarily the concentration of carbonate ions in solution, is an important measure of the cation binding capacity of fresh water9 Alkalinity and pH are important variables influencing the toxicity of metal ions to fish (Lauren and McDonald, 1985)9 Transformations of chemicals dissolved in water can occur by hydrolysis, photolysis and oxidation (Connell, 1997). Photolytic transformation of some photoactive compounds, such as polynuclear aromatic hydrocarbons (PAHs) can increase toxicity to fish (Oris and Geisy, 1987)9 However, in general, hydrolysis is the most important transformation process affecting toxicity9 Hydrolysis occurs with both organic and inorganic chemicals, but hydrolysis of metal salts is a very important process governing toxicity to fish: Me ~+ + H 2 0 - -

M e ( O H ) (n-l)+ + H +

e.g.

Zn 2+ + H 2 0

Zn(OH) § + H +

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If a metal dissolves in distilled water, without any ligands being present, the metal will still form hydroxy complexes through hydrolysis. Metals differ in their stability constants as hydroxy complexes, so some metals form stable hydroxy complexes at relatively low pH, while others require higher pH. In general, trivalent cations such as A13§ form significant levels of hydroxy complexes (e.g. A1OH 2+, AI[OH2] +) at the pH of natural water (Campbell, 1995). These hydroxy complexes are relatively non-toxic relative to the free cation, so hydrolysis, which is pH dependent, is an important factor influencing toxicity. Water 'hardness', which is essentially the concentration of calcium and magnesium cations in water, affects fish toxicity for physiological reasons. Calcium may competitively bind to proteins on cell membranes. This binding may reduce the transport of toxic cations (e.g. copper, cadmium) across membranes such as the gill; thus reducing toxicity (Campbell and Stokes, 1985). Figure 37.1 illustrates the ameliorative effect of calcium ion concentration on the toxicity of aluminum to early life stages of rainbow trout (Oncorhynchus rnykiss). Since high calcium concentrations usually co-occur with high carbonate concentrations in natural waters, it is often difficult to separate the effects of hardness and alkalinity on toxicity to fish (Welsh et al., 1996). Overall, if we accept that the toxicity of ionic chemicals is usually dependent upon the concentrations of the free ion in solution, then various factors that affect speciation of ions will affect toxicity; including pH, alkalinity, hardness, and concentrations of

ILl

100

m El: iii

8o

x Iii

organic ligands (Playle et al., 1992, 1993). Figure 37.2 illustrates the speciation processes that affect the toxicity of ionic chemicals to fish in water. The toxicity of non-ionizable chemicals, such as organic compounds, is affected to a lesser extent by water quality conditions such as pH, alkalinity and hardness. However, dissolved and particulate organic material in water can alter the toxicity of organic compounds by acting as ligands for hydrophobic substances.

Biological interactions A chemical can be toxic to a fish in two possible ways. It may affect tissues on the surface of the organism (e.g. gill epithelium) or the chemical may enter the organism and cause toxicity. The epithelial and endothelial integument of fish is usually thickened and relatively impermeable to chemicals, except in the gill tissues, which are specialized for gas exchange, and in the gastrointestinal tract. Thus, branchial or gastrointestinal uptake routes are the most efficient mechanisms for uptake of toxic chemicals into fish. 'Bioaccumulation' of chemicals represents the uptake and retention of chemical from the environment into fish via any pathway (e.g. food, water), whereas, 'bioconcentration' represents uptake and retention of a chemical directly from water into fish. A toxic chemical must pass through cell membrane barriers to reach 'target' organs or tissues. Therefore, uptake of chemicals by fish can be influenced by both the lipophilicity and molecular size of the chemical (Protic and Sabljic, 1989). The membranes consist of a bimolecular layer of phospholipids,

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i

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50

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100

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150

AI (#g L-I ) Figure 37.1 Percent mortalities of eggs and fry of rainbow trout exposed to four concentrations of aluminum (25, 50, 100, 150pg L i) under conditions of 1,4 and 10rag L~-I of calcium (unpublished data by the author).

Particulate Figure 37.2 Speciation processes that influence the toxicity of chemical ions in tests with fish, including binding to suspended particulates, binding to dissolved organic and inorganic ligands, formation of colloids and formation of toxic free ions.

with integral and peripheral proteins. Because lipids are relatively non-polar, lipophilic chemicals will pass readily through them. Generally, the more lipophilic the chemical, the more readily it passes through the lipid-rich cell membranes into the organism (Gobas and Russell, 1992), but some large 'superhydrophobic' molecules cannot pass readily through membranes because of their molecular size (Bruggeman et al., 1984). The condition of the fish, and especially the fat content, can often influence the toxicity of hydrophobic chemicals since these compounds will tend to partition into fats, keeping them away from the target tissues and organs. Inorganic contaminants in the elemental form (e.g. Hg ~ can diffuse through lipid membranes, but inorganic cations and anions must pass through the cell membrane by routes that do not involve the lipid component of the membrane. There are several models proposed for this process, but the most likely mechanism involves binding of ions with amino acid or protein ligands within the membrane matrix, and passive diffusion of the complex through the membrane, or 'carrier mediated diffusion' (Astrue, 1989). For some chemicals, the rate of uptake is strongly influenced by the physiology of the fish. Gill ventilation rates and dietary intake are governed by the metabolic rates of fish. There is considerable variation in the metabolism of fish; from fast-swimming pelagic predators to slow-swimming benthivores, so toxicity thresholds may vary considerably, depending on the fish species tested. In poikilothermic organisms such as fish, metabolism changes with the water temperature, so temperature may be an important factor influencing toxicity (Hodson and Blunt, 1986). Similarily, dissolved oxygen concentrations may influence gill ventilatory rates. Early life stages of fish tend to have higher rates of metabolism than later life stages (McKim, 1985). Freshwater fish and saltwater fish have very different physiological mechanisms for regulating levels of essential ions such as Na § K § C1- (Evans, 1980). As illustrated in Figure 37.3, freshwater fish are constantly pumping ions into the body and excreting water out. Therefore, some metal cations and anions can be taken into the body along with essential ions (Na § C1 ). In salt water, aquatic organisms are constantly pumping ions out into the surrounding water, and imbibing large quantities of water. There may be transport of chemicals into the body along with the water in saltwater fish. Anadromous fish (e.g. salmon) and catadromous fish (e.g. eels) go through changes in ionoregulatory physiology when they move from

Water (a) Freshwater fish

Response 9inward active transport of ions 9copious urine production (hypo-osmotic)

~- Ions

-9Water (b) S a l t w ~ "

Response 9 outward active transport of ions 9continuous water intake 9small volume of urine (iso-osmotic)

--I Ions

Figure 37,3 Flux of essential ions and water and mechanisms for ionoregulation and osmoregulation in freshwater and saltwater fish.

0

X I

0 r0

@ m x

fresh water to salt water and vice versa. There can be large differences in the toxicity of chemicals to these fish species, depending upon the salinity of the water and the degree of acclimation of the test species (McDonald et al., 1989). Similar physiological adaptations must be taken into account in toxicity tests with euryhaline fish, which can tolerate a wide range of salinities. Fish possess metabolic pathways capable of transforming chemicals. Oxidation of chemicals through the activity of cytochrome P-450 monooxygenases is an important biotransformation pathway in fish (Stegeman and Hahn, 1994), and binding of chemicals to proteins or other large biomolecules (i.e. conjugation) is another transformation process (Di Giulio et al., 1995). In terms of affecting toxicity, biotransformation can result in: (i) inactivation of the chemical; (ii) activation of the chemical to form reactive metabolites; or (iii) maintenance of activity. Any factors that influence the rates of metabolic transformations of chemicals, such as water temperature, the sex and reproductive state of the fish, nutrition and the acclimation period, will often influence toxicity.

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The life stage of the fish is a major determinant of toxicity. In general, the most sensitive period for chemically induced toxicity in fish is the embryolarval or early juvenile stages. This is illustrated by the data on the median lethal concentration in Japanese medaka exposed to aqueous solutions of nonylphenol (Table 37.3), where the LCs0 for embryolarval stages and adults exposed to nonylphenol were 4601agL -l (Gray and Metcalfe, 1997) and 1400 lagL -1 (Yoshimura, 1986), respectively. Acute exposure of early life stages (ELS) of fish to a chemical generally provides a good estimate of the MATC obtained in chronic toxicity tests with later life stages (McKim et al., 1985). The high rate of metabolism of ELS may partially explain this sensitivity, but it is largely due to the fact that eggs, larvae and fry are going through many developmental events over a short space of time - several involving biochemical or hormonal events that depend on critical timing. Fish species differ widely in their sensitivity to the toxic effects of chemicals, as illustrated in Table 37.3. In fact, there may be greater variability in toxicity associated with the test species than associated with differences in test conditions (Sprague, 1985). Some of this variability can be attributed to differences in metabolic rates and physiology of the fish, or the presence of alloenzyrne genotypes that confer sensitivity or tolerance (Gillespie and Guttman, 1989). In some cases, whole fish taxa are more tolerant than other taxa. For instance, Spear and Pierce (1979) observed that centrarchids (i.e. Sunfish) are about 15 times more tolerant of copper that salmonids or minnows. It is important to note that previous exposure to chemicals may increase resistance to toxicity in fish (Duncan and Klaverkamp, 1982).

u.l

Toxicity testing with fish Throughout this volume, there are detailed descriptions of the anatomy of fish and some of the pathological conditions that can occur as a result of exposure to chemicals. A comprehensive review of all of the possible 'endpoints' of toxicity that can be used with fish is beyond the scope of this chapter. However, as illustrated in Table 37.4, quantitative estimates of chemical toxicity to fish vary according to the toxic endpoint and species tested. In general, studies with adult fish using lethality as a toxic endpoint and acute (short-term) exposures are the least sensitive tests, but they are the most commonly used methods for assessing toxicity because they are relatively rapid, simple and inexpensive. Toxicological studies are generally divided into four categories, depending on the duration of exposure: 1. Acute toxicity studies which involve either a single administration of the chemical to an organism, or continuous exposure over a very short period (e.g. 24-96 h). 2. Short-term toxicity studies which involve continuous, repeated or 'pulsed' exposure to a chemical over a period of about 10% of the life span. 3. Chronic toxicity studies which involve continuous or repeated exposures over a period of the entire life span of the test organism. If exposures occur over only a major portion of the life span, some researchers may classify these studies as long-term toxicity studies.

Test species

LC5o (~Lg L 1)

Atlantic salmon (Salmo salar) - juvenile

17 a 130 b

Sheepshead minnow (Cyprinodon variegatus)

210a

Winter flounder (Pseudopleuronectes americanus)

Fathead minnow (Pimephales promelas)

300 ~

Japanese medaka (Oryzias latipes) - embryolarvae

460 c

Japanese medaka (Oryzias latipes) - adult

1,400 d

Cod (Gadus morhua)

3,000a

aBrooke (1993); bTalmage (1994); CGray and Metcalfe (1997); ~foshimura (1986).

Japanese medaka (adult)

Lethality Lethality

1400 (LCso)~ 194 (LCso)b

Rainbow trout Japanese medaka (males)

Induction of intersex in gonad

50 (LOEC)c

Rainbow trout (males)

Inhibited testicular development

54 (LOEC)d

Induction of vitellogenin (VtG)

20 (LOEC)d

Induction of VtG mRNA precursor

14 (EC50)e

Inhibition of growth

10 (LOEC)b

Rainbow trout ~

(1986); bBrooke (1993); CGray and Metcalfe (1997); djobling et al. (1996); eLech et ol. (1996).

4. Life cycle toxicity studies which involve continuous or repeated exposures over at least two generations of the test organism - usually 'embryo to embryo'. Of course, chronic toxicity studies are very expensive and time consuming in comparison to acute toxicity studies. In order to reduce costs, chronic studies are typically done with few doses or concentrations. These generally include a control and at least a low dose and a high dose. Acute toxicity tests usually involve a large number of doses in order to determine more precisely the thresholds for a toxic response.

20 ~ Cumulative m o r t a l i t i e s ~

15

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tO

0

5

0

~ -

-~

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10

t

~ ,,

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100

,

I

Concentration (pg mL-1)

Figure 37.4 Cumulative mortalities observed over 17

Lethality testing In lethality testing, the median lethal concentration or dose (LCs0 or LDs0) is the usual method of reporting results, and a time period must be specified for the parameter (e.g. 96-h LCs0). This parameter is determined by observing whether there is increasing lethality in organisms exposed to a range of test doses or concentrations (i.e. a 'dose-response' relationship). Toxicant concentrations usually span a range which includes a concentration causing zero mortality, and a concentration causing complete mortality. Good lethality estimates are possible when at least one concentration or dose (and preferably two) kill some, but not all of the test organisms in the maximum observation period (i.e. 'partial mortalities'). When lethality is monitored, a plot of the frequency of mortalities in the test population generally forms an S-shaped curve (Figure 37.4). The central portion of the curve may be sufficiently straight to estimate a median response (e.g. LCs0), but if the curve can be 'straightened', a better estimate can be made of

@

300

days between fertilization of eggs and swim-up (start of exogenous feeding) in embryos and larvae of Japanese medaka (n - 20) in a control treatment and a treatment with aqueous solutions of 2,3,7,8-TCDD (30pg mL 1). (Data froFn Metcalfe eta/., 1997.)

i=n x "13 r'n

I'n

z

i= -

O m i'--

the 50% response point. A common method of curve straightening is the 'probit' method, where probit units are assigned to deviations around the mean (Table 37.5). There are other statistical methods for analyzing lethality data, including moving average interpolation, non-parametric methods (e.g. trimmed Spearman-Karber method) and binomial confidence intervals (Ellerseik and La Point, 1995). The basic data for lethality tests are tabulations of percent cumulative mortalities at each concentration, as observed at different times throughout the assay. These data can be used to calculate lethal concentrations. The concentration or dose at which 50% of the population dies at time t can be determined by using a computer program that uses a probit transformation model, or by plotting data on probit paper. In order

to illustrate the methods used in acute lethality testing, consider the data in Table 37.6. For each test concentration, there are several observation times. In the example, probit transformed percent cumulative mortality data are used to calculate LCs0~ and 95% confidence limits around these estimates for each observation time over 1 to 18 days. The LCs0~ are then plotted against observation time to determine a 'lethality line', as shown in

Figure 37.5. Note that the lowest concentration of the chemical (i.e. 4 pg m L - i ) caused less than 50% mortality (i.e. 8 of 20 fish) at the maximum observation time of 18 days. This point is plotted against the maximum observation time as 3 . A lethality line allows the 30

,: i,, o',

Japanese medaka embryo larval toxicity i

f

,

-

10

TABLE 37.5: Probit units used to transform cumu-

r

E

lative toxicity data for estimates of median toxic

cO

responses

o 0.1

-3

2

2.3

-2

3

15.9

-I

4

1.9

50.0

0

5

,--I

84.1

I

6

97.7

2

7

99.9

3

8

0

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I-

O C/I U.I

a o

I 6

i Li[ I0

L

L J J J ; i~] 100 LCs0 (pg mL ~)

1 200

Figure 37.5 A lethality line plotted as LCso vs. observation time from the data presented in Table 37.6 on the lethality of 2,3,7,8-TCDD to embryolarvae of Japanese medaka. The lowest concentration causing less than 50% mortality at the maximum observation time is plotted as a filled circle.

TABLE 37.6: Cumulative mortality data for the lethality of 2,3,7,8-TCDD ( 0 - 2 0 0 p g m L

1) to

embryolarvae of Japanese medaka ( n - 20) over 18 days between fertilization of eggs and swimup. The LCsos plus upper and lower 9 5 % confidence limits (in brackets) were calculated for each observation period by probit analysis

._I

< l-

z L~

13_

x

Observation Time (days) I

0

0

0

0

2

2

0

0

2

4

3

0

0

2

4

4

> 200

5

6

> 200

6

12

176 (155-204)

4

0

2

3

4

I0

18

69 (45-I 19)

6

0

3

4

5

12

20

47 (31-75)

8

0

5

8

10

18

20

20 (31-75)

12

0

8

9

13

20

20

12 (6-19)

14

0

8

9

16

20

20

10 (5-16)

16

0

8

9

20

20

20

9 (4-13)

18

0

8

10

20

20

20

8 (4-12)

Hatch

Swim-up Source: Original data are from the study published by Metcalfe et al. (1997).

researcher to extract the maximum amount of information from the lethality test. Instead of estimating an LCs0 from a single observation time (which may have anomalous data), all of the observation times can be utilized. From a single lethality line, we can get estimates of the LCs0s at 4 days, 10 days, 15 days, etc. We can also estimate a 'lethal threshold', or 'incipient LCs0' as the highest concentration at which less than 50% of the animals die. This is calculated as the mean of the highest concentration showing < 50% mortality (i.e. 4 pg mL -1) and the lowest LCs0 in the bioassay (i.e. 8 pg mL 1). The lethal threshold is sometimes used as an estimate of the MATC, although it is preferable to estimate this latter parameter from chronic toxicity data.

Sublethal toxicity testing It is beyond the scope of this chapter to review in detail all of the toxicological models for evaluating sublethal effects in fish. The rapid developments in this area of fish toxicology, particularily for evaluating molecular, biochemical, immunological and histological responses to chemical exposure, have been documented over the past 15 years in several reviews (Thomas, 1990; Hinton and Lauren, 1990; Stegeman and Hahn, 1993; Arkoosh et al., 1994; Bunton, 1996; Di Giulio et al., 1995; Metcalfe, 1998). A summary of the most commonly used methods for evaluating sublethal responses of fish to chemical exposure is provided in Table 37.7, along with a limited number of references that illustrate the methods used in the sublethal tests. The experimental models vary in the duration of exposures and the complexity and time required to evaluate the endpoints. For instance, responses such as induction of hepatic ethoxyresorufin-deethylase (EROD) or inhibition of brain acetylcholinesterase (AchE) can be evaluated after acute exposures of fish for 24-72 h, and the biochemical methods required to evaluate these responses are relatively rapid and simple. However, chronic exposures over several months may be required to evaluate endpoints such as carcinogenicity or alterations to reproductive potential in fish. In addition, some sublethal endpoints are indicators of exposure to specific classes of chemicals, such as the induction of DNA adducts by aromatic hydrocarbons or the induction of vitellogenin (VtG) synthesis in the livers of male fish by estrogenic substances. Other sublethal responses can occur in fish exposed to many chemical classes, as well as environmental

stressors (e.g. disease, nutrition, temperature, reproductive status). Obviously, it is important to control for all environmental stressors that may influence sublethal responses in these experimental models. Recently, there has been a great deal of interest in endocrine disrupting (modulating) substances, particularly those that alter sexual development and reproductive success in fish and wildlife (Hose and Guillette, 1995). There are a variety of sublethal endpoints that have been used as indicators of endocrine disruption in fish, including: (i) altered serum steroid levels; (ii) synthesis of egg-yolk protein (vitellogenin) in male fish; (iii) alterations to the development of gonadal tissues; and (iv) effects on reproductive success. However, it is fair to say that the process of identifying the most appropriate in vivo models and fish species for assessing the endocrine modulating potential of chemicals is ongoing (Ankley, 1998). For some classes of endocrine disrupting substances, such as those that exhibit anti-androgenic activity, including p,p'-DDE and the fungicide, vinclozolin (Monosson et al., 1997), there are few in vivo models currently available for determining endocrine modulating effects on fish.

Bioassays A'bioassay' is a test in which the quantity of material is determined by the response of a living organism to it. For instance, we may evaluate the concentration of toxic constituents in an industrial effluent by the lethal or sublethal toxic response of organisms to it. The term 'bioassay' should not be confused with toxicity test. Bioassays with fish and other aquatic organisms are commonly used for regulatory purposes to ensure that toxic substances in industrial effluents and wastewaters are not being discharged in excessive quantities. There are standard methods required by regulatory agencies for these bioassay procedures (e.g. ASTM, 1991). Bioassays are also useful for 'Toxicity IdentificationandEvaluation'(TIE),whichisaprocess of systematic treatment of environmental samples (e.g. pH adjustment, filtration, subfractionation) followed by bioassays for toxicity to identify the chemicals in the sample that are responsible for lethal or sublethal effects (US, EPA, 1991). TIE approaches with fish bioassays have been used to identify carcinogens in sediments (Balchetal., 1995), embryotoxic compounds in extracts from Lake Ontario trout (Harris et al., 1994a) and the EROD-inducing compounds in a lampricide formulation (Hewittetal., 1997).

--I 0

X

0 0

O m x m

m

Z 21, r-

O m

TABLE 37.7: Sublethal toxicity endpoints with fish species Endpoint

Reference

(I) Enzyme/protein induction

(a) CYP4501A monooxygenases: EROD

Hodson et al. (1997)

AHH

Janz and Metcalfe (1991)

mRNA

Williams et al. (1997)

(b) Glutathione:

Nishimoto et al. (1995)

(c) Metallothionein:

Hamilton and Mehrle (1986)

(11) Enzyme inhibition

(a) AchE:

Richmonds and Dutta (1992)

(b) Phosphatases:

Sharma (1990)

(111)Enzyme release

Lysosomal:

Dixon et al. (1985)

Hepatic:

Dixon et al. (1987)

(IV) Disturbance in metabolism

Adenine nucleotides

Vetter et al. (1996)

0

Bilirubin

Mattsoff and Oikari (1987)

0

(V) Ionoregulation

McDonald et al. (1989)

X

0

I

(VI) Endocrine disruption

(a) Retinoid:

Palace and Brown (1994)

(b) Thyroid:

Ruby et al. (1993)

(c) Catecholamine:

Black and Nilsson (1990)

(d) Sex steroid:

MacLatchy and Van Der Kraak (1995);

(e) Vitellogenin:

Jobling et al. (1996)

Miranda et al. (1992)

I.lJ

0 ._]

(VII) Reproduction

< z

I.I.I

Arcand-Hoy and Benson (1998); Donaldson (1990) (VIII) Gonad development

Gray and Metcalfe (1997); Gimeno et al. (1997) (IX) Oxidative damage x

Di Giulio et al. (1993); Scarano et al (1994) (X) Growth

Benton et al. (1994) (Xl) Behavior

(a) Courtship:

Schroder and Peters (1988)

(b) Avoidance:

Morgan et ol. (1991)

(c) Swimming:

Little and Finger (1990)

(d) Ventilatory:

Diamond et al. (1990)

(XII) Immunomodulation

Anderson and Zeeman (1995) Zeeman and Brindley (1993) Zelikoff et al. (1991) Arkoosh et al. (1994) (XIII) Carcinogenesis

Bunton (1996); Metcalfe (1989)

"ndpoint

Reference

(XIV) Genotoxicity (b) DNA damage:

Stein et al. (1994) Pandrangi and Petras (1995)

(c) Clastogenic:

AI-Sabti and Metcalfe (1995)

(a) DNA adducts:

(XV) Histopathology Hinton and Lauren (1990) Metcalfe (1998)

Toxic equivalency factors (TEFs) Many of the toxic responses observed in vertebrates exposed to planar halogenated aromatic hydrocarbons (HAHs) such as 2,3,7,8-TCDD are mediated through binding to a specific cellular protein called theAh (aryl hydrocarbon) receptor. One group of proteins synthesized as a result of this molecular response is the hepatic monooxygenases belonging to the cytochrome P4501AI (CYPIAI) subgroup (e.g. EROD, AHH). The degree of induction of CYP4501AI monooxygenases in fish has been used as a biochemical indicator of the toxic potency of PCB congeners (Janz and Metcalfe, 1991; Harris et al., 1994b). In order to standardize estimates of the toxic potency of planar HAHs, Safe et al. (1990) proposed toxic equivalency factors (TEFs) based upon the potencies of various HAHs relative to a reference compound, 2,3,7,8TCDD for severalAh-mediated responses in mammalian models. As outlined above, much of the data describing the structure-activity relationships for the toxicity of planar HAHs has been determined with mammalian in vivo and in vitro models. Since the toxicity of coplanar PCBs and other planar HAHs may vary between vertebrate taxa, there have been several studies focused on determining TEFs for fish. The most complete set of TEF data was developed from embryolarval lethality tests with salmonid species (Walker and Peterson, 1991; Zabel et al., 1995). The TEF data summarized in Table 37.8 were calculated by dividing the LDs0 for embryolarval mortalities obtained by in ovo injection of a planar H A H congener (nmol) by the LDs0 (nmol) for 2,3,7,8-TCDD. These TEFs indicate that several dioxin and dibenzofuran congeners and nonortho (coplanar) PCB congeners are extremely embryotoxic to fish. The PCB congeners with chlorine

substitution in ortho positions all have very low toxic potencies in fish relative to the non-ortho PCB congeners. It is likely that the 'TEF approach' for total planar HAHs will be increasingly used to regulate discharges or to legislate remedial action for dioxin, dibenzofuran and PCB contaminants in the environment.

Overview Although teleost in vitro models with cell lines, primary cell cultures and tissue explants are being used increasingly, there are several reasons why in vivo models will remain in use for toxicity testing of aquatic contaminants. As described earlier in this chapter, the uptake, metabolism and clearance of chemicals in fish are complex processes that cannot be reproduced easily with in vitro models. Chemical toxicity to fish is often affected by external factors, such as photoperiod, temperature, salinity, reproductive status, disease and exposure to other external stressors, and these factors cannot be replicated in in vitro tests. In addition, biochemical, physiological and endocrine systems in fish are often complex and are controlled by positive and negative feedback processes that involve several different tissues and organs. Once again, these conditions are difficult to replicate in in vitro models. In vitro toxicity tests are very useful to initially screen chemicals for toxic potential or to evaluate the mechanisms of toxicity. However, it is safe to assume that researchers and regulators will continue to require in vivo toxicity testing with fish for definitive evidence of the toxicity of chemicals to aquatic organisms. Indeed, regulatory agencies in some countries require in vivo toxicity testing with 'native' fish species before toxicity data are accepted. Techniques for in vivo toxicity testing with fish continue to be

0 X R o o

0 m x m

m

Z r-"

O m ~,~

TABLE 37.8: Toxic equivalency factors (TEFs) for selected polychlorinated dibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs) and PCB congeners as determined from embryolarval mortality data for salmonids Congener

TEF

PCDDs:

Congener

TEF

Non-ortho PCBs: 1.0

Congener 126

0.004 9

1,2,3,7,8-PeCDD

0.659

Congener 81

0.000 62

1,2,3,4,7,8-HxCDD

0.263

Congener 77

0.000 18

1,2,3,6,7,8-HxCDD

0.020

Congener 169

0.000 04

1,2,3,4,6,7,8-HpCDD

0.0015

2,3,7,8-TCDD

Ortho PCBs:

PCDFs: 2,3,4,7,8-PeCDF

0.339

Congener 28

< 0.000 009

1,2,3,4,7,8-HxCDF

0.240

Congener 118

< 0.000 003

1,2,3,7,8-PeCDF

0.032

Congener 105

< 0.000 002

2,3,7,8-TCDF

0.030

Congener 156

< 0.000 001

Source: Zabel et al., 1995.

0 ,=,I 0

_u x 0 I'-

O Url .-.I LLI

a o

developed and refined - especially for sublethal toxicity tests using biochemical indicators. However, we need much more basic research on the biochemical and physiological bases of toxicity in fish in order to evaluate the significance of these types of toxicity data.

References

._1

< z

L,I.I

e~ u.l X

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-4 0 xm 0 f,0

O m x m

m

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0 0 r""

O ,,,,/ O X O I--

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a o

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