OCS Study 

MMS 2005-054 

 

Coastal Marine Institute 

Evaluating Sublethal Effects of Exposure 

to Petroleum Additives on Fishes 

Associated with Offshore Platforms 

Cooperative Agreement

Coastal Marine Institute 

Louisiana State University 

U.S. Department of the Interior

Minerals Management Service 

Gulf of Mexico OCS Region 
OCS Study 

 MMS 2005-054 

Coastal Marine Institute 

Evaluating Sublethal Effects of Exposure 

to Petroleum Additives on Fishes 

Associated with Offshore Platforms 

Editors 

Donald M. Baltz 

Edward J. Chesney 

December 2005 

Prepared under MMS Contract 

1435-01-99- CA-30951-16802 

by 

Coastal Marine Institute 

Louisiana State University 

Baton Rouge, Louisiana 70803 

Published by 

U.S. Department of the Interior Cooperative Agreement 

Minerals Management Service Coastal Marine Institute 

Gulf of Mexico OCS Region Louisiana State University 
iii

DISCLAIMER 

This report was prepared under contract between the minerals Management Service (MMS) and 

Coastal marine Institute. This report has been technically reviewed by the MMS, and it has been 

approved for publication. Approval does not signify that the contents necessarily reflect the 

view and policies of the MMS, nor does mention of trade names or commercial products 

constitute endorsement or recommendation for use. However, it is exempt from review and 

compliance with the MMS editorial standards. 

REPORT AVAILABILITY 

Extra copies of the report may be obtained from the Public Information Office (Mail Stop 5034) 

at the following address: 

 U.S. Department of the Interior 

 Minerals Management Service 

 Gulf of Mexico OCS Region 

 Public Information Office (MS 5034) 

 1201 Elmwood Park Boulevard 

 New Orleans, Louisiana 70123-2394 

CITATION 

Suggested citation: 

Baltz, D. M. and E. J. Chesney. 2005. Evaluating Sublethal Effects of Exposure to Petroleum 

Additives on Fishes Associated with Offshore Platforms. U.S. Department of the 

Interior, Minerals Management Service, Gulf of Mexico OCS Region, New Orleans, LA. 

OCS Study MMS 2005-054. 76 pp. 

ABOUT THE COVER 

The Florida pompano Trachinotus carolinus is commonly caught by anglers around petroleum 

platforms where it is exposed to contaminants associated with the production and transport of oil 

and gas. The swimming performances of juvenile pompano were tested in swimming tunnels 

before and after sublethal exposures to ethylene glycol and methanol to evaluate ecological 

effects on individuals.
v

ACKNOWLEDGEMENTS 

We appreciate the critical comments on earlier manuscripts by Jaye Cable, Kenny Rose, Larry 

Rouse, and several anonymous reviewers, John Hawke for assistance with husbandry, and many 

members of the Louisiana Universities Marine Consortium staff for all manner of assistance, 

especially Russell Martin for fabricating the swimming chambers. Field and laboratory work 

would not have been possible without the help of Marcus Bradley, Cody Short, Murt Verburg, 

Ted Switzer, Ken Riley, and Chris Salmonsen and the captain of the R/V Acadiana, Craig 

LeBoeuf. We are also grateful to the Minerals Management Service for funding (contract 1435-

01-99-CA-30951). 
vii

ABSTRACT 

Recently, new technology has pushed petrochemical exploration into increasingly deeper water 

(>305 m) at increased risks to marine fauna. One risk is from chemical additives used to 

enhance deepwater production, such as ethylene glycol and methanol, which are used during the 

production and treatment of petroleum to prevent the formation of gas hydrates in deepwater 

wells and pipelines. Deepwater petroleum production requires that a number of additives, such 

as ethylene glycol and methanol, be transported and stored offshore in large quantities posing the 

risk of spills. Although these additives may not be highly toxic per se, it is important to evaluate 

and understand the risks to marine organisms, especially fishes that are often strongly associated 

with oil production facilities. Juvenile Florida pompano Trachinotus carolinus were used in 

three separate controlled experiments to test the sublethal effects of 3.0% ethylene glycol (EG), 

1.07% methanol (MeOH) and a combination of the two chemicals (EG + MeOH) on swimming 

performance of individual fish. Florida pompano swimming performance was evaluated by 

comparing differences in pre- and post-exposure critical swimming speeds (Ucrit) for each 

individual to quantify sublethal effects. The experimental protocols included identical fasting, 

exposure, acclimation, and swimming experience for each group and required up to18 days for 

the experimental trial. An additional ethylene glycol exposure experiment was conducted on a 

second species, juvenile Atlantic spadefish Chaetodipterus faber. Ethylene glycol toxicity tests 

on juvenile Florida pompano and subsequent probit analysis identified the LC50 as 5.63 % 

(volume per volume) at 30 practical salinity units and 25 °C. Behavioral observations of 

treatment and control groups during 24-h exposures and 15-h recovery trials showed that 2.1 % 

ethylene glycol was the lowest concentration at which individuals displayed lethargic behavior 

relative to controls after 24 h. Fish exposed to concentrations of 3.1 % or less became lethargic 

or distressed relative to controls, but showed signs of recovery after 15 h in clean seawater. The 

mean (± SE) Ucrit of juveniles (23.1±4.73 g) was evaluated before and after exposure to a 3.0 % 

concentration. The mean Ucrit declined significantly by 13.5 % (P<0.0002) from 95.9 ±2.37 cm/s 

in pre-exposure trials to 83.0 ±3.45 cm/s in post-exposure trials. Exposure to ethylene glycol did 

not impact all individuals to the same degree, and smaller fish experienced a greater percent 

decrease in swim performance (%R) than did larger individuals (%R = - 53.5 + [1.7 x Body 

Mass], R2

 = 0.3148, df = 13, P<0.0295). We also evaluated the toxicity and sublethal effects of 

methanol on the swimming performance of juvenile Florida pompano. A 24-h static exposure

identified the LC50 of methanol as 1.28% (volume per volume, % v/v) at 30 practical salinity 

units and 25°C. The mean (± SE) Ucrit of juveniles (20.5 ± 4.59 g) was evaluated before and 

after exposure to a 1.07% concentration of methanol and showed that Ucrit was significantly 

reduced (P < 0.0002) from 90.1 ± 1.35 cm/s to 84.2 ± 1.36 cm/s in post-exposure trials. 

Following exposure and a 17-h recovery period in clean seawater, the mean 6.5% decline in 

performance of the methanol treatment group contrasted sharply with a mean increase of 4.0% in 

the control group, indicating that conditioning and/or training effects were significantly 

surpassed by the negative sublethal effects of methanol exposure. In a combination experiment 

(i.e., Control, EG, MeOH and EG + MeOH) on juvenile Florida pompano swimming 

performance, single exposures to ethylene glycol and the combination of ethylene glycol and 

methanol significantly reduced Ucrit by 13.0 and 42.0%, respectively. In this experiment, no 

detectable differences in Ucrit were found for Florida pompano exposed to methanol or for the 

controls. Juvenile Atlantic spadefish were used to test the single effects of 3.0% ethylene glycol 

on swimming performance of individuals using the same protocol developed for pompano. 
viii

Treatment fish experienced a 6.9% reduction in Ucrit compared to pre-exposure swimming 

performance and a 17.9% reduction compared to the controls. The reduced ability of Florida 

pompano and Atlantic spadefish to sustain high prolonged and burst performance levels could 

have affected an individual's ability to avoid predators and feed effectively. 
ix

TABLE OF CONTENTS 

LIST OF TABLES ....................................................................................................................... xi

LIST OF FIGURES ...................................................................................................................xiii

1.0 INTRODUCTION............................................................................................................. 1

2.0 SWIMMING PERFORMANCE OF JUVENILE FLORIDA POMPANO 

EXPOSED TO ETHYLENE GLYCOL ......................................................................... 5

2.1 METHODS ........................................................................................................................ 6 

2.1.1 Collection and maintenance of fish......................................................................... 6

2.1.2 Acute toxicity tests................................................................................................... 7

2.1.3 Behavior and recovery experiments........................................................................ 8

2.1.4 Swimming chamber design and calibration............................................................ 9

2.1.5 Measuring critical swimming speed ....................................................................... 9

2.1.6 Statistical analysis ................................................................................................ 11

2.2 RESULTS ........................................................................................................................ 12 

2.2.1 Acute toxicity tests................................................................................................. 12

2.2.2 Behavior and recovery experiments...................................................................... 12

2.2.3 Critical swimming speed....................................................................................... 13

2.3 DISCUSSION ................................................................................................................... 16 

2.3.1 Individual variation in Ucrit................................................................................... 16

2.3.2 Size effects............................................................................................................. 17

2.3.3 Mode of action ...................................................................................................... 17

2.3.4 Ethylene glycol toxicity......................................................................................... 18

2.3.5 Marine exposures.................................................................................................. 19

2.3.6 Threshold of tolerance .......................................................................................... 19

2.3.7 Ecological consequences of sublethal effects ....................................................... 20

3.0 TOXICITY AND SUBLETHAL EFFECTS OF METHANOL ON SWIMMING 

PERFORMANCE OF JUVENILE FLORIDA POMPANO ...................................... 21

3.1 METHODS ...................................................................................................................... 22 

3.2 RESULTS ........................................................................................................................ 25 

3.2.1 Acute toxicity and recovery tests........................................................................... 25

3.2.2 Critical swimming speed....................................................................................... 26

3.3 DISCUSSION ................................................................................................................... 28 

3.3.1 Methanol toxicity and mode of action................................................................... 28

3.3.2 Sublethal effects on Ucrit........................................................................................ 29

3.3.3 Individual variation in Ucrit................................................................................... 29

3.3.4 Marine exposures.................................................................................................. 30

3.3.5 Ecological consequences of sublethal effects ....................................................... 31

4.0 SWIMMING PERFORMANCE OF JUVENILE FLORIDA POMPANO AFTER 

SUBLETHAL EXPOSURE TO ETHYLENE GLYCOL AND METHANOL: 

SYNERGISTIC EFFECTS............................................................................................ 33

4.1 MATERIALS AND METHODS ........................................................................................... 34 

4.1.1 Swimming tunnel design ....................................................................................... 34
x

4.1.2 Collection and exposure ....................................................................................... 35

4.1.3 Swimming performance ........................................................................................ 36

4.1.4 Plasma samples..................................................................................................... 36

4.1.5 Statistical analysis ................................................................................................ 38

4.2 RESULTS ........................................................................................................................ 39 

4.2.1 Swimming performance ........................................................................................ 39

4.2.2 Plasma samples..................................................................................................... 41

4.3 DISCUSSION ................................................................................................................... 41 

4.3.1 Combination exposure .......................................................................................... 42

4.3.2 Individual variation .............................................................................................. 43

4.3.3 Ecological effects.................................................................................................. 43

5.0 SWIMMING PERFORMANCE OF JUVENILE ATLANTIC SPADEFISH 

EXPOSED TO SUBLETHAL CONCENTRATIONS OF ETHYLENE GLYCOL 45

5.1 MATERIALS AND METHODS ........................................................................................... 47 

5.1.1 Respirometer design.............................................................................................. 47

5.1.2 Collection and exposure ....................................................................................... 48

5.1.3 Swimming performance ........................................................................................ 49

5.1.4 Plasma concentration ........................................................................................... 49

5.1.5 Statistical analysis ................................................................................................ 50

5.2 RESULTS ........................................................................................................................ 50 

5.2.1 Swimming performance ........................................................................................ 50

5.2.2 Plasma samples..................................................................................................... 53

5.3 DISCUSSION ................................................................................................................... 54 

5.3.1 Ethylene glycol...................................................................................................... 55

5.3.2 Modes of action..................................................................................................... 55

5.3.3 Individual variation .............................................................................................. 55

5.3.4 Spills...................................................................................................................... 56

5.3.5 Ecological effects.................................................................................................. 56

5.4 CONCLUSION.................................................................................................................. 57 

6.0 GENERAL SUMMARY ................................................................................................ 59

7.0 LITERATURE CITED .................................................................................................. 61
xi

LIST OF TABLES 

Table 2-1. Behavior classifications used to describe the activity patterns of Florida pompano 

exposed to several ethylene glycol concentrations for 24 h with a 15-h recovery 

period following exposure. ......................................................................................8

Table 2-2. Calculated lethal concentrations (LC; % volume per volume) and associated 95% 

confidence limits for juvenile Florida pompano exposed to ethylene glycol for 

24h..........................................................................................................................12 

Table 2-3. Comparative mortality and behavior of juvenile Florida pompano after 24-h 

exposure to treatment (2.06-4.12 %) and control (0%) levels of ethylene glycol 

and after a 15-h of recovery in clean seawater. Five fish were exposed to each 

concentration..........................................................................................................13 

Table 2-4. Estimated LC50 values (mean +/- SE) for several fish species exposed to ethylene 

glycol......................................................................................................................18

Table 3-1. Calculated lethal concentrations (LC, % v/v) values and associated 95% 

confidence limits for juvenile Florida pompano exposed to methanol for 24 h. ...25 

Table 4-1. Individual fork length (fl mm), body mass (g), and pre- and post-exposure ucrit

(cm/s, bl/s) values for treatment (methanol, ethylene glycol, and combined 

[m+eg]) and control groups of juvenile florida pompano. Statistical summaries 

are means (±1se) for size variables and lsmeans for absolute and relative ucrit

values. ....................................................................................................................37 

Table 5-1. Individual Body Mass (g), Fork Length (mm) and pre-exposure and post-exposure 

Ucrit (cm s-1, BL s-1) for treatment and control groups of Atlantic spadefish.........52 
LIST OF FIGURES 

Figure 2-1. Individual variation in the critical swimming speeds (Ucrit) of juvenile Florida 

pompano (N=15) before and after a 24-h exposure to 3.0 % ethylene glycol. ......14

Figure 2-2. Critical swimming speeds (Ucrit; body length/s) of juvenile Florida pompano 

before and after a 24-h exposure to 3.0 % ethylene glycol (treatment fish, 1-15) 

and before and after sham exposure (control fish, 16-21). ....................................15 

Figure 2-3. The relationship between body mass of Florida pompano (N=15) and percentage 

reduction in critical swimming speed (Ucrit) after exposure to 3.0 % ethylene 

glycol......................................................................................................................16 

Figure 3-1. The regression of methanol concentration in blood plasma, control, and exposure 

waters following exposure to concentrations ranging from 0.99 to1.49% v/v. 

Points are estimates of pooled samples from survivors in control and treatment 

groups.....................................................................................................................26 

Figure 3-2. Critical swimming speeds, in Body Lengths per second (BL/s), of juvenile Florida 

pompano before and after a 24-h exposure to 1.07% methanol (treatment fish, N = 

16) and before and after sham exposure (control fish, N = 6). ..............................27 

Figure 3-3. Size-related changes in mean critical swimming speeds (BL/s) of juvenile Florida 

pompano in treatment (T) and control (C) groups before and after exposure and 

sham exposure to 1.07% methanol for large (T: 106-119 mm, N = 5; C: 103-105 

mm, N= 2), medium (T: 100-105 mm, N = 5; C: 100-102 mm, N= 2), and small 

(T: 85-99 mm, N = 6; C: 92-99 mm, N= 2) individuals. .......................................28 

Figure 4-1. Critical swimming speeds (Ucrit, LSMean (±1 SE)) of juvenile Florida pompano 

before and after 24-h exposures to 1.07% (v/v) methanol (N = 10), 3.0% (v/v) 

ethylene glycol (N = 12), the combined chemicals (N = 7), and sham-exposed 

controls (N = 10)....................................................................................................40 

Figure 4-2. Individual variation in the critical swimming speeds (Ucrit) of juvenile Florida 

pompano before and after 24-h exposures to 1.07% methanol (N=10), 3.0% 

ethylene glycol (N = 12), the combined chemicals (N = 7), and sham-exposed 

controls (N = 10)....................................................................................................41 

Figure 5-1. Mean (± SE) critical swimming speeds (Ucrit) of juvenile Atlantic spadefish before 

and after a 24-hr exposure to 3.0 % ethylene glycol (N = 25) and sham controls 

(N = 6)....................................................................................................................51 

Figure 5-2. Critical swimming speeds (Ucrit) of juvenile Atlantic spadefish before and after a 

24-hr exposure to 3.0 % ethylene glycol (N = 25) and sham controls (N=6)........53 

Figure 5-3. Individual variation in the critical swimming speeds (Ucrit) of juvenile Atlantic 

spadefish (N = 25) before and after exposure to 3.0 % ethylene glycol and sham 

controls (N = 6)......................................................................................................54

xiii
1

1.0 INTRODUCTION 

D. M. Baltz1, 2 and E. J. Chesney3 

1

Department of Oceanography and Coastal Sciences, Louisiana State University, Baton Rouge, 

LA 70803-7503 

2

Coastal Fisheries Institute, Louisiana State University, Baton Rouge, LA 70803-7503 

3

Louisiana Universities Marine Consortium, Chauvin, LA 70344 

A tremendous amount of research has been directed towards understanding the effects of 

offshore oil development in the northern Gulf of Mexico (Gallaway et al. 1981; Grizzle 1986; 

Stanley and Wilson 1990; Grossman et al. 1987; Boehm et al. 2001, Hymel et al. 2002). 

Recently, declining inshore production, rising oil prices and new technology have expanded 

petrochemical exploration into increasingly deeper water (>300 m) at increased risks to the 

marine fauna including the release of contaminants and noise pollution associated with the 

construction of offshore oil platforms. One type of risk is from chemical additives that are used 

to enhance deep-water production. Oil and gas platforms are a source of anthropogenic inputs 

including contaminated waters that are separated from crude oil at the extraction point (i.e., 

produced formation water), drilling-fluid chemicals, oil-based and water-based drilling muds and 

cuttings, and production additives, some of which are toxic to the fauna that reside on and near 

oil production structures (Holdway 2002). 

Additives such as ethylene glycol and methanol are often transported long distances from shore 

by underwater pipelines up to 96 km in length and by ships to deep-water production platforms 

where they are stored in volumes of up to 300,000 L per platform (Boehm et al. 2001). These 

additives are used to prevent the formation of gas hydrates in deep-water wells and pipelines 

(Anonymous 1996, 2000; Herzhaft and Dalmazzone 2000; Boehm et al. 2001). Although the 

possibility of a large spill is unlikely, the risk remains plausible with increased usage, transport 

and storage of large volumes of ethylene glycol and methanol (Boehm et al. 2001). Although we 

are not aware of any instance were both ethylene glycol and menthanol are used at the same 

production facility, multiple other chemicals and proprietary mixtures of chemicals are 

commonly used together in production and transport (Boehm et al. 2001). 

Little attention has been paid towards understanding chronic or acute sublethal effects associated 

with low-level, localized discharges of toxicants, which may reduce an individual fish’s ability to 

avoid predators, feed, reproduce, and resist diseases and parasites. Effects at the level of the 

organism may, in turn, lead to effects at the population and community levels (Weis et al. 1999). 

It is, therefore, important to quantify sublethal effects at the individual level as a prelude to 

understanding population and community responses to these impacts. 

Ethylene glycol and methanol are toxic to fishes to varying degrees. Median lethal 

concentrations of ethylene glycol have been reported for various brackish and freshwater 
2

organisms (Bridie et al. 1979; Adbelghani et al. 1989; Pillard 1995; Green and Kocan 1997). 

However, ethylene glycol LC50 values have been reported for only a few freshwater fishes 

including the rainbow trout Oncorhynchus mykiss, fathead minnow Pimephales promelas

(Pillard 1995; Greene and Kocan 1997), and bluegill sunfish Lepomis macrochirus (Abdelghani 

et al. 1989). Methanol LC50 values have been reported for various freshwater and two salttolerant or saltwater fishes by Bengtsson et al. (1984) for the bleak Alburnus alburnus and by 

Portmann and Wilson (1971) for the hooknose Agonus cataphractus . 

We used critical swimming speed (Brett 1964) to quantify the capacity of individual fish for 

prolonged swimming activity. Swimming performance has been used to determine tolerances of 

fishes to varying environmental stresses and pollutants (Cairns 1966; Sprague 1971; Nikl and 

Farrell 1993; Beaumont et al.1995; Heath 1995; Kolok et al. 1998; Hymel et al. 2002). The 

measurement of swimming performance was first suggested as an important criterion in the 

determination of sublethal effects of toxicants on fishes by Cairns (1966) and has since been 

used to determine tolerances of fishes to a variety of environmental factors and pollutants (Nikl 

and Farrell 1993; Beaumont et al. 1995; Heath 1995; Kolok et al. 1998; Hymel et al. 2002). 

Swimming performance is associated with the ability of fish to migrate, feed, escape predation 

and maintain position in a current and involves the integrated effects of numerous physiological 

processes (Beamish 1978). Selyle (1950) first defined stress as the sum of all the physiological 

responses by which an animal tries to maintain or re-establish a normal metabolism in the face of 

a physical or chemical force. Stress and resisting a stressor are energy draining processes 

(Schreck 1982; Barton and Schreck 1987) and while responding to stress an organism should 

have less energy available to devote to other life functions (Schreck 1990). Thus estimates of 

swimming performance provide an index of stress. 

Juvenile Florida pompano Trachinotus carolinus and Atlantic spadefish Chaetodipterus faber

were used to evaluate the sublethal effects of ethylene glycol and methanol because they occur in 

the vicinity of offshore petroleum production platforms and pipelines in the northern Gulf of 

Mexico. Pompano and spadefish are important commercial and sport fishes. These species were 

also chosen because they are excellent fish models for many similar species such as amberjack, 

crevalle jack and other species that reside offshore and are strongly associated with reefs and 

other structures including pipelines and platforms. Pompano apparently have a protracted 

spawning season in the northern Gulf of Mexico during the spring and late summer (Iverson and 

Berry 1969; Finucane 1969; Nelson and Murphy 2001) and may spawn year round in the tropical 

regions of the Gulf of Mexico (Berry and Iverson 1967). Spadefish are believed to spawn from 

May through September, principally in the northern Gulf of Mexico (Ditty et al. 1993). Both 

species are believed to spawn offshore and juveniles usually migrate inshore to feed and the 

young of the year move back offshore when temperatures cool (Finucane 1969; Johnson 1978; 

Gilbert and Parsons 1986; Hayse 1989; Nelson and Murphy 2001). 

The goals of this study were to understand the separate and combined sublethal effects of 

ethylene glycol and methanol on the swimming performance of individual fishes with Florida 

pompano and Atlantic spadefish serving as the fish models. When testing for the sublethal 

effects of contaminants it is common practice to first find an appropriate LC50 for the species of 

interest. Hymel et al. (2002) and Baltz et al. (2005) found LC50 values for ethylene glycol and 

methanol on juvenile Florida pompano. In genral exposure concentrations were based on LC50, 

behavior, and recovery experiments detailed in Hymel et al. (2002) and Baltz et al. (2005). 
3

Because of the difficulty of collecting numbers of similarly sized Atlantic spadefish, we did not 

conduct an LC50 range finding experiment for that species. 

Chapters 2 through 5 are standalone documents describing four sets of experiments that are or 

will be published. Chapters 2 and 3 report on the toxicities and sublethal effects on swimming 

performance of ethylene glycol (Hymel et al. 2000) and methanol (Baltz et al. 2005), 

respectively, on Florida pompano. Chapter 4 reports on significant reductions in critical 

swimming speed after Florida pompano were exposed to combinations of both chemicals in a 

randomized experimental design to test for any additive, synergistic, or antagonistic sublethal 

effects that ethylene glycol and methanol might have on swimming performance. And Chapter 5 

reports on a single exposure of Atlantic spadefish to ethylene glycol in a randomized 

experimental design to examine potential sublethal effects on a marine species with a different 

physiology and ecology. 
5

2.0 SWIMMING PERFORMANCE OF JUVENILE FLORIDA POMPANO 

EXPOSED TO ETHYLENE GLYCOL 

M. E. Kay1

, D. M. Baltz1,2, E. J. Chesney3, M. A. Tarr4 and A. S. Kolok5 

1

Department of Oceanography and Coastal Sciences, Louisiana State University, Baton Rouge, 

LA 70803-7503 

2

Coastal Fisheries Institute, Louisiana State University, Baton Rouge, LA 70803-7503 

3

Louisiana Universities Marine Consortium, Chauvin, LA 70344 

4

Department of Chemistry, University of New Orleans, New Orleans, LA 70148 

5

Biology Department, University of Nebraska at Omaha, Omaha, NE 68182

Marine fishes face an array of environmental contaminants that may reduce their ability to 

perform in ecological situations. The effects of factors producing direct mortality in populations, 

such as fishing, are often easier to quantify than sublethal impacts. Though not responsible for 

direct mortality, sublethal impacts may affect scope for activity (Cech 1990), swimming 

performance (Brett 1964, Wedemeyer et al. 1990), growth, reproduction, and survival of 

individuals (Little et al. 1990), with consequences for population size and biomass. 

Quantification of sublethal effects at the individual level is an important step in predicting how 

individuals and populations may respond ecologically to these impacts. 

Marine fisheries are potentially at risk from releases of chemical additives from deepwater 

petroleum production operations. With the development of new technologies, oil and gas 

production is now expanding into greater ocean depths (> 305 m), thus posing a new array of 

environmental risks to marine fishes. Pure chemicals like ethylene glycol and methanol may be 

transported in small supply lines (i.e., umbilicals) up to 96.5 km in length, stored in large 

volumes and used at high rates. This expansion has created a need for the ecological evaluation 

of new practices and chemicals used in petroleum production and transport (Boehm et al. 2001, 

Minerals Management Service 2002). One such contamination risk comes from the use of 

ethylene glycol in the production of petroleum products from deepwater wells to prevent the 

formation of gas hydrates (Anonymous 1996, 2000; Herzhaft and Dalmazzone 2000; Boehm et 

al. 2001), which can potentially clog pipelines. 

The environmental effects of ethylene glycol have been widely studied in relation to its use as an 

antifreeze, coolant, and deicing agent (Barber 1999). Median lethal concentrations (LC50) have 

been reported for various freshwater and brackish-water organisms (Abdelghani et al. 1989; 

Pillard 1995; Greene and Kocan 1997). The toxicity of ethylene glycol to fishes has been 

assessed on only a few species, including rainbow trout Oncorhynchus mykiss, fathead minnow 
6

Pimephales promelas ( Pillard 1995; Greene and Kocan 1997), bluegill sunfish Lepomis 

macrochirus (Abdelghani et al. 1989), and goldfish Carassius auratus (Bridie et al. 1979). 

Sublethal exposure levels can lead to important individual effects measurable by diminished 

performance (Sprague 1971), which in turn creates an enhanced risk of mortality or an impaired 

ability to perform normal life functions (Schreck 1990). Sublethal effects may potentially lead to 

declining numbers in a population (Heath 1995). Changes in critical swimming speed (Ucrit) 

have been used to indicate fish performance in many studies assessing sublethal effects of 

contaminants (Nikl and Farrell 1993; Beaumont et al. 1995; Heath 1995; Kolok et al. 1998). 

Reduced swimming performance may have ecological significance, such as impairing the ability 

of larval fishes to avoid predators and catch prey, but has not been demonstrated for larger 

individuals. Reduced swimming performance and prey capture ability has been demonstrated in 

fish larvae exposed to sublethal concentrations of methylmercury (Zhou et al. 1996; Samson et 

al. 2001). Vulnerability to predation is dynamic, and exposure to contaminants could increase 

such vulnerability by effecting swimming speed (Beauvais et al. 2000), perception (Baker and 

Montgomery 2001), and behavior (Weis et al. 1999, 2001, Scholz et al. 2000, Wibe et al. 2001). 

The Florida pompano Trachinotus carolinus was chosen for assessment of acute and sublethal 

effects of exposure to ethylene glycol because the species occurs in the vicinity of transport 

pipelines and deepwater wells in the Gulf of Mexico. A marine fish of considerable economic 

importance to sport and commercial fisheries (Gilbert 1986) the Florida pompano is found in 

Atlantic and Gulf of Mexico waters from Massachusetts to Brazil (Hoese and Moore 1977). 

Juveniles are abundant in and near the surf zone during summer months (Bellinger and Avault 

1970). Adults can be found in inshore and offshore waters and spawn on the continental shelf 

(Gilbert 1986). 

We addressed our goal in assessing the acute and sublethal effects of ethylene glycol on juvenile 

Florida pompano swimming performance by identifying a short-term LC50 and a range of lower 

sublethal concentrations and by testing for changes in Ucrit following sublethal exposure. The 

objectives were to provide an estimate of the 24-h LC50, to identify sublethal exposure levels 

from which individuals could apparently recover after short-term exposures expected in open 

marine systems (Boehm et al. 2001), and to evaluate sublethal effects by comparing changes in 

individual swimming performance before and after exposure to ethylene glycol. We found that 

the mean swimming performance was significantly reduced by 14 % in individuals after a 24-h 

exposure to a 3.0 % concentration of ethylene glycol and a 17-h recovery period in clean 

seawater. These results provide a clearer understanding of the possible ecological consequences 

from sublethal spills or leakage in the marine environment.

2.1 Methods 

2.1.1 Collection and maintenance of fish 

Juvenile Florida pompano ranging in body mass from 0.5 to 16.0 g were collected for toxicity, 

behavior, and swimming experiments. Individuals were collected in May-June 1999 (0.5 to 6.0 

g), and again in August 1999 (2.0 to 16.0 g), from Port Fourchon Beach and Isle Dernieres, 

Louisiana, with a 3.05 m (10 ft) beach seine. Fish were held for about 4 weeks in a 5,000-L 

recirculating seawater system at 30 practical salinity units (psu) before testing. During holding, 
7

the fish were treated with 0.2 mg/L copper (Cutrine-Plus) to eliminate ectoparasites, especially 

Amyloodinium (Cheung 1993). Water quality parameters, including pH, salinity, ammonia level, 

and copper concentrations were monitored several times per week during acclimation and daily 

during experiments. Amquel (HOCH2SO3Na) was added to holding systems to control the 

ammonia level when it exceeded 0.5mg/L. Fish were fed a commercially produced pellet feed 

(0.25 g day-1 individual-1; 45 % protein, 8 % lipid) except within the 15-17 h prior to and during 

acute toxicity tests, behavioral experiments, and swimming experiments. 

2.1.2 Acute toxicity tests 

Twenty-four hour static tests were conducted to determine the LC50 of ethylene glycol. Lower 

concentrations were then tested to allow identification of a sublethal concentration for 

subsequent swimming experiments. Three LC50 experiments were conducted in a system of 

eight conical fiberglass tanks. Each tank was filled to 70 L with ultra violet (UV)-sterilized, 30-

psu seawater. The tanks were individually aerated and maintained at 25.0±1 °C with aquarium 

heaters. Four to six logarithmically spaced exposure concentrations plus a control were included 

in each test, and the exposure concentrations were randomly assigned to the tank array. A 'blank' 

tank without fish was also included to monitor the mixing and concentration of the chemical over 

the 24-h. Ten fish were randomly assigned to each of the exposure and control tanks before the 

addition of ethylene glycol (Fisher Scientific, certified). We maintained a 20-min interval 

between the sequential dosing of each tank to allow for processing of the fish at the end of the 

experiment. The activity patterns of the fish in the treatment groups relative to the controls were 

recorded during the second and third tests and in subsequent behavior tests, and were classified 

according to the criteria in Table 2-1. Water samples (1.5 mL) for gas chromatographic (GC) 

analysis, taken periodically from each tank during the second and third tests, showed that the 

ethylene glycol exposure solutions were accurately estimated, well mixed, and stable throughout 

the 24-h experiment (Kay 2000). 

The initial LC50 range-finding test was conducted on Florida pompano with six ethylene glycol 

concentrations ranging from 0.56 to 10.00 % volume per volume. In fish from the first 

experiment, the mean (± SE) BM was 7.6 ± 0.4 g and fork length (FL) was 71.4 ± 1.1 mm. Four 

exposure concentrations, ranging from 3.59 to 7.16 %, were used in the second experiment. The 

mean body mass of fish in the second experiment was 8.6 ± 1.1 g, and the mean FL was 72.7 ± 

1.3 mm. Because the body size of the fish was consistent between the two experiments, and the 

experiments were conducted only 6 d apart, we pooled the data to determine LC values with the 

Environmental Protection Agency’s Probit Analysis Program, Version 1.5 (Lazaorchak 1994). 

We collected blood samples from each fish to analyze for the effects of ethylene glycol exposure. 

At the end of each experiment, blood was drawn from the caudal vessels of surviving 

individuals, anesthesia with tricaine methanesulfonate (MS-222; ≤ 50 mg/L), into heparinized 

capillary tubes for GC analyses. After a 10-min centrifugation at 12,800±1,200 revolutions per 

minute (RPM), the plasma samples from individuals of each treatment group was pooled and 

frozen until GC analyses were performed. 
8

Table 2-1. Behavior classifications used to describe the activity patterns of Florida pompano 

exposed to several ethylene glycol concentrations for 24 h with a 15-h recovery 

period following exposure.

Classification Definition 

Normal activity Swimming activity was similar to control fish 

Slightly lethargic First signs of decreased swimming activity relative 

 to control fish 

Severely lethargic Significantly decreased swimming activity 

relative to control fish 

Severely stressed No longer able to swim; still operculating 

Dead No longer operculating 

Recovering Initial signs of improvement following exposure, 

but activity still decreased relative to control fish 

Completely recovered Swimming activity is completely improved 

and is similar to control fish 

2.1.3 Behavior and recovery experiments 

Following the LC50 tests, two behavior and recovery experiments were conducted prior to the 

swimming experiments to identify a sublethal concentration of ethylene glycol from which 

juvenile Florida pompano could apparently recover. In recovery experiment 1, the mean BM of 

the fish was not determined, but the fish were similar in size to the second experiment conducted 

15 d later. In recovery experiment 2, the mean (± SE) body mass of the fish was 16.9 ± 1.0 g and 

the mean FL was 93.6 ± 1.7 mm. Exposure of the fish to ethylene glycol was conducted in the 

same manner as in the LC50 tests. Behavior and mortality were recorded periodically during the 

24-h exposure, as well as during a 15-h recovery period in seawater containing no ethylene 

glycol. At 24 h, fish were switched to clean tanks for further observation. The 15-h recovery 

period approximated the post-exposure time interval before the swimming tests would be 

initiated. In the first experiment, three fish per treatment tank were exposed to 0.0, 1.0, 1.25, and 

1.5 % concentrations of ethylene glycol. In the second experiment, five fish per tank were 

exposed to concentrations equivalent to 100, 75, 50, and 0 % of the LC1 (i.e., 4.1, 3.1, 2.1, and 0 

% ethylene glycol respectively). Behavior was recorded during both experiments and blood 

plasma samples for GC analysis were collected and pooled by exposure level from all surviving 

fish after the second experiment (N = 5 survivors for 0%, 5 for 2.1 %, 4 for 3.1 %, and 3 for 4.1 

%). 
9

2.1.4 Swimming chamber design and calibration 

Three Blazka-type swimming respirometers were used in simultaneous swimming performance 

tests. Each unit consisted of two concentric acrylic cylinders, with the inner cylinder having an 

inside diameter of 20.32 cm. The volume of each unit was 104 L, and the maximum water 

velocity was 100 cm/s. Water flow within the swimming chamber was generated by a propeller 

at one end of the inner cylinder and was controlled by a variable-speed DC motor. A plastic grid 

was placed at each end of the inner cylinder to contain fish and to dampen turbulence in the 

swimming chamber. 

We individually calibrated each swimming chamber to determine the relationship between RPM 

and water velocity (cm/s) by two separate methods, one intrusive and one unintrusive. In the 

intrusive method, a Marsh-McBirney model 2000 portable flowmeter was inserted into the 

swimming chamber near the outflow end of the inner tube. A non-contact tachometer (Monarch 

model 100) was used to measure RPM of the propeller shaft. Revolutions were increased by 

100-RPM increments until a value of 1,700-1,900 RPM was reached. Water velocity (cm/s) was 

recorded at each increment. The unintrusive method involved video analysis of particles (brine 

shrimp eggs) moving along the central axis of the inner chamber. A video camera positioned 

perpendicularly to the flow recorded moving particles as RPM was increased by 100-RPM 

increments. The velocity (cm/s) of a randomly chosen particle at each corresponding RPM was 

determined using the Optimas 6.5 Imaging System. The regression relationship between RPM 

and water velocity for each method was determined independently for each swimming chamber 

(Kay 2000). The linear models produced from the video analyses: 

Swimming chamber 1: Velocity=0.062(RPM)-4.0099; (r

2

=0.9452, P<0.001), 

Swimming chamber 2: Velocity=0.0627(RPM)-1.1029; (r

2

=0.9715, P<0.001), and 

Swimming chamber 3: Velocity=0.0571(RPM)+0.0582; (r

2

=0.9886, P<0.001). 

The models were used to solve for the RPM setting at each desired water velocity increment used 

in the swimming trials. 

2.1.5 Measuring critical swimming speed 

Critical swimming speed before and after sublethal exposure to ethylene glycol was measured 

and compared for individual fish (Kolok et al. 1998). To keep track of individuals, 21 juvenile 

Florida pompano (mean ± SD: 23.1±4.73 g; 102.8±6.21 mm FL) were marked 11 weeks before 

the start of the experiment. Fish were anesthetized using MS-222 (≤ 50 mg/L) before marking. 

A unique scheme of colored latex marks (Northwest Marine Technology, Inc.) was applied by 

injection to the dorsal, caudal, and anal fins of each individual. After injection, the fish were 

held in the 5,000-L recirculating seawater system. 

Several days prior to the experiment, the marked fish were randomly separated into groups of 

three, which were tested simultaneously in the three swimming chambers. Two of the seven 

groups were randomly assigned to the control, and the remaining five groups were assigned to 
10

treatments. Each group was randomly assigned to one of seven 60-L, interconnected holding 

tanks in a recirculating system with a total volume of 550 L. The holding system was equipped 

with a biofilter and UV sterilizer to maintain water quality. Salinity in the holding system was 

maintained at 30±1 psu, and temperature was maintained at 25.0±1 °C using aquarium heaters in 

each tank. Holding tanks were individually aerated with air stone. 

The 16-d protocol for the swimming experiment ensured identical fasting, holding, and 

acclimation periods for each group of fish (Kay 2000). The protocol allowed for a 17-h 

acclimation to the swimming chambers following the 24-h exposure and included a 41-h fasting 

period (i.e., 24-h exposure and 17-h acclimation) before pre- and post-exposure swimming trials. 

All exposures and sham exposures were carried out in one of three round fiberglass tanks, filled 

to 60 L with 30-psu seawater. The exposure tanks were individually aerated, and the water 

temperature was maintained at 25.0±1 °C with aquarium heaters. All fish were sham-exposed 

before the first swimming trial, but only the control groups were sham-exposed before the 

second swimming trial. Treatment groups were exposed to a 3.0 % ethylene glycol solution 

before the second swimming trial. To confirm concentrations, we collected water samples for 

analyses from the exposure tank at the beginning and end of each exposure period to confirm 

concentrations. All exposed and sham-exposed fish were placed in a bucket of clean sea water 

for several minutes before being placed in the swimming chambers. 

During the 17-h acclimation period, the three swimming chambers were connected to a larger 

recirculating system that included a UV sterilizer and a biofilter. The entire swimming system 

was housed in a 25°C temperature-control room. The total recirculating volume of the system 

was about 765 L. After exposure, each fish was placed in one of the swimming chambers for the 

acclimation interval at a velocity of 5 cm/s. Before the start of a swimming trial, each swimming 

chamber was cleared of bubbles and sealed. Individuals were initially assigned randomly to a 

swim chamber for the pre-exposure swimming trial, but were re-tested in the same swimming 

chamber for the post-exposure swimming trial. 

The experimental protocol tested for changes in the swimming performance of individuals 

exposed to ethylene glycol by determining pre- and post-exposure critical swimming speeds 

(Ucrit) of each individual. The water velocity was set at 10 cm/s at time zero. During the tests, 

the velocity of each swimming chamber was increased by 10 cm/s every 10 min, with no rest 

interval between steps. Each swimming test was terminated when the fish could no longer 

continue swimming and made contact with the back grid for about 2-4 s. A pre- and postexposure Ucrit value was calculated for each individual following Brett (1964): 

Ucrit = ui + (ti/tii × uii) 

where ui is the highest velocity maintained for the full prescribed period (10 min), uii is the 

highest velocity increment, ti is the time the fish swam in the fatigue velocity interval, and tii is 

the prescribed period of swimming. All fish that failed to fatigue after 10 min of swimming at 

the swimming chamber’s maximum speed (100 cm/s) were assigned a Ucrit value of 100.0 cm/s. 

After each post-exposure swimming trial, the fish were anesthetized with MS-222 (≤ 50 mg/L), 

and body masses and fork lengths were measured. Blood was drawn from the caudal vessels into 

heparinized capillary tubes, and centrifuged in the same manner as the LC50 blood samples. 

Plasma was combined for the three fish from each group and frozen for GC analysis. 
11

2.1.6 Statistical analysis 

The Ucrit values were expressed in absolute (cm/s) and relative (body lengths per second, BL/s) 

terms. Statistical analyses were performed on the relative values of Ucrit to standardize for small 

variations in individual size. A two-factor analysis of variance (ANOVA) with a randomized 

block design tested for differences in pre- and post-exposure relative Ucrit values within and 

between treatment and control groups. Pre- and post-exposure relative Ucrit values were paired 

for each individual, with the fish identification number as the random blocking variable. 

Differences between the least square means were calculated for all possible combinations of the 

treatment and control groups. Residuals were tested for normality, and no transformations were 

required. We used SAS computer software (SAS Institute 1996) to test the model, Yijk = μ + τ1i 

+ τ2j + βk + (τ1 τ2)ij + εijk, where τ1 is the trial treatment variable (pre or post), τ2 is the group 

treatment variable (treatment or control) and β is the random blocking variable (individual fish). 

We evaluated individual variation in the performance of treatment fish by testing for the 

influence of size on absolute Ucrit. We also used a linear regression analysis (SAS Institute 1996) 

to test for a relationship between body mass and percentage reduction in the absolute Ucrit of 

treatment fish. 
12

2.2 Results 

2.2.1 Acute toxicity tests 

The 24-h LC50 value was identified for juvenile Florida pompano over a series of experiments 

that included 11 concentrations ranging from 0% to 10.0 % (volume per volume) ethylene 

glycol. In the second and third experiments, mortality of all 10 fish was observed at 

concentrations above 7.16 %, and no sublethal behavioral effects were observed at 

concentrations below 1.0 %. At concentrations of 1.0-7.16 %, 21 individuals exhibited lethargic 

behavior, 31 were severely stressed and 10 died. Combined data from the second and third 

experiments resulted in a 24-h LC50 estimate of 5.63 % (Table 2-2). The lowest experimental 

concentration at which mortalities were observed was 4.52 %. The simple linear regression of 

ethylene glycol plasma concentrations on exposure concentrations from eight tanks with 

surviving individuals was positive and significant (F=59.167; df = 1, 6; P<0.0003; r2

=0.91). 

Table 2-2. Calculated lethal concentrations (LC; % volume per volume) and associated 95% 

confidence limits for juvenile Florida pompano exposed to ethylene glycol for 

24h.

Concentration 95 % Confidence limits

LC level Lower Upper 

1 4.12 2.776 4.650

5 4.51 3.390 4.955

10 4.74 3.765 5.135

15 4.90 4.036 5.266

50 5.63 5.216 6.084

85 6.47 6.005 7.890

90 6.68 6.157 8.459

95 7.02 6.379 9.397

99 7.69 6.796 11.479 

2.2.2 Behavior and recovery experiments 

Although slightly to severely lethargic behavior was observed in treatment fish at ethylene glycol 

concentrations of 1.0% and 1.77 % during the LC50 experiments, sublethal behavioral effects 

were not observed at any of the concentrations (0, 1.0, 1.25, and 1.50 %) used in the first 

behavior-recovery experiment. This possible improvement in tolerance may have been due to an 

increase of 8.93 g in the mean body mass of the fish by the time of this behavioral experiment. 

Noticeable differences in behavior were observed between fish exposed to each concentration 

used in the second behavior-recovery experiment. A 2.1 % concentration was the lowest at 

which lethargic behavior was observed at 24 h (Table 2-3). A 4.1 % concentration produced 

severe stress at 24 h and the death of two fish before the end of the 15-h recovery period. At the 

3.1 % concentration, the death of one fish was observed only 7 h into the experiment. At 24 h, 

three of the remaining fish were severely lethargic and disoriented and the fourth was in severe 
13

stress. All four fish, however, improved to a slightly lethargic state by the end of the 15-h 

recovery period. Because the next highest concentration (4.1 %) had produced no deaths at 24 h, 

it was concluded that the single mortality at the 3.1 % concentration might have been due to the 

random inclusion of a stressed, highly susceptible, or unhealthy individual. Therefore, we 

selected a 3.0 % concentration as the suitable sublethal exposure for the critical swimming 

experiments. 

Table 2-3. Comparative mortality and behavior of juvenile Florida pompano after 24-h 

exposure to treatment (2.06-4.12 %) and control (0%) levels of ethylene glycol 

and after a 15-h of recovery in clean seawater. Five fish were exposed to each 

concentration.

Ethylene glycol 

Concentration 

Mortality and behavior 

After 24-h exposure 

Mortality and behavior 

 After 15-h recovery

0.00 All normally active All normally active

2.06 All severely lethargic All slightly lethargic

3.09 3 severely lethargic; 

1 severely stressed; 1 dead 

All severely lethargic

4.12 2 severely lethargic; 

 3 severely stressed 

2 severely lethargic; 

 1 severely stressed; 2 dead

Gas chromatographic analyses of initial and final water samples from the exposure tanks showed 

that analyzed concentrations were within 0.3 % of the calculated concentrations. Ethylene glycol 

concentrations in blood plasma collected at 15-h post-exposure from all surviving fish exposed 

to concentrations of 0, 2.1, 3.1 and 4.1 % were 0, 0.37, 0.40 and 0.89 %, respectively, based on 

GC analysis. The linear regression of plasma concentration on exposure concentration showed a 

marginally significant increase in plasma concentration as exposure concentration increased 

(F=16.156, df =2, P<0.0567, r2=0.8898). 

2.2.3 Critical swimming speed 

A 14 % reduction in relative swimming performance was observed in fish exposed to 3.0 % 

ethylene glycol. The mean (± SE) critical swimming speed of treatment fish (N=15) was 

95.9±1.13 cm/s (9.3±0.17 BL/s) in the pre-exposure trial and 83.0±3.45 cm/s (8.0±0.26 BL/s) in 

the post-exposure trial. The mean critical swimming speed of control fish (N=6) was 94.1±3.42 

cm/s (9.2±0.17 BL/s) in the pre-exposure trial, and 98.2±1.28 cm/s (9.6±0.23 BL/s) in the sham 

post-exposure trial. These critical swimming speeds are somewhat conservative because fish 

that did not fatigue after 10 min of swimming at 100 cm/s were assigned a Ucrit value of 100.0 

cm/s. Six of the control fish (three pre- and three post-exposure) and seven of the treatment fish 

(six pre- and one post-exposure) were assigned a Ucrit value of 100 cm/s. 

Least square means from the ANOVA confirmed the negative effects of ethylene glycol on 

swimming performance. Pre- and post-exposure Ucrit differed significantly in the treatment 

group (P<0.0002), but not in the control group (P<0.3492). Nevertheless, a slight increase of 0.4 
14

BL/s in the control group may have been due to training or conditioning effects. No difference 

in pre-exposure Ucrit was detected between the treatment and control groups (P<0.6693), but a 

significant difference in post-exposure Ucrit between the treatment and control groups (P<0.0004) 

was found. In addition, we detected a significant difference between the pre-exposure 

performance of the control group and the post-exposure performance of the treatment group 

(P<0.0058) but not between the pre-exposure performance of the treatment group and the postexposure performance of the control group (P<0.4843). 

Pre- and post-exposure swimming performance varied considerably among individuals (Figure 

2-1). Of the 15 treatment fish, only one individual showed an increase in relative Ucrit of 1.4 

BL/s after exposure (Figure 2-2). For treatment fish, performance declined and variability 

increased after exposure. In the control group, one individual decreased its performance by 0.3 

BL/s, three individuals improved, and two individuals outperformed the maximum test velocity 

in both the pre- and post-exposure trials. 

50

60

70

80

90

100

Pre-exposure Post-exposure

Ucrit (cm/s)

Figure 2-1. Individual variation in the critical swimming speeds (Ucrit) of juvenile Florida 

pompano (N=15) before and after a 24-h exposure to 3.0 % ethylene glycol. 
15

0.0

2.0

4.0

6.0

8.0

10.0

12.0

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21

Treatment Fish Control Fish

Ucrit (BL/s)

Pre

Post

Figure 2-2. Critical swimming speeds (Ucrit; body length/s) of juvenile Florida pompano 

before and after a 24-h exposure to 3.0 % ethylene glycol (treatment fish, 1-15) 

and before and after sham exposure (control fish, 16-21). 

The post-exposure reduction (%R) in absolute Ucrit was not randomly distributed among the 

exposed fish, but was linearly related to body mass (%R = - 53.5 + [1.7 x body mass], R2

 = 

0.3148, df = 13, P<0.0295). Smaller fish experienced a greater percent decrease in swimming 

performance than did larger fish (Figure 2-3). 

Calculated exposure concentrations were verified using GC analysis. Ethylene glycol 

concentrations did not differ between mean initial and final water samples taken from each 24-h 

exposure period (P<0.0897). Gas chromatographic analysis showed that there was a mean (± 

SE) ethylene glycol concentration of 3.0±0.04 % for the treatment exposure solutions. Analysis 

of pooled blood plasma collected approximately 20 h post-exposure (N=5) indicated a mean (± 

SE) ethylene glycol concentration of 0.3±0.02 % for the exposed fish. No ethylene glycol was 

detected in the pooled plasma of the control fish (N=2). 
16

y = - 53.512 + 1.7444x 

R2

 = 0.3154

-50.0

-40.0

-30.0

-20.0

-10.0

0.0

10.0

20.0

30.0

15 20 25 30 35 40

Body mass (g)

% reduction in Ucrit

Figure 2-3. The relationship between body mass of Florida pompano (N=15) and percentage 

reduction in critical swimming speed (Ucrit) after exposure to 3.0 % ethylene 

glycol. 

2.3 Discussion 

A 14 % reduction in the relative swimming performance of juvenile Florida pompano was 

observed in individuals exposed to 3.0 % ethylene glycol for 24 h and allowed to recover for 17 

h before testing. Other sublethal effects included lethargic and disoriented behavior at 

concentrations below 3.0 %. Exposure did not affect the performance of all individuals to the 

same degree. Variation in critical swimming velocity was negatively correlated with body size 

and swimming performance of smaller individuals was significantly reduced compared to that of 

larger individuals. 

2.3.1 Individual variation in Ucrit 

Though mean critical swimming speed of the treatment group decreased by 14 % after exposure, 

considerable individual variation in performance was detected in both the treatment and control 

groups. Swimming performance studies examining sublethal effects of toxicants commonly 

compare mean performances of control and treatment groups, but such an approach fails to 

address within-group variation. Experimental design that addresses within-group variation are 

valuable because swimming performance of individual fish is repeatable over time (Kolok 1992; 
17

Kolok et al. 1998; Reidy et al. 2000). The two-factor ANOVA with pre- and post-exposure 

blocks uses this repeatable source of individual variation to analyze treatment effects. 

Few prior studies have quantified the effect of a stressor by measuring the pre- and postexposure Ucrit of individual fish pre- and post-exposure. In juvenile largemouth bass 

Micropterus salmoides exposed to a 10 oC decline in water temperature (from 20 oC to 10 oC) 

between swim-performance trials, Ucrit was reduced by a consistent amount in all individuals 

(Kolok 1992). Similarly, Kolok and Farrell (1994) found that the Ucrit of northern pikeminnow 

Ptychocheilus oregonensis after a surgical procedure was reduced from pre-surgery values by a 

consistent amount in all fish. The results of a study on fathead minnows (Kolok et al. 1998) 

contrast with those mentioned above. In that study, swimming performance was evaluated 

before and after an exposure to field-collected sediments contaminated with heavy metals and 

hexachlorobenzene. Post-exposure Ucrit varied widely, with some individuals experiencing 

profound reductions in Ucrit and others maintaining their performance level regardless of the 

exposure. Our results resembled those of Kolok et al. (1998), in that the response of individual 

fish varied greatly, from no effect to a severe effect. This differential response of individuals 

was not, however, randomly distributed, but was at least partially correlated with body mass 

(Figure 2-3). 

2.3.2 Size effects 

Size effects among the Florida pompano used in this study were detected during the toxicity and 

behavioral tests through behavioral observations made during and after exposure. Behavioral 

effects of ethylene glycol appeared at lower concentrations in the smaller fish used in the LC50 

tests than in the larger fish used in the behavioral experiments. These results suggest that 

toxicity may decrease with increasing size in Florida pompano, in contrast to results a study on 

rainbow trout in which ethylene glycol was more acutely toxic to larger individuals (Greene and 

Kocan 1997). Size effects often vary from one species to another for a given toxicant (Sprague 

1990). 

It is unclear why ethylene glycol was more toxic to the smaller fish than the larger ones. One 

possibility is that the rate of ethylene glycol uptake is greater in smaller fish due to their surfacearea-volume-ratio. An alternative hypothesis would be that more of the toxic intermediates build 

up in smaller fish. 

2.3.3 Mode of action 

The reduction in swimming performance of Florida pompano in this study may have been due to 

the buildup of ethylene glycol metabolic products. Studies on mammals have shown that 

ethylene glycol is sequentially oxidized in the liver (Chou and Richardson 1978) and involves 

several metabolites, with glycolate (glycolic acid) playing a major role in ethylene glycol 

toxicity. A greater toxicity for metabolic intermediates is also supported by Greene and Kocan 

(1997). In that study, inhibition of the enzyme alcohol dehydrogenase (the enzyme responsible 

for the first step of ethylene glycol metabolism) reduced ethylene glycol toxicity in rainbow 

trout, while inhibition of aldehyde dehydrogenase (the enzyme responsible for the second step in 

ethylene glycol metabolism) increased ethylene glycol toxicity. 
18

Ethylene glycol may alternatively decrease Ucrit by evaluating metabolic rate, and decrease the 

energy available for swimming. Jones (1971) proposed several possible factors, including an 

inability to exchange enough oxygen at the gills, to deliver enough oxygen to the tissues, to 

remove metabolic products, or to activate enzymic processes. A reduction in the ability to 

transfer oxygen across the gills would effectively decrease the scope for active metabolism, 

thereby reducing the fish’s capacity for sustained swimming activity. Our study does not 

indicate whether gill damage directly influenced the reduction in swimming performance of 

exposed individuals. Histological analysis showed a definite effect of 10.0 % ethylene glycol 

exposure, but effects at lower concentrations may have originated from the use of MS-222 before 

sacrifice (J. Hawke, LSU Veterinary School, personal communication). 

2.3.4 Ethylene glycol toxicity 

The estimated 24-h LC50 value (5.63 %) for Florida pompano exposed to ethylene glycol cannot 

be directly compared to 96-h estimates for bluegill sunfish (2.75%; Abdelghani et al. 1990), 

fathead minnow (6.83 to 7.29 %; Pillard 1995; Greene and Kocan 1997), and rainbow trout (6.08 

%; Greene and Kocan 1997; Table 2-4). However, in our study the LC50 was achieved in onefourth the time at comparable concentrations on much larger individuals indicating that Florida 

pompano are relatively more sensitive than the minnow and trout tested previously. The toxicity 

values of different studies can also be compared by using an index of toxicity to estimate 1 h 

LC50 values (Boehm et al. 2001). Thus, assuming a linear-response model index (i.e., 

concentration x time), the LC50 for a 1-hr exposure is four to five times lower for Florida 

pompano, indicating that ethylene glycol was considerably more toxic to the marine species in 

this study than to some freshwater fishes that have been tested. Tolerance to many toxicants is 

believed to vary less among fishes (often five- or tenfold) than among other taxonomic groups, 

such as invertebrates (Sprague 1990). 

Table 2-4. Estimated LC50 values (mean +/- SE) for several fish species exposed to ethylene 

glycol. 

Species Mean body 

mass (g) 

Test duration 

(h) 

LC50 %) Reference

Bluegill sunfish 

 (Lepomis macrochirus) 

0.9±0.60 96 h 2.75 Abdelghani 

 et al. 1990

Fathead minnow 

(Pimephales promelas) 

(=7 days old) 96 h 7.29 Pillard 1995

Fathead minnow 0.1±0.01 

0.1±0.01 

96 h 

96 h 

6.83 

6.95 

Greene and 

 Kocan 1997

Rainbow trout 

(Oncorhynchus mykiss) 

0.7±0.20 

1.5±0.50 

96 h 

96 h 

6.08 

5.65 

Greene and 

 Kocan 1997

Florida pompano 

(Trachinotus carolinus)

8.0±3.20 24 h 5.63 Kay et al. 2000 

(present study)
19

2.3.5 Marine exposures 

The high LC50 values of ethylene glycol reported in this and other studies on fishes generally 

reflect the relatively low toxicity of this chemical. Marine fishes would most likely be exposed 

to lethal concentrations only in the event of highly localized, large-scale releases into the 

environment. Marine releases of considerable magnitude have been reported (Boehm et al. 

2001). For example, 13,523 L of ethylene glycol leaked into the Gulf of Mexico from an 

underwater pipeline in June of 1998 (National Response Center 2002). Exposures of marine 

fishes are likely brief because of the open nature of marine systems and the relatively rapid 

biodegradation of ethylene glycol (half life ~ 32 d, see Boehm et al. [2001]). Biooxidation of 

ethylene glycol in the laboratory was 39 % in 5 d and 96 % in 20 d (Conway et al. 1983); 

dispersion and dilution of ethylene glycol in the marine environment should be even more rapid. 

Other factors affecting the toxicity of ethylene glycol must also be considered, such as additive 

or synergistic effects with other chemicals. The toxicity of ethylene glycol increases in 

freshwater fishes when sublethal levels of thiram, a component of an agricultural seed-protectant, 

are present (Greene and Kocan 1997), and when it is combined with other chemicals in a deicing 

agent (Pillard 1995). Boehm et al. (2001) simulated an ethylene glycol spill of 402,902 L 

(106,447 gal) from a deepwater supply line near the shallow end (water depth, 113 m) and found 

that the predicted exposure concentrations were an order of magnitude lower than the lowest 

published 48-h LC50 values (see Pillard [1995]; 34,400 mg/L for a freshwater crustacean, 

Ceriodaphnia dubia). However, we found a relatively lower LC50 for Florida pompanoas well 

as substantial and significant sublethal effects on swimming performance below the LC1 level 

(Table 2-2). 

2.3.6 Threshold of tolerance 

The variation between individuals in the present study suggests that some fish may have a higher 

threshold of tolerance to ethylene glycol toxicity. Many studies have estimated threshold values 

for fish exposed to contaminants by testing swimming speed at a range of exposure 

concentrations (Peterson 1974; Nikl and Farrell 1993). Because this study tested fish at only one 

ethylene glycol concentration, we were unable to calculate a mean threshold value for this 

species. Nevertheless, we can estimate, based on significant reduction in swimming speed, that 

the threshold lies at or below the 3.0 % ethylene glycol concentration (53 % of the estimated 

LC50 for slightly smaller fish). 

Though adverse effects on critical swimming speed were only quantified at an exposure 

concentration of 3.0 %, Florida pompano exposed to slightly lower concentrations (1.0-2.1 %) 

exhibited changes in behavioral patterns. Changes in behavior patterns may have ecological 

implications for individuals, and potentially for populations subject to chronic exposure. 

Sublethal behavioral effects were not observed at ethylene glycol concentrations below 1.0 % in 

the LC50 experiments, or below 2.1 % in the behavior and recovery experiments (which used 

somewhat larger fish). In a preliminary swimming trial conducted prior to this study, a 1.5 % 

ethylene glycol exposure did not produce an obvious reduction in the critical swimming speeds 

of two out of three individuals. Future studies may want to examine the effects of ethylene 

glycol concentrations between 1.5 % and 3.0 % on critical swimming speed to determine a 

threshold tolerance value this species. 
20

2.3.7 Ecological consequences of sublethal effects 

Most studies of petroleum-related contaminants have concentrated on the mortality effects 

associated with major oil spills (Howarth 1991; Kennicutt et al. 1996; Peterson et al. 1996), but 

much remains unknown about the effects of low-level exposure on marine organisms. Sublethal 

effects of contaminant exposure are inherently more difficult to assess and extrapolate to the 

field. We have shown that individual marine fish can experience reduced swimming 

performance when exposed to sublethal levels of ethylene glycol. The implications of these 

results include adverse effects on the ability to perform in ecological situations. While normal 

cruising speeds of Florida pompano are not near the high critical swimming speeds observed in 

this study, the capacity for burst swimming and endurance were undoubtedly reduced by 

ethylene glycol exposure, with direct bearing on the efficiency of prey capture and predator 

avoidance. 

An understanding of sublethal effects is important for the accurate prediction of the impacts to 

habitat and fish populations because many more organisms are likely to be exposed to sublethal 

concentrations of contaminants than to lethal concentrations as a spill disperses and because 

many may experience chronic low levels of one or more contaminants over time. For example, 

in a study of fishes surrounding oil production platforms, Grizzle (1986) found histological 

changes in the gills that possibly reflected chronic, low-level exposure to petroleum or petroleum 

additives. 

Extrapolation of sublethal effects from the individual level to the population level remains 

problematic (Rose et al. 1993), particularly when exposure causing sublethal effects for one sizeclass are much more toxic (perhaps fatal) to smaller individuals. Mortality can be highly 

compensatory in fish populations, especially in early life history stages. High levels of mortality 

from an episodic event, such as a contaminant spill, may not necessarily have a significant effect 

at the population level, if significant compensation in the mortality process occurs (Houde 1989). 

For example, individual-based models of contaminant exposures demonstrate that chronic 

exposures are more likely than episodic exposures to affect recruitment of striped bass Morone 

saxatilis (Rose et al. 1993). 

Nevertheless, episodic exposures could have a greater impact on recruitment if they occur at or 

shortly after a spawning peak or in a discreet nursery area, such as an estuary. Sublethal impacts, 

such as those documented for swimming performance of Florida pompano, are even more 

difficult to extrapolate to the population level without sophisticated models and details of 

exposure conditions (time, contaminant concentration, life stage, and ecological interactions). 

Currently, deepwater oil production uses ethylene glycol and other pure chemicals and 

proprietary chemical mixtures as additives to facilitate production. These chemicals and 

mixtures will also need to be evaluated individually and in concert to achieve an accurate 

assessment of the potential sublethal and lethal effects of contaminant exposures to marine fish 

populations. We found a substantial and significant sublethal effect on individual swimming 

performance from exposure to ethylene glycol concentrations below the LC1 level that is cause 

for concern. 
21

3.0 TOXICITY AND SUBLETHAL EFFECTS OF METHANOL ON SWIMMING 

PERFORMANCE OF JUVENILE FLORIDA POMPANO 

D. M. Baltz1,2, E. J. Chesney3, M. A. Tarr4, A. S. Kolok5 and M. J. Bradley3

1

Department of Oceanography and Coastal Sciences, Louisiana State University, Baton Rouge, 

LA 70803-7503 

2

Coastal Fisheries Institute, Louisiana State University, Baton Rouge, LA 70803-7503 

3

Louisiana Universities Marine Consortium, Chauvin, LA 70344 

4

Department of Chemistry, University of New Orleans, New Orleans, LA 70148 

5

Biology Department, University of Nebraska at Omaha, Omaha, NE 68182 

Marine fishes face an ever-changing array of environmental contaminants that may reduce their 

ability to perform in ecological situations. In the northern Gulf of Mexico, the development of 

new technologies has allowed expansion of oil and gas production into deep water (> 305 m) and 

has created the need to evaluate fishery-related risks associated with new practices and chemicals 

used in petroleum production and transport (Minerals Management Service, 

www.gomr.mms.gov; Boehm et al. 2001). Marine fishes are potentially at risk from accidental 

releases of chemical additives used in deepwater wells to prevent the formation of gas hydrates 

(Anonymous 1996, 2000; Herzhaft and Dalmazzone 2000; Boehm et al. 2001), which can 

potentially clog pipelines. A variety of chemicals, such as ethylene glycol (Hymel et al. 2002) 

and methanol, may be transported in small supply lines up to 96.5 km in length, stored in large 

volumes and used at high rates. Though not responsible for producing direct mortality because 

of relatively low concentrations or toxicities, sublethal impacts of spills or leaks may affect 

scope for activity (Cech 1990), swimming performance (Brett 1964, Wedemeyer et al. 1990), 

growth, reproduction, and survival of individuals (Little et al. 1990). Quantification of sublethal 

effects at the individual level is an important step in predicting how individuals and populations 

may respond ecologically to single and multiple contaminants. 

The environmental effects of methanol have been widely studied in relation to its use as an 

industrial organic solvent and reactant in organic procedures (Tephly 1991). Median lethal 

concentrations (LC50) have been reported for various marine algae and invertebrates (Boehm et 

al. 2001). The toxicity of methanol to fishes has been assessed on relatively few species, 

primarily freshwater (Gillette et al. 1952, Nishiuchi and Hashimoto 1967, Buzzell et al. 1968, 

Juhnke and Luedemann 1978, Linden et al. 1979, Johnson and Finley 1980, Call et al. 1983, 

Veith et al. 1983, Bengtsson et al. 1984, Ewell et al. 1986, Mayer and Ellersieck 1986, Poirier et 

al. 1986, Tsuji et al. 1986, Robertson et al. 1988) and two marine or salt-tolerant species, 
22

hooknose Agonus cataphractus (Portmann and Wilson 1971) and bleak Alburnus alburnus

(Tarkpea and Svanberg 1982). 

Chemical spills in aquatic systems can lead to environmental effects that range from direct 

mortality from lethal exposure levels to sublethal effects as contaminants spread and are diluted. 

Important effects of sublethal exposure on individuals cannot be predicted from toxicity tests 

alone; however, sublethal effects may be evaluated by measures of individual performance 

(Sprague 1971) that relate to Darwinian fitness (Plaut 2001). For example, diminished 

swimming performance can lead to increased risks of mortality or an impaired ability to perform 

normal life functions (Schreck 1990) and subsequently add to the risk of a decline in a 

population (Heath 1995). Changes in critical swimming speed (Ucrit) have been used as a 

performance indicator in many studies assessing sublethal effects of contaminants on fishes 

(Beaumont et al. 1995; Heath 1995; Kolok et al. 1998; Nikl and Farrell 1993). Swimming 

performance has ecological significance and has been shown to affect the ability of larval fishes 

to avoid predators and catch prey (Zhou et al. 1996; Samson et al. 2001), but the same has not 

been demonstrated for larger individuals (Plaut 2001). Vulnerability to predation is dynamic and 

can be affected by exposure to contaminants, resulting in effects on mobility (Beauvais et al. 

2000), perception (Baker and Montgomery 2001), and behavior (Weis et al. 1999, 2001; Scholz 

et al. 2000; Wibe et al. 2001). 

The Florida pompano Trachinotus carolinus was chosen for assessment of acute and sublethal 

effects of exposure to methanol because the species occurs throughout the coastal zone in the 

vicinity of petroleum production structures and transport pipelines. Juveniles are abundant in 

and near the surf zone during summer (Bellinger and Avault 1970). Adults can be found in 

inshore and offshore waters and are believed to spawn on the continental shelf (Gilbert 1986). It 

is also representative of the jack family (Carangidae) with many species that reside near 

deepwater structures in the Gulf of Mexico and are of considerable economic importance to 

recreational and commercial fisheries (Gilbert 1986). 

We assessed the acute and sublethal effects of methanol on juvenile Florida pompano by 

identifying a short-term LC50 and a range of lower sublethal concentrations and by testing for 

changes in Ucrit following sublethal exposure. The objectives were to estimate the static 24-h 

LC50, use it to identify sublethal exposure levels from which individuals could recover after 

short-term exposures expected in open marine systems (Boehm et al. 2001), and then to evaluate 

within-individual changes in swimming performance before and after sublethal exposure. 

3.1 Methods 

Collection and maintenance of fish.--Juvenile Florida pompano were collected for experiments in 

August 1999 from Port Fourchon Beach and Isle Dernieres, Louisiana, with a 3.05 m (10 ft) 

beach seine. Individual body mass at the time of capture ranged from 2.0 to 16.0 g. Fish were 

held for at least 4 weeks in a 5,000-L recirculating seawater system at 30 practical salinity units 

(psu) before testing to allow them to acclimate to captivity and recover from the stress of 

capture. During their initial quarantine, the fish were briefly treated with 0.2 mg/L copper 

(Cutrine-Plus) to eliminate ectoparasites, especially Amyloodinium, a parasitic dinoflagellate 

known to cause mass mortality in captive fishes (Cheung 1993). The recirculating holding 

system was constantly aerated and water quality parameters, including pH, salinity, and 

ammonia level, were monitored several times a week and daily during experiments, and adjusted 
23

as necessary to maintain suitable conditions. Amquel (HOCH2SO3Na) was added to holding 

systems to reduce the ammonia level when it exceeded 0.5 mg/L. Fish were fed a commercially 

produced feed (0.25 g/day per individual; 45% protein, 8% lipid) except within the 17-h prior to 

and during acute toxicity and swimming experiments. 

Acute toxicity and recovery tests.-- The results from a preliminary toxicity experiment were used 

to select a range of concentrations, 0.99 to 1.51% methanol (volume per volume, % v/v), that 

was logarithmically divided into the six concentrations, 0.99, 1.07, 1.17, 1.27, 1.38, and 1.51%. 

A 24-h static LC50 experiment was conducted in a system of eight conical fiberglass tanks. 

Each tank was filled to 70 L with ultra-violet (UV) sterilized, 30 psu seawater. The tanks were 

individually aerated, maintained at 25.0±1°C with aquarium heaters, and covered with clear 

acrylic lids to reduce volatilization of methanol. The exposure concentrations were randomly 

assigned to the tank array. A ‘blank’ tank without fish was also included to monitor the mixing 

and concentration of the chemical over the 24-h period. Ten fish were randomly assigned to 

each of the exposure and control tanks and they were introduced after the addition of methanol 

(Fisher Scientific, Certified ACS). The mean (± SE) body mass and fork length (FL) were 33.2 ± 

2.4 g and 110 ± 3.1 mm FL. We maintained a 20-min interval between sequential introductions 

to allow for processing of fish at the end of the experiment. The behavior of the fish in the 

treatment groups, relative to the controls, was observed and classified following Hymel et al. 

(2002). Water samples (1.5 mL) for gas chromatography (GC) analysis, taken periodically from 

each tank during the tests, showed that the methanol exposure solutions were accurately 

estimated, well mixed, and stable throughout the 24-h experiment. Data from the experiment 

was used to estimate LC values using the EPA Probit Analysis Program, Version 1.5 (EPA 

1994). 

We collected blood samples from surviving fish to analyze the effects of methanol exposure. At 

the end of the experiment, individuals were anesthetized with tricaine methanesulfonate (MS222; <50 mg/L), blood was drawn from the caudal vessels, and collected into heparinized 

capillary tubes for GC analyses. After a 10-min centrifugation at 12,800 ± 1,200 revolutions per 

minute (RPM), plasma samples from individuals of each treatment and control group were 

pooled and frozen until GC analyses were performed. Prior to all statistical analyses of LC50 

data and blood samples, estimated methanol concentrations were measured by GC analysis. 

Following the LC50 tests, a behavior-recovery experiment was conducted prior to the swimming 

experiment to identify a sublethal concentration from which juveniles could apparently recover 

over 15 h in clean seawater (Hymel et al. 2002). 

 

Critical swimming speed.-- Three Blazka-type swimming respirometers were used to 

simultaneously test swimming performances of 3 fish (Hymel et al. 2002). Critical swimming 

speed before and after sublethal exposure to methanol was measured and compared for 

individual fish (Kolok et al. 1998). To keep track of individuals, 24 juveniles (mean ± SE; 20.5 

± 0.9 g body mass; 100.8 ± 1.6 mm FL) were marked two weeks before the start of the 

experiment. Fish were anesthetized (MS-222 < 50 mg/L) and a unique combination of colored 

latex marks (Northwest Marine Technology, Inc.) were applied by injection to their dorsal, 

caudal, and anal fins. After injection, the fish were held in the 5,000-L recirculating seawater 

system. 

Several days prior to the experiment, the marked fish were randomly separated into groups of 

three, which were tested simultaneously in the swimming chambers. Two of the eight groups 
24

were randomly assigned to the control, and the remaining six groups were assigned to the 

treatment. Each group was randomly assigned to one of eight 60-L, interconnected holding tanks 

in a recirculating system with a total volume of 550 L. The holding system was equipped with a 

biofilter and UV sterilizer to maintain water quality. Salinity in the holding system was 

maintained at 30 ± 1 psu, and temperature was maintained at 25 ± 1°C. Holding tanks were 

individually aerated with air stones. 

The protocol for the swimming experiment required 27 days to ensure identical fasting, holding, 

and acclimation periods for each group of fish, and all pre- and post-evolutions occurred at an 

interval of seven days. The protocol allowed for a 17-h acclimation to the swimming chambers 

following the 24-h exposure and included a 41-h fasting period (i.e., 24-h exposure and 17-h 

acclimation) before pre- and post-exposure swimming trials. All exposures and sham exposures 

were carried out in one of three round fiberglass tanks, filled to 60 L with 30 psu seawater. The 

exposure tanks were individually aerated, and the water temperature was maintained at 25 ± 1°C. 

All fish were sham exposed before the first swimming trial, but only the control groups were 

sham exposed before the second swimming trial. Treatment groups were exposed to a 1.07% 

methanol solution before the second swimming trial. To confirm concentrations, we collected 

water samples for analyses from the exposure tank at the beginning and end of each exposure 

period. All exposed and sham-exposed fish were placed in a bucket of clean seawater for several 

minutes before being placed in the swimming chambers. 

During the 17-h acclimation period, the three swimming chambers were connected to a larger 

recirculating system that included a UV sterilizer and a biofilter. The total recirculating volume 

of the system was about 765 L. The entire swimming system was housed in a 25°C temperaturecontrolled room. After exposure, each fish was placed in one of the swimming chambers for the 

acclimation interval and the water circulated at a velocity of 5 cm/s. Before the start of a 

swimming trial, each swimming chamber was cleared of bubbles and sealed. For the preexposure swimming trials individuals were randomly assigned to a swim chamber, but were retested in the same swimming chamber for the post-exposure trials. 

The experimental protocol tested for changes in the swimming performance of individuals 

exposed to methanol by determining pre- and post-exposure Ucrit of each individual. At the start 

of each trial, water velocity was set at 10 cm/s. During the tests, the velocity of each swimming 

chamber was increased by 10 cm/s every 10 min, with no rest interval between steps. Each 

swimming test was terminated when the fish could no longer continue swimming and made 

contact with the back grid for about 2-4 s. A pre- and post-exposure Ucrit value was calculated 

for each individual following Brett (1964): 

Ucrit = ui + (ti/tii x uii), 

where ui is the highest velocity maintained for the full prescribed period (10 min), uii is the 

highest velocity increment, ti is the time the fish swam in the fatigue velocity interval, and tii is 

the prescribed period of swimming. A fish that failed to fatigue after 10 min of swimming at the 

swimming chamber’s maximum velocity (100 cm/s) was assigned a Ucrit value of 100.0 cm/s. 

After each post-exposure swimming trial, the fish were anesthetized with MS-222 (< 50 mg/L), 

and then weighed (g) and measured (FL mm). 
25

Statistical analysis. --The Ucrit values were expressed in absolute (cm/s) and relative (body 

lengths per second, BL/s) terms. Statistical analyses were performed on the relative values of 

Ucrit to standardize for small variations in individual size. We used the Mixed Procedure in SAS 

(Littell et al. 1996) to evaluate relative Ucrit (BL/s) with Group (i.e., treatment and control) and 

Trial (i.e., pre- and post-exposure performances of an individual) and their interaction 

(Group*Trial) as fixed factors and individual fish within groups as the randomized factor. 

Posterior testing was accomplished by comparing differences in least square means among six 

Group*Trial comparisons. Residuals were tested for normality and no transformations were 

required. The analysis was performed using SAS computer software (SAS Institute, Version 

6.12 1996) to test the model, Yijk = μ + t1i + t2j + $ k + (t1t2)ij + ,ijk, where t1 is the ‘Trial’ 

variable (pre or post), t2 is the ‘Group’ variable (treatment or control), t1t2 is the interaction, and 

$ is the random blocking variable (individual fish). We evaluated individual variation in the 

performance of treatment fish by testing for the influence of size on absolute Ucrit. We also used 

a linear regression analysis (SAS Institute 1996) to test for a relationship between body mass and 

percentage reduction in absolute Ucrit of treatment fish. 

3.2 Results 

3.2.1 Acute toxicity and recovery tests 

The 24-h LC50 value was identified for juvenile Florida pompano over a series of seven test 

concentrations determined by GC analysis: 0.00, 0.99, 1.13, 1.22, 1.36, 1.46, and 1.49% (% v/v) 

methanol. Probit analysis of data from the experiment yielded a 24-h LC50 estimate of 1.28% 

(Table 3-1). Mortality of all 10 fish was observed at a concentration of 1.49%. The lowest 

experimental concentration at which mortalities were observed was 1.13%. The simple linear 

regression (Figure 3-1) of pooled methanol plasma concentrations (% v/v) on exposure 

concentrations from five treatment levels with surviving individuals and the control was positive 

and significant ([in plasma] = -0.02 + 0.41[in water]; R2

 = 0.67; N = 6, P < 0.045). 

Table 3-1. Calculated lethal concentrations (LC, % v/v) values and associated 95% 

confidence limits for juvenile Florida pompano exposed to methanol for 24 h. 

 Exposure Concentration 95% Confidence limits

LC level Lower Upper 

 

LC1 0.958 0.757 1.058 

LC5 1.044 0.877 1.128 

LC10 1.093 0.947 1.168 

LC15 1.127 0.997 1.197 

LC50 1.284 1.213 1.353 

LC85 1.464 1.384 1.632 

LC90 1.509 1.419 1.716 

LC95 1.580 1.470 1.852 

LC99 1.722 1.568 2.143 
26

Noticeable differences in behavior were observed in fish exposed to control and treatment 

concentrations used in the behavior-recovery experiment (i.e., 0.0, 0.98, 1.07 and 1.17% as 

verified by GC analysis). Most treatment fish recovered to a normally active state after a 15-h 

recovery period in clean seawater, but fish in the 1.17% concentration showed residual effects 

(i.e., minor discoloration). Therefore, we selected 1.07% as the suitable sublethal exposure for 

the critical swimming experiment. 

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

0 0.2 0.4 0.6 0.8 1 1.2 1.4 1.6

Methanol % in water

Methanol % in plasma

Y = -0.0554 + 0.4645X, R2=0.7819, P ≤ 0.02

Figure 3-1. The regression of methanol concentration in blood plasma, control, and exposure 

waters following exposure to concentrations ranging from 0.99 to1.49% v/v. 

Points are estimates of pooled samples from survivors in control and treatment 

groups. 

3.2.2 Critical swimming speed 

The mixed-model analysis of relative Ucrit (i.e., BL/s) found a significant interaction between 

pre- and post-exposure trials of control and treatment groups. Two treatment fish were excluded 

from the analysis: one performed poorly in pre- and post-exposure trials achieving only 1.68 and 

1.74 BL/s, and another swam normally in the pre-exposure trial (9.17 BL/s) but later died during 

exposure. Mean differences for the fixed effects Group (F = 0.60, df = 1 & 20, P > 0.448) and 

Trial (F = 1.03, df = 1 & 20, P > 0.3223) were not significant, but the interaction between Group 

and Trial was significant (F = 15.32, df = 1 & 20, P < 0.0009). The least square means for the 

pre- and post-control trials increased from 8.76 ± 0.26 to 9.11 ± 0.26 BL/s. The least square 

means for the pre- and post-treatment trials decreased from 9.01 ± 0.16 to 8.42 ± 0.16 BL/s. 

Only two comparisons of least square means among the six pair-wise were significant. The least 

square means for the control post-trial (9.11 BL/s) was significantly (P < 0.0381) higher than the 

treatment post-trial (8.42 BL/s). The least square means for the treatment post-trial (8.42 BL/s) 

was significantly (P < 0.0001) lower than the treatment pre-trial (9.01 BL/s) confirming the 

negative effects of methanol on swimming performance. Although there was no detectable 
27

difference between pre- and post-exposure Ucrit in the control group (P < 0.1047), a slight 

increase of 0.34 BL/s in the post-exposure trial may have been due to training or conditioning 

effects. 

A 6.5% reduction in relative swimming performance was observed in treatment fish after 

exposure to 1.07% methanol. This value is a conservative underestimate because one treatment 

fish that did not fatigue after 10 min of swimming at 100 cm/s was assigned a pretreatment Ucrit

value of 100.0 cm/s. The mean (± SE) absolute Ucrit of treatment fish (N = 16) was 90.1 ± 1.35 

cm/s in the pre-exposure trial and 84.2 ± 1.36 cm/s in the post-exposure trial. The mean absolute 

Ucrit of control fish (N = 6) was 87.7 ± 2.54 cm/s in the pre-exposure trial and 91.2 ± 2.75 cm/s in 

the sham post-exposure trial. 

Pre- and post-exposure swimming performance varied considerably among individuals (Figure 

3-2). Of the16 treatment fish, only two individuals showed a mean increase in relative ucrit of 

0.18 and 0.35 bl/s following exposure (Figure 3-2). For treatment fish, mean performance 

declined between pre- and post-exposure trials. In the control group, four individuals increased 

their performance by 0.27 to 0.72 bl/s, and two individuals decreased in performance by 0.03 and 

0.10 bl/s. Although the post-exposure reduction (%r) in absolute ucrit did not impact all 

individuals to the same degree, variation among treatment fish was unrelated to individual body 

mass (g) (%r = 6.9 – [0.02 x body mass], n = 16, r = -0.02, p > 0.05) and three size groups 

showed similar declines (Figure 3-3). 

6.00

7.00

8.00

9.00

10.00

11.00

6.00 7.00 8.00 9.00 10.00 11.00

Pre-Exposure BL/s

Post-Exposure BL/s

Treatment Control Linear (Treatment) Linear (Control)

Figure 3-2. Critical swimming speeds, in Body Lengths per second (BL/s), of juvenile Florida 

pompano before and after a 24-h exposure to 1.07% methanol (treatment fish, N = 

16) and before and after sham exposure (control fish, N = 6). 
28

7.50

8.50

9.50

10.50

Pre-Exposure Post-Exposure

Ucrit (BL s-1)

Large T Medium T Small T

Large C Medium C Small C

Figure 3-3. Size-related changes in mean critical swimming speeds (BL/s) of juvenile Florida 

pompano in treatment (T) and control (C) groups before and after exposure and 

sham exposure to 1.07% methanol for large (T: 106-119 mm, N = 5; C: 103-105 

mm, N= 2), medium (T: 100-105 mm, N = 5; C: 100-102 mm, N= 2), and small 

(T: 85-99 mm, N = 6; C: 92-99 mm, N= 2) individuals.

3.3 Discussion 

3.3.1 Methanol toxicity and mode of action 

The estimated 24-h static LC50 value (1.28%) for juvenile Florida pompano exposed to 

methanol cannot be directly compared to many previous studies (Gillette et al. 1952, Nishiuchi 

and Hashimoto 1967, Buzzell et al. 1968, Juhnke and Luedemann 1978, Linden et al. 1979, 

Johnson and Finley 1980, Call et al. 1983, Veith et al. 1983, Bengtsson et al. 1984, Ewell et al. 

1986, Mayer and Ellersieck 1986, Poirier et al. 1986, Tsuji et al. 1986, Robertson et al. 1988) 

because of different exposure types and test durations, and neither of the studies in sea water 

used similar protocols (Portmann and Wilson 1971, Tarkpea and Svanberg 1982). Three of four 

studies reporting values for 24-h LC50 tests on bluegill (2.41%), fathead minnow (3.75%), 

medaka Oryzias latipes (> 1.26%), and rainbow trout (2.57%) are considerably higher than the 

1.28% LC50 for Florida pompano found in this study. Although osmoregulation in fish has 

been well studied, the interaction between salinity and toxicity of xenobiotics is relatively 

unexplored (Ferguson and Hogstrand 1998, El-Alfy et al. 2001). To counteract water gain in 

fresh water and water loss in sea water, freshwater fishes do not drink significantly while marine 

fishes do (Ferguson and Hogstrand 1998, Moyle and Cech 2000), enhancing the intestinal 

pathway for uptake of waterborne contaminants. This may account, in part, for the higher 
29

toxicity of methanol in marine fishes. Moreover, Florida pompano also appear to be more 

sensitive than two other species tested in saltwater, bleak (a euryhaline freshwater species, 

3.54%, 96-hr static, tested at 7 ppt) and hooknose (a marine species, 1.26 to 4.17%, 48- and 96-

hr renewal). In our study the LC50 was achieved in one-half to one-fourth the time at 

comparable or lower concentrations indicating that Florida pompano are relatively more 

sensitive than bleak or hooknose. Toxicity estimates from different studies can be roughly 

compared with an index of toxicity (IT) to estimate 1-hr LC50 values (Boehm et al. 2001). Thus, 

assuming a linear-response model index (i.e., concentration x time), the LC50 for a 1-hr 

exposure is two to twelve times lower for Florida pompano (IT = 30.72), indicating that 

methanol was considerably more toxic to pompano than to the euryhaline bleak (IT = 340), the 

marine hooknose (IT = 60 to 400) and most of the freshwater fishes that have been tested (IT = 

48 to 357). Only the study on medaka (Tsuji et al. 1986) yielded results similar to those for 

pompano. 

The literature is conflicting on the mode of action directly responsible for the toxic effects of 

methanol. In vivo methanol is quickly metabolized into formaldehyde, which has been 

suggested as the metabolite that produces the toxic effect (Potts and Johnson 1952, Koivusalo 

1979). A more recent study indicates that a second metabolic product, formate, is responsible for 

the toxicity associated with methanol exposure (Tephly 1991). Formate causes the depletion of 

plasma bicarbonate and the development of metabolic acidosis (Tephly 1991). The buildup of 

formate and subsequent changes in metabolic equilibrium may reduce a fish’s ability for 

prolonged or burst swimming; however, some species are able to oxidate formate into carbon 

dioxide at a faster rate, and thus have a higher tolerance to methanol exposure (Tephly 1991). 

3.3.2 Sublethal effects on Ucrit

Individuals exposed to a methanol concentration between the LC05 and LC10 levels exhibited 

sublethal effects on Ucrit after being allowed to recover in clean seawater for 17 h. A 6.5% 

reduction (9.01 to 8.42 BL/s) in the relative swimming performance of exposed juveniles 

contrasts sharply with the 4.0% improved performance of the control group. Thus, the 

conditioning and/or training effects were surpassed by the significant negative sublethal effects 

of the methanol exposure. 

3.3.3 Individual variation in Ucrit

Though mean critical swimming speed of the treatment group decreased by 6.5% after exposure, 

considerable individual variation in performance was detected in both the treatment and control 

groups. Swimming performance studies examining sublethal effects of toxicants commonly 

evaluate performances by comparing the means of separate control and treatment groups, but 

such an approach can overlook within-group variation. Experimental designs that address 

within-group variation are valuable because swimming performance of individual fish is 

repeatable over time (Kolok 1992; Kolok et al. 1998; Reidy et al. 2000). Our two-factor 

ANOVA with pre- and post-exposure measurements used this repeatable source of individual 

variation for analysis of treatment effects that were controlled for training and conditioning. The 

design allows both the comparison of pre-treatment versus post-treatment performance in the 
30

same animal and the measurement of sublethal impacts of a contaminant (i.e., methanol), 

features not commonly seen in toxicological studies. 

Few prior studies have quantified the effect of a stressor by measuring the pre- and post exposure 

Ucrit of individual fish. In juvenile largemouth bass Micropterus salmoides exposed to a 10oC 

decline in water temperature (from 20oC to 10oC) between swim-performance trials, Ucrit was 

reduced by a consistent amount in all individuals (Kolok 1992). Similarly, Kolok and Farrell 

(1994) found that the Ucrit of northern pikeminnow Ptychocheilus oregonensis after a surgical 

procedure was reduced from pre-surgery values by a consistent amount in all fish. The results of 

a study on fathead minnows (Kolok et al. 1998) contrast with those mentioned above. In that 

study, swimming performance was evaluated before and after an exposure to field-collected 

sediments contaminated with heavy metals and hexachlorobenzene. Post-exposure Ucrit varied 

widely, with some individuals experiencing profound reductions in Ucrit and others maintaining 

their performance level regardless of the exposure. Our results resembled those of Kolok et al. 

(1998) and Hymel et al. (2002), in that the response of individual fish following exposure to 

methanol (and a recovery period) varied greatly, from no effect on performance to a severe 

effect. However, in contrast with Hymel et al.’s (2002) finding of a size-dependent effect of 

ethylene glycol on juvenile Florida pompano, we were unable to detect a size effect for 

methanol. 

3.3.4 Marine exposures 

The LC50 values of methanol reported in this and other studies on fishes generally reflect the 

relatively low toxicity of this chemical. Marine fishes are most likely to be exposed to lethal 

concentrations only in the event of highly localized, large-scale releases into the environment. 

Large offshore storage volumes of methanol in the northern Gulf of Mexico (up to 380,800 L) 

pose significant risk of a spill (Boehm et al. 2001). Marine releases of considerable magnitude 

have been reported (Boehm et al. 2001). Nine methanol spills were reported from platforms in 

the northern Gulf of Mexico between March 1994 and January 2002 (National Response Center 

2002) and other types of incidents were more common. For example, 39,742 L were released in 

June 1997 (National Response Center 2002a) and 11,355 L of methanol leaked into the Gulf of 

Mexico from an underwater injection line in November of 1997 (National Response Center 

2002b). Exposures of marine fishes are most likely to be brief because of rapid dispersion and 

mixing in an open marine system and relatively rapid biodegradation of methanol (half life ~6.2 

d, see Boehm et al. 2001). While dispersion and dilution of methanol in the marine environment 

would be expected to be rapid, other factors affecting the toxicity of methanol must also be 

considered, such as additive or synergistic effects with other chemicals. Boehm et al. (2001) 

simulated a methanol spill of 363,360 L into the surface waters over the continental shelf and 

found that the predicted exposure concentrations were two orders of magnitude lower than the 

lowest published 96-h LC50 values (see Tarkpea and Svanberg [1982]; 12,539 mg/L) for a 

benthic marine harpacticoid copepod Nitroca spinipes. However, at a comparable concentration, 

we found a short-term LC50 (24-h, ~12,800 mg/L) for Florida pompano and substantial and 

significant sublethal effects on swimming performance between the LC05 and LC10 levels 

(Table 3-1). 
31

3.3.5 Ecological consequences of sublethal effects 

Most studies of petroleum-related contaminants have concentrated on the mortality effects 

associated with major oil spills (Howarth 1991; Kennicutt et al. 1996; Peterson et al. 1996), but 

much remains unknown about the effects of low-level exposures on marine organisms. Sublethal 

effects of contaminant exposure are inherently more difficult to assess and extrapolate to the 

field. We have shown that individual marine fish experience reduced Ucrit (mostly aerobic) 

swimming performance when exposed to sublethal levels of methanol or ethylene glycol (Hymel 

et al. 2002). The implications of these results include adverse effects on the ability to perform 

aerobically or anaerobically in ecological situations (Plaut 2001). Ucrit was developed primarily 

as a measure of aerobic capacity, even though the fish is pushed to use anaerobic metabolism and 

fatigue. While maximum (sensu Sepulveda and Dickson 2000) cruising speeds of Florida 

pompano are probably not near the high critical swimming speeds observed in this study, the 

capacity for sprint (s, anaerobic), burst (min, anaerobic), and critical (hr, aerobic) swimming 

modes (sensu Reidy et al. 2000) are all probably affected to some degree by methanol exposure. 

Each mode has direct relevance to foraging, prey capture, predator avoidance, trawl avoidance, 

strong currents, migrations, or spawning. However, all three modes of swimming have been 

tested in only one species, Atlantic cod Gadus morhua. Reidy et al. (2000) found that for 

healthy cod, Ucrit and sprint performances are positively correlated, but Ucrit and burst 

performances are negatively correlated, and sprint and burst performances are uncorrelated. 

Morphological and physiological tradeoffs among aerobic and anaerobic swimming types in 

other species and how contaminant exposure affects their sprint, burst, and endurance swimming 

performances remains unstudied (Plaut 2001, Nelson et al. 2002). 

An understanding of sublethal effects is important for the accurate prediction of the impacts to 

habitat and fish populations because many species may experience chronic low levels of one or 

more contaminants over time. For example, in a study of fishes surrounding oil drilling 

platforms Grizzle (1986) found histological changes in the gills that possibly reflected chronic, 

low-level exposure to drilling contaminants. 

Extrapolation of sublethal effects from the individual level to the population level remains 

problematic (Rose et al. 1993), particularly when sublethal effects vary across size classes. 

Mortality can be highly compensatory in fish populations, especially when it falls on early life 

history stages. High levels of mortality from an episodic event, such as a contaminant spill, may 

not necessarily have a significant effect at the population level, if there is significant 

compensation in the mortality process (Houde 1989). For example, individual-based models of 

striped bass Morone saxatilis demonstrate that chronic exposures are more likely than episodic 

exposures to affect recruitment (Rose et al. 1993). Nevertheless, episodic exposures could have 

a greater impact on recruitment if they occur at or shortly after a spawning peak or in a discrete 

nursery area (Rose et al. 1993). It is also important to evaluate sublethal effects because they are 

likely additive (and may be synergistic) and need to be considered when evaluating the effects of 

contaminant spills. Currently, deepwater oil facilities use methanol and other pure chemicals 

and proprietary chemical mixtures as additives to facilitate petroleum production. These 

chemicals and mixtures should be evaluated individually and in concert to explore additive and 

synergistic effects to achieve an accurate assessment of their potential sublethal and lethal effects 

on marine fish populations. We found a substantial and significant sublethal effect on individual 

swimming performance from exposure to methanol concentrations below the LC10 level that 

should be incorporated into risk assessments. 
33

4.0 SWIMMING PERFORMANCE OF JUVENILE FLORIDA POMPANO 

AFTER SUBLETHAL EXPOSURE TO ETHYLENE GLYCOL AND 

METHANOL: SYNERGISTIC EFFECTS 

M. A. Stead1

, D. M. Baltz1, 2, E. J. Chesney3, M. A. Tarr4, A. S. Kolok5, and B. D. Marx6

1

Department of Oceanography and Coastal Sciences, Louisiana State University, Baton 

Rouge, LA 70803-7503 

2

Coastal Fisheries Institute, Louisiana State University, Baton Rouge, LA 70803-7503 

3

Louisiana Universities Marine Consortium, Chauvin, LA 70344 

4

Department of Chemistry, University of New Orleans, New Orleans, LA 70148 

5

Biology Department, University of Nebraska at Omaha, Omaha, NE 68182 

6

Experimental Statistics, Louisiana State University, Baton Rouge, LA 70803-5606 

Although a tremendous amount of research has been directed towards understanding the 

effects of offshore oil development on fish communities in the northern Gulf of Mexico 

(Gallaway et al. 1981; Grizzle 1986; Stanley and Wilson 1990; Grossman et al. 1997; 

Boehm et al. 2001, Hymel et al. 2002), little attention has been paid to sublethal effects 

associated with low-level, localized discharges of toxicants, which may reduce an 

individual’s ability to avoid predators, feed, reproduce, and resist diseases and parasites. 

Quantification of the sublethal effects of exposure to one or more contaminants at the 

individual level is important to our understanding of how populations and communities 

may respond to these impacts (Weis et al. 1999). Oil and gas platforms are a source of 

anthropogenic inputs including contaminated waters that are separated from crude oil at 

the extraction point (i.e., produced formation water), drilling-fluid chemicals, oil-based 

and water-based drilling muds and cuttings, and production additives, some of which are 

toxic to the faunas that utilize these structures (Holdway 2002). 

 

Additives such as ethylene glycol and methanol are often transported long distances from 

shore by underwater pipelines up to 96 km in length and by ships to deep-water 

production platforms. There they are stored in volumes of up to 300,000 L per platform 

(Boehm et al. 2001), and are used to prevent the formation of hydrates. Hydrates are 

solid structures that can permanently clog pipelines (Anonymous 1996; Herzhaft and 

Dalmazzone 2000). They form when water molecules crystallize around guest molecules 

in natural gas and other products (Abdelghani et al. 1989; Tephly 1991; Herzhaft and 

Dalmazzone 2000) due to the cooling effect of seawater at extreme depths. 

 In this study, critical swimming speed (Ucrit) was used to quantify the capacity of a 

marine fish, Florida pompano Trachinotus carolinus, for prolonged swimming activity 
34

(Brett 1964). Florida pompano were used as a ‘fish model’ because they occur from the 

beach to offshore in the vicinity of petroleum production platforms and pipelines in the 

northern Gulf of Mexico, they are representative of a number of reef associated species 

(i.e., family Carangidae), they range widely in the western Atlantic and the Gulf, and they 

are highly prized commercial and sport fish (Gilbert 1986). 

 

This study extends previous research by testing the combined sublethal effects of 3.0% 

(v/v) ethylene glycol and 1.07% (v/v) methanol on swimming performance. Swimming 

performance has been used to determine tolerances of fishes to varying environmental 

stresses and pollutants (Cairns 1966; Sprague 1971; Nikl and Farrell 1993; Beaumont et 

al.1995; Heath 1995; Kolok et al. 1998; Hymel et al. 2002). It is associated with the 

ability of individuals to carry out essential activities, integrates the effects of numerous 

physiological processes (Beamish 1978), and relates to Darwinian fitness (Plaut 2001). 

Stress responses are energy draining processes (Schreck 1982; Barton and Schreck 1987) 

that reduce energy available for other life functions (Schreck 1990). Thus, changes in 

swimming performance can provide a measure of stress (Cech 1990). Hymel et al. 

(2002) and Baltz et al. (2005) found significant reductions of 14% and 6.5%, 

respectively, in critical swimming performance, but the synergistic effects of ethylene 

glycol and methanol were more profound than the single exposures. Mean performance 

was synergistically reduced by 41.1% after exposure to both chemicals in combination. 

4.1 Materials and Methods 

4.1.1 Swimming tunnel design 

Three Blazka-type swimming respirometers were used to simultaneously test the 

swimming performance of three individual fish (Blazka et al. 1960). The swim tunnel 

design and techniques developed by Hymel et al. (2002) to quantify swimming 

performance (Ucrit) are similar to those used in previous swimming experiments (Brett 

1964; Parsons 1994; Gregory and Wood 1999). The inner chamber had a diameter of 

20.3 cm with square plastic grids at both ends to contain the fish and reduce turbulent 

flow characteristics, and the total volume of each swimming tunnel was 104 L. Propeller 

speed was adjusted using a hand-held tachometer to set revolutions per minute (RPM) on 

a variable-speed D.C. motor to obtain desired velocities up to a maximum of 100 cm/s 

based on three linear regressions of RPM on Velocity (cm/s) developed by Hymel et al. 

(2002): 

Swimming tunnel 1: Velocity = 0.062(RPM) -4.0099 (r2 = 0.9452, P < 0.001); 

Swimming tunnel 2: Velocity = 0.0627(RPM) -1.1029 (r2

 = 0.9715, P < 0.001); 

and 

Swimming tunnel 3: Velocity = 0.0571(RPM) +0.0582 (r2 = 0.9886, P < 0.001). 
35

4.1.2 Collection and exposure 

Juvenile Florida pompano were collected at Port Fourchon Beach, Louisiana, using a 

3.05 m (10 ft) beach seine and then transferred to holding facilities at Louisiana 

Universities Marine Consortium. Pompano were initially quarantined for one month to 

recover from the stress of capture and treated with 0.2 ppm copper solution (CutrinePlus) to eliminate ectoparasites, principally Amylodinium species. They were held for a 

total of eight weeks in a 5,000 L recirculating seawater system at a salinity of 30 ± 1 

practical salinity units (psu). Throughout the holding and experimental phases of this 

study fish were fed 0.25 g per individual of 45.0% protein pellet fish food and water 

quality was monitored once daily for pH, salinity, and ammonia concentrations. 

For each experiment twenty-four fish were marked with a latex elastomer (Northwest 

Marine Technologies, Inc.) to distinguish individuals. Individuals were anesthetized with 

tricaine methanesulfonate (MS-222; <50 mg/L) and elastomer injected in the dorsal, 

caudal, and anal fins. Following the marking process, fish were allowed to recover for 

two weeks. 

Twenty-four hours prior to the start of each swimming experiment, marked fish were 

randomized into eight groups of three individuals and each group was randomly assigned 

to one of eight 60 L holding tanks. Two groups were randomly assigned to each 

treatment (ethylene glycol, methanol, ethylene glycol + methanol, and control). The 

recirculating holding tank system had a total volume of 550 L and was maintained at a 

salinity of 30 ± 1 psu. A biofilter and a UV light sterilizer were used to maintain water 

quality in the system. Each holding tank was aerated and kept at a constant temperature 

of 25 ± 1°C using 200 W aquarium heaters. 

The experimental protocol included identical fasting, exposure, acclimation, and 

swimming experience for each group and required 18 days. Groups were exposed or 

sham exposed (i.e., without chemical) in one of three round fiberglass tanks filled with 60 

L of 30 ± 1 psu seawater at a temperature of 25 ± 1°C and aerated with two air stones. 

Fish were acclimated for 17 h in the swimming tunnels following a 24-h exposure and 

fasted during the entire 41-h period of exposure (24 h) and acclimation (17 h). 

Swimming performance before and after sublethal exposure to 3.0% ethylene glycol 

(Hymel et al. 2002), 1.07% methanol (Baltz et al. 2005), or the combination of the two 

chemicals (3.0% ethylene glycol and 1.07% methanol) was measured for individual fish. 

These concentrations were selected from dose response curves because they have 

adverse, but not lethal, effects on swim performance (Hymel et al. 2002, Baltz et al. 

2005). Sham exposures were conducted in exactly the same manner as real exposure 

trials except for the addition of the treatment chemical. During the protocol all groups 

were sham exposed once, except the two control groups were sham exposed twice. 

Homogenous mixing of treatment chemicals was tested by taking water samples from 

exposure tanks at the beginning and end of each 24-h exposure period for gas 

chromatography (GC) analysis. GC analyses and paired t-tests (SAS Institute Version 8) 

of exposure concentrations showed that there were no significant differences (P > 0.05) 

between initial and final exposure tank concentrations for ethylene glycol, methanol or 

combination treatments. 
36

Immediately after exposure, groups were placed in a bucket of clean seawater and 

transferred to swimming tunnels that were set up in a constant temperature room at 25 ± 

1°C, and acclimated in clean 30 psu seawater. During the 17-h acclimation, swimming 

tunnels were connected to a recirculating system that included a biofilter and a UV 

sterilizer and velocity in each swimming tunnel was set at 5 cm/s. After acclimation, air 

bubbles were removed and swimming tunnels isolated before the start of the swimming 

trials. Individual fish were tested and re-tested in the same swimming tunnel during the 

pre- and post-exposure phases of the swimming trials. 

4.1.3 Swimming performance 

Swimming performance was tested by determining differences in pre- and post-exposure 

critical swimming speed for individual fish. Velocities in the swimming tunnels were 

increased by 10 cm/s every 10 min, without intervening rest periods. Swimming trials 

ended when the fish fatigued and was pinned against the back grid for two to three 

seconds. Pre- and post-exposure critical swimming speed (Ucrit) was calculated for each 

trial using the equation formulated by Brett (1964): 

Ucrit =ui+(ti/tii x uii); 

where ui is the highest velocity maintained for the prescribed period, uii is the velocity 

increment, ti is the time that fish swam at the fatigue velocity, and tii is the prescribed 

period of swimming. Fish that did not fatigue after swimming for 10 min at the 

maximum velocity of 100 cm/s were assigned a Ucrit value of 100 cm/s. Dissolved 

oxygen in mg/L was measured in each swimming tunnel at the beginning and ending of 

each swimming trial with a YSI Model 85 handheld oxygen, salinity and temperature 

meter to verify that the fish were not stressed by low oxygen concentrations. 

4.1.4 Plasma samples 

At the end of the post-exposure swimming trials, fish were anesthetized using <50 mg/L

of MS-222, and Body Mass (BM) to the nearest 0.01 g and Fork Length (FL) to the 

nearest mm were measured (Table 4-1). Anesthetized fish were sacrificed and blood was 

extracted and pooled for each treatment and control group. Blood was drawn from both 

the caudal and gill vessels into heparinized capillary tubes and samples were placed in an 

IECMicro-MB Centrifuge for 15 min at 12,700 RPM to separate red blood cells from 

plasma. The blood plasma samples were frozen at –80°C until the GC analysis of 

ethylene glycol and methanol. 
37

Table 4-1. Individual fork length (fl mm), body mass (g), and pre- and post-exposure 

ucrit (cm/s, bl/s) values for treatment (methanol, ethylene glycol, and 

combined [m+eg]) and control groups of juvenile florida pompano. 

Statistical summaries are means (±1se) for size variables and lsmeans for 

absolute and relative ucrit values. 

Florida pompano Pre-exposure Ucrit Post-exposure Ucrit

Treatment 

FL 

(mm) Body Mass (g) cm/s BL/s cm/s BL/s 

Methanol 102.00 23.32 96.93 9.50 89.95 8.82 

Methanol 98.00 19.29 99.07 10.11 96.97 9.89 

Methanol* 107.00 25.67 100.00 9.35 100.00 9.35 

Methanol 90.00 15.24 83.85 9.32 74.17 8.24 

Methanol 92.00 16.11 94.18 10.24 91.62 9.96 

Methanol 91.00 15.50 88.13 9.68 89.03 9.78 

Methanol 91.00 14.71 83.88 9.22 82.08 9.02 

Methanol 93.00 15.66 90.83 9.77 85.02 9.14 

Methanol 94.00 17.11 87.40 9.30 89.07 9.48 

Methanol 110.00 26.51 94.93 8.63 91.10 8.28 

Methanol 108.00 24.70 98.48 9.12 99.41 9.20 

Methanol* 102.00 22.14 100.00 9.80 100.00 9.80 

Mean or LSMean 98.17 19.66 92.01 9.47 89.08 9.17 

SE 2.12 1.30 5.39 0.51 5.39 0.51 

 

Ethylene glycol 111.00 27.45 93.53 8.43 95.33 8.59 

Ethylene glycol 102.00 24.31 100.00 9.80 91.30 8.95 

Ethylene glycol 99.00 19.79 100.00 10.10 84.97 8.58 

Ethylene glycol 112.00 28.70 74.18 6.62 82.23 7.34 

Ethylene glycol 100.00 21.34 100.00 10.00 91.20 9.12 

Ethylene glycol 108.00 24.03 91.97 8.52 83.32 7.71 

Ethylene glycol 91.00 15.00 92.86 10.20 89.03 9.78 

Ethylene glycol 97.00 18.51 91.63 9.45 88.88 9.16 

Ethylene glycol 89.00 15.08 92.65 10.41 62.32 7.00 

Ethylene glycol 90.00 14.78 81.43 9.05 68.12 7.57 

Ethylene glycol 91.00 17.41 89.73 9.86 78.68 8.65 

Ethylene glycol 89.00 14.57 95.83 10.77 41.30 4.64 

Mean or LSMean 98.25 20.08 91.41 9.44 79.15 8.10 

SE 2.48 1.46 4.86 0.44 4.86 0.44 

 

M+EG 106.00 26.81 100.00 9.43 86.05 8.12 

M+EG 109.00 24.66 100.00 9.17 65.48 6.01 

M+EG 87.00 13.32 83.32 9.58 44.22 5.08 

M+EG 89.00 14.44 74.26 8.34 71.25 8.01 

M+EG 92.00 14.20 90.48 9.83 72.23 7.85 
38

Florida pompano Pre-exposure Ucrit Post-exposure Ucrit

Treatment 

FL 

(mm) Body Mass (g) cm/s BL/s cm/s BL/s 

M+EG 89.00 16.50 86.03 9.67 6.92 0.78 

M+EG 88.00 14.78 86.95 9.88 14.78 1.68 

M+EG* 87.00 14.87 86.10 9.90 0.00 0.00 

Mean or LSMean 94.29 17.82 90.45 9.32 53.29 5.26 

SE 2.65 1.59 5.74 0.56 5.74 0.56 

 

Control 105.00 27.30 94.95 9.04 100.00 9.52 

Control 105.00 22.30 99.12 9.44 100.00 9.52 

Control* 108.00 25.60 100.00 9.26 100.00 9.26 

Control 112.00 27.92 100.00 8.93 96.85 8.65 

Control 108.00 25.50 94.05 8.71 99.50 9.21 

Control 89.00 16.13 88.08 9.90 59.83 6.72 

Control 95.00 16.34 87.08 9.17 91.17 9.60 

Control 94.00 16.64 86.78 9.23 91.53 9.74 

Control 88.00 16.22 85.10 9.67 92.65 10.50 

Control 92.00 16.28 84.43 9.18 82.82 9.00 

Control 94.00 17.81 89.06 9.48 92.95 9.89 

Mean or LSMean 99.09 20.73 90.66 9.29 90.53 9.25 

SE 2.48 1.44 5.10 0.47 5.10 0.47 

* Surviving individuals excluded from all statistical analyses 

4.1.5 Statistical analysis 

A repeated measures design was used to determine the differences in pre- and postexposure Ucrit for the replicated Florida pompano swimming performance tests. Although 

a 1-way ANOVA used to examine the sizes of individuals randomized among four 

treatments was unable to detect size differences among treatments (d.f. = 3, 35; F = 0.39; 

P> 0.76), we included size as a covariate and report Ucrit in absolute (cm/s) and relative 

(body lengths per second, BL/s) terms. Statistical analyses of Ucrit in absolute and 

relative terms, with and without size (mm FL) as a covariate, yielded similar results. 

Replicate experiments were conducted in August, 2001 and September, 2001 to increase 

total sample size and treatments were intermingled to avoid pseudoreplication (Hurlbert 

1984). There were two experimental replicates (1 or 2). Exposure tank number (1, 2 or 

3) and fish were the experimental units for the chemical/sham treatment and trial time, 

respectively. The Mixed procedure of SAS (SAS Institute 1996) was used to test the 

model Yijklm = μ + ρi + αj + αρij + T(αρ)k(ij) + τl + ατjl + F(ραT)m(ijk) + ßL ijkm + εijklm for 

swimming performance tests where ρ i is the random effect of replication (experiment 1 

or 2), αj is the fixed treatment effect (j=1, 2, 3, 4), αρij is the random interaction of 

replication by treatment, T(αρ)k(ij) is the random effect of tank (1, 2 or 3) nested within 

experiment by treatment, τl is the fixed trial effect (pre- or post-exposure), ατjl is the fixed

interaction between treatment and trial, and F(ραT) m(ijk) is the random effect of fish 
39

nested within tank by experiment by treatment, and ß is the slope associated with the 

covariate length (L) for each fish. The distribution of the model’s residuals showed 

strong symmetric and unimodal behavior. However, due to a few observations in the 

tails, the Shapiro-Wilk’s test rejected normality (P < 0.01). As the F-tests are robust for 

such residual distributions (Glass et al. 1972), we do not present analyses using 

transformed responses. In post-exposure testing, differences between least-square means 

(LSMeans) were calculated for all possible treatment comparisons. 

Nine fish were excluded from the statistical analysis resulting in a final sample size of 39 

individuals. One control and three individuals belonging to the combination treatment 

group died in the exposure tanks during the pre-exposure (sham) experience. Following 

the combination exposure (actual experience), one individual died during acclimation to 

the swim tunnel and another individual would not swim Another three individuals (i.e., 

1 control and 2 methanol fish) out-swam the capacity of the swim tunnel during both preand post-exposure trials. 

4.2 Results 

4.2.1 Swimming performance 

In the experiments designed to test for potential additive, antagonistic or synergistic 

effects of simultaneous sublethal exposures, the combination groups exposed to both 

methanol and ethylene glycol performed significantly more poorly than single exposure 

and control groups (Figure 4-1). 

No detectable difference was observed for the main effect of treatment (df = 3, 3, F = 

4.59, P > 0.12); however, the trial (df = 1, 34, F = 29.36, P< 0.0001) and the treatment by 

trial interaction (df = 3, 34, F = 10.05, P < 0.0001) were highly significant, while the size 

covariate was not (df = 1, 34, F = 3.53, P = 0.0689). The significant interaction indicated 

the pre- and post-exposures to ethylene glycol, methanol, the combination, and controls 

differed widely in effects on swimming performance. The LSMeans from the Mixed 

procedure (Figure 4-1) showed evidence of a significant difference between pre- and 

post-exposure swimming performance of pompano exposed to 3.0% ethylene glycol (P < 

0.0071) and combination treatment (P < 0.0001). There was no evidence that pre- and 

post-exposure Ucrit differed for pompano exposed to 1.07% methanol (P > 0.53) or 

controls (P > 0.97). 

Individuals exposed to 1.07% methanol exhibited a 3.2% mean reduction in Ucrit (Figure 

4-1) but the reduction was not detectably different from zero (P > 0.53). Mean (±1 SE) 

pre-exposure Ucrit was 92.01 ± 5.39 cm/s (9.47 ± 0.53 BL/s) and the post- exposure Ucrit

was 89.08 ± 5.39 cm/s (9.17 ± 0.53 BL/s). Seven of the 10 pompano exhibited a mean 

decrease in Ucrit of 5.17%, three fish showed a minor increase of 1.29% in Ucrit (Figure 4-

2). 
40

Figure 4-1. Critical swimming speeds (Ucrit, LSMean (±1 SE)) of juvenile Florida 

pompano before and after 24-h exposures to 1.07% (v/v) methanol (N = 

10), 3.0% (v/v) ethylene glycol (N = 12), the combined chemicals (N = 7), 

and sham-exposed controls (N = 10). 

Individuals exposed to ethylene glycol exhibited a mean 13.4% decrease in Ucrit (Figure 

4-1). Mean (±1 SE) pre-exposure Ucrit was 91.41 ± 4.86 cm/s (9.45 ± 0.44 BL/s) and 

mean post-exposure Ucrit was 79.15 ± 4.86 cm/s (8.10 ± 0.44 BL/s). Ten of the 12 fish 

exhibited a mean 16.74% reduction in Ucrit, while the remaining two fish showed a mean 

increase in Ucrit of 6.39% after exposure to ethylene glycol (Figure 4-2). 

Individuals exposed to both ethylene glycol and methanol in combination registered the 

largest drop in performance of all treatment groups (Figure 4-1), a mean decrease in Ucrit

of 42.1%. The mean (±1 SE) pre-exposure Ucrit of combination exposed fish was 90.45 ± 

5.74 cm/s (9.32 ± 0.56 BL/s) and mean post-exposure Ucrit was 53.29 ± 5.74 cm/s (5.26 ± 

0.56 BL/s). A noticeable amount of individual variation in Ucrit was observed for the 

combination treatment. Four of the seven fish exhibited a mean decrease of 64.10% in 

Ucrit, while the remaining three fish showed a relatively smaller decrease of 12.72% after 

exposure (Figure 4-2). 

Individuals in control groups performed at relatively equal levels after both shamexposures (Figure 4-1) with a mean overall decrease of 0.1%. Mean (±1 SE) preexposure Ucrit was 90.66 ± 5.10 cm/s (9.29 ± 0.47 BL/s) and mean post-exposure Ucrit

was 90.53 ± 5.10 cm/s (9.25 ± 0.47 BL/s). Seven of the 10 control fish exhibited a mean 

increase in Ucrit of 5.06%, while three fish showed a mean decrease of 12.38% (Figure 4-

2). 

45

55

65

75

85

95

Control Methanol Ethylene

Glycol

Combined

 Ucrit cm/s

Preexposure

Postexposure
41

Figure 4-2. Individual variation in the critical swimming speeds (Ucrit) of juvenile 

Florida pompano before and after 24-h exposures to 1.07% methanol 

(N=10), 3.0% ethylene glycol (N = 12), the combined chemicals (N = 7), 

and sham-exposed controls (N = 10). 

4.2.2 Plasma samples 

GC analysis of plasma samples pooled from each group was used to detect traces of 

ethylene glycol or methanol in each treatment and control group. After 17 h of 

acclimation, up to 100 min of experimental swimming trials, and 0.5 h of 

postexperimental processing, traces of ethylene glycol were found in fish exposed to the 

combination (N = 3 groups) and single ethylene glycol treatments (N = 4 groups); 

however, no trace of methanol was detected in plasma samples from fish exposed to the 

combination (N = 3 groups) or single methanol (N = 4 groups) treatments. Pooled 

plasma collected from juvenile Florida pompano groups (N = 7 groups) exposed to 3.0% 

ethylene glycol or combination treatment contained a mean percent (±1 SE) ethylene 

glycol concentration of 0.623 ± 0.04%. 

4.3 Discussion 

The simultaneous exposure to sublethal concentrations of ethylene glycol and methanol 

significantly reduced the Ucrit of juvenile Florida pompano. The 41.1% reduction in 

performance caused by exposure to the combination of contaminants was profound 

compared to the reductions exhibited by fish exposed solely to ethylene glycol (a 13.4% 

reduction) or methanol (a 3.2% reduction). The reduction in Ucrit for Florida pompano 

exposed to 3.0% ethylene glycol (a 13.4% reduction) was similar to the 14.0% reduction 

Methanol 

0

20

40

60

80

100

Ucrit cm/s

Ethylene Glycol

0

20

40

60

80

100

Preexposure Postexposure

Methanol + 

Ethylene Glycol

0

20

40

60

80

100

Preexposure Postexposure

Ucrit cm/s

Control

0

20

40

60

80

100
42

observed by Hymel at el. (2002); however, the 3.2% reduction for methanol was 

somewhat less than the significant 6.5% reduction observed by Baltz et al. (2005). In this 

study sublethal exposure to 1.07% methanol had a short-term effect on behavior but did 

little to impair the swimming performance after 17 h of recovery. The lack of detectable 

level of methanol in the plasma samples suggests that pompano completely metabolized 

methanol (probably into formaldehyde and formate products) after ~ 20 h of acclimation 

and swimming (Potts and Johnson 1952; Koivusalo 1970; Tephly 1991); however, 

methanol was detectable in plasma in similarly exposed individuals that were not 

exercised (Baltz et al. 2005). In contrast, the mean percentage (±1 SE) of ethylene glycol 

still present in pompano plasma samples ~ 20 h after exposure, acclimation and exercise 

was 0.623 ± 0.04%. 

 

4.3.1 Combination exposure 

Compared to fish exposed solely to the single concentration of ethylene glycol, 

performance declined by three fold in fish exposed to the combination of ethylene glycol 

and methanol. Similar joint action or additive interactions occur when toxicants have 

similar modes of action, but act independently, resulting in the toxicity of the mixture 

amounting to the sum of the toxicities of the individual toxicants present (Bliss 1939; 

Sprague 1970; Marking 1985). Interactive action can be either synergistic (more than 

additive) or antagonistic (less then additive). These interactions occur when one toxicant 

alters the toxicity of another toxicant present. The synergistic interaction observed in our 

study was similar to that of Greene and Kocan (1997) who found the toxicity of ethylene 

glycol increased in freshwater fish when sublethal levels of thiram, a chemical used in an 

agricultural seed-protectant, was present. 

A discharge or spill often involves more then one chemical component and the 

interaction between two or more chemicals can result in toxicity that is greater than the 

sum of the toxicities of the individual components (Greene and Kocan 1997). Both 

ethylene glycol and methanol are competitors for aldehyde dehydrogenase (ALDH) in 

humans and may inhibit each other from being completely metabolized and increase the 

toxic effects after exposure (Dawidek-Pietryka et al. 1998). The metabolic pathways of 

ethylene glycol and methanol are in fact so similar that treatment for ethylene glycol or 

methanol poisoning is the administration of ethanol, another ALDH competitor 

(Dawidek-Pietryka et al. 1998). If the metabolic pathways are similar in pompano, then 

combining the concentrations of the two would lead to more or less the same number of 

toxic intermediate compounds as would the doubling of either concentration without 

mixing them together. In one case the only toxic intermediate would be formaldehyde or 

glycoaldehyde, while in the other it would be both; nevertheless, if we assume that the 

toxic intermediates are more or less equally toxic, then we would expect an additive 

rather than a synergistic effect (i.e., one toxic aldehyde + a second toxic aldehyde 

probably causes the equivalent toxicity of two aldehydes of the same chemical species). 

Thus, it is interesting that the metabolites probably increased more or less arithmetically, 

while the effect on the fish was not arithmetic, but synergistic. 
43

4.3.2 Individual variation 

Florida pompano mean Ucrit decreased after exposure to ethylene glycol, methanol and 

combination treatments by 13.4, 3.2, and 41.1%, respectively. However, there was 

considerable individual variation in Ucrit within treatment and control groups following 

exposure and sham-exposure (Figure 4-2). In previous swimming performance studies, it 

was common practice to compare Ucrit means of different treatment and controls groups 

in a one-way ANOVA design, but this approach fails to address within-group variation. 

Within-group variability made it difficult to find significant differences between 

treatment and control group means, which can limit the usefulness of swimming 

performance tests (Kolok et al. 1998). In this and related studies (Hymel et al. 2002; 

Baltz et al. 2005) swimming performance of individual fish before and after exposure 

was used in a split-plot design with blocks to take advantage of the within-group 

variability to analyze changes in swimming performance (Kolok 1992). 

Previous studies revealed conflicting results concerning the performance variability 

among individuals after stressful treatments. Kolok et al. (1998) found large within 

group variation in Ucrit during swimming performance tests, but this variation was 

repeatable over replicate experiments. In contrast Kolok and Farrel (1994) found that the 

reduction in Ucrit of northern pikeminnow Ptychocheilus oregonensis after a surgical 

technique was fairly consistent. The results of our study resembled those of Kolok et al. 

(1998) in that the individual performance in Ucrit after exposure to ethylene glycol, 

methanol, and combination treatments varied between slight increases and striking 

reductions. The variability in this study implies that individual fish may have lower or 

higher tolerances to ethylene glycol, methanol or combination treatments. This variation 

may also account in part for the non-significant reduction (3.2%) due to methanol 

exposure that we found in contrast to earlier, significant findings (6.5%) by Baltz et al. 

(2005). 

4.3.3 Ecological effects 

Although Boehm et al. (2001) found that realistic models of chemical spills for ethylene 

glycol or methanol would result in exposures well below the lowest median lethal 

concentration (LC50) values for other species found in the literature, pompano appear to 

be more sensitive to the two chemicals than any species they examined (Hymel et al. 

2002; Baltz et al. 2005). An episodic spill of ethylene glycol and methanol could have 

severe ecological implications at individual, community, or population levels. 

Nevertheless, it is difficult to forecast how the effects of exposure at the individual level 

will translate to populations or communities (Weis et al. 1999). Fish populations and 

recruitment are likely to exhibit compensatory responses (Houde 1989) to an episodic 

stress or mortality such as a spill; therefore, a spill may not necessarily affect a species at 

the population level if the exposure is not sustained or widespread (Rose et al. 1993). But 

a spill during critical periods of life history could reduce recruitment and add an 

additional stress to populations already burdened by overfishing, reduced habitat quality 

and other contaminants (Rose et al. 1993). 
44

It is also important to evaluate sublethal effects because they are likely additive (and may 

be synergistic) and need to be considered when evaluating or forecasting the effects of 

contaminant spills. Currently, deepwater oil facilities use methanol and other pure 

chemicals and undisclosed proprietary chemical mixtures as additives to facilitate 

petroleum production. These pure chemicals and mixtures should be evaluated 

individually and in concert to explore additive and synergistic effects to achieve an 

accurate assessment of their potential sublethal and lethal effects on marine fish 

populations. We found a substantial and significant synergistic effect on individual 

swimming performance from exposure to methanol concentrations below the LC10 level 

and to ethylene glycol concentrations below the LC01 level that should be incorporated 

into risk assessments. An understanding of sublethal and additive and synergistic effects 

is important for the accurate prediction of the impacts to habitat and fish populations 

because many species may experience chronic low levels of one or more contaminants 

over time. For example, in a study of fishes surrounding oil drilling platforms Grizzle 

(1986) found histological changes in the gills that possibly reflected chronic, low-level 

exposure to drilling contaminants. 

Most studies of petroleum-related contaminants have concentrated on the mortality 

effects associated with major oil spills (Howarth 1991; Kennicutt et al. 1996; Peterson et 

al. 1996), but much remains unknown about the effects of low-level exposures on marine 

organisms. Sublethal effects of contaminant exposure are inherently more difficult to 

assess and extrapolate to the field. We have shown that individual marine fish experience 

reduced Ucrit (mostly aerobic) swimming performance when exposed to sublethal levels 

of methanol (Baltz et al. 2005) or ethylene glycol (Hymel et al. 2002). The implications 

of these results include adverse effects on the ability to perform aerobically or 

anaerobically in ecological situations (Plaut 2001). The Ucrit procedure was developed 

primarily as a measure of aerobic capacity, even though the fish is pushed to use 

anaerobic metabolism and fatigue (K. Dickson, Cal. State Univ. Fullerton, personal 

communication). While maximum (sensu Sepulveda and Dickson 2000) cruising speeds 

of Florida pompano are probably not near the high pre-exposure Ucrit values observed in 

this study, the capacity for sprint (seconds, anaerobic), burst (minutes, anaerobic), and 

critical (hours, aerobic) swimming modes (sensu Reidy et al. 2000) are all probably 

affected to some degree by contaminant exposure. Each mode has direct relevance to 

foraging, prey capture, predator avoidance, trawl avoidance, strong currents, migrations, 

or spawning. However, all three modes of swimming have been tested in only one 

species, Atlantic cod Gadus morhua. Reidy et al. (2000) found that for healthy cod, Ucrit

and sprint performances are positively correlated, but Ucrit and burst performances are 

negatively correlated, and sprint and burst performances are uncorrelated. Morphological 

and physiological tradeoffs among aerobic and anaerobic swimming types in other 

species and how contaminant exposure affects their sprint, burst, and endurance 

swimming performances remains unstudied (Plaut 2001, Nelson et al. 2002). What is 

clear is that exposure to contaminants at sublethal concentrations can reduce swimming 

performance of fishes and synergistic effects are likely. 

 
45

5.0 SWIMMING PERFORMANCE OF JUVENILE ATLANTIC SPADEFISH 

EXPOSED TO SUBLETHAL CONCENTRATIONS OF ETHYLENE GLYCOL 

M. A. Stead1

, D. M. Baltz1, 2, E. J. Chesney3, M. A. Tarr4, A. S. Kolok5, and B. D. Marx6

1

Department of Oceanography and Coastal Sciences, Louisiana State University, Baton Rouge, 

LA 70803-7503 

2

Coastal Fisheries Institute, Louisiana State University, Baton Rouge, LA 70803-7503 

3

Louisiana Universities Marine Consortium, Chauvin, LA 70344 

4

Department of Chemistry, University of New Orleans, New Orleans, LA 70148 

5

Biology Department, University of Nebraska at Omaha, Omaha, NE 68182 

6

Experimental Statistics, Louisiana State University, Baton Rouge, LA 70803-5606 

There are an estimated 4,000 oil and gas platforms in the northern Gulf of Mexico that, as 

artificial reefs, constitute most of the known hard substrate off the Louisiana and Texas coasts 

(Stanley and Wilson 2000). Oil and gas platforms extend throughout the water column and are 

thought to affect benthic, demersal and pelagic fishes (Gallaway et al. 1981; Continental Shelf 

Associates 1982). They attract available recreational fishes and may enhance production of other 

species (Bohnsack 1989; Grossman et al. 1997). They also create fishing opportunities, reduce 

user conflicts, save time and fuel, make locating fish more predictable, and increase public 

access (Stone 1985; National Academy Press 1988). The success of platforms as artificial reefs 

is indicated by the concentration of fishes and recreational fishing around these structures off the 

coast of Louisiana (Stanley and Wilson 1990). Unfortunately, oil and gas platforms are also a 

source of anthropogenic inputs including produced formation water, drilling fluid chemicals, oilbased and water-based drilling muds and cuttings and production additives, some of which are 

toxic to the fauna that utilizes these structures (Holdway 2002). 

As petrochemical exploration moves into increasingly deeper water (>300 m), facilitated by new 

technology, there are risks to the marine fauna including the release of contaminants (Boehm et 

al. 2000) and the noise pollution associated with the construction and operation of offshore oil 

platforms (Richardson 1995). New chemical additives and combinations of additives are used to 

enhance deep-water production. For example ethylene glycol is used during the production and 

treatment to prevent the formation of gas hydrates in deep-water wells and pipelines 

(Anonymous 1996, 2000; Herzhaft and Dalmazzone 2000; Boehm et al. 2001). It is often 

transported long distances through underwater pipelines and by ships to deep-water production 

platforms. Although the possibility of a large spill may be low, the risk remains plausible with 
46

increased usage, transport and storage of volumes of up to a 300,000 L of ethylene glycol per 

platform (Boehm et al. 2001). 

Ethylene glycol is a common chemical used by both industry and consumers. It is a common 

ingredient in antifreeze, is used as a deicing agent on bridges and airport runways, and can 

prevent the formation of hydrates in deep-water pipelines (Abdelghani et al. 1990; Herzhaft and 

Dalmazzone 2000). Hydrates are solids that form when water molecules crystallize around 

“guest” molecules in natural gas and other petroleum products (Herzhaft and Dalmazzone 2000). 

Hydrate formation is particularly common in underwater pipelines due to the cooling effect of 

sea water on pipeline contents at extreme depths (Anonymous 1996). Hydrates form during 

drilling and extraction process, but also during transport in deep-water pipelines and can result in 

permanent blockage (Herzhaft and Dalmazzone 2000). With intensified activity in deeper water, 

larger volumes of ethylene glycol are being transported, stored and used by the offshore industry. 

Ethylene glycol is toxic to both mammals and fishes to varying degrees (Bridie et al. 1979; 

Adbelghani et al. 1989; Pillard 1995; Green and Kocan 1997 and Hymel et al. 2002). 

Although research has been directed towards understanding the effects of offshore oil 

development on the northern Gulf of Mexico (Gallaway et al. 1981; Grizzle 1986; Stanley and 

Wilson 1990; Grossman et al. 1987; Boehm et al. 2001, Hymel et al. 2002), little attention has 

been paid to understanding the chronic or acute sublethal effects associated with low-level, 

localized discharges of toxicants. Such sublethal effects may reduce an individual fish’s ability 

to avoid predators, feed, reproduce, and resist diseases and parasites, and may, in turn, lead to 

effects at population and community levels (Weis et al. 1999). Therefore, it is important to 

quantify effects at the individual level to enhance our understanding of how populations and 

communities may respond. 

Induced changes in swimming performance were first suggested as an important tool for 

assessing the sublethal effects of toxicants on fishes by Cairns (1966) and have since been used 

to determine tolerances of fishes to varying environmental factors, stresses and pollutants (Nikl 

and Farrell 1993; Beaumont et al. 1995, Heath 1995; Kolok et al. 1998; Hymel et al. 2002). 

Swimming performance is an indicator of the ability of an individual to capture prey, escape 

predation, migrate and maintain position in a current and involves the integrated effects of 

numerous physiological processes (Beamish 1978). Selyle (1950) first defined stress as the sum 

of all the physiological responses by which an animal tries to maintain or re-establish a normal 

metabolism in the face of a physical or chemical force. Stress response and resistance to a 

stressor are energy draining processes (Schreck 1982; Barton and Schreck 1987) and by 

responding to stress an organism should have less energy available to devote to other life 

functions (Schreck 1990). Estimation of swimming ability can, therefore, provide an index of 

stress caused by exposures to chemical additives that can reduce an individual’s scope for 

activity (Cech 1990). 

In this study, critical swimming speed (Brett 1964) was used to quantify the capacity of 

individual fish for prolonged swimming activity. Prolonged swimming is a higher level of 

activity that can be maintained for minutes to hours depending on velocity, but eventually results 

in muscular fatigue (Beamish 1978). Burst swimming is extremely intense activity that can only 

be maintained for seconds (Hartwell and Otto 1991). Burst swimming would probably be a 
47

better ecological index, but is more difficult to assess in laboratory experiments. Atlantic 

spadefish (Chaetodipterus faber) were selected to evaluate the sublethal effects of ethylene 

glycol because they often congregate in the vicinity of deep-water pipelines and gas and oil 

wells. The Atlantic spadefish is the only member of the family Ephippidae in the Atlantic Ocean 

(Hayse 1989). Spadefish are one of the most common fish found near oil platforms and are 

sometimes caught by sport fishermen, but little is known about their life history. Spadefish feed 

on a variety of hydroids, epifaunal amphipods, and anthozoans. A study on the stomach contents 

by Hayse (1989) found the cannonball jelly (Stomolophus meleagris) to be the dominant 

organism by all numeric indicators. Spadefish are believed to spawn from May through 

September, principally in the northern Gulf of Mexico (Ditty et. al 1993). Juveniles migrate 

inshore during summer and young of the year move into deeper offshore waters once 

temperatures cool. Sexual maturity is reached after 2 years at a size of 135 mm total length 

(Johnson 1978). 

The goal of this study was to extend previous research on the effects of ethylene glycol on 

Florida pompano (Hymel et al. 2000) by testing the effects of a 3.0 % (volume/volume) exposure 

on swimming performance of another marine fish species. Atlantic spadefish were exposed to the 

same concentration of ethylene glycol in an experimental design developed by Hymel et al. 

(2002) wherein fish exposed to 3.0 % ethylene glycol concentration displayed lethargic behavior 

but improved after the recovery period and few mortalities were observed. Ethylene glycol did 

reduce the swimming performance of juvenile Atlantic spadefish by an average of 7.3 % after 

exposure and by 18.4 % compared to the improvement exhibited by the control group. 

 

5.1 Materials and Methods 

The purpose of this experiment was to evaluate potential effects that sublethal concentrations of 

ethylene glycol have on the swimming performance of a juvenile Atlantic spadefish. Spadefish 

were used to test the effects of a 24-hr exposure to 3.0 % ethylene glycol on the swimming 

performance of individual fish. Swimming performance was evaluated by comparing differences 

in pre- and post-exposure critical swimming speeds for each individual. Spadefish were exposed 

to the same concentration of ethylene glycol used in previous experiments by Hymel et al. 

(2002). The swimming performance experiment was repeated a second time to increase sample 

size. Logistical and time constraints prevented us from conducting a single swimming 

performance trial with a larger sample size without altering the protocol used in previous 

research on ethylene glycol (Hymel et al. 2002). 

5.1.1 Respirometer design 

Three Blazka-type swimming respirometers were used to simultaneously test swimming 

performance of three individual fish (Blazka et al. 1960). The respirometer design and 

techniques developed by Hymel et al (2002) to quantify critical swimming speed (Ucrit) are 

similar to those used in previous swimming experiments (Brett 1964; Parsons 1994; Gregory and 

Wood 1999; Hymel et al. 2002). The inner respirometer chamber had a diameter of 20.3 cm and 

the total volume of each respirometer was 104 L. Plastic square grids at both ends of the inner 

chamber contained the fish during swimming trials and reduced turbulent flow characteristics. A 
48

propeller driven by a variable-speed D.C. motor provided water velocities up to a maximum of 

100 cm s-1. Propeller speed was adjusted using a hand-held tachometer to set motor revolutions 

per minute (RPM) to obtain desired velocities for individual chambers based on system 

calibrations by Hymel et al. (2002): 

 Swimming chamber 1: Velocity = 0.062(RPM) – 4.0099; (r2

 = 0.9452, P < 0.001), 

 Swimming chamber 2: Velocity = 0.0627(RPM) – 1.1029; (r2

 = 0.9715, P < 0.001), 

 Swimming chamber 3: Velocity = 0.0571(RPM) + 0.0582; (r2 = 0.9886, P < 0.001). 

5.1.2 Collection and exposure 

Atlantic spadefish were captured by trawls in Terrebone Bay, Louisiana. Fish were maintained 

in live-wells and then transferred to Louisiana Universities Marine Consortium. Spadefish were 

held in captivity for one month to recover from the stress of capture and treated with 0.2 ppm 

copper solution (Cutrine-Plus) to eliminate ectoparasites, principally Amylodinium sp. They 

were held for a total of eight weeks in a 5000 L recirculating seawater system at a salinity of 30 

± 1 practical salinity units (psu). Two weeks before the start of the swimming experiment, 1.3 g 

of Praziqantel dissolved in 20 ml of 95.0 % ethanol was added to an isolated 1000 L tank 

containing experimental fish to remove gill flukes. Throughout the holding and experimental 

phases of this study spadefish were fed 1.5 g per individual of frozen shrimp once daily and 

water quality was monitored once daily for pH, salinity, and ammonia concentrations. 

For each experiment, eighteen fish were uniquely marked with latex elastomer (Northwest 

Marine Technologies, Inc.) to distinguish individuals and later randomized among treatments 

into eight holding tanks with three individuals per tank. Individuals anesthetized with ≤ 50 mg/L 

of MS-222 and elastomer injected in the dorsal, caudal, and anal fins. Following the marking 

process, spadefish were allowed to recover for two weeks. Twenty-four hours prior to the start 

of swimming experiment trials, marked fish were randomized into six groups of three individuals 

and each group was randomly assigned to one of eight 60 L holding tanks. Five groups were 

randomly assigned as treatment groups and the remaining group was designated as the control 

group. The recirculating holding system had a total volume of 550 L and was maintained at a 

salinity of 30 ± 1 psu. A biofilter and a UV light sterilizer were used to maintain water quality in 

the system. Each holding tank was aerated and kept at a constant temperature of 25 ± 1 °C using 

200 W aquarium heaters. Groups were exposed or sham exposed in one of three round fiberglass 

tanks filled with 60 L of 30 ± 1 psu sea water at a temperature of 25 ± 1 °C and aerated with two 

air stones. 

The experimental protocol included identical fasting, exposure, acclimation, and swimming 

experiences for each group and required 15 days. Fish were acclimated for 17 hr in the 

swimming respirometers following the 24-hr exposures, and were fasted during the entire 41-hr 

period of exposure (24 hr) and acclimation (17 hr). Swimming performance for Atlantic 

spadefish before and after exposure to 3.0 % ethylene glycol concentration was measured for 

individual fish (Hymel et al. 2002). Sham exposures were conducted in exactly the same manner 

as real exposure trials except for the addition of the treatment chemical. During the protocol all 

groups were sham exposed once, except the control group was sham exposed twice. Treatment 
49

groups were exposed for 24 hrs before the post-exposure swimming test. Homogenous mixing 

of treatment chemical was tested by taking water samples from exposure tanks at the beginning 

and end of each 24-hr exposure period for gas chromotography (GC) analysis. The GC analysis 

and paired t-test (SAS Institute Version 8) of exposure concentrations revealed no significant 

difference between 3.0 % ethylene glycol initial and final exposure tank concentrations. After 

exposure, groups were placed in a bucket of clean sea water and transferred to swimming 

respirometers that were set up in a constant temperature room at 25 ± 1 °C. During the 

acclimation, respirometers were connected to a recirculating system that included a biofilter and 

a UV sterilizer and flow in each respirometer was set at a velocity of 5 cm s-1. All air bubbles 

were removed and respirometers were isolated before the start of the swimming trials. 

Individual fish were tested and retested in the same respirometers during the pre- and postexposure phases of the swimming trials. 

5.1.3 Swimming performance 

Swimming performance was evaluated by determining differences in pre- and post-exposure 

critical swimming speed for individuals. Velocities in the respirometers were increased by 10 

cm s-1 every 10 min, with no rest period between 10-min swimming intervals. Swimming trials 

ended when the fish fatigued and was momentarily pinned against the back grid for two to three 

seconds. Pre- and post-exposure critical swimming speed (Ucrit) was calculated for each trial 

using Brett’s (1964) equation: 

Ucrit = ui+(ti/tii x uii); 

where ui is the highest velocity maintained for the prescribed interval, uii is the velocity 

increment, ti is the time that fish swam at the fatigue velocity, and tii is the prescribed interval of 

swimming. Fish that did not fatigue after swimming for 10 min at the maximum velocity of 100 

cm s

-1 were assigned a Ucrit value of 100 cm s-1. Dissolved oxygen in mg L-1 was measured in 

each respirometer at the beginning and ending of each swimming trial with a YSI Model 85 

handheld oxygen, conductivity, salinity and temperature meter to verify that the fish were not 

stressed by low oxygen concentrations. 

5.1.4 Plasma concentration 

Blood plasma from each fish was analyzed for concentrations of ethylene glycol using GC 

analysis. At the end of the post-exposure swimming trials, fish were anesthetized using ≤ 50 

mg/L of MS-222, and Body Mass (BM) to the nearest 0.01 g and Fork Length (FL) to the nearest 

mm were measured. Fish were sacrificed and blood was extracted and pooled for three 

individuals in each treatment and control group. Blood was drawn from both the caudal and gill 

vessels into heparinized capillary tubes. Blood samples were placed in an IECMicro-MB 

Centrifuge for 15 min at 12,700 RPM to separate red blood cells from plasma. The plasma from 

the each group was frozen at –80 °C until GC analysis. 
50

5.1.5 Statistical analysis 

A split plot design was used to determine the differences in pre- and post-exposure Ucrit for the 

Atlantic spadefish swimming performance tests. Ucrit values were standardized from cm s-1 to 

body lengths per second (BL s-1) to account for variation in individual size. Replicate 

experiments were conducted in September, 2000 and January 2002 to increase the sample size to 

a total of N = 36 individuals. Experiment replicates (1 or 2), exposure tank number (1, 2 or 3) 

and fish were used as subplots. The Mixed procedure of SAS (SAS Institute Version 8) was 

used to test the model Yijk = μ + ρi + αj + (αρ)ij + T(αρ) + τk + (ατ)jk + (ρτ)ik + F(Tρα )ijk + ε 

for Atlantic spadefish swimming performance tests where ρ i is the random effect of replication 

(experiment 1 or 2), αj is the fixed treatment effect, (αρ)ij is the random interaction of replication 

by treatment, T(αρ) is the random interaction of tank (1, 2 or 3) nested within experiment by 

treatment, τk is the fixed trial effect (pre- or post-exposure), (ατ)jk is the fixed interaction between 

treatment and trial, and F(Tρα )ijk is the random interaction of fish nested within tank by 

experiment by treatment. The Shapiro-Wilks test (W = P < 0.0225; W = 0.9484) indicated 

evidence the residuals were not normally distributed for the Atlantic spadefish swimming 

performance tests, however residuals plots exhibited a normal symmetrical distribution with a 

few outliers. The F-test is fairly robust in considering non-normal distributions and no other 

transformations were deemed necessary (Glass et al. 1972). In post-exposure testing, differences 

between least-square means (LSMeans) were calculated for all possible treatment comparisons. 

5.2 Results 

5.2.1 Swimming performance 

The treatment groups exposed to ethylene glycol performed significantly more poorly than 

control groups (Figure 5-1). Five treatment fish did not survive the swimming performance 

experiment leaving 25 treatment and 6 control fish for the analysis. The overall F-test (df = 32, F 

= 2.63, P < 0.0050), the main effects of treatment (df = 1, F = 315.62, P < 0.0359) and trial (df = 

1, F = 0.0593, P < 0.0893) and the treatment by trial interaction (df = 1, F = 3.15, P < 0.0861) 

were all significant at the α = 0.10 level. The significant interaction (P < 0.0861) indicated the 

pre- and post-exposures by treatment differed widely in effects on swimming performance. The 

Least Square Means (LSMeans) from the Mixed procedure showed evidence of a significant 

difference between post-exposure Ucrit for treatment groups and post-exposure control groups 

(P < 0.0085). No detectable difference in Ucrit was observed within pre- and post-exposure 

groups for treatment groups (P < 0.0921), control (P < 0.2699), or between the pre-exposure 

treatment and control groups (P < 0.7462). 
51

Figure 5-1. Mean (± SE) critical swimming speeds (Ucrit ) of juvenile Atlantic spadefish 

before and after a 24-hr exposure to 3.0 % ethylene glycol (N = 25) and sham 

controls (N = 6). 

Sublethal exposure to 3.0 % v/v ethylene glycol reduced the swimming performance of juvenile 

Atlantic spadefish by 7.3 % compared to their initial performance (P < 0.0921) and by 18.4 % 

compared to the final performance (P < 0.0085) of the controls. The mean (± SE) pre-exposure 

critical swimming speed of treatment fish (Table 5-1) was 66.90 ± 5.54 cm s-1 (6.59 ± 0.36 BL 

s

-1). Mean post-exposure critical swimming speed of treatment fish was 62.44 ± 5.54 cm s-1

(6.14 ± 0.36 BL s-1). Values for pre- and post-exposure Ucrit of treatment fish may be slightly 

conservative due to one fish (1 out of 31) that did not fatigue after 10 min at 100 cm s-1 and was 

subsequently assigned a Ucrit value of 100 cm s-1 in both pre- and post-trials. 

5

5.5

6

6.5

7

7.5

8

8.5

Ucrit (BL s-1)

Control

Treatment

 Pre Post
52

Table 5-1. Individual Body Mass (g), Fork Length (mm) and pre-exposure and post-exposure 

Ucrit (cm s-1, BL s-1) for treatment and control groups of Atlantic spadefish. 

Spadefish Pre-exposure Post-exposure 

 Ucrit Ucrit

Treatment 

Fish FL (mm) BM (g) cm s-1 BL s-1 cm s-1 BL s-1

1 

2 

3 

4 

5 

6 

7 

8 

9 

10 

11 

12 

13 

14 

15 

16 

17 

18 

19 

20 

21 

22 

23 

24 

25 

97 

110 

83 

93 

105 

88 

108 

88 

98 

76 

116 

119 

119 

101 

115 

95 

116 

115 

110 

134 

68 

89 

136 

84 

109 

34.8 

54.3 

16.1 

25.9 

41.5 

22.0 

48.9 

18.6 

34.7 

15.5 

39.0 

71.3 

59.9 

33.3 

48.6 

26.8 

52.4 

52.6 

49.0 

93.5 

9.8 

26.8 

95.9 

20.0 

46.1 

69.33 

87.93 

40.25 

58.55 

68.38 

55.97 

82.07 

42.43 

69.45 

58.03 

61.18 

84.78 

75.98 

51.87 

66.85 

60.28 

59.85 

77.05 

100.0 

87.92 

43.57 

60.17 

84.12 

53.87 

65.25 

7.15 

7.99 

4.85 

6.30 

6.51 

6.36 

7.60 

4.82 

7.09 

7.64 

5.27 

7.12 

6.38 

5.14 

5.81 

6.35 

5.16 

6.70 

9.09 

6.56 

6.41 

6.76 

6.19 

6.41 

5.99 

53.40 

67.75 

26.63 

57.08 

15.82 

55.38 

69.17 

33.00 

60.12 

50.73 

63.80 

76.05 

72.75 

55.05 

73.90 

62.75 

59.13 

79.98 

100.0 

89.02 

42.17 

71.47 

85.92 

54.47 

77.92 

5.51 

6.16 

3.21 

6.14 

1.51 

6.29 

6.40 

3.75 

6.13 

6.68 

5.50 

6.39 

6.11 

5.45 

6.42 

6.61 

5.10 

6.95 

9.09 

6.64 

6.20 

8.03 

6.32 

6.48 

7.15 

Mean 

SE 

102.9 

3.40 

41.5 

4.48 

66.90 

5.54 

6.59 

0.36 

62.44 

5.54 

6.13 

0.36 

Control Fish FL (mm) BM (g) cm s-1 BL s-1 cm s-1 BL s-1

1 

2 

3 

4 

5 

6 

112 

120 

96 

109 

112 

139 

58.2 

61.3 

31.1 

42.9 

39.6 

105.7 

82.43 

81.85 

69.22 

73.33 

58.58 

98.08 

7.36 

6.82 

7.21 

6.73 

5.23 

7.06

87.30 

90.04 

77.80 

83.92 

89.48 

68.22 

7.79 

7.53 

8.10 

7.70 

7.99 

4.91 

Mean 

SE

114.7 

5.82

56.5 

0.89

77.25 

8.24

6.73 

0.55

 82.85 

 8.24

7.34 

0.55
53

Individuals in control groups performed at a higher level after their second sham exposure 

(Figure 5-1), showing a mean 8.9 % increase in Ucrit in the second trial. The mean pre-exposure 

critical swimming of the control groups (Table 5-1) was 82.85 ± 8.24 cm s-1 (6.73 ± 0.56 BL s-1) 

and the mean post-exposure Ucrit was 77.25 ± 8.24 cm s-1 (7.34 ± 0.56 BL s-1). Individual 

variation was observed between pre- and post-exposure swimming performance of Atlantic 

spadefish (Figures 5-2 and 5-3). Following exposure 10 of the 25 treatment fish showed a mean 

7.2 % increase in Ucrit while the remaining 15 treatment fish exhibited a mean 17.4 % decrease in 

Ucrit. Five of the six control fish showed a 19.1 % increase in Ucrit, while one control fish 

exhibited a 30.0 % decrease in Ucrit following sham-exposure. 

Figure 5-2. Critical swimming speeds (Ucrit) of juvenile Atlantic spadefish before and after a 

24-hr exposure to 3.0 % ethylene glycol (N = 25) and sham controls (N=6). 

5.2.2 Plasma samples 

GC analysis of plasma samples pooled for each treatment and control group was used to detect 

traces of ethylene glycol after 17 hr of acclimation, up to 100 min experimental swim time and ~ 

0.5 hr of postexperimental processing. Pooled plasma from juvenile Atlantic spadefish treatment 

groups (N = 10 groups) contained a mean (± SE) ethylene glycol concentration of 0.891 ± 0.09 

%. 

0.0

2.0

4.0

6.0

8.0

10.0

12.0

 C o n t r o l T r e a t m e n t 

Ucrit (BL s-1 )

P r e-exposure

P o s t-exposure
54

Figure 5-3: Individual variation in the critical swimming speeds (Ucrit) of juvenile Atlantic 

spadefish (N = 25) before and after exposure to 3.0 % ethylene glycol and sham 

controls (N = 6). 

5.3 Discussion 

Sublethal exposure to ethylene glycol significantly reduced the swimming performance of 

juvenile Atlantic spadefish by 7.3 % compared to their initial performance and by 18.4 % 

compared to the final performance of the controls. In the event of exposure, the inability of 

spadefish to swim at prolonged and burst swimming performance levels could have an effect on 

individuals due to reduced ability to avoid predators and feed effectively. The overall mean 

reduction of 7.3 % in critical swimming speed of Atlantic spadefish exposed to 3.0 % ethylene 

glycol was roughly half the 14.2 % reduction observed for Florida pompano. 

Treatment

1

2

3

4

5

6

7

8

9

10

Pre Post

Ucrit (BL s-1)

Control

1

2

3

4

5

6

7

8

9

10

Pre Post

Ucrit (BL s-1)
55

5.3.1 Ethylene glycol 

The severity of ethylene glycol toxicity varies from species to species and has been shown to be 

five times more toxic to humans than to poultry (Beasley 1980). In primates, ethylene glycol 

causes depression of the nervous system that is similar to ethanol intoxication followed by 

drowsiness, coma, respiratory failure, convulsions, and renal damage (Merck 1983). Hymel et 

al. (2002) found that exposure to as little as 1.0-2.1 % of ethylene glycol causes changes in the 

swimming behavior of Florida pompano after exposure. 

 

The 7.3 % and 18.4 % reductions in Ucrit compared to initial performance and post-treatment 

control means, respectively, may be a result of a buildup of metabolic products. In mammals 

ethylene glycol is oxidized in the liver (Chou and Richardson 1978) and involves several 

metabolic products, with glycolic acid believed to be the metabolic product that is responsible 

for the toxic effects of the chemical. The buildup of metabolic products and increased metabolic 

cost of removing ethylene glycol from the body may have directly caused or contributed to the 

decrease in swimming performance of exposed fish. 

5.3.2 Modes of action 

In addition to the buildup of metabolic products, the increased metabolic cost of removing 

ethylene glycol from the body and reduced gas exchange from gill damage may reduce 

swimming performance. Several theories have been proposed in the literature on how exposure 

to a toxicant, such as ethylene glycol, results in a reduction in swimming performance. A 

decrease in critical swimming speed has been attributed to impaired transport or exchange of 

respiratory gases (Sprague 1971; Satchell 1984). Several factors may hinder the swimming 

performance of fishes: the inability to supply adequate oxygen to the gills; the inability to supply 

oxygen to the tissues; the inability to remove metabolic products; and the inability to provide 

adequate substrate or to activate enzymatic processes (Jones 1971). Wilson and Taylor (1993) 

found acute levels of copper exposure in freshwater rainbow trout caused a disruption of ion 

regulation and progressive hypoxia and plasma acidosis resulting from physical damage to the 

gills. It was unclear whether gill damage was a result of the ethylene glycol, as examination of 

spadefish gill lamelle revealed signs of damage and fusion; however, this could have also been 

caused by the use of MS-222 before sacrifice. 

5.3.3 Individual variation 

Considerable within group variation was observed for individuals in the swimming performance 

tests (Figure 5-2, 5-3). Traditionally in swimming performance studies, it is common practice to 

compare the mean Ucrit of different treatment and control groups, in a one-way ANOVA design, 

but this approach fails to address within-group variation. In the traditional approach, high 

within-group variation makes it difficult to find significant differences between treatment and 

control group means, which can limit the usefulness of swimming performance tests (Kolok 

1998). In this and related studies (Hymel et al. 2002; Baltz in prep) swimming performance of 

individual fish before and after exposure was used in a split-plot design with blocks to take 
56

advantage of within-group variability to analyze changes in Florida pompano swimming 

performance (Kolok 1992). 

Previous studies revealed conflicting results concerning the performance variability among 

individuals after treatment. Kolok (1998) found large within group variation in Ucrit during 

swimming performance tests, but this variation was repeatable over replicate experiments. In 

contrast Kolok and Farrel (1994) found that the reduction in Ucrit of northern pikeminnow 

(Ptychocheilus oregonensis) after a surgical technique was fairly consistent. The results of this 

study resembled those by Kolok (1998) where individual performance in Ucrit after exposure to 

ethylene glycol varied between slight increases and traumatic reductions. The variability in this 

study implies that individual fish may have lower or higher tolerances to ethylene glycol. This 

study used only one concentration of ethylene glycol and it was not possible to calculate 

tolerance values for individuals. 

5.3.4 Spills 

A large localized spill of ethylene glycol poses a risk with large offshore storage volumes up to 

416,400 L per platform (Boehm et al. 2001). Significant spills have been reported in recent 

years, 13,510 L of ethylene glycol leaked from an underwater pipeline into the northern Gulf of 

Mexico on June 29, 1998 (National Response Center 2002 b). Fish would most likely be 

exposed for only a short period of time due to the rapid dilution and biodegradation of ethylene 

glycol in the natural environment. The half-life of ethylene glycol in the water column and 

marine sediments is estimated at 32.50 days. Boehm et al. (2001) modeled a possible spill of 

ethylene glycol (416,400 L) and found the predicted exposure concentrations would be an order 

of magnitude smaller than the lowest 48-hr LC50 (34,400 mg/L or 3.13 %) for the water flea 

(Ceriodaphnia dubia). However, marine fish may be more sensitive to the toxic effects of 

ethylene glycol. If it is assumed that LC50 decreases linearly with time of exposure, then LC50 

values in studies using different test durations can be grossly compared with a toxicity index 

(concentration x time) (French et al. 1996a; Boehm et al. 2001). Using the index of toxicity 

(Index) to estimate 1-hr LC50 found by Hymel et al. (2002) for Florida pompano (Index = 135) 

would be lower than those found by Boehm et al. (2001) for the water flea (Index = 150). 

Sublethal concentrations (Index = 72) used in this experiment would be half as large as values 

found for the water flea suggesting that a spill could affect juvenile Atlantic spadefish to a 

greater extent than predicted. 

5.3.5 Ecological effects 

An episodic spill of ethylene glycol could have ecological implications at individual, 

community, or population levels. The results of this study showed that sublethal exposure to 

small concentrations of ethylene glycol can greatly reduce the swimming performance of an 

individual. Nevertheless, it is difficult to forecast how the effects of exposure at the individual 

level translate to the population or community level. The reduced ability of an individual to feed 

and avoid predators could have effects on a population. Although post-exposure critical 

swimming speeds were well above those used in normal swimming speeds, the ability for 
57

prolonged or burst swimming speeds would be reduced in the event of exposure, which may 

result in the reduced ability to capture prey or avoid predators. However, mortality in fish 

populations can be compensatory; therefore, a spill may not necessarily affect a species at the 

population levels if the exposure is not sustained or widespread (Houde 1989). But a spill during 

peak spawning period could reduce recruitment and add an additional stress to populations 

already burdened by overfishing, pollution and reduced habitat quality. An in-depth model 

including parameters on food resources, avoidance behavior, dispersion patterns of ethylene 

glycol and life histories of marine fishes may be the best way to understand the implications of 

sublethal exposures at the population level. 

5.4 Conclusion 

The results of this study showed conclusively that a single exposure to ethylene glycol 

significantly reduced the swimming performance of juvenile Atlantic spadefish. Although 

Boehm et al. (2001) found that realistic chemical spills of ethylene glycol to be well below the 

lowest LC50 values for other species found in the literature, at least two marine fish, Florida 

pompano and Atlantic spadefish appear to be more sensitive to the toxic effects of the chemical. 

The reduced ability of Atlantic spadefish to swim at high prolonged and burst performance levels 

could have effects on an individual’s ability to avoid predators and feed effectively. Effects at 

the level of the organism may, in turn, lead to effects at the population and community levels 

(Weis et al. 1999). Exposure to ethylene glycol could ultimately affect population and 

community structure in the vicinity of a spill if recurrent or combined with other population level 

effects. 
59

6.0 GENERAL SUMMARY 

D. M. Baltz1, 2 and E. J. Chesney3 

1

Department of Oceanography and Coastal Sciences, Louisiana State University, Baton Rouge, 

LA 70803-7503 

2

Coastal Fisheries Institute, Louisiana State University, Baton Rouge, LA 70803-7503 

3

Louisiana Universities Marine Consortium, Chauvin, LA 70344 

The results of the juvenile Florida pompano study showed that exposure to 1.07 % methanol and 

17 hr recovery did little to impair their swimming performance. Sublethal exposure to methanol 

had only a short-term effect on swimming behavior but little effect on measured swimming 

performance. The lack of detectable levels of methanol in the plasma samples suggests that 

pompano completely metabolized methanol (probably into formaldehyde and formate products) 

during the ~ 20-hr swimming and recovery process. In contrast, the mean percentage (± SE) of 

ethylene glycol still present in pompano plasma samples ~ 20 hours after exposure was 0.623 ± 

0.04 % (v/v). 

The reduction in swimming performance of juvenile Florida pompano (a 13.0 % reduction) 

exposed to ethylene glycol was substantially higher then the reduction in performance of Atlantic 

spadefish (a 6.9 % reduction). The severity of ethylene glycol toxicity varies from species to 

species and has been shown to be more toxic to humans than poultry (Beasley 1980). Spadefish 

may be more efficient at buffering against or removing the metabolite responsible (glycolic acid) 

for the toxic effects of ethylene glycol. 

Pompano exposed to the combination of ethylene glycol and methanol exhibited the most 

profound reduction in swimming performance. Fish that were exposed to the combination of 

ethylene glycol and methanol performed 250.0% worse then fish exposed to the single 

concentration of ethylene glycol. Methanol and ethylene glycol may become more toxic when 

fish are exposed to a combination of both chemicals. Similar joint action or additive interactions 

occur when toxicants have similar modes of action but act independently resulting in the toxicity 

of the mixture amounting to the sum of the toxicities of the individual toxicants present (Bliss 

1939; Sprague 1970; Marking 1985). Interactive action can be either synergistic (more then 

additive) or antagonistic (less then additive). These interactions occur when one toxicant alters 

the toxicity of another toxicant present. The results of the Florida pompano study showed that a 

synergistic interaction could have occurred after juveniles were exposed to the combination of 

ethylene glycol and methanol. This interpretation is similar to that of Greene and Kocan (1997) 

who found the toxicity of ethylene glycol increases in freshwater fish when sublethal levels of a 

chemical used in an agricultural seed-protectant (thiram) is present. The extensive toxic effects 

observed might have been the result of interactions between the various components of the 

formulation, referred to as synergism (Marking 1985). 
60

A discharge or spill often involves more then one chemical component and the interaction 

between two or more chemicals can result in toxicity that is greater than the sum of the toxicities 

of the individual components (Greene and Kocan 1997). Both ethylene glycol and methanol are 

competitors for aldehyde dehydrogenase (ALDH) and may inhibit each other from being 

completely metabolized and increase the toxic effects after exposure (Pietryka et al. 1998). The 

metabolic pathways of ethylene glycol and methanol are in fact so similar that treatment of 

humans for ethylene glycol and methanol poisoning is the administration of ethanol, another 

ALDH competitor (Katarzyna et al. 1998). 

Both fish species we tested exhibited a high amount of individual variation in swimming 

performance after exposure to ethylene glycol or methanol. Kolok (1998) found large within 

group variation in Ucrit during swimming performance tests, but this variation was repeatable 

over replicate experiments. In contrast Kolok and Farrel (1994) found that the reduction in Ucrit

of northern pikeminnow Ptychocheilus oregonensis after a surgical technique was fairly 

consistent. The results of both the Florida pompano and Atlantic spadefish swimming 

performance studies resembled those of Kolok (1998) in that the individual reduction in Ucrit

after exposure to ethylene glycol, methanol, and the combination treatment was substantial but 

repeatable. The variability in these studies implies that individual fish may have lower or higher 

tolerances to ethylene glycol, methanol or combination treatments. 

Future studies should concentrate on accumulating more data on the sublethal and lethal effects 

of ethylene glycol, methanol, and other additives on a variety of species associated with 

deepwater offshore oil and gas platforms (i.e., encrusting organisms, planktonic species, etc.). 

This data could aid in developing a working model to better understand potential impacts on 

predator-prey interactions, chemical avoidance, behavior and feeding activity of fishes after 

exposure. Quantifying the sublethal and lethal effects of ethylene glycol and methanol at all 

levels of the reef ecosystem would be a good first step in understanding the effects of deepwater 

oil development on fisheries in the northern Gulf of Mexico. 

Due to lack of available fish, finding an appropriate LC50 value of an ethylene glycol and 

methanol mixture and testing their potential and additive or synergistic effects on the swimming 

performance of Atlantic spadefish was not possible. LC50 values would allow comparisons of 

the toxic effects of ethylene glycol and methanol on species with different physiologies and 

would be important in understanding the overall toxicity of a substance. Answering the question 

of how the exposure to methanol and the combination of ethylene glycol and methanol would 

affect a species with different physiology would aid in understanding the overall effect of a 

localized spill and begin to extend our understanding from the population to community level. 

Another objective of future studies should be to find LC50 values for several marine fish at risk 

of exposure. 

Only one sublethal concentration was used to evaluate the sublethal effects of ethylene glycol, 

and methanol on Florida pompano and Atlantic spadefish. Although behavior and recovery trials 

by Hymel (2002) and Baltz et al (2005) were used to find sublethal test concentrations; testing a 

range of sublethal concentrations on swimming performance would help in identifying the 

effects of differing exposure ranges. 
61

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The Department of the Interior Mission

As the Nation's principal conservation agency, the Department of the Interior has responsibility 

for most of our nationally owned public lands and natural resources. This includes fostering 

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As a bureau of the Department of the Interior, the Minerals Management Service's (MMS) 

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lands, and distribute those revenues. 

Moreover, in working to meet its responsibilities, the Offshore Minerals Management Program 

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The MMS strives to fulfill its responsibilities through the general guiding principles of: (1) being 

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