Environ. Sci. Technol. XXXX, xxx, 000–000
Oxygen Flux As an Indicator of
Physiological Stress in Fathead
Minnow (Pimephales promelas)
Embryos: A Real-Time Biomonitoring
System of Water Quality
B R I A N C . S A N C H E Z , †,‡
HUGO OCHOA-ACUÑA,§
D . M A R S H A L L P O R T E R F I E L D , * ,‡,| A N D
M A R Í A S . S E P Ú L V E D A * , †
Department of Forestry and Natural Resources, 715 West State
Street, Purdue University, West Lafayette, Indiana 47907, The
Bindley Bioscience Center Physiological Sensing Facility,
Purdue University, West Lafayette, Indiana 47907,
Department of Comparative Pathobiology, Purdue University,
725 Harrison Street, West Lafayette, Indiana 47907, and
Departments of Agricultural & Biological Engineering, and
Horticulture & Landscape Architecture, Purdue University, 225
South University Street, West Lafayette, Indiana 47907
Received November 16, 2007. Revised manuscript received
June 5, 2008. Accepted June 19, 2008.
The detection of harmful chemicals and biological agents in
real time is a critical need for protecting freshwater ecosystems.
We studied the real-time effects of five environmental
contaminants with differing modes of action (atrazine, cadmium
chloride, pentachlorophenol, malathion, and potassium
cyanide) on respiratory oxygen consumption in 2-day postfertilization fathead minnow (Pimephales promelas) eggs. Our
objective was to assess the sensitivity of fathead minnow eggs
using the self-referencing micro-optrode technique to detect
instantaneous changes in oxygen consumption after brief
exposures to low concentrations of contaminants. Oxygen
consumption data indicated that the technique is indeed sensitive
enough to reliably detect physiological alterations induced
by four of the five contaminants. After 2 h of exposure, we
identified significant increases in oxygen consumption upon
exposure to pentachlorophenol (100 and 1000 µg/L), cadmium
chloride (0.0002 and 0.002 µg/L), and atrazine (150 µg/L). In
contrast, we observed a significant decrease in oxygen flux
after exposures to potassium cyanide (44 and 66 µg/L) and atrazine
(1500 µg/L). No effects were detected after exposures to
malathion (200 and 340 µg/L). Our work is the first step in
development of a new technique for physiologically coupled
biomonitoring as a sensitive and reliable tool for the detection
of environmental toxicants.
* Address correspondence to either author. E-mail: porterf@
purdue.edu (D.M.P.), mssepulv@purdue.edu (M.S.S.). Phone: 765494-1190(D.M.P.), 765-496-3428(M.S.S.).
†
Department of Forestry and Natural Resources.
‡
The Bindley Bioscience Center Physiological Sensing Facility.
§
Department of Comparative Pathobiology.
|
Departments of Agricultural & Biological Engineering, and
Horticulture & Landscape Architecture.
10.1021/es702879t CCC: $40.75
Published on Web 08/06/2008
 XXXX American Chemical Society
Introduction
The use of biological early warning systems to continuously
monitor water quality has been investigated since the early
1970s (1). Systems have exploited stress responses to
contaminants in freshwater and saltwater bacteria (2, 3),
cladocerans (4), amphipods (5), freshwater and marine
bivalves (6, 7), and several species of fish (8–10). The
quantification of stress in fish models is often measured in
terms of changes in activity (9), opercular movement (1, 11),
and electric organ discharge (8). One of the more sophisticated systems was developed by Shedd et al. (12) whereby
they recorded irregularity in body movement, cough rate,
and ventilatory rate and depth of bluegill (Lepomis macrochirus) exposed to test water (10, 12). Similar to other
biomonitors, pairs of bulk electrodes in each test chamber
were used to measure changes in impedance patterns, which
were recorded and associated with aberrant behavior due to
degraded water quality. This system has been used to monitor
water quality in the field for a long period (2 y) (13).
Respiratory activity of a fish is often the first physiological
response to be affected by the presence of contaminants in
the aquatic environment (14). Although many biological early
warning systems monitor abnormal opercular movement
(i.e., ventilation depth, rate, and coughing) as an indicator
of respiratory stress, a more direct measurement of stress in
this sense necessitates the quantification of oxygen consumed
by the fish. Although oxygen consumption is not often used
as a bioindicator of pollution-associated stress in biological
early warning systems, it has recently been applied to a system
utilizing the cladoceran, Daphnia magna, with relatively good
success (4).
The paucity of biomonitoring systems based on oxygen
consumption to date may have been due to the unavailability
of technology that would allow measuring these fluxes in
real time. Fiber optic biosensing technology has yielded the
ability to measure minute fluxes of oxygen in real time at the
cellular level (15). By sequentially measuring the concentration of oxygen at different distances from the cell or organism,
this technology allows measurement of oxygen flux (i.e.,
consumption) in real time. Although measurement of oxygen
flux in extracellular microgradients using polarographic
microelectrodes is not new (16–18), the use of optically
transduced sensors (optrodes) for this purpose is rather novel.
Optrodes are superior because their measurements do not
consume oxygen, are not disturbed by external electromagnetic interference, and do not require the use of a reference
electrode (15, 19). Because of these advantages, optrodes
provide an optimal platform for developing sensitive and
reliable biological early warning systems based on measuring
oxygen consumption.
The objective of this study was to evaluate the performance
of a biosensing approach based on measuring the respiratory
responses of fathead minnow (Pimephales promelas) embryos
as quantified by their real-time oxygen consumption (pmol/
cm2/s). Specifically, we wanted to determine if our proposed
technique was sensitive enough to detect instantaneous
changes in oxygen consumption after exposure to low
concentrations of contaminants.
Materials and Methods
Fathead Minnows. We chose to use fathead minnows
because they are a commonly used model organism in
ecotoxicological studies and are relatively easy to breed in
captivity. Fathead minnow embryos are ideal for this
application, because although they are complex organisms
VOL. xxx, NO. xx, XXXX / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
A
FIGURE 1. Diagram of the self-referencing oxygen optrode system. From the test medium, an optrode is coupled to a digital signal
processor (DSP)-based frequency domain lifetime fluorometer (Presens, Regansburg, Germany). The fluorometer excites the
fluorophore and measures its luminescence decay time in the frequency domain as a phase angle (Φ) shift. The output from the
fluorometer is transferred to a DC-coupled lock-in differential amplifier (Applicable Electronics, Inc. Forestdale, MA) that subtracts
the baseline signal (offset potential) from the optrode baseline signal (offset subtraction) prior to amplification. The mV output values
are digitally logged within the Automated Scanning Electrode Technique software (ASET, Science Wares, East Falmouth, MA)
installed on a personal computer. Analog-to-digital (A/D) and digital-to-analog (D/A) converters are incorporated into the system to
allow effective communication between the personal computer and the differential amplifier. The zoom microscope and optrode are
operated with 2- and 3-dimensional (2D and 3D) stepper motors controlled through the software via motor controllers and interfaced
to the personal computer through two parallel ports. A frame grabber linked to a video camera mounted on the zoom microscope
enables the operator to visualize the position of the optrode within the ASET on-screen display.
during testing (2 d postfertilization), they are immobile,
thereby allowing accurate flux measurements to be taken.
There is also considerable evidence showing that early life
stages of fish are the most sensitive to many, perhaps most,
toxic agents (20–23). As such, the system should prove
responsive to low levels of aquatic contamination.
Pairs of breeding fathead minnows were held in static 9.4
L aquaria that underwent 50% water changes three times
per week. Water was maintained at 25 °C and a 16:8 h (light/
dark) photoperiod was imposed. Breeding pairs were fed
frozen brine shrimp (Artemia salina) twice daily until
satiation. Each aquarium was supplied with a semicircle of
polyvinylchloride (PVC) pipe matrix (7.6 cm diameter, 12 cm
length) on which the fish were able to spawn and fertilize
eggs. Aquaria were checked for eggs daily and when present,
the matrix was removed and placed in a separate 9.4 L
aquarium until used for experimental trials.
Oxygen Optrode. The setup of our system is illustrated
in Figure 1. The basic components of the self-referencing
optrode setup include a microscope with a head-stage sensor
amplifier driven by a translational motion control system.
The microscope allows visualization of the position of the
optrode relative to the cell or organism being studied (Figure
2). Although initial positioning of the probe is done by visual
observation, final positioning was done by controlling the
probe position with a computer-driven motion control
system. The microscope, camera, and motion control system
were mounted on an antivibration table within a Faraday
cage. The motion control system allowed the optrode tip to
be moved through the gradient at a known frequency and
between known points (commonly 10-50 µm apart).
The tip of the probe is located on the end of a fiber optic
cable that is approximately 5 µm in diameter. It is coated
with an immobilized, quenchable fluorophore, platinum
tetrakis (pentafluorophenyl) porphyrin (PtTFPP). The probe
B
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. xxx, NO. xx, XXXX
is housed within a tapered pipet held by the motion control
system. The probe is coupled to a digital signal processor
(DSP)-based lifetime fluorometer (Presense, Regansburg,
Germany), which excites the fluorophore, measures its
luminescence decay time and phase angle (Φ), and transfers
the analog signal to a DC-coupled, lock-in amplifier
(10-100×). The amplifier discretely subtracts the baseline
signal (offset potential) from the optrode baseline signal
(offset subtraction) prior to amplification. The mV output
values are digitally logged within the automated scanning
electrode technique software (Science Wares, East Falmouth,
MA) installed on a personal computer. The zoom microscope
and optrode are operated with 2- and 3-dimensional (2D
and 3D) stepper motors controlled through the software.
Calculation of Oxygen Flux. Oxygen flux was calculated
using the Fick Equation:
J ) -D(∆C ⁄ ∆r)
2/s),
(1)
D is the diffusion coefficient of
where J is flux (mol/cm
oxygen in water (0.0000242 cm2/s at 25 °C; ref 24), ∆C is the
difference in oxygen concentration at two points in a
concentration gradient adjacent to the embryo chorion (mol/
mL), and ∆r is the distance between the two points within
the gradient (cm). The two measurement points were (1)
directly adjacent to the chorion, and (2) 20 µm perpendicular
to the chorion (i.e., ∆r ) 0.02 mm). Prior to each trial, the
optrode underwent a two-point calibration at 100% oxygen
saturation in deionized water and 0% oxygen concentration
in nitrogen-bubbled deionized water.
Experimental Trials. Fathead minnow eggs (2 d postfertilization) were removed from the matrix and adhered to
the bottom of a BD Falcon tissue culture dish (35 × 10 mm,
Franklin Lakes, NJ). Each dish had been previously treated
with 350 µL of 0.1 mg/mL poly L-lysine solution, a commonly
used adhesive in cell and tissue culturing, and allowed to dry
FIGURE 2. Image of fathead minnow embryo and oxygen optrode during a monitoring event.
for a minimum of 6 h (25). Prior to use, the dishes were
rinsed 5 times with a phosphate-buffered saline solution (150
mM NaCl, 2.2 mM NaH2PO4, 8.1 mM Na2HPO4) followed by
a single rinse with deionized water. The dish was filled with
4.9 mL of deionized water (pH ) 6.5, free of CaCO3). Eggs
were numbered 1 through 5 and pre-exposure oxygen
consumption measurements were taken from eggs 1 through
4 in sequence for 10 min each to yield their pre-exposure
oxygen consumption measurements. Egg 5 was typically
monitored continuously for 1 h prior to exposure and the
final 10 min of data were used for egg 5 pre-exposure statistical
analysis. Therefore, oxygen consumption of each egg was
always measured when exposed to uncontaminated water
(i.e., served as its own control) before being exposed to a
given chemical. After 1 h, 100 µL of a spike solution was
added to the 4.9 mL of deionized water to yield the target
chemical concentration in 5 mL of solution. Data on egg 5
were typically recorded for another 2 h, and the final 10 min
of data were used for egg 5 postexposure statistical analysis.
Eggs 1 though 4 were subsequently monitored in reverse
order for 10 min each to yield their postexposure flux
measurements.
Experimental Chemicals and Concentrations. We chose
four common environmental contaminants (atrazine; cadmiumchloride,CdCl2;pentachlorophenol,PCP;andmalathion)
that vary in mode of toxic action, in order to assess the utility
of our technique (see Table 1). We also utilized potassium
cyanide (KCN), a known inhibitor of oxygen consumption,
as a negative control (26). Carbonyl cyanide m-chlorophenylhydazone (CCCP, 10 mM, 1023 mg/L) and rotenone (5
µM, 3944 mg/L) were also used in successive exposure events
to demonstrate the system’s ability to detect instantaneous
oxygen consumption changes upon embryo exposure to
chemicals with mitochondrial effects known to both increase
and decrease oxygen consumption, respectively (26). Our
test concentrations were based on those used by other realtime water quality monitoring systems reported in the
literature or on established water quality guidelines for
atrazine, CdCl2, PCP, malathion, and KCN (Table 1). Additional concentrations of KCN (44 µg/L and 66 µg/L) were
included post hoc (see Results).
Statistical Analysis. We used SAS statistical analysis
software (SAS Institute Inc., Cary, NC) for all analyses. We
calculated mean oxygen consumption and standard errors
for each embryo for the pre- and postexposure measurements. We conducted a repeated measures analysis of
variance (ANOVA) to test for significant differences (R ) 0.05)
among oxygen consumption rates measured during 10 min
of pre-exposure and 10 min of postexposure for each embryo
on a chemical by dose basis. This was done because each
embryo served as its own control. Oxygen consumption
values were recorded every 10 s; therefore approximately 60
points per condition were compared.
Results
As can be seen in Figure 3, our system was able to detect
CCCP and rotenone in water in near-real time. Introduction
of 5 µM CCCP in the test media resulted in an almost
immediate increasing trend in oxygen consumption until
adding 10 mM of rotenone resulted in a decrease in oxygen
consumption. A similar time trend was observed with the 66
µg/L KCN test solution (Figure 4A). In this situation, a
decrease in oxygen consumption began soon after exposure.
VOL. xxx, NO. xx, XXXX / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
C
FIGURE 3. Oxygen consumption time course for a single fathead
m-chlorophenylhydrazone (CCCP; 5 µM) and rotenone (10 mM) in succession.
minnow
embryo
exposed
to
carbonyl
cyanide
TABLE 1. Mode of Toxic Action, Concentration, and Basis for Choosing the Concentration of Each Test Chemical Used in the
Present Study
chemical
mode of
toxic action
concentrations
(µg/L)
basis
atrazine
oxidative stress
150, 1500*
*maximum 1 h concentration not to be
exceeded more than once every 3 years for the
protection of aquatic life (73)
cadmium chloride
calcium metabolism
interference
0.0002, 0.002*
*criterion maximum concentration for the
protection of aquatic life (74)
pentachlorophenol
uncoupler of
oxidative
phosphorylation
100, 1000*
*1-day human health advisory level (75)
malathion
acetylcholinesterase
inhibitor
200*, 340**
*1-day human health advisory level (75),
**bluegill LC50 concentration used in van der
Shalie et al. (10)
potassium
cyanide
electron transport
disruption
5.2*, 22**
*criteria continuous concentration and
**criteria maximum concentration for the
protection of aquatic life (76)
FIGURE 4. Pre- and postexposure oxygen consumption for five fathead minnow embryos exposed to 66 µg/L potassium cyanide (A).
Bars represent average 10 min pre- (white bars) and 10 min postexposure (cross-hatched) values (pmol/cm2/s ( 1 standard error) for
each embryo. Postexposure measurements were taken 30 min after commencing exposure. Oxygen consumption time course for a
single fathead minnow embryo (B). Potassium cyanide (66 µg/L) was added (arrow) 60 min after the trial began and measurements
were taken for an additional 30 min.
All eggs demonstrated significantly different oxygen consumption after 30 min (Figure 4A).
Results from the tests comparing oxygen consumption
before and after exposures are presented in Figure 5. No
D
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. xxx, NO. xx, XXXX
embryo mortality was observed within the timeframes
evaluated. Except for malathion and the two lowest doses of
KCN, the system was capable of detecting the test chemicals
within 2 h of exposure. Significant increases in oxygen
FIGURE 5. Average percent change between 10 min pre- and 10 min postexposure oxygen flux values (pmol/cm2/s) for five embryos
exposed to atrazine, cadmium chloride (CdCl2), pentachlorophenol (PCP), malathion, and potassium cyanide (KCN) exposed to
chemical concentrations (µg/L) indicated above each bar. Postexposure measurements were taken approximately 2 h after spiking
for all chemicals and concentrations with the exception of 66 µg/L KCN (30 min). Asterisks indicate a significant overall difference
(repeated measures ANOVA, r ) 0.05) between pre- and postexposure values.
consumption after 2 h of exposure were detected for the low
dose of atrazine (150 µg/L), CdCl2 (0.0002 and 0.002 µg/L),
and PCP (100 and 1000 µg/L). We detected a significant
decrease in oxygen consumption after 2 h of exposure to the
high dose atrazine (1500 µg/L). There was no significant
change in oxygen consumption after 2 h of exposure to
malathion (200 and 340 µg/L) or the lowest doses of KCN
(5.2 and 22 µg/L). We also meaured oxygen consumption
after exposure to 44 and 66 µg/L KCN. Postexposure
measurements were recorded after 2 h for the 44 µg/L
exposure and 30 min for the 66 µg/L exposure. The exposure
time for the 66 µg/L trial was reduced because the real-time
data suggested that a significant decrease had occurred prior
to the typical 2 h exposure period. Decreases in oxygen flux
for the 44 and 66 ug/L doses of KCN were significant. We
detected time by treatment effects for atrazine and KCN
indicating that the oxygen consumption response was
different between the doses of these chemicals.
Discussion
We demonstrated that our proposed system may be significantly more sensitive than other systems now in use. Time
to alarm reported for the system developed by van der Schiale
et al. (10) for some chemicals is rather prolonged. For
example, their system required 12.25 h to detect 250 µg/L
PCP (10). In constrast, we were able to detect a lower
concentration (100 µg/L) of PCP in 2 h. We were also able
to detect a comparable concentration of cyanide (66 vs 60
µg/L) as quickly as 15 min after exposure compared to their
first alarm time of 30 min. Furthermore, in comparison to
several other biomonitor and general stress response studies,
our system has proven more responsive to physiological
alterations due to the presence of contaminants.
Atrazine is a widely used herbicide that blocks photosynthetic electron transport, thereby inhibiting photosynthesis (28). It has been shown to elicit mitochondrial
dysfunction and malformation in fish in previous studies
(29, 30). It is also known to increase the formation of reactive
oxygen species (ROS) in rat heart mitochondria (31) and the
production of the superoxide dismutase enzyme in bluegill
(32). While the decrease in oxygen flux at the high dose
concentration (1500 µg/L) may be a direct indication of
mitochondrial damage to the extent that oxidative phosphorylation is in decline, the disparity in the increase in
oxygen flux at the low dose concentration (150 µg/L) may
indicate a general stress response with a subsequent increase
in energetic allocation to combating the presence of the
chemical. An increase in metabolic rate and oxygen consumption in response to contaminants has been documented
for embryos, larvae, and adults of other vertebrate species
(33, 34). Although Hussein et al. (35) noted a nonquantitative
increase in opercular activity and respiration of adult Nile
tilapia (Oreochromis niloticus) and catfish (Chrysichthyes
auratus) in response to atrazine exposure (3 and 6 mg/L),
another study reported no significant increase in respiration
rates of red drum (Sciaenops ocellatus) larvae after 96 h
exposures to up to 80 µg/L (36).
The inverse response in oxygen flux between the low and
high atrazine concentrations in the present study may also
indicate the presence of a hermetic response. This phenomenon is not uncommon in ecotoxicological research (37),
but would necessitate further experimentation to establish
a causative factor. Despite this disparity, our data clearly
indicate that our technique for detecting oxygen consumption
changes of fathead minnow embryos in response to low
concentrations of atrazine is effective as a monitoring system
of atrazine concentrations in drinking water. Atrazine’s 1-to10 d Health Advisory Level (HAL), the concentrations in
drinking water that should not be exceeded on average for
a 1-10 day period, is 100 ug/L. Our system was able to detect
a similar atrazine concentration after a 2-h exposure. Other
studies have demonstrated lower sensitivity to this chemical.
For instance, Liu et al. (30) were not able to detect significant
differences in several metabolic parameters in grass carp
(Ctenopharygodon idella; i.e., intracellular O2- and H2O2,
cellular ATP depletion, and mitochondrial membrane potential disruption) until 24-36 h postexposure to 37.8 mg/L
atrazine. Our test concentrations were approximately 25 and
250 times lower.
Cadmium is a persistent environmental contaminant that
accumulates in the gills, kidneys, and liver of fish (38). It is
known to impair respiration and mitochondrial activity of
VOL. xxx, NO. xx, XXXX / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
E
fish and mammals. For instance, Yilmaz et al. (39) observed
that guppies (Poecilia reticulata) exposed to 28.5 mg/L Cd
experienced difficulty breathing and attempted to breathe
atmospheric air. Increased gill tissue respiration of winter
flounder (Pseudopleuronectes americanus) was documented
upon exposure to 5 µg/L Cd for 60 d (40) and oxygen
consumption rates of grass carp were elevated after 96 h
exposure to 500 µg/L Cd (41). Conversely, oxygen consumption rates of striped bass (Morone saxatilis) decreased after
30 d exposure to 0.5, 2.5, and 5 µg/L Cd, but were not different
from controls at 90 d or after a 30 d recovery period (42). The
gill respiration capacity of wild yellow perch (Perca flavescens)
was shown to decrease in relation to an increase in liver Cd
concentration (43). Cadmium has also been shown to induce
the expression of genes associated with ROS production and
mitochondrial metabolism in zebrafish (Danio rerio; 44) and
goldfish (45). Achard-Joris et al. (46) observed that the gene
encoding for cytochrome oxidase (complex IV of the electron
transport chain) in marine bivalves was up-regulated in
response to 14.61 µg/L Cd after 10-14 d of exposure. This
was interpreted as a compensatory response to restore a
decrease in mitochondrial activity (46). Cadmium has also
been reported to stimulate ROS production and inhibit the
electron transport chain (47) which Oh and Lim (48) tied to
a collapse in the mitochondrial membrane potential and the
release of cytochrome C into the cytosol. Additionally, these
authors noted rapid and transient ROS production by human
HepG2 cells within 30 min of exposure to 1.79 mg/L Cd. Our
results show that oxygen consumption increased after a 2 h
exposure to considerably lower concentrations of CdCl2
compared to previous studies (0.0002 µg/L). The detection
capability of our system for cadmium suggest that it could
easily detect exceedances of even the rather low ambient
water quality criteria developed by the U.S. EPA for this
element. Assuming a water hardness of 50 mg/L CaCO3, this
criterion equals 0.012 µg/L, which is 60 times higher than the
lowest concentration detected by our system.
Pentachlorophenol is a biocide commonly used as an
antifungal agent for wood preservation (49). It is a strong
uncoupler of oxidative phosphorylation (50), increasing the
rate of mitochondrial oxygen consumption and inhibiting
the production of ATP (27). Juvenile largemouth bass
(Micropterus salmoides) had reduced growth rates and food
conversion efficiencies at or above 25 µg/L (51), and decreased
feeding rates and hyperactivity at 67 µg/L (52). Female
rainbow trout (Oncorhynchus mykiss) produced fewer viable
oocytes upon exposure to 22 µg/L PCP for 18 d (53).
Pentachlorophenol has also been shown to increase oxygen
consumption of mature sockeye salmon (O. mykiss) exposured to 20 µg/L for 12-14 h (54) and of river puffer fish
(Takifugu obscurus) (50 µg/L for 120 h; 55). Our results are
consistent with these findings since a significant oxygen flux
difference was detected after 2 h of exposure to 100 µg/L
PCP. The U.S. EPA has set a Maximum Contaminant Level
(MCL) for PCP in drinking water of 1 µg/L, whereas the 1-d
HAL has been set at 1,000 µg/L. More importantly, our
detection time was 10.25 h faster than that reported for the
bluegill model at 250 µg/L (10).
Our results indicate that no difference in oxygen flux was
experienced by fathead minnow embryos within 2 h of
exposure to either malathion concentration. Supplemental
experimentation did not detect a difference after 8 h (p )
0.96, data not presented). Malathion is an organophosphate
insecticide that inhibits cholinesterase activity by binding to
its esteratic site. This hinders the enzyme’s ability to scavenge
acetylcholine from synaptic clefts and thereby perpetuates
uncontrolled nervous signaling (56). In fish, exposure to
organophosphate insecticides has been shown to inhibit
cholinesterase activity (57) and interfere with behavior (58)
and reproduction (59). The effect of malathion exposure on
F
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. xxx, NO. xx, XXXX
the respiratory activity of fish has been addressed in few
studies and the results are inconclusive. Two studies found
that Mozambique tilapia (Tilapia mossambica) exposed to
122 and 200 µg/L for 48 h experienced an increase in oxygen
consumption through hour 12-24, thereafter undergoing a
decrease in oxygen consumption through hour 48 (60, 61).
Another study found no measurable effect on the respiration
of red drum larvae after 96 h exposure to 1 and 10 µg/L
malathion (62). McKim et al. (63) noted decreases in heart
rate and oxygen utilization with a concomitant increase in
ventilation volume, but no detectable changes in oxygen
consumption of rainbow trout exposed to 296 µg/L malathion
for 24-48 h. The respiration by isolated rat liver mitochondria
exposed to malathion (64 µg/L - 27 mg/L) for 20 min was
likewise not affected (64). These authors concede, however,
that the enzymes required to convert malathion into its
activated oxygen analog (i.e., malaoxon, 65) were not present
in the assay. Nguyen et al. (66) detected no growth inhibition
of larva African catfish (Clarias gariepinus) upon exposure
to 5 mg/L malathion when eggs were in the late blastula
stage (3 h postfertilization) but significant growth inhibition
was observed if larvae were exposed at the 2-4 cell stage
(0-1 h postfertilization) and larval stage (g10 h posthatching).
The authors suggested that the chorion hindered the influx
of malathion after it had undergone water-hardening (assumed to be complete 3 h postfertilization). Our observed
lack of a significant response upon malathion exposure may
be analogous to this suggested effect of chorion hardening.
The bluegill model detected a significant decrease in ventilatory depth after 88.5 h of exposure to 340 µg/L malathion
(10).
Cyanide’s mode of toxic action toward mitochondria is well
documented (26). It acts on cytochrome oxidase (complex IV),
competing with oxygen for the heme a3 site. This effectively
inhibits the final transfer of electrons to oxygen, destroys the
mitochondrial membrane potential, and stalls the production
of ATP (50). Brown bullheads (Ictalurus nebulosus) exposed to
increasing concentrations of cyanide (200-1800 µg/L) over a
9 h period experienced an increase in oxygen consumption
and heart rate through 3 h of exposure, followed by a decrease
in both parameters until death at 9 h (67). Brook trout (Salvelinus
fontinalis) exposed to 5 µg/L cyanide over 29 d had their
swimming performance reduced by 50% (68), while those
exposed to 25 µg/L for 5 h underwent an inhibition in oxygen
intake (69). Oxygen consumption by liver tissues of white-tailed
damselfish (Dascyllus aruanus) was lower for those exposed to
25 and 50 mg/L cyanide for pulses of 10 and 60 s after 2.5 weeks
of unstressed recovery. Conversely, liver tissue of fish exposed
to cyanide in the same manner, but stressed during recovery,
had elevated oxygen consumption rates 2.5 weeks after the
exposure (70). Oxygen consumption by unfertilized chinook
salmon (O. tshawytscha) ceased upon exposure to 132 mg/L
KCN (71). The calculated LC50 (96 h) of cyanide for fathead
minnow eggs at 25 °C is 121-202 µg/L (72). Our results indicate
that significant decreases in oxygen flux by fathead minnows
are detectable within 2 h at concentrations at or above 44 µg/L
and that a dose-response relationship exists. This suggests that
the classic mode of toxic action by cyanide on cytochrome
oxidase may be occurring in fathead minnow embryos. The
bluegill model (10) detected a change in movement within 30
min of exposure to 60 µg/L sodium cyanide. We determined
an effect for a comparable dose in our study (66 µg/L) at 30 min
as well, and a re-examination of the data reveals that a significant
decrease in oxygen flux for egg 5 was detected as early as 15
min after exposure (p < 0.01).
We demonstrated that our proposed system is effective
at detecting the presence of environmentally relevant
concentrations of four of five chemicals tested. We believe
this is because oxygen consumption measurements provide
a robust indicator of whole animal stress and concomitant
water quality. Our findings encourage further development
of this technology and its ultimate use in water quality
monitoring programs and real-time early warning systems.
Acknowledgments
This work was funded by Purdue University Center for the
Environment and a Graduate Assistance in Areas of National
Need (GAANN) fellowship awarded to the primary author by
the U.S. Department of Education. We thank Nathan Barton,
Monica Hensley, Sonia Johns, Geoff Laban, Brett Lowry, Reid
Morehouse, and Bob Rode for helping with the maintenance
of reproductive fathead minnows. We also thank Rameez
Chatni and Aeraj ul Haque for their guidance in operating
the micro optrode and associated equipment.
Literature Cited
(1) Cairns, J., Jr.; Dickson, K. L.; Sparks, R. E.; Waller, W. T. A
preliminary report on rapid biological information for water
pollution control. J. Water Pollut. Control Fed. 1970, 42, 685–
703.
(2) Munkittrick, K. R.; Power, E. A.; Sergy, G. A. The relative sensitivity
of Microtox, Daphnid, rainbow trout, fathead minnow acute
lethality tests. Environ. Toxicol. Water Qual. Int. J. 1991, 6, 35–
62.
(3) Cho, J.; Park, K.; Ihm, H.; Park, J.; Kim, S.; Kang, I.; Lee, K.;
Jahng, D.; Lee, D.; Kim, S. A novel continuous toxicity test system
using a luminously modified freshwater bacterium. Biosens.
Bioelectron. 2004, 20, 338–344.
(4) Martins, J. C.; Saker, M. L.; Teles, L. F. O.; Vasconcelos, V. M.
Oxygen consumption by Daphnia magna Straus as a marker of
chemical stress in the aquatic environment. Environ. Toxicol.
Chem. 2007, 26, 1987–1991.
(5) Maltby, L.; Clayton, S. A.; Wood, R. M.; McLoughlin, N. Evaluation
of the Gammarus pulex in situ feeding assay as a biomonitor
of water quality: robustness, responsiveness, and relevance.
Environ. Toxicol. Chem. 2002, 21, 361–368.
(6) Sluyts, H.; Van Hoof, F.; Cornet, A.; Paulussen, J. A dynamic
new alarm system for use in biological early warning systems.
Environ. Toxicol. Chem. 1996, 15, 1371–1323.
(7) Kramer, K. J.; Jenner, H. A.; de Zwart, D. The valve movement
response of mussels: a tool in biological monitoring. Hydrobiologia 1989, 188/189, 433–443.
(8) Thomas, M.; Florion, A.; Chrétien, D.; Terver, D. Real-time
biomonitoring of water contamination by cyanide based on
analysis of the continuous electric signal emitted by a tropical
fish Apteronotus albifrons. Water Res. 1996, 30, 3083–3091.
(9) Gerhardt, A. Whole effluent toxicity with Oncorhynchus mykiss
(Walbaum 1792): survival and behavioral responses to a dilution
series of a mining effluent in South Africa. Arch. Environ.
Contam. Toxicol. 1998, 35, 309–316.
(10) van der Shalie, W. H.; Shedd, T. R.; Widder, M. W.; Brenan, L. M.
Response characteristics of an aquatic biomonitor used for rapid
toxicity detection. J. Appl Toxicol. 2004, 24, 387–394.
(11) Cairns, J., Jr.; van der Shalie, W. H. Biological monitoring part
I: early warning systems. Water Res. 1980, 14, 1179–1196.
(12) Shedd, T. R.; Widder, M. W.; Leach, J. D.; van der Schalie, W. H.;
Bishoff, R. C. Apparatus and method for automated biomonitoring of water quality. US Patent 6,393,899 B1, 28, 2000..
(13) Shedd, T. R.; van der Shalie, W. H.; Widder, M. W.; Burton, D. T.;
Burrows, E. P. Long-term operation of an automated fish
biomonitoring system for continuous effluent acute toxicity
surveillance. Bull. Environ. Contam. Toxicol. 2001, 66, 392–399.
(14) Heath, A. G. Water Pollution and Fish Physiology, 2nd ed.; Lewis
Publishers: New York, 1995.
(15) Porterfield, D. M.; Rickus, J. L.; Kopelman, R. Non-invasive
approaches to measuring respiratory patterns using a PtTFPP
based, phase-lifetime, self-referencing oxygen optrode. Proc.
SPIE 2006, 6380, 63800S1–63800S8.
(16) Land, S. C.; Porterfield, D. M.; Sanger, R. H.; Smith, P. J. S. The
self-referencing oxygen-selective microelectrode: detection of
trans-membrane oxygen flux from single cells. J. Exp. Biol. 1999,
202, 211–218.
(17) Porterfiled, D. M.; Smith, P. J. S. Characterization of trans-cellular
oxygen flux and proton fluxes from Spirogyra grevelleana using
self-referencing microelectrodes. Protoplasma 2000, 212, 80–88.
(18) Porterfield, D. M. Measuring metabolism and biophysical flux
in the tissue, cellular and sub-cellular domains: Recent developments in self-referencing amperometry for physiological
sensing. Biosens. Bioelectron. 2007, 22, 1186–1196.
(19) Klimant, I.; Meyer, V.; Kühl, M. Fiber-optic oxygen microsensors,
a new tool in aquatic biology. Limnol. Oceanogr. 1995, 40, 1159–
1165.
(20) McKim, J. M. Evaluation of tests with early life stages of fish for
predicting long-term toxicity. J. Fish Res. Board Can. 1977, 34,
1148–1154.
(21) Luckenbach, T.; Kilian, M.; Triebskorn, R.; Oberemm, A. Fish
early life stage tests as a tool to assess embryotoxic potentials
in small streams. J. Aquat. Ecosyst. Stress Recov. 2001, 8, 355–
370.
(22) Eaton, J. G.; McKim, J. M.; Holcombe, G. W. Metal toxicity to
embryos and larvae of seven freshwater fish species. I. Cadmium.
Bull. Environ. Contam. Toxicol. 1978, 19, 95–103.
(23) Kristensen, P. Sensitivity of embryos and larvae in relation to
other stages in the life cycle of fish: a literature review. In
Sublethal and Chronic Effects of Pollutants on Freshwater Fish;
Müller, R., Lloyd, R., Eds.; FAO: Oxford, UK, 1995; pp155-166..
(24) Fluid properties, Section 6. In CRC Handbook of Chemistry and
Physics, 87th ed.; Lide, D. R., Ed.; CRC Press: Boca Raton, FL,
2006.
(25) Banker, G.; Goslin, K. Culturing Nerve Cells, 2nd ed.; MIT Press:
Cambridge, MA, 1998.
(26) Albaum, H. G.; Tepperman, J.; Bodansky, O. The in vivo
inactivation by cyanide of brain cytochrome oxidase and its
effect on glycolysis and on the high energy phosphorous
compounds in brain. J. Biol. Chem. 1946, 164, 45–51.
(27) Tzagoloff, A. Mitochondria. In Cellular Oganelles; Siekevitz, P.,
Ed.; Plenum Press: New York, 1982.
(28) Duke, S. O. Overview of herbicide mechanisms of action.
Environ. Health Perspect. 1990, 87, 263–271.
(29) Biagianti-Risbourg, S.; Bastide, J. Hepatic perturbations induced
by a herbicide (atrazine) in juvenile grey mullet Liza ramada
(Mugilidae, Teleostei): An ultrastructural study. Aquat. Toxicol.
1995, 31, 217–229.
(30) Liu, X.; Shao, J.; Xiang, L.; Chen, X. Cytotoxic effects and apoptosis
induction of atrazine in a grass carp (Ctenopharyngodon idellus)
cell line. Environ. Toxicol. 2006, 21, 80–89.
(31) Stolze, K.; Nohl, H. Effect of xenobiotics on the respiratory activity
of rat heart mitochondria and the concomitant formation of
superoxide radicals. Environ. Toxicol. Chem. 1994, 13, 499–502.
(32) Elia, A. C.; Waller, W. T.; Norton, S. J. Biochemical responses
of bluegill sunfish (Lepomis macrochirus, Rafinesque) to atrazine
induced oxidative stress. Bull. Environ. Contam. Toxicol. 2002,
68, 809–816.
(33) Rowe, C. L.; McKinney, O. M.; Nagle, R. D.; Congdon, J. D.
Elevated maintenance costs in an anuran (Rana catesbeiana)
exposed to a mixture of trace elements during the embryonic
and early larval periods. Physiol. Zool. 1998, 71, 27–35.
(34) Hopkins, W. A.; Rowe, C. L.; Congdon, J. D. Elevated trace
element concentrations and standard metabolic rate in banded
water snakes (Nerodia fasciata) exposed to coal combustion
wastes. Environ. Toxicol. Chem. 1999, 18, 1258–1263.
(35) Hussein, S. Y.; El-Nasser, M. A.; Ahmed, S. M. Comparative
studies on the effects of herbicide atrazine on freshwater fish
Oreochromis niloticus and Chrysichthyes auratus at Assiut, Egypt.
Environ. Contam. Toxicol. 1996, 57, 503–510.
(36) Alvarez, M. C.; Fuiman, L. A. Environmental levels of atrazine
and its degradation products impair survival skills and growth
of red drum larvae. Aquat. Toxicol. 2005, 74, 229–241.
(37) Davis, J. M.; Svendsgaard, D. J. U-shaped curves: their occurrence
and implications for risk assessment. J. Toxicol. Environ. Health
1990, 30, 71–83.
(38) Kay, J.; Thomas, D. G.; Brown, M. W.; Cryer, A.; Shurben, D.;
Solbe, J. F. G.; Garvey, J. S. Cadmium accumulation and protein
binding patterns in tissues of the rainbow trout Salmo gairdneri.
Environ. Health Perspect. 1986, 65, 133–139.
(39) Yilmaz, M.; Gül, A.; Karaköse, E. Investigation of acute toxicity
and the effect of cadmium chloride (CdCl2 · H2O) metal salt on
behavior of the guppy (Poecilia reticulata). Chemosphere 2004,
56, 375–380.
(40) Calabrese, A.; Thurberg, F. P.; Dawson, M. A. ; Wensloff, D. R.
Sublethal physiological stress induced by cadmium and mercury
in the winter flounder, Pseudopleuronectes americanus. In
Sublethal Effects of Toxic Chemicals on Aquatic Animals;
Koeman, J. H., Strik, J. J. T. W. A., Eds.; Elsevier Publishing
Company: Amsterdam, 1975; Vol. 1, pp 5-21..
(41) Espina, S.; Saliban, A.; Dı́az, F. Influence of cadmium on the
respiratory function of the grass carp (Ctenopharygodon idella).
Water Soil Air Pollut. 2000, 119, 1–10.
(42) Dawson, M. A.; Gould, E.; Thurber, F. P.; Calabrese, A.
Physiological response of juvenile striped bass Morone saxatilis,
VOL. xxx, NO. xx, XXXX / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
G
(43)
(44)
(45)
(46)
(47)
(48)
(49)
(50)
(51)
(52)
(53)
(54)
(55)
(56)
(57)
(58)
H
9
to low levels of cadmium and mercury. Chesapeake Sci. 1977,
18, 353–359.
Couture, P.; Kumar, P. R. Impairment of metabolic capacities
in copper and cadmium contaminated wild yellow perch (Perca
flavescens). Aquat. Toxicol. 2003, 64, 107–120.
Gonzales, P.; Baudrimont, M.; Boudou, A.; Bourdineaud, J. P.
Comparative effects of direct cadmium contamination on gene
expression in gills, liver, skeletal muscles and brain of the
zebrafish (Danio rerio). BioMetals 2006, 19, 225–235.
Gargiulo, G.; Arcamone, N.; de Girolamo, P.; Andreozzi, G.;
Antonucci, R.; Esposito, V.; Ferrara, L.; Battaglini, P. Histochemical study of the effects of cadmium uptake on oxidative
enzymes of intermediary metabolism in kidney of goldfish
(Carassius aruatus). Comp. Biochem. Physiol. 1996, 113C, 177–
183.
Achard-Joris, M.; Gonzales, P.; Marie, V.; Baudrimont, M.;
Bourdineaud, J. P. Cytochrome c oxydase subunit I gene is upregulated by cadmium in freshwater and marine bivalves.
BioMetals 2006, 19, 237–244.
Wang, Y.; Fang, J.; Leonard, S. S.; Rao, K. M. K. Cadmium inhibits
the electron transfer chain and induces reactive oxygen species.
Free Radical Biol. Med. 2004, 36, 1434–1443.
Oh, S. H.; Lim, S. C. A rapid and transient ROS generation by
cadmium triggers apoptosis via caspase-dependent pathway
in HepG2 cells and this is inhibited through N-acetylcysteinemediated catalase upregulation. Toxicol. Appl. Pharmacol. 2006,
212, 212–223.
Ruckdeschel, G.; Renner, G. Effects of pentachlorophenol and
some of its known and possible metabolites on fungi. Appl.
Environ. Microbiol. 1986, 51, 1370–1372.
Weinbach, E. C. The effect of pentachlorophenol on oxidative
phosphorylation. J. Biol. Chem. 1954, 210, 545–550.
Johansen, P. H.; Mathers, R. A.; Brown, J. A. Effect of exposure
to several pentachlorophenol concentrations on growth of
young-of-year largemouth bass, Micropterus salmoides, with
comparisons to other indicators of toxicity. Bull. Environ.
Contam. Toxicol. 1987, 39, 379–384.
Brown, J. A.; Johansen, P. H.; Colgan, P. W.; Mathers, R. A.
Impairment of early feeding behavior of largemouth bass by
pentachlorophenol exposure: a preliminary assessment. Trans.
Am. Fish. Soc. 1987, 116, 71–78.
Nagler, J. J.; Aysola, P.; Ruby, S. M. Effect of sublethal pentachlorophenol on early oogenesis in maturing female rainbow
trout (Salmo gairdneri). Arch. Environ. Contam. Toxicol. 1986,
15, 549–555.
Farrell, A.; Gamperl, A. K.; Birtwell, I. K. Prolonged swimming,
recovery and repeat swimming performance of mature sockeye
salmon Oncorhynchus nerka exposed to moderate hypoxia and
pentachlorophenol. J. Exp. Biol. 1998, 201, 2183–2193.
Kim, W. S.; Jeon, J. K.; Lee, S. H.; Huh, H. T. Effects of
pentachlorophenol (PCP) on the oxygen consumption rate of
the river puffer fish Takifugu obscurus. Mar. Ecol.: Prog. Ser.
1996, 143, 9–14.
Marrs, T. C. Organophosphate poisoning. Parmacol. Ther. 1993,
53, 51–66.
Beauvais, S. L.; Jones, S. B.; Brewer, S. K.; Little, E. E. Physiological
measures of neurotoxicity of diazinon and malathion to larval
rainbow trout (Oncorhynchus mykiss) and their correlation with
behavioral measures. Environ. Toxiocol. Chem. 2000, 19, 1875–
1880.
Pavlov, D. D.; Chuiko, G. M.; Gerasimov, Y. V.; Tonkopiy, V. D.
Feeding behavior and brain acetylcholinesterase activity in
bream (Abramis brama L.) as affected by DDVP, an organo-
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. xxx, NO. xx, XXXX
(59)
(60)
(61)
(62)
(63)
(64)
(65)
(66)
(67)
(68)
(69)
(70)
(71)
(72)
(73)
(74)
(75)
(76)
phosphorous insecticide. Comp. Biochem. Physiol. 1992, 103C,
563–568.
Bhattacharga, S. Target and non-target effects of anticholinesterase pesticides in fish. Sci. Total Environ. 1993, 859–866.
Sahib, I. K. A.; Rao, K. S. J.; Rao, K. V. R. Effect of malathion
exposure on some physical parameters of whole body and on
tissue cations of teleost Tilapia mossmbica (Peters). J. Biosci.
1981, 3, 17–21.
Basha, S. M.; Rao, K. S. P.; Rao, K. R. S. S.; Rao, K. V. R. Respiratory
potentials of the fish (Tilapia mossambica) under malathion,
carbaryl and lindane intoxication. Bull. Environ. Contam.
Toxicol. 1984, 32, 570–574.
Alvarez, M. C.; Fuiman, L. A. Ecological performance of red
drum (Sciaenops ocellatus) larvae exposed to environmental
levels of the insecticide malathion. Environ. Toxicol. Chem. 2006,
25, 1426–1432.
McKim, J. M.; Schmieder, P. K.; Niemi, G. J.; Carlson, R. W.;
Henry, T. R. Use of respiratory-cardiovascular responses of
rainbow trout (Salmo gairdneri) in identifying acute toxicity
syndromes in fish: Part 2, malathion, carbaryl, acrolein, and
benzaldehyde. Environ. Toxicol. Chem. 1987, 6, 313–328.
Haubenstricker, M. E.; Holodnick, S. E.; Mancy, K. H.; Brabec,
M. J. Rapid toxicity testing based on mitochondrial respiratory
activity. Bull. Environ. Contam. Toxicol. 1990, 44, 675–680.
Casarett, L. J.; Doull, J. Toxicology, the Basic Science of Poisons,
5th ed.; Macmillan: New York, 1996.
Nguyen, L. T. H.; Janssen, C. R.; Volckaert, A. M. Susceptibility
of embryonic and larval African catfish (Clarias gariepinus) to
toxicants. Bull. Environ. Contam. Toxicol. 1999, 62, 230–237.
Sawyer, P. L.; Heath, A. G. Cardiac, ventilatory and metabolic
responses of two ecologically dissimilar species of fish to
waterborne cyanide. Fish Physiol. Biochem. 1988, 4, 203–219.
LeDuc, G. Cyanides in water: toxicological significance. In
Aquatic Toxicology, Vol. 2; Weber, L. J., Ed.; Raven Press: New
York, 1984; Vol. 15; pp 3-224..
U.S. Environmental Protection Agency. Ambient Water Quality
Criteria for Cyanides; Report 440/5-80-037; Washington, DC,
1980.
Hanawa, M.; Harris, L.; Graham, M.; Farrell, A. P.; BendallYoung, L. I. Effects of cyanide exposure on Dascyllus aruanus,
a tropical marine fish species: lethality, anaesthesia and
physiological effects. Aquar. Sci. Cons. 1998, 2, 21–34.
Holcomb, M.; Cloud, J. G.; Woolsey, J.; Ingermann, R. L. Oxygen
consumption in unfertilized salmonid eggs: an indicator of egg
quality. Comp. Biochem. Physiol. 2004, 138A, 349–354.
Smith, L. L., Jr.; Broderius, S. J.; Oseid, D. M.; Kimball, G. L.;
Koenst, W. M.; Lind, D. T. Acute and Chronic Toxicity of HCN
to Fish and Invertebrates; Report 600/3-79-009; U.S. Environmental Protection Agency: Washington, DC, 1979.
U.S. EPA. Ambient Aquatic Life Water Quality Criteria for Atrazine
- Revised Draft; 822-R-03-023; U.S. EPA Office of Water:
Washington, DC, 2003.
U.S. EPA. Update of Ambient Water Quality Criteria for Cadmium; 822-R-01-001; U.S. EPA Office of Water: Washington,
DC, 2001.
U.S. EPA. Drinking Water Standards and Health Advisories; 822R-06-013; U.S. EPA Office of Water: Washington, DC, 2006.
U.S. EPA. National Recommended Water Quality Criteria; U.S.
EPA Office of Water: Washington, DC, 2006.
ES702879T