Keywords: fish, organophosphate, chlorpyrifos, cholinesterase inhibition, mudskipper, .... are highly terrestrial with a number of specializations for amphibious.
Cholinesterase inhibition in mudskippers (Periophthalmus novaeguineaensis) by chlorpyrifos
Minor Project by Florita Flores Degree enrolled in Masters of Applied Science School of Marine and Tropical Biology, James Cook University, Townsville, QLD 4811 Supervisors: Andrew Negri (AIMS) and Mark McCormick (JCU) Submitted February 2007
Word Count: 6927
Abstract:
Sublethal effects of cholinesterase (ChE) activity of mudskippers
(Periophthalmus novaeguineaensis) were evaluated in a static, renewable system using organophosphate insecticide, chlorpyrifos. A significant negative relationship between body ChE activity and total length was found (r2 = 0.353; p < 0.0001), with larger fish exhibiting lower ChE activities. Due to the effect of fish size on ChE activity, fish of similar size were used for all subsequent experiments. After 96 h exposures, mortality was highest at nominal chlorpyrifos concentrations of 128 µg L-1 and 256 µg L-1. The inhibition of ChE was dose dependent and ChE activity was greatly reduced after 96 h of exposure of chlorpyrifos at sublethal concentrations.
The lowest concentration that
significantly inhibited ChE activity after 96 h was 4 µg L-1 and concentrations higher than 100 µg L-1 caused more than 75% inhibition. The IC50 at 96 h was approximately 12 ± 16.032 µg L-1. Rate of ChE inhibition was considerably greater than the rate of recovery and the rate of recovery was dependent on the concentration of the initial exposure and time; exposure to a higher concentration required a longer recovery time. Full recovery for P. novaeguineaensis was relatively quick. Fish exposed at 64 µg L-1 required one week in clean water for ChE activity levels to return to control ChE activity levels. ChE inhibition in body tissue serves as a valuable biomarker to assess the sublethal effects of chlorpyrifos on the tropical fish P. novaeguineaensis. Keywords: fish, organophosphate, chlorpyrifos, cholinesterase inhibition, mudskipper, Periophthalmus novaeguineaensis, insecticide, toxicity
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Introduction Like many countries, Australia has a growing population, with much of this growth occurring along the coastal margins, particularly along the coasts of Victoria, New South Wales, south-west Western Australia and Queensland (DEH 2001). Queensland’s rising population and the associated increase in its agricultural sector has resulted in substantial increases in the use of agrochemicals (Lacher & Goldstein 1997). Increasing loads of sediments, fertilizers and pesticides may be entering waterways via agricultural runoff, posing a potential threat to several aquatic ecosystems, including the Great Barrier Reef (GBR) (Fabricius 2005).
Many pesticides also pose incidental threats to non-target organisms (Odenkirchen & Eisler 1988, Bretaud et al. 2000, Galloway & Handy 2003), which may lead to detrimental effects on the entire ecosystem. Historically, organochlorines were the most heavily used insecticide in Australia, but were banned from use by the 1990s due to concerns of their high persistence and toxicity in the environment (Radcliffe 2002). Since then the organophosphates have become the most widely used insecticide class in Australian agriculture (Radcliffe 2002).
Although organophosphates are effective
insecticides and are less persistent in the environment, they lack target specificity and have been associated with high acute toxicity to a multitude of non-target organisms (Fulton & Key 2001, Galloway & Handy 2003). Organophosphates act by inhibiting acetylcholinesterase (AChE), the enzyme responsible for breaking down acetylcholine (ACh), which is the major chemical transmitter between neurons in all animals (Emden & Service 2004). When this enzyme is inactivated, ACh accumulates at the nerve synapse leading to continuous stimulation of nerves and muscles. This produces rapid twitching of voluntary and involuntary muscles, followed by paralysis and eventually death. The reactivation of AChE after inhibition by organophosphates is so slow that it is considered irreversible; therefore, recovery can only occur with new enzyme synthesis (Bocquene & Galgani 1998, Connell et al. 1999, Fulton & Key 2001).
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The most important organophosphate insecticide currently used in Australian agriculture is
chlorpyrifos.
Chlorpyrifos
[O,O-diethyl
O-(3,5,6-trichloro-2-pyridyl)
phosphorothioate] is a prominent broad-spectrum insecticide heavily used in Queensland for over 30 years (NRA 2000). It is the most heavily used insecticide in the sugarcanegrowing areas with an annual estimated usage of 74 tonnes of active ingredient (Hamilton & Haydon 1996). The cane growing areas of Queensland lie within 26 major drainage basins and river catchments of the Great Barrier Reef World Heritage Area (Humphrey & Klumpp 2003). Catchment-sourced pollutants, including sediments and their associated pesticides, are transported from agricultural areas to nearby marine environments (Haynes & Michalek-Wagner 2000). Due to agricultural runoff, biota, including fish, are at risk from exposure to chlorpyrifos.
Typical symptoms of organophosphate
intoxication in fish are decreased swimming ability, tremors, and paralysis (Odenkirchen & Eisler 1988).
In recent years, biomarkers have become an increasingly important tool in assessing the contaminant exposure on resident biota. Biomarkers can be defined as quantitative measurements reflecting an interaction between a biological system and a potential hazard, which may be chemical, physical or biological (van der Oost et al. 2003). Biomarkers are often cellular, biochemical, molecular or physiological changes which can be related to exposure and/or effects of environmental chemicals (Lam & Gray 2003, van der Oost et al. 2003). They are useful prognostic and diagnostic tools to assess the presence of toxicants in the environment with potential effects on exposed individuals. They also help to indicate sublethal effects of contaminant exposure and provide an integrated response to exposure, including water quality and fitness of fish, via rates of growth, feeding, reproduction and survivorship.
Cholinesterase inhibition has been widely used as a biomarker to detect exposure to and effects of organophosphate poisoning (Bocquene & Galgani 1998, Chambers et al. 2002, Sarkar et al. 2006, Walker 2006). Though chlorpyrifos may degrade rapidly in the environment, inhibition of ChE persists longer, from days to weeks (Fulton & Key 2001). Use of this biomarker needs to be conducted on local species to accurately assess the
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effects of contaminants on these populations and their local habitats. Although several relevant toxicity studies have been conducted in Australia, environmental managers still rely heavily on international toxicity data for water quality and risk assessment analyses (Warne et al. 1998). Few studies have been conducted on pesticide use and fate in tropical aquatic environments (Peters et al. 1997), and even fewer are relevant to the aquatic habitats of tropical Queensland catchment areas (Hamilton & Haydon 1996).
Mudskippers are small fish, abundant in tropical estuarine habitats and spend much of their life in burrows within mudflats and mangrove sediments. Organic contaminants have been found in coastal waters of northern Queensland (Haynes et al. 2000) suggesting that mudskippers may be exposed to these contaminants. Few toxicity studies have been conducted on mudskippers with most work conducted on Boleophthalmus dussumieri (PAN 2006). In Australia however, work conducted on mudskippers has focused mainly on their behavior and distribution (Nursall 1981, Lee et al. 2005). Periophthalmus novaeguineaensis was selected for its potential as a regional bioindicator of chemical contamination due to its widespread occurrence and abundance locally. The main objective of this study was to investigate the effects of chlorpyrifos on ChE activity in a locally abundant tropical estuarine species, P. novaeguineaensis.
Materials and Methods
Study species There are 36 known species of mudskippers (Family: Gobiidae; Subfamily: Oxudercinae) (Murdy 1989, Zhang et al. 2003, Swanson & Gibb 2004). Mudskippers are euryhaline, inhabiting soft bottom intertidal areas and mangrove swamps of the Indo-west Pacific (Murdy 1989). They are economically important in China, Taiwan and India where they are cultured for human consumption (Clayton 1993). In Malaysia, the raw flesh of mudskippers is considered an aphrodisiac. However, mudskippers are mainly a trade aquarium fish.
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Figure 1. Study species: Periophthalmus novaeguineaensis.
Periophthalmus novaeguineaensis (Fig. 1) have been found in southern Irian Jaya, Indonesia, and from Port Hedland to Townsville, Australia (Fig. 2) (Murdy 1989). P. novaeguineaensis are highly terrestrial with a number of specializations for amphibious life, including aerial vision, aerial respiration, and terrestrial locomotion (Zhang et al. 2003).
Figure 2. Known localities (∆) of Periophthalmus novaeguineaensis in Australia. Map of Australia courtesy of Google Earth with localities adapted from Murdy (1989).
Study site Wild mudskippers, Periophthalmus novaeguineaensis, were collected using handnets at Bald Rock Creek and Salmon Creek, near Townsville, Queensland, Australia (19°15′S, 146°50′E; Fig. 3). This catchment is essentially free from agricultural activity and nearby mangroves are not sprayed for mosquito control.
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Figure 3. Map of site (∆; 19°17′S, 147°02′E) where mudskippers were collected. Mudskippers were collected in creeks near Townsville, Australia. Maps courtesy of Google Earth with modification.
Fish maintenance Mudskippers were maintained in 120 L flow-through aquaria under ambient conditions (water temperature (27.2 ± 0.7°C), salinity (38 ppt), photoperiod (12 h light:dark) and acclimatized to laboratory conditions for at least 1 week prior to experiments. Holding tanks were placed on a 10° incline to provide a dry feeding area for the mudskippers. Fish were fed bloodworms once daily to satiation and debris and feces were siphoned from the tanks every 48 h.
Experiment 1: 96 h range-finding A range-finding test was carried out to determine the concentration range of chlorpyrifos to be used in the definitive sublethal exposure test. Four replicate fish were used for each treatment. Fish between 35-74 mm total length (TL) were placed in 4.8 L glass aquaria with lids. Seawater was filtered to 1 µm and dissolved oxygen and temperature were measured daily (YSI model 55, Yellow Springs, Co.) and salinity was measured with a refractometer. Gentle aeration was provided to ensure adequate dissolved oxygen levels. Seven treatments were examined, including an experimental control (no contaminant), solvent control and five nominal treatment concentrations: 0.1 µg L-1, 1 µg L-1, 10 µg L1
, 100 µg L-1, and 500 µg L-1. A stock solution was prepared by dissolving chlorpyrifos -6-
in analytical grade acetone.
Technical grade chlorpyrifos [O,O-diethyl O-(3,5,6-
trichloro-2-pyridyl) phosphorothioate] of 99.5% purity was provided by DowElanco Ltd., (Australia). A 1.5 mg ml-1 solution of chlorpyrifos in acetone was prepared and used as stock solution for concentrations: 10 to 500 µg L-1. The stock solution was then diluted 10-fold in acetone to obtain a 10-1 stock solution to be used for the lower concentration treatments. Fish were not fed for the duration of the experiment. Aliquots of the solvent were added to the treatment chambers, such that the solvent volumes were equal to the volume found in the highest pesticide concentration treatment. Chlorpyrifos stock was pipetted into the treatment chambers and gently stirred with a glass pipette and water changes (90%) were performed every 24 h. Mortality was recorded daily and dead individuals were immediately removed. The exposures were terminated at 96 h, fish were killed in an ice slurry and then kept frozen at -80°C before enzymatic assay.
Experiment 2: Fish size-ChE activity relationship In order to determine whether a relationship exists between size of fish and ChE activity, 63 mudskippers of varying lengths (TL: 28-94 mm) were analyzed. Fish were sacrificed and stored at -80°C before enzymatic assay. Fish were measured with vernier calipers to the nearest 0.1 mm. All fish lengths provided are total lengths (TL).
Experiment 3: 96 h exposure concentration-ChE activity relationship Fish Since a correlation was found between ChE activity in body tissues and total length in P. novaeguineanensis, fish selected for this experiment were restricted to a narrow length range (n = 120; TL = 43-73 mm). Fish were maintained on a 12:12 h light:dark regime. Water quality parameters were taken as per Experiment 1. At the end of the acute toxicity test, remaining fish were sacrificed in an ice slurry and stored at -80°C until enzymatic assay. Exposure procedures were followed as per Experiment 1. Based on the results of the range finding experiment, the nominal chlorpyrifos concentrations used were 1 µg L-1, 4 µg L-1, 16 µg L-1, 64 µg L-1, 128 µg L-1, and 256 µg L-1. A control test chamber without toxicant and a solvent control were also included in the experiment. Treatments were
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performed in triplicate (n = 15 fish/treatment) and test chambers were arranged randomly (Fig. 4). Five fish were randomly removed from the stock aquaria and placed in each test chamber. The exposures were terminated at 96 h when fish were sacrificed and kept frozen at -80°C until enzymatic assay.
Figure 4. Treatment tanks utilized for the acute toxicity testing of chlorpyrifos on P. novaeguineaensis (n = 120 fish).
Experiment 4: 96 h exposure and recovery kinetics The exposure-recovery experiment followed similar methods employed in the exposure procedures described in Experiment 1. Fish (67.4 ± 8.4 mm TL; n = 144) were placed in 50 L glass aquaria with lids (Fig. 5a). Nominal chlorpyrifos concentrations used were 4 µg L-1 and 64 µg L-1. These concentrations were selected to investigate whether recovery was a function of concentration, thus, a high and a low concentration were used. 64 µg L-1 was used as the high concentration since previous experiments showed that low mortality was associated with that concentration. A control test chamber without toxicant and a solvent control were also included in the experiment. Treatments were performed in randomly distributed duplicate tanks each containing 18 fish.
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(a) (b) Figure 5 (a). Acute toxicity test chambers for P. novaeguineaensis (n = 144 fish) (b). Holding tanks for P. novaeguineaensis used for the recovery period post chlorpyrifos exposure.
Following 96 h exposure to chlorpyrifos remaining fish were placed into corresponding plastic 60 L tanks (Fig. 5b) with flow-through water where they would remain for the duration of the recovery. Fish in each tank were fed approximately 3 g bloodworms every second day during recovery. Food was reduced to reflect the number of remaining fish after sampling. ChE activities were measured shortly after sampling to determine whether activity reached levels prior to exposure.
Sampling began 8 hours after
exposure. Four fish from each treatment were randomly sampled at 8 h, 24 h, and 96 h post-exposure. During the recovery period, three to four fish from each treatment were randomly sampled at days 2, 7, 17, 27, 41, and 55. The study was terminated after 55 days. Fish were stored at -80°C for later enzymatic assay.
Cholinesterase activity determination Cholinesterase inhibition assay The cholinesterase (ChE) assay was performed following the protocol of Ellman (1961) with slight modifications for use with an automated microplate reader (Bio-Tek Instruments, Inc.) Measurements were made in quadruplicate and all samples were processed in random order. Twenty fish were sacrificed to optimize tissue concentration for the cholinesterase assay.
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Acetylcholinesterase (AChE, E.C.3.1.1.7) and non-specific butyrylcholinesterase (BChE, E.C.3.1.1.8) are the two forms of cholinesterases likely to be present in marine fish muscle (Sturm et al. 2000). Tissue homogenates were not treated with enzyme-specific ChE inhibitors, therefore, the term ‘cholinesterase’ (ChE) was used throughout this thesis.
Principle of determination of ChE activity acetylcholinesterase ⎯ ⎯ ⎯⎯→ thiocholine + acetate acetylthiocholine ⎯⎯
thiocholine + dithiobisnitrobenzoate (DTNB) → TNB yellow color
When using acetylthiocholine as the substrate, the method of ChE activity determination is based on the measurement of the rate of production of thiocholine as acetylthiocholine is hydrolyzed. Thiol reacts with 5:5-dithiobis-2-nitrobenzoate ion (I)5 to produce the yellow anion of 5-thio-2-nitro-benzoic acid (II) (Fig 7; (Ellman et al. 1961):
+
+
+
enzyme H 2O + (CH 3 ) 3 N CH 2 CH 2 SCOCH 3 ⎯⎯ ⎯→(CH 3 ) 3 N CH 2CH 2 S − + C H 3 COO − + 2 H +
+
+
(CH 3 ) 3 N CH 2 CH 2 S − + RSSR ⎯ ⎯→(CH 3 ) 3 N CH 2 CH 2 SSR + RS − (I) (II)
R
= O2N COO-
Figure 6. Chemical reaction of acetythiocholine hydrolyzed into thiocholine and acetate. Thiocholine along with dithiobisnitrobenzoate produces the yellow color of 5-thio-2-nitro-benzoic acid. Figure adapted from Ellman et al. (1961).
The effect of tissue homogenate concentration on ChE was investigated. Optimal tissue concentration was found to be a 1:5 (w/v) dilution of the extract from body tissue; hence, in all subsequent analysis this dilution was used. Heads were dissected behind the operculum and discarded. Bodies were homogenized 1:5 (w/v) in 0.02M phosphate buffer (pH 7.0) containing 0.1% Triton X 100 using a Potter-Elvehjem homogenizer.
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Extracts were placed in pre-cooled centrifuge tubes and centrifuged at 10,000 × g for 20 min at 4°C. The supernatant from all samples were placed in deep-well plates and stored at -20°C. For cholinesterase activity measurements, 300 µl of 0.02M phosphate buffer (pH 7), 20 µl of 0.01M dithiobisnitrobenzoate (DTNB), and 10 µl of tissue extracts (200 mg ml-1) were added to the microtitre plates (Fig. 6). Extracts were incubated for 5 min before the addition of 10 µl of 0.1M acetylthiocholine (ACTC) to start the reaction. ChE activity was measured via a microplate reader (Bio-Tek Instruments, Inc.) at 405 nm at 25 ± 1°C for 10 min. Enzyme activity is given as the amount of enzyme which catalyzes the hydrolysis of 1 µmole of acetylcholine per minute per mg protein (µmol ACTC min-1 mg protein-1), and is calculated using the following equation (Bocquene & Galgani 1998): ChE activity (µmol ACTC min-1 mg protein-1) =
∆A405 × VolT × 1000 1.36 × 104 × lightpath × Vol s × [ protein ]
where: ∆A405 = change in absorbance (OD) per min, corrected for spontaneous hydrolysis
VolT = total assay volume (0.340 ml) 1.36 x 104 = extinction coefficient of TNB (M-1 cm-1) lightpath = microplate well depth (1 cm)
Vols = sample volume (in ml)
[ protein] = concentration of protein in the enzymatic extract (mg ml-1).
Figure 7. Microtitre plate used for cholinesterase and protein assays. Each plate contains 96 microwells.
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Protein assay Protein was determined spectrophotometrically using the Bio-Rad DC protein assay kit (Richmond, CA, USA) with bovine serum albumin (BSA) as standard, based on the method of Lowry et al. (1951). Statistical Analysis Statistica software 7.0 (StatSoft Inc, Tulsa, OK, USA) was used for all statistical analysis. Cochran’s test was used to investigate homogeneity of variances and a residual analysis was used to check normality and when necessary data were log transformed. General linear regression analysis was used to explore the relationship between ChE levels and size of fish. Tank and treatment differences of cholinesterase levels were compared by nested one-way analysis of variance (ANOVA). Results were considered statistically significant if p ≤ 0.05.
If significantly different, the Tukey post-hoc multiple
comparisons test was carried out to distinguish among means. ChE activity data was log transformed to meet the assumptions of normality and homogeneity of variances implicit for analysis of variance.
Percentages form a binomial, rather than a normal, distribution; therefore, percent mortality data were square root arcsine transformed to meet the assumption of normality implicit in analysis of variance (Zar 1999). For percentage mortality data, each fish was treated as an independent replicate.
For the recovery data, data was log transformed to meet the assumptions of normality and homogeneity of variances implicit for ANOVA.
A student’s t-test was performed
between the experimental and solvent controls. Type III sums of squares ANOVA was used due to an unbalanced design and to investigate the interactions between tank, treatment and time. ChE activity levels were analyzed by using two-way ANOVA with time and treatment as the factors and significant differences were analyzed with the Tukey post-hoc multiple comparisons test.
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Results
Water temperature, salinity and dissolved oxygen were monitored daily during each acute toxicity test (Table 1) and not found to differ between treatments.
Table 1. Mean water quality parameters during exposure studies.
Parameter Mean Min-Max 25.0 24.8 – 25.1 Temperature (°C) 38 38 Salinity (ppt) 4.3 – 5.4 Dissolved oxygen (mg L-1) 4.8
Optimization for enzymatic assays A linear increase in ChE activity with tissue homogenate concentration was observed for both body and head tissue (Fig. 8a). Greater ChE activity levels were found in body than head tissue. Mean ChE activity was also examined with the highest activity found from body tissue at a tissue concentration of 200 mg ml-1 of buffer (Fig. 8b). At the other concentrations, ChE activity from body tissue was similar to the activity obtained from head tissue.
Body Head
0.6 0.5
0.10 Mean AChE activity
∆ OD(405) min-1
Body Head
0.12
0.4 r2 = 0.973 y = 0.001x + 0.050
0.3
r2 = 0.996 y = 0.001x + 0.014
0.2 0.1
0.08 0.06 0.04 0.02
0.0
0.00 50
100
150
200
250
300
350
Tissue concentration (mg ml-1)
0
50
100
150
200
250
300
350
Tissue concentration (mg ml-1)
(a)
(b)
Figure 8. (a) Increase in absorbance per minute (± SE) as a function of muscle tissue and brain tissue concentration in the homogenate. (b) Mean ChE activity (± SE) as a function of muscle and brain tissue concentration in the homogenate.
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Experiment 1: 96 h range-finding
It was found with increasing chlorpyrifos concentration mortality also increased. No mortality was observed in both the experimental and solvent controls (Fig. 9). Two fish from 0.1 µg L-1 and 1.0 µg L-1 treatment chambers died within 24 h. Fish exposed to 500 µg L-1 moved their operculum more frequently than other fish. At approximately 65 h, fish in the 500 µg L-1 test chamber experienced typical organophosphorous intoxication, including tremors, decreased swimming ability, and inability to stay upright. Fish were not able to swim or move consistently, and spinal curvature was observed. Three of the four fish exposed to 500 µg L-1 chlorpyrifos died before completion of the 96 h toxicity test, which was the highest mortality rate observed for the experiment.
Percentage mortality
80
60
40 o Coln2tr co
l
n tr co
ol
20
0 nt tal lv e o l en im ol so ontr r c pe tr ex con
0.1
1
10
10
0
0 50
Treatment
Figure 9. Percentage mortality of P. novaeguineaensis exposed to chlorpyrifos at varying concentrations (in µg L-1) for 96 h.
Experiment 2: Size-ChE activity relationship There was a significant negative linear relationship between the length of fish and ChE activity (y = 0.364 – 0.004x; p < 0.000; Fig. 10). Highest ChE activity was seen in smallest fish while lower activity levels were observed in larger sized fish. Fish of total length between 45-90 mm exhibited similar levels of ChE activities.
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AChE activity (µmol ACTC min-1 mg protein-1)
0.6 0.5 0.4 0.3 0.2 0.1 0.0 20
30
40
50
60
70
80
90
100
TL (mm)
Figure 10. Relationship between total length (mm) of fish and ChE activity in P. novaeguineaensis (n = 63 fish). y = 0.364 – 0.004x; r2 = 0.352; p < 0.000.
Experiment 3: 96 h exposure concentration-ChE activity relationship There were no tank differences on ChE activity (F7,16 = 1.388; p = 0.167; Table 2) while treatment (chlorpyrifos concentration) had a significant effect on ChE activity (F1,7 = 14.690; p < 0.000; Table 2).
Table 2. Significance levels for nested ANOVA of effects of tank and treatment on ChE activity of P. novaeguineaensis in 96 h sublethal exposure experiment. Significant p values are in bold.
Effect df MS F p 7 0.015 14.690 < 0.000 Treatment 0.167 Tank(Treatment) 16 0.001 1.388 87 0.001 Error
Cholinesterase activity was inhibited with increasing chlorpyrifos concentration (Fig. 11). The experimental control exhibited the highest ChE activity while the lowest ChE activity levels were found in fish exposed to 128 and 256 µg L-1. There were no significant differences between the two controls or the controls to the 1 µg L-1 treatment, while a significant difference was observed between the experimental control and all other treatments (Fig. 11). ChE activity levels found in fish exposed to the solvent control was not significantly different to the ChE levels of fish exposed to 1 to 16 µg L-1.
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Mean AChE activity (µmol ACTC min-1 mg protein-1)
0.14 0.12 0.10 ∗
0.08
∗ ∗
0.06
∗
0.04
∗
0.02 0.00 ta l en rim trol e p ex con
nt lv e l so ntro o c
1
4
16
64
8 12
6 25
Treatment
Figure 11. Mean ChE activity (± SE) post 96 h sublethal toxicity test. Fish (n = 15 fish/treatment) were exposed to varying levels of chlorpyrifos (in µg L-1). Significance testing was performed on log transformed data. * denotes significance from experimental control. No significant difference was seen between the experimental control and solvent control.
Since no significant difference was found between the experimental and solvent controls (Fig. 11), inhibition of ChE activity was plotted relative to the experimental control. A sigmoidal relationship was observed between increasing chlorpyrifos concentration and decreasing ChE activity (y = -2.463 + 100.000x0.502/12.3620.502 + x0.502; r2 = 0.979; Fig. 12). Inhibition of ChE activity increased with increasing chlorpyrifos concentration. A sharp increase in inhibition of ChE activity was observed up to about 12 µg L-1 then started to plateau with higher concentrations. More than 75% inhibition of ChE activity was observed at a chlorpyrifos concentration of 128 µg L-1 and a similar inhibition was observed when fish were exposed to twice that concentration at 256 µg L-1. The IC50 value for P. novaeguineaensis, which is the concentration at which 50% of ChE has been inhibited, was approximately 12 ± 16.032 µg L-1 (Fig. 12).
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% ChE inhibition (% experimental control)
100
80
60
40 IC50
20
0 0
50
100
150
200
250
300
Chlorpyrifos concentration (µg L-1)
Figure 12. The change in the inhibition of ChE (± SE) relative to the experimental control with increasing chlorpyrifos concentration (in µg L-1) using mudskippers (n = 111 individuals). y = -2.463 + 100.000x0.502/12.3620.502 + x0.502; r2 = 0.979.
Mortality was highest in the 128 µg L-1 and 256 µg L-1 treatments where about half of the treatment fish died before completion of the experiment (Fig. 13). No mortality was recorded in the experimental control group. Two fish were found dead in the solvent control at 48 and 92 h of exposure. Mortality was also observed at the low chlorpyrifos concentrations, 1 µg L-1and 4 µg L-1; however, no mortality was observed in the 16 µg L1
treatment. Mortality was not significantly different between the treatments (F1,7 =
1.984; p = 0.122; Table 3).
60
Percentage mortality
50 40 30 20 10 0
tal en i m ro l r pe t ex con
nt lve l so ntro co
1
4
16
64
8 12
6 25
Treatment
Figure 13. Percentage mortality (± SE) of P. novaeguineaensis (n = 15 individuals/treatment) exposed to chlorpyrifos at differing concentrations (in µg L-1) for 96 hours.
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Table 3. Significance levels for one-way ANOVA of effects of a 96 h chlorpyrifos exposure on mortality of P. novaeguineaensis. Percentage mortality data were arcsine square root transformed prior to analysis.
Effect df MS F p Treatment 7 2.376 1.984 0.122 16 1.198 Error
Experiment 4: 96 h exposure and recovery kinetics During the exposure period, one fish from the experimental control and solvent control were found dead. Two fish were found dead in the 64 µg L-1 treatment. No deaths occurred during the recovery period of the experiment.
There was no significant difference in ChE activity between the experimental control and solvent control (p = 0.955; Table 4); therefore, subsequent analyses will not include solvent control data.
Table 4. Significance levels of student’s t-test between log ChE activity levels in P. novaeguineaensis of experimental and solvent controls of the recovery experiment.
Effect df t-value p log ChE activity 65 0.057 0.955
Treatment (chlorpyrifos concentration) (F1,2 = 31.73; p = 0.009) and time (F2,8 = 7.98; p < 0.000) had significant effects on ChE activity (Table 5). There were no tank (within treatment) differences on ChE activity (F16,3 = 0.99; p = 0.413; Table 5); therefore, data from replicate tanks within treatments were combined to investigate significant differences between treatments. A significant interaction was also observed for tank nested within time and treatment (F3,24 = 1.78; p = 0.046; Table 5).
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Table 4. Significance levels using Type III sums of squares ANOVA of effects of tank, treatment, and time on ChE activity of P. novaeguineaensis. Treatment and time were fixed factors. Significance factors are in bold. Data were log transformed prior to analysis.
Effect Effect df MS F p Fixed 2 0.727 31.73 0.009 Treatment Fixed 8 0.196 7.98 < 0.000 Time Fixed 16 0.063 2.57 Treatment*Time 0.018 Random 3 0.024 0.99 0.413 Tank(Treatment) Tank(Treatment*Time) Random 24 0.025 1.78 0.046 47 0.014 Error
ChE activities decreased with increased time of exposure (Fig. 14).
There was no
inhibition observed for both treatments after 8 h of exposure. Inhibition was detected for fish exposed to 64 µg L-1 after 24 h, while inhibition was not observed until 96 h of exposure for the 4 µg L-1-treated fish. By 96 h fish from both treatments exhibited a significant reduction in ChE activity relative to the experimental control (Fig. 14).
After the transfer of fish to pesticide-free seawater, ChE activities increased with time and recovery commenced (Fig. 14). By day 17, ChE activity levels for fish treated with 4 µg L-1 were not significantly different with ChE activities of control fish. Recovery for fish exposed to 64 µg L-1 took longer and by day 41, ChE activity was still significantly different than ChE activities of control fish. After 55 d of recovery, ChE activity of fish exposed at 64 µg L-1 was not significantly different from control ChE activity.
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experimental control 4 mg L-1 64 mg L-1
Mean ChE activity
0.20
0.20
0.16
0.16
0.12
0.12
0.08
0.08
* *
0.04
0.04
*
*
0.00 0
1
2
3
4
* 5
0
10
20
30
40
50
0.00 60
Recovery Time (days)
Exposure Time (days)
Figure 14. Mean ChE inhibition of P. novaeguineaensis exposed to 4 and 64 µg L-1 chlorpyrifos for 96 hrs relative to experimental control. Two-way ANOVA was performed on log transformed data and Tukey’s post-hoc comparison test was used to investigate significance between treatment and control ChE activity levels at that time point. * denotes significance between treatment ChE activity from experimental control ChE activity level at that time point.
Discussion
Signs of chlorpyrifos intoxication observed in P. novaeguineaensis included tremors, decreased swimming ability, and an inability to stay upright, which are typical signs of organophosphate poisoning as observed in other fish species (Odenkirchen & Eisler 1988, Zinkl et al. 1991). Spinal curvature was observed in mudskippers exposed to high concentrations of chlorpyrifos.
Skeletal deformities have also been associated with
chlorpyrifos in other fish species (Jarvinen et al. 1983, Harden & Curtis 2002) but this was not tested in the current project. The range finding study assessed the concentrations needed to investigate sublethal effects of chlorpyrifos on cholinesterase (ChE) activity. No mortality was observed at the 100 µg L-1 treatment while 75% mortality was exhibited in the 500 µg L-1 treatment indicating that concentrations lower than 500 µg L-1 should be implemented in the sublethal exposures. Therefore, the highest concentration used in the 96 h sublethal exposure test was 256 µg L-1.
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Greater ChE activity was found in body tissue of P. novaeguineaensis versus activity found in the head. This may be due to the presence of higher amounts of cholinesterase in body tissue.
Galgani et al. (1992) found both acetylcholinesterase (AChE) and
butyrylcholinesterase (BChE) in muscle tissue of dab Limanda limanda and found that overall, a higher level of cholinesterase was present in the brain than in muscle tissue. Another possibility of the greater ChE activity in body tissue is the presence of more than one type of cholinesterase (ChE). Several studies have found both AChE and BChE in muscle tissue of several marine fish species (Galgani et al. 1992, Sturm et al. 1999a) and freshwater fish (Sturm et al. 2000). However, in several teleost fish, brain tissue contains predominantly AChE (Galgani et al. 1992, Monteiro et al. 2005, Osten et al. 2005, Whitehead et al. 2005). Possible additional cholinesterases within the muscle tissue, therefore, may contribute to the greater activity observed in the body.
There are different forms of cholinesterases in fish tissues which may exhibit different sensitivities to anticholinesterase agents (Monteiro et al. 2005, Osten et al. 2005). Two other
esterases,
butyrylcholinesterase
(BChE
or
pseudocholinesterase)
and
carboxylesterase (aliesterase), are also targets of inhibition by organophosphates (Chambers et al. 2002).
Studies have shown that BChE is more sensitive to
organophosphates than AChE (Sturm et al. 2000, Fulton & Key 2001). Techniques, such as the Ellman method, do not distinguish between AChE and BChE. Therefore, if the tissue also contains BChE, then the inhibitory effects of chlorpyrifos on the target enzyme, AChE, may be overestimated. Consequently, characterization studies should be employed to better understand the kinetics of chlorpyrifos on the possible different cholinesterases of mudskippers.
A negative linear relationship was observed between total length of the fish and ChE activity such that larger fish exhibited lower ChE activities than smaller-sized fish. A similar relationship was also found by Flammarion et al. (2002) using chub (Leuciscus cephalus) and Sturm et al. (1999b) with three-spined sticklebacks. According to Dutta et al. (1995), juvenile catfish brain cholinesterase was more sensitive to organophosphate inhibition than brain cholinesterase from adults. On the other hand, Galgani et al. (1992)
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did not find a significant effect of size on ChE activity in dab (Limanda limanda). Nevertheless, differences in susceptibility to chlorpyrifos and ChE sensitivity at various developmental stages may exist in P. novaeguineaensis. Therefore, mudskippers of similar size were used for the sublethal toxicity tests since size, age and developmental stage may influence ChE activity. In addition, size of fish may influence toxicity in static tests since absorption by the fish may decrease the concentration available (Barron & Woodburn 1995).
Although no chlorpyrifos was present in the solvent control from the 96 h sublethal exposure test, mortality was still observed. Previous studies have shown that acetone itself has negligible toxicity to fish (Weiss 1961). The presence of acetone, though, may decrease the oxygen content in the test water due to the breakdown of acetone (van der Wel & Welling 1989) influencing the survival of the test animals. However, mean values of dissolved oxygen (DO) were not considerably different among the treatments suggesting minimal or no effect of low DO on the fish. In addition, mudskippers are able to withstand and tolerate high ammonia (Polgar 2007) and anoxic environments (Ip et al. 2005, Lee et al. 2005). The dissolved oxygen concentration did not appear to fall below the critical level for mudskippers; however, a lower concentration of acetone should be used for further investigations.
Sex of fish was not determined which may explain some of the variability in inhibition of ChE activity. Van der Wel and Welling (1989) found that male guppies always had higher ChE activities than females. However, other studies, including Flammarion et al. (2002) and Galgani et al. (1992), did not find a sex effect on ChE activity using chub (Leuciscus cephalus) and dab (Limanda limanda), respectively. Determining whether sex influences ChE activity in P. novaeguineaensis is a possible focus for future studies.
Organophosphates, such as chlorpyrifos, are known as cholinesterase inhibitors. When AChE is inactivated, ACh accumulates and leads to continual stimulation of nerves and muscles, which produces rapid twitching of voluntary and involuntary muscles. This rapid twitching may lead to paralysis and eventually death.
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Several studies have
investigated the relationship between chlorpyrifos and mortality in fish (Jarvinen & Tanner 1982, Borthwick et al. 1985, Mayer & Ellersieck 1986, Mayer 1987). Fulton and Key (2001) reviewed ChE inhibition in fish and found that in certain species when brain ChE inhibition reached 70 to 80%, mortality was likely to occur while other studies found no distinct relationship. Brain ChE was not investigated in this study, thus, a relationship cannot be determined between ChE inhibition and lethality in P. novaeguineaensis.
Knowledge about the relationship between ChE inhibition and
lethality in fish is not well established (Boone & Chambers 1996, Ferrari et al. 2004). However, it was evident that inhibition of ChE activity was dose dependent with higher concentrations exhibiting highest mortality and inhibition.
Humphrey and Klumpp
(2003) also observed that mortality of rainbowfish (Melanotaenia splendida splendida) increased with increasing chlorpyrifos concentration.
Considering the
percentage
mortality
results,
an
approximate
LD50
for
P.
novaeguineaensis was around 250 µg L-1. A review of the toxicity of chlorpyrifos by Barron and Woodburn (1995) found that chlorpyrifos was acutely toxic to fish, with LC50 values ranging widely from lower than 1 µg L-1 to greater than 1000 µg L-1 so the sensitivity of P. novaeguineaensis to chlorpyrifos falls within the range observed in several other fish species. The lowest observed effective concentration (LOEC) for P. novaeguineaensis, which is the lowest concentration for which the effect is different from that of the controls, was 4 µg L-1, and the IC50 was approximately 12 ± 16.032 µg L-1. The error term for the IC50 value for P. novaeguineaensis is comparatively high due to the initial sharp slope. The LD50 was over 50 times the concentration of the LOEC and over 20 times the concentration of the IC50. Cholinesterase inhibition, as a biomarker, is a good indicator of organophosphate intoxication due to its high sensitivity to sublethal exposures.
Uptake of chlorpyrifos (as indicated by ChE inhibition) was faster for fish exposed to 64 ug L-1 than fish exposed to the lower concentration. Inhibition was not evident until 96 h exposure for the 4 µg L-1-treated fish.
Mac and Seelye (1981) suggested that the
presence of acetone may increase the solubility and availability of a contaminant, thus,
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influencing uptake.
However, equivalent concentrations of acetone were used.
A
laboratory study conducted by Lalah et al. (2003) found that uptake of chlorpyrifos was rapid, detecting chlorpyrifos in rabbitfish (Seganus stellatus) and snapper fish (Luthrimus fulviflama) after only 4 hours.
It is predicted that chlorpyrifos is not a threat to groundwater via leaching and has minimal surface runoff potential (Racke 1993, Galloway & Handy 2003). However, temporal factors, such as heavy rainfall during the wet season, may generate high enough discharge to reach and extend to the aquatic environment. Areas of soil erosion are particularly important with the potential of pollutants entering aquatic and marine environments. Though runoff potential is proposed to be minimal, several fish kills have been reported possibly due to termiticidal treatments containing chlorpyrifos running off and contaminating aquatic systems (NRA 2000). Low concentrations of chlorpyrifos, however, have been detected in fish during field studies (Barron & Woodburn 1995). Reported concentrations of chlorpyrifos found in fish ranged from 0.004 to 0.0004 mg kg-1 (USEPA 1992). In addition, half-life values of chlorpyrifos in fish is relatively short with half-life values of 14 h for three-spined stickleback Gasterosteus aculeatus and 66 h for rainbow trout Oncorhynchus mykiss (Barron & Woodburn 1995). Due to rapid hydrolytic degradation, chlorpyrifos oxon is most likely short-lived in the environment (Racke 1993). Also, chlorpyrifos has a great tendency to partition from aqueous into organic phases in the environment due to its hydrophobicity and low water solubility (< 2000 µg L-1) (Racke 1993). Studies have shown that toxicity of chlorpyrifos has been dependent on the species (Barron & Woodburn 1995, Heath 2000), and that differences exist in uptake and metabolism of organophosphates between species (Galloway & Handy 2003).
The first day of sampling of the recovery period was 2 days post-chlorpyrifos exposure. It appeared that fish exposed to the higher concentration, 64 µg L-1, showed further ChE inhibition 2 d after transference to pesticide-free seawater while fish exposed to the 4 µg L-1 did not experience any additional inhibition after transference. The further decrease in ChE activity seen in the 64 µg L-1-treated fish may be due to residual chlorpyrifos in
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the tissues of the fish, which was still inhibiting enzyme activity. Due to the higher dosage, chlorpyrifos residues may have been present in the tissues of the 64 µg L-1treated fish 2 days post exposure while no chlorpyrifos residues remained in the fish exposed to the lower concentration. Residues may have been present in the lower concentration-treated fish immediately after exposure, though due to the time between sampling periods, any residual chlorpyrifos may already have been excreted or metabolized. Lalah et al. (2003) found that although recovery occurred in snapper fish and rabbitfish, it was relatively slow so that high concentrations of chlorpyrifos were still evident within the tissues of exposed fish. Though chlorpyrifos may degrade rapidly in the environment, inhibition of AChE persists longer, from days to weeks (Fulton & Key 2001). The metabolism of organophosphates is complex and not completely understood (Galloway & Handy 2003) and further investigation is required.
Recovery of ChE activity began quickly in mudskippers noticing an increase in activity one week after exposure.
Organophosphate insecticides are considered irreversible
inhibitors of acetylcholinesterase so it has been suggested that recovery of ChE activity is due to the production of new enzymes (Fulton & Key 2001). P. novaeguineaensis demonstrated a capacity for the regeneration of new ChE. However, recovery rates were considerably much longer than the rate of inhibition. Recovery of ChE activity occurred when fish were transferred to pesticide-free seawater. For mudskippers, recovery of ChE activity commenced at one week in clean water. In addition, the recovery for fish exposed to the higher concentration (64 µg L-1) was relatively slower than fish exposed to 4 µg L-1. ChE activities for fish exposed at 64 µg L-1 were statistically significantly different from control fish at day 7 while no significant difference was observed between ChE levels of fish exposed at 4 µg L-1 and controls at that time point. Other studies found even slower recoveries of ChE activity after exposure to organophosphates with inhibition still detected after 30 d in pesticide-free water (Ferrari et al. 2004, Van Cong et al. 2006). Recovery of ChE activity was a function of concentration of prior exposure such that fish exposed to higher chlorpyrifos concentrations required a longer recovery time than those exposed to lower concentrations.
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Several studies have shown that
toxicity of chlorpyrifos greatly depends on the dose, route and extent of exposure to the toxin (Galloway & Handy 2003).
Many of the cholinesterase inhibitors can be metabolized and excreted quickly and new acetylcholine can be synthesized so that recovery can occur (Connell et al. 1999). The slow rate of recovery of ChE suggests that new enzyme is not being synthesized at a significant rate compared to the rate of inhibition. The slow rate of recovery also suggests continued inhibition or irreversible inhibition and may be due to slow synthesis of new enzyme.
Measuring ChE activity provides a useful method in determining whether mudskippers have been exposed to organophosphates. Besides inhibition of ChE, organophosphate exposure may lead to behavioral and physiological differences. Depression of ChE activity by organophosphates has been associated with decreased food intake and decreased growth (Kumar & Chapman 1998). In addition, Dutta et al. (1992) found that along with ChE inhibition, the optomotor response of the bluegill sunfish was impaired when exposed to an organophosphate insecticide. Optomotor function is crucial in search for food, and locating and avoiding predators, thus, organophosphate exposure greatly influences the survival of exposed individuals. If exposed, these individuals may pose detrimental effects not only to themselves but also to other organisms within that ecosystem. Several studies have shown that chlorpyrifos is highly toxic to fish (Barron & Woodburn 1995, Carr et al. 1997), thus, bioaccumulation of chlorpyrifos should also be considered. High concentrations within exposed organisms may pose serious risks to its predators and in consequence may impact the interconnectivity of the ecosystem. Therefore, determining whether individuals have been exposed to organophosphates by means of biomarkers is important for the sustainability of the entire ecosystem.
Conclusion
ChE inhibition was dose dependent with higher chlorpyrifos concentrations yielding greater depression of ChE activity. Full recovery of ChE activity was relatively quick
- 26 -
requiring only one week in clean water for fish exposed at 64 µg L-1. Inhibition of ChE is a useful biomarker for detecting organophosphate insecticide exposure on this species by enabling detection of both recent and past contamination; however, characterization of cholinesterases is necessary as sensitivity of the different cholinesterases may vary. Biomarkers provide useful diagnostic tools of pesticide exposure, illustrating effects at much lower concentrations than found for traditional toxicity testing, e.g. LD50. Detection of ChE inhibition weeks after initial exposure is advantageous in determining exposure to organophosphate insecticides. Investigating the enzymatic inhibitory effects of chlorpyrifos is important since inhibition of ChE may produce behavioral and physiological alterations that may affect the animal’s fitness.
Exposure to
organophosphates not only affects the exposed individual but also can have consequences on the entire ecosystem.
Acknowledgements: I would like to acknowledge A. Negri and C. Humphrey for their supervision and assistance with this research project and report, and to M. McCormick for his guidance with statistical analyses. I would also like to thank DowElanco (Australia) Ltd. for providing chlorpyrifos. Many thanks to the mudskipper hunting crew: A. Negri, C. Humphrey, and S. Boyle, and to W. Purnell for his support throughout this endeavor.
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