ABSTRACT
Previous work suggested that Ca2+ and Zn2+ share a common uptake pathway in rainbow trout gills. We here report on relationships between the kinetic variables for unidirectional Ca2+ influx and unidirectional Zn2+ influx during a 1 month exposure of freshwater rainbow trout to Zn2+ (150 μg l-1=2.3 μmol l-1 as total zinc, Zn). Initial exposure to Zn2+ caused a large competitive inhibition of Ca2+ influx, as indicated by a threefold increase in apparent Km for Ca2+ (measured in the presence of Zn2+). There was also a smaller non-competitive inhibition (50 % decrease in Jmax) of the Ca2+ transport system, which was abolished after 1–2 weeks of exposure. The Km, measured in the absence of Zn2+, decreased dramatically (i.e. elevated affinity) on days 1–4 but increased thereafter; both true and apparent Km finally stabilized significantly above control levels. However, the Km values for Ca2+ (<200 μmol l-1) were low relative to the Ca2+ level in the water (1000 μmol l-1), and therefore the changes did not influence the actual Ca2+ influx of the fish, which tracked Jmax. In contrast, water [Zn2+] (2.3 μmol l-1 as total Zn) was close to the reported apparent Km (3.7 μmol l-1) for Zn2+ influx in the presence of 1000 μmol l-1 Ca2+. Unidirectional Zn2+ influx increased during the first week of exposure to waterborne Zn2+, followed by a persistent reduction to about 50 % of control levels, effects that may be largely explained by the observed changes in true Km for Ca2+. We speculate that the initial response of the fish to elevated [Zn2+] is to compensate for a reduced availability of Ca2+ by markedly increasing the affinity of a dual Ca2+/Zn2+ transporter. Once the Ca2+ influx is ‘corrected’ by restoration of functional transport sites (Jmax), the system is tuned to limit the influx of Zn2+ by a persistent reduction in the affinities for both ions.
The changes in influx characteristics for Ca2+ and Zn2+ were correlated with internal physiological alterations indicative of adaptation to Zn2+ and increased metabolic cost. Depressed plasma [Ca] was corrected within 1 week, and there were no effects on whole-body [Ca] or [Zn]. A slight accumulation of Zn in the gills was associated with increased branchial metallothionein levels. Rates of protein synthesis and degradation in the gills were initially increased and whole-body growth was transiently impaired, effects which were reversed after 18 days of exposure. Sublethal challenge with Zn2+ (at 450 μg l-1=6.9 μmol l-1 as total Zn) always depressed plasma [Ca] in control fish, but by 1 month of exposure to Zn2+ at 150 μg l-1 (as total Zn), experimental fish were resistant to challenge. However, the fish did not acquire increased survival tolerance (LT50) to a lethal concentration of Zn2+ (4 mg l-1=61 μmol l-1 as total Zn).
Introduction
Zinc is a micronutrient for fish, but elevated concentrations of waterborne Zn2+ can impair the branchial uptake of Ca2+ (Spry and Wood, 1985, 1988; Sayer et al. 1991; Hogstrand et al. 1994). In brown trout (Salmo trutta), a lethal concentration of waterborne Zn2+ caused a net loss of Ca2+ by affecting both influx and efflux variables (Sayer et al. 1991). Recently, we found that a sublethal level of waterborne Zn2+ competitively inhibited the branchial Ca2+ influx of freshwater rainbow trout (Oncorhynchus mykiss), suggesting that Zn2+ and Ca2+ may compete for the same uptake sites (Hogstrand et al. 1994). These data confirm and extend previous findings that Ca2+ is a competitive inhibitor of Zn2+ influx through the gills (Spry and Wood, 1989). A simple competitive interaction between Ca2+ and Zn2+ at the gill surface could also explain the well-known protective effect of increased water hardness on fish exposed to Zn2+ (Pagenkopf, 1983).
Disturbances of branchial ionoregulation are common effects of metal exposure, although the underlying mechanisms may differ between metals (Wood, 1992). During chronic exposure to sublethal concentrations of metals, fish often adapt and normal rates of ion transport are restored (McDonald and Wood, 1993). However, adaptation to an elevated Zn2+ concentration does not seem to be associated with a complete restoration of normal Ca2+ influx kinetics: during prolonged exposure (60 days) of hardwater-acclimated rainbow trout to waterborne Zn2+, the apparent Km for Ca2+ influx was chronically elevated and Zn2+ influx was reduced (Hogstrand et al. 1994). Bradley et al. (1985) also observed that the rate of Zn accumulation in gills was lower in Zn2+-exposed rainbow trout. Hogstrand et al. (1994) reasoned that the Km for Ca2+ was kept elevated to reduce the influx of Zn2+ and that this, in itself, could serve as a mechanism of adaptation. However, the influx of Zn2+ was only measured at the end of the 60-day exposure period and a relationship between Zn2+ influx and the affinity of the Ca2+ transporter has yet to be demonstrated.
The first major objective of the present study was, therefore, to examine the relationship between the kinetic variables of unidirectional Ca2+ influx and the influx of Zn2+ in freshwater rainbow trout. In our previous study (Hogstrand et al. 1994), the Ca2+ influx kinetics of Zn2+-exposed fish was analyzed, but only in the presence of the inhibitor (Zn2+). Therefore, it could not be determined whether the chronically elevated Km for Ca2+ was due to the continuing competitive inhibition by Zn2+ or whether the properties of the carrier(s) per se had also changed. To address this question, the Ca2+ influx kinetics of rainbow trout chronically exposed for 1 month to Zn2+ (150 μg l-1=2.3 μmol l-1 as total Zn) was studied both in the presence and absence of the inhibitor (Zn2+), yielding values of the apparent Jmax and Km as well as the true Jmax and Km for Ca2+ influx. The rate of Zn2+ influx into the whole animal and the Zn accumulation rate in gill tissue were also analyzed throughout the experimental period. These measurements allowed a continuous comparison between the kinetic variables for Ca2+ influx and the influx of Zn2+. As measures of the degree of adaptation to an elevated Zn2+ concentration, plasma and whole-body levels of Ca and levels of Zn in the gills and whole body were monitored during the exposure. Periodic challenges with a higher, but still sublethal, concentration of Zn2+ were performed to assess whether the Ca2+ regulatory system was becoming resistant to the hypocalcaemic effect of Zn2+ challenge. Acute toxicity tests (LT50) to Zn2+ were conducted regularly to give evidence for any change in tolerance to a lethal concentration of Zn2+. Finally, the branchial intracellular sequestering capacity for Zn was assessed by analysing the concentrations of the metal-binding protein metallothionein (MT), using a specific and sensitive radioimmunoassay (Hogstrand and Haux, 1990).
The second major objective of this study was to assess the metabolic cost associated with adaptation to sublethal levels of Zn2+. It is generally believed that adaptation to any physiologically unfavourable environment is linked to increased metabolic expenditure, so that less energy is available for other processes such as feeding, growth and reproduction (Bayne et al. 1979; Widdows et al. 1990; Calow, 1991). Wilson et al. (1994a,b) have recently provided evidence for such energetic limitations in rainbow trout adapting to sublethal Al3+ concentrations. McDonald and Wood (1993) have further postulated that adaptation to waterborne metals involves a damage repair process in the gills, such that increased cost would be induced at this site. While damage repair would probably be associated with an increased rate of protein synthesis, evidence to date is equivocal. Indeed, Hogstrand et al. (1994) observed a slightly reduced protein synthesis rate in the gills of Zn2+-exposed rainbow trout. However, in that study, the rate of protein synthesis was measured after increased tolerance to Zn2+ had developed (23 days), by which time any damage repair process may already have been completed. To investigate these questions, in the present study, the growth of fish on a fixed ration of food was monitored throughout the period of Zn2+ exposure, while protein and RNA concentrations in the gills, together with protein synthesis and degradation rates, were determined on days 9 and 18 of the exposure using the technique developed by Houlihan et al. (1986).
Materials and methods
Experimental animals
Approximately 1600 juvenile rainbow trout Oncorhynchus mykiss (Walbaum) (4.90±0.92 g; mean ± S.D., N=96) were obtained from a local fish hatchery (Spring Valley Trout Farm, Petersburg, Ontario). The fish were held in two 264 l fibreglass tanks (800 per tank), each supplied with dechlorinated, aerated Hamilton city tapwater ([Na+]=0.6 mmol l-1; [Cl-]=0.7 mmol l-1; [Ca2+]=1.0 mmol l-1; [HCO3-]= 1.9 mmol l-1; pH=7.9–8.2) at a flow rate of 900 ml min-1 and a temperature of 9 °C. Fish were fed dry trout pellets (Martin’s Feed Mill Ltd, Ontario) at a ration of 1 % of their body mass per day.
Chronic Zn2+ exposure
After an acclimation period of 14 days, one tank was equipped with a dosing system that added Zn2+, as ZnSO4.7H2O (BDH Chemicals), from a stock solution. The other served as control tank. The flow rate of the added Zn2+ stock (45 mg l-1=690 μmol l-1) was maintained at 3 ml min-1, by a peristaltic pump, to reach a concentration of total Zn in the tank of 150 μg l-1 (2.3 μmol l-1; measured range 132–184 μg l-1). Using the MINEQL+ computer program for chemical equilibrium management (Schecher and McAvoy, 1991), it was calculated that the nominal [Zn2+] of the exposure water was 1.4 μmol l-1. The day the Zn2+ exposure started is referred to as ‘day 0’ throughout the text. Water flow rate and dosing rate were checked daily and adjusted if necessary, and feeding was maintained at 1 % of the body mass per day. The experiment continued for 31 days, over May and June, during which the water temperature gradually increased from 9 to 13 °C due to ambient conditions. No mortality occurred during the Zn2+ exposure.
Water samples from the exposure tank were taken for analysis of total Zn every 1–2 days. Samples were acidified with HNO3 (trace metal analysis grade, BDH Chemicals, was used in all procedures) to a final acid concentration of 1 % (w/w), and Zn was measured with an atomic absorption spectrophotometer (AAS; Varian AA-1275), using an air/acetylene flame.
Ca2+ uptake kinetics
The kinetics of unidirectional Ca2+ influx was determined on days -2 (i.e. 2 days before the treatment started in the experimental group), 17 and 31 in the control group, and on days 0 (i.e. after 4 h of exposure to Zn2+), 1, 4, 7, 15 and 29 in the Zn2+-exposed group. Kinetic analysis was performed as described by Hogstrand et al. (1994). In brief, eight fish from the control group were put into each of six polypropylene flux bags, containing 3 l of synthetic water ([NaCl]=0.7 mmol l-1; [KHCO3]=1.9 mmol l-1; pH=8.0) with a designated concentration of Ca2+ reached by adding an appropriate volume of a 45Ca(NO3)2 stock solution (52 mmol l-1; specific activity 35 kBq mol-1). The six flux bags represented a geometric series of increasing [Ca2+], with approximate values of 52, 104, 207, 413, 820 and 1615 μmol l-1. In the flux medium, 99.4 % of the total Ca content was calculated to be present as Ca2+ (MINEQL+; Schecher and McAvoy, 1991). For the Zn2+-exposed fish, two kinetic series (six flux bags each) were run in parallel. One series was performed identically to those for the control group (i.e. in the absence of Zn2+), while the other was run in the presence of Zn2+ (150 μg l-1=2.3 μmol l-1 as total Zn in each flux bag). The former yielded values of true Km and Jmax (no competitor present), and the latter yielded values of apparent Km and Jmax (i.e. determined in the presence of the competitor). On day 0 (first 4 h of exposure), only the latter series was performed. On all other days, both series were performed. The Zn2+-exposed fish that were tested for Ca2+ influx in the absence of Zn2+ were rinsed in dechlorinated Hamilton tapwater for 5 min before they were introduced to the flux bags.
The unidirectional Ca2+ influx was calculated from the appearance of 45Ca radioactivity in the whole body over a 4 h period. The fish were killed, rinsed in 10 mmol l-1 Ca(NO3)2 to displace surface-bound 45Ca, and processed for counting as described by Hogstrand et al. (1994). Lineweaver–Burk plots were used to obtain values of Jmax and Km for the Michaelis–Menten equation. Note that the Lineweaver–Burk transformation yields asymmetrical S.E.M. values. The Michaelis–Menten equation was then used to interpolate the actual unidirectional influx of Ca2+, Jin, for fish at the acclimation concentration of Ca2+, 1.0 mmol l-1. The body masses of the 48–96 fish used for analysis of Ca2+ influx were measured for estimation of growth in the two groups.
Influx of Zn2+ and Zn accumulation rate in gills
The unidirectional influx of Zn2+, at the Zn2+ concentration of the exposure tank (150 μg l-1=2.3 μmol l-1 as total Zn), was measured on days -2 (i.e. 2 days before the treatment started in the experimental group), 17 and 31 in the control group, and on days 0 (i.e. first 24 h of Zn2+ exposure), 1, 4, 7, 15 and 29, as described by Spry and Wood (1989) with the modifications used by Hogstrand et al. (1994). Black polypropylene bags were used as flux compartments. Each of the bags was filled with 20 l of dechlorinated Hamilton tapwater (see above) with 150 μg Zn l-1 (2.3 μmol l-1) added as ZnSO4. 5 min before the fish were introduced to the bags, 2.6 MBq of carrier-free 65Zn was added to each bag. Ten fish from each tank were transferred to the flux bags and were then held there for 24 h. At the end of the flux period, the fish were lightly anaesthetized (MS 222, 20 mg l-1) and individually transferred for 1 min to a beaker with tapwater containing MS 222 (1.0 g l-1), and ‘cold’ Zn2+ (1.0 mg l-1 as total Zn) to displace surface-bound 65Zn. The fish was blotted dry and a terminal blood sample (100 μl) was withdrawn with a heparinized Hamilton syringe from the caudal vessels. The blood was then centrifuged at 14 000 g for 3 min to obtain the plasma fraction. The gills were dissected out and the soft tissue was scraped from the cartilaginous tissue with two microscope slides. Plasma and gills were assayed for 65Zn activity in a -γ-counter (MINAXI -γ Auto-Gamma 5000 Series, Canberra-Packard). Water samples were similarly counted for 65Zn activity and total Zn in water was measured by atomic absorption spectroscopy (AAS) (Varian 1275). The influx rate of Zn2+ was calculated by the method established for rainbow trout by Spry and Wood (1989). The procedure was based on the relationship between the steady-state activity of 65Zn in plasma after 24 h of exposure and the influx of Zn2+ (in nmol kg-1 h-1). The calculation is not affected by the endogenous level of Zn in plasma (Spry and Wood, 1989). The accumulation rate of Zn in the gills was calculated from the 65Zn activity in the gills, corrected for counts from blood trapped in the tissue (Munger et al. 1991), the specific activity of 65Zn in the water and the mass of the soft gill tissue.
The data presented on Ca2+ and Zn2+ influx are likely to represent influx via the branchial pathway, because oesophageal ligation experiments have eliminated the intestinal route as a significant route of Ca2+ and Zn2+ uptake under identical conditions to the flux experiments performed here (Perry and Wood, 1985; Spry, 1987).
Levels of metallothionein, Cu, Zn and Ca
The levels of metallothionein (MT), Cu and Zn in gills, Ca in plasma and whole-body concentrations of Ca and Zn were analyzed at different points during the experiment. Cu in gills was measured because MT binds both Cu and Zn, and therefore an altered Zn concentration could change the Cu level. Exposed fish were sampled on days 1, 4, 7, 15 and 29, whereas controls were sampled on days -2 (i.e. 2 days before the exposure started in the experimental group), 17 and 28. On each sampling occasion, 10 fish were killed by a blow to the head and a blood sample from each fish was taken from the caudal vessels with a heparinized Hamilton syringe. The blood sample was centrifuged (14 000 g, 3 min) and the plasma was separated, frozen in liquid nitrogen and stored at -70 °C for later analysis of Ca content by AAS. Gills were dissected out and the soft tissue of the gill filaments was scraped from the cartilaginous tissue with two microscope slides. The gill samples and the carcasses were frozen in liquid nitrogen and stored at -70 °C.
The frozen gill scrapings from fish sampled on days -2, 1, 15, 17, 28 and 29 were later weighed (typically, 50 mg of tissue was obtained from each fish), thawed and homogenized individually in 1.0 ml of 50 mmol l-1 Tris–HCl, pH 8.0, at 0 °C, using a glass–Teflon homogenizer. A 450μl sample from each homogenate was stored at -20 °C for subsequent analysis of Zn and Cu. The remainder was centrifuged at 10 000 g for 20 min at 4 °C; the supernatant was decanted, frozen in liquid nitrogen and stored at -70 °C until used for measurement of [MT]. Levels of MT were analyzed with a double-antibody radioimmunoassay, RIA, using rabbit antiserum raised against MT from perch, Perca fluviatilis, as the first antibody, 125I-labelled rainbow trout MT as tracer, and goat anti-rabbit IgG as the second antibody (Hogstrand and Haux, 1990). The MT (I and II) from rainbow trout used as a tracer was purified according to Olsson and Haux (1985), with the modifications described by Hogstrand and Haux (1990) for perch MT. A 10 000 g supernatant prepared from liver of cadmium-injected rainbow trout was used as a MT standard. The MT content of the standard was calibrated against a standard curve prepared from purified rainbow trout MT (Hogstrand and Haux, 1990). The working range of the RIA was 1.5–15 pmol of rainbow trout MT per assay tube, which corresponds to 92–920 pmol g-1 tissue wet mass.
For analysis of metals, gill homogenates were digested in acid-washed glass tubes for 1 h with 5 vols of 70 % HNO3 at 120 °C. The samples were cooled to room temperature, 0.75 vols of H2O2 was added, and the samples were then evaporated to dryness at 120 °C. Finally, 5 ml of 1 % HNO3 was added to the digestion tubes and Cu and Zn were analyzed by AAS as described above. No solid material was present in the final solution.
Whole carcasses (whole body minus gills) of fish sampled on days -2 (i.e. 2 days before the start of the exposure), 1, 15, 17, 28 and 29 were digested in 20 ml of 35 % HNO3 at 85 °C for 2 h, using graduated plastic tubes with screw caps (Falcon). The tubes were cooled, 2.0 ml of 30 % H2O2 was added, and the digestion was then continued at 85 °C for another 12 h. The digest was then made up to the nearest 5 ml with deionized water (Nanopure). Samples for Zn analysis were further diluted 25 times with deionized water (Nanopure) and analyzed by AAS. Ca analysis was performed on digests diluted 3000 times with 0.2 % LaCl2.
Sublethal Zn2+ challenge
The effect of a sublethal Zn2+ challenge on plasma Ca levels was studied on days -2 (i.e. 2 days before the start of exposure in the experimental group), 17 and 31 for control fish and on days 1, 15 and 29 for pre-exposed fish. In each challenge, 10 fish were transferred to a polypropylene bag containing 20 l of dechlorinated Hamilton tapwater with Zn2+ (450 μg l-1= 6.9 μmol l-1 as total Zn), added as ZnSO4. The bag was fitted with an air-line and submerged in a waterbath at the same temperature as the holding tanks. After 24 h, the fish were killed by a blow to the head and a blood sample was withdrawn from the caudal vessels. The blood was centrifuged (14 000 g, 3 min) and the plasma separated and frozen in liquid nitrogen. Samples were stored at -70 °C until analysis of Ca content by AAS. The plasma [Ca] of the challenged fish was compared with that of untreated control fish. Sampling of untreated control fish is described above.
Acute toxicity tests
Median lethal time (LT50) tests of both Zn2+-exposed fish and controls were performed on days 1, 4, 7, 15 and 30. On these days, 10 fish from each group were transferred to a 14 l black acrylic box. The test chamber was equipped with air-lines and centrally divided by a plastic mesh to allow separate exposure of the two groups to identical water conditions. To ensure efficient water turnover, the inflow and the drain were at opposite ends of the chamber. The water was from the same source and was at the same temperature as the water in the holding tanks. A ZnSO4 stock (160 mg Zn l-1=2.45 mmol l-1) was added at 3.0 ml min-1 to the incoming water (120 ml min-1) by a peristaltic pump and mixing funnel, yielding a measured total Zn concentration in the test chambers averaging 4 mg l-1 (61 μmol l-1). [Zn] in the water was monitored daily. The concentrations of total Zn (in mg l-1) in the water at the different test occasions were 4.1±0.1 (N=4), 3.9±0.5 (N=5), 4.1±0.4 (N=5), 4.0±0.1 (N=3) and 4.1±0.3 (N=7) (mean ± S.D.), respectively. Mortality was monitored throughout the first day of exposure, and four times daily thereafter, for up to 10 days (226 h). The fish were not fed during the toxicity test. LT50 values ±95 % confidence limits were calculated from plots of probit mortality against log time by the methods of Litchfield (1949).
Protein turnover
Rates of protein synthesis and degradation, protein content and total RNA level in gills were analyzed on days 9 and 18. One day before the start of the Zn2+-exposure (i.e. day -1), 20 fish from each group were lightly anaesthetized (MS 222), weighed to the nearest 0.01 g and labelled individually by freeze-branding (Mighell, 1969). Branded fish were put in a secluded compartment of each tank. The compartments were made from 8 l polypropylene buckets, each with a mesh bottom and lid. The buckets were submerged in the tanks and the incoming water was directed through the buckets into the tanks. Fish for protein turnover determination were fed daily and independently of the other fish in the tanks, at a rate of 1 % of their body mass (i.e. at the same rate as the rest of the fish). On day 9, all of the branded fish were quickly and gently dried and weighed. Ten fish from each group were then injected via the caudal vessels with 150 mmol l-1 phenylalanine, pH 7.5 (1 ml per 100 g fish), containing 3.7 MBq ml-1 of L-2,6-[3H]phenylalanine. Following injection, the fish were placed into individual 250 ml black beakers with a fitting lid and air-line, containing the same water as before. After 1 h, fish were killed by a blow to the head. The whole gill basket was immediately dissected out and frozen in liquid nitrogen. The samples were then stored at -70 °C for later analysis. On day 18, the remaining 10 fish from each group were weighed, injected with radioactive phenylalanine, and sampled as described above.
Gill content of protein and RNA and fractional rate of protein synthesis (Ks) in gills were analyzed as detailed in Houlihan et al. (1986) and Hogstrand et al. (1994). The protein degradation (Kd) rate was calculated by subtracting the fractional protein synthesis rate of the gills from the fractional growth rate of the gills multiplied by measured protein content.
Statistical methods
Owing to known temporal fluctuations in Ca2+ influx in juvenile rainbow trout (see Wagner et al. 1985; Hogstrand et al. 1994), statistical tests were only applied to samples taken within 3 day periods. Significant differences between the control and the Zn2+-exposed groups in Ca2+ influx, and the Km and Jmax values for Ca2+ influx, were evaluated by Student’s t-tests (two-tailed, unpaired). Statistical tests on these data were based on N=6 (six Ca2+ concentrations) per treatment, rather than N=48, because the eight fish in each test bag were not independent, but rather were exposed to identical water conditions in a single flux chamber. The Mann–Whitney U-test was used to detect differences between Zn2+-exposed fish and controls for plasma [Ca] and tissue levels of Cu, Zn, MT and total RNA, Zn2+ influx, growth and protein turnover rate. Groups were considered significantly different at P<0.05. Within a treatment group, the significance of trends over time was assessed by analysis of variance (ANOVA). The Tukey HSD test was used to extend the ANOVA with post-hoc comparisons of means. In the acute toxicity tests, the groups were considered to be statistically different if the 95 % confidence limits for the LT50 values did not overlap.
Results
Plasma Ca content was initially depressed by Zn2+ exposure (at 150 μg l-1=2.3 μmol l-1 as total Zn) but recovered within a week (Fig. 1). After 29 days of the experiment, the plasma [Ca] of exposed fish was significantly higher than that of the controls. There was also a slight time-dependent variation in plasma [Ca] within the control group; on day 17, the concentration of Ca in plasma was significantly lower than on day -2 (ANOVA, Tukey HSD, P<0.01). There were no differences in the whole-body Ca level that could be attributed to Zn2+ exposure (Table 1). However, there was a general increase in the whole-body Ca content of both groups over the experimental period (ANOVA, P<0.001).
There was a significant (ANOVA, P<0.05) increase in the Zn content of the gills of both groups of fish over the experimental period. Exposure to Zn2+ increased the Zn concentration in the soft tissue of the gills relative to control values, but this elevation in branchial Zn was only significant at the end of the experiment (day 29; Table 1). The whole-body content of Zn and branchial Cu levels remained unchanged throughout the experiment (Table 1). In the control group, the MT concentration of the gill was higher at the first sampling point on day -2 than in subsequent samplings (ANOVA, Tukey HSD, P<0.01). An increased level of branchial MT relative to controls was found after 15 and 30 days of Zn2+ exposure, indicating an increased capacity of the gills to immobilize Zn (Table 1). A 24 h sublethal challenge with Zn2+ (450 μg l-1=6.9 μmol l-1 as total Zn) always significantly decreased the plasma [Ca] in fish that had not been previously exposed to Zn2+ [Control (challenged); Table 2]. At day 1, the plasma [Ca] was also reduced in fish that had been pre-exposed to Zn2+ (150 μg l-1=2.3 μmol l-1 as total Zn), but not to the same extent as in controls. After 29 days of exposure, the pre-exposed fish were able to withstand the Zn2+ challenge without developing hypocalcaemia (Table 2).
In spite of the increased Ca2+ regulatory capability demonstrated by the sublethal Zn2+ challenge, the Zn2+-exposed fish did not acquire an increased tolerance to a lethal concentration of Zn2+ (Table 3). After 30 days of exposure to Zn2+ (at 150 μg l-1=2.3 μmol l-1 as total Zn), these fish had a lower LT50 when challenged with a lethal [Zn2+] (4 mg l-1=61 μmol l-1 as total Zn) than the controls. Before that, there was no significant difference in LT50 between the groups.
Ca2+ influx through the gills obeyed Michaelis–Menten kinetics throughout the exposure period (see Fig. 1 from Hogstrand et al. 1994, for a typical relationship). The most pronounced effects of chronic Zn2+ exposure were on the kinetic variables (Km, Jmax) of Ca2+ influx. Initially, exposure to Zn2+ caused a 50 % reduction in the apparent Jmax for Ca2+ influx (in the presence of Zn2+) on day 0 during the first 4 h of Zn2+ exposure; i.e. a loss of available Ca2+ uptake sites (Fig. 2). This reduction in Jmax persisted at 24 h (day 1) and was seen even when the treatment was withdrawn during the measurement of Ca2+ influx (in the absence of Zn2+); i.e. both apparent Jmax and true Jmax were reduced (Fig. 2). Both Jmax values were restored within a week and significantly elevated values of both apparent and true Jmax were recorded on day 15. At the end of the experiment, the Jmax values returned to the control level.
Acute exposure of ‘Zn2+-naive’ fish to Zn2+ (150 μg l-1=2.3 μmol l-1 as total Zn) on day 0 increased the apparent Km for unidirectional Ca2+ influx by a factor of three; i.e. a large decline in Ca2+ affinity (Fig. 3). However, the apparent Km returned to control values in fish that were exposed to Zn2+ for 24 h prior to the measurement of Ca2+ influx. 3 days later (day 4), the apparent Km was found to be 4.5 times initial control values. After this peak, the apparent Km declined and levelled out at a value about three times higher than that of the control for the remainder of the experiment. The true Km for Ca2+ influx, as measured in the absence of the inhibitor (Zn2+) in the water, decreased drastically after 24 h of exposure to Zn2+ from 50 to 0.75 μmol l-1; i.e. a large increase in Ca2+ affinity. This effect was still present 3 days later, but after day 4 the true Km increased and peaked on day 7. Later in the experiment, the true Km stabilized at a level significantly above the control value (Fig. 3).
What is important for the fish, during chronic sublethal exposure to Zn2+, is the actual influx of Ca2+at the [Ca2+] of the ambient medium (1.0 mmol l-1). The actual Ca2+ influx (Jin) in the presence of Zn (150 μg l-1=2.3 μmol l-1 as total Zn) was calculated from the Michaelis–Menten equation, using the measured values of apparent Jmax and Km (Fig. 4). The actual influx of Ca2+ (Fig. 4) closely followed the apparent Jmax (Fig. 2), showing an initial decrease, followed by an ‘overshoot’ and a subsequent return to control values.
The unidirectional branchial influx of Zn2+ was measured for both Zn2+-exposed fish and controls at the experimental concentration of Zn2+ (150 μg l-1=2.3 μmol l-1 as total Zn; Fig. 5). It was found that the influx of Zn2+ was only about 0.3 % of that of Ca2+ (compare with Fig. 4). The Zn2+ influx of the control group fluctuated during the experimental period and was higher on day 17 than on days -2 and 31 (ANOVA, Tukey HSD, P<0.01). On day 1, the Zn2+ influx of pre-exposed fish did not differ from the control value for day -2. However, on days 15 and 29, the Zn2+ influx of pre-exposed fish was found to be half that of the control on days 17 and 31, respectively. The accumulation of Zn in the soft tissue of the gills, during the 24 h flux period, is shown in Fig. 6. With the exception for day 15, when the accumulation of Zn in gills of pre-exposed fish was the same as in controls, the Zn accumulation followed the Zn2+ influx. The Zn accumulation in gills of control fish exhibited an upward trend during the experiment and was significantly higher on days 17 and 31 than on day -2 (ANOVA, Tukey HSD, P<0.01).
The growth of fish over the 1 month Zn2+ exposure, determined from the masses of the 48–96 fish randomly sampled at the same times as the Ca2+ kinetics measurements, was similar for both groups (Fig. 7). However, the Zn2+-exposed fish showed no increase in mass between days 7 and 15, with the consequence that Zn2+-exposed fish sampled on day 15 were smaller than the controls sampled on day 17.
The fractional protein synthesis rate in the gills, Ks, was significantly elevated in Zn2+-exposed fish on day 9, but after 18 days of exposure this difference was no longer present (Fig. 8A). The protein degradation rate of the gills, Kd, was approximately twice as high on day 9 as on day 18, and on day 9 the Zn2+-exposed fish had significantly higher Kd than the control fish (Fig. 8B). There were no differences in the RNA content of the gills between groups or over time, suggesting that the maximal capacity for protein synthesis was not affected by the Zn2+ exposure (Fig. 8C). On both sampling occasions (days 9 and 18), Zn2+-exposed fish had a lower branchial protein content than the control fish (Fig. 8D). The protein content within each group was not different between the sampling dates.
Discussion
We have previously shown that Zn2+ is a competitive inhibitor of branchial Ca2+ uptake in freshwater-adapted rainbow trout (Hogstrand et al. 1994). This finding was substantiated in the present study, which demonstrates very large changes in Km (Fig. 3) and smaller changes in Jmax. This study provides details on the nature of the effects of Zn2+ on the Ca2+ influx kinetics and on their consequences for Ca2+ and Zn2+ uptake. During the first days of Zn2+ exposure, there was a loss of available Ca2+ uptake sites (decreased apparent Jmax; Fig. 2) that was reflected in a reduced Jin for Ca2+ (Fig. 4) and a lowered plasma [Ca] (Fig. 1). The inhibition of Ca2+ transporting capacity was not ameliorated when Zn2+ was removed from the water (true Jmax), which suggests a persistent binding of Zn2+ to allosteric inhibitory sites and/or permanent inactivation of Ca2+ carriers. Within 7 days of the exposure, the number of functional uptake sites for Ca2+ was restored and the plasma [Ca] seemed to be normalized within the first few days. There was a dramatic increase in affinity of the Ca2+ carriers (decreased true Km) observed after 1 and 4 days of Zn2+ exposure (Fig. 3), a response that was also observed in our previous study (Hogstrand et al. 1994). We speculate that the increased affinity for Ca2+ is a compensatory response to the loss of functional uptake sites and to the competitive inhibition of Ca2+ influx by Zn2+. However, from day 7 and through the rest of the experiment, the true Km was increased compared with the control. This delayed decrease in affinity for Ca2+ might seem contradictory considering that the fish were already suffering a disturbance of Ca2+ homeostasis, but it may have served to limit the uptake of Zn2+. It has previously been found not only that Zn2+ competitively inhibits the branchial Ca2+ influx (Hogstrand et al. 1994) but also that Ca2+ is a competitive inhibitor of Zn2+ influx at the gills (Spry and Wood, 1989). This information suggests that Ca2+ and Zn2+ partially or fully share a branchial uptake route (Spry and Wood, 1989; Hogstrand et al. 1994). Therefore, a decreased affinity of the carriers for Ca2+ could serve to reduce the Zn2+ influx.
Most of the Ca2+ transporting capacity (Jmax) was restored during the first week of exposure, which would help the fish to maintain plasma [Ca] homeostasis even with a decreased affinity of the transporting sites. Furthermore, at the acclimation Ca2+ concentration of the water (1.0 mmol l-1), the influx of Ca2+ was largely determined by the Jmax (Figs 2 and 4) and the elevated Km had a negligible effect on Jin. Zn2+ influx, however, should have been greatly influenced by a change in the affinity of the carrier, because the total Zn concentration in the exposure tank (150 μg l-1=2.3 μmol l-1) was below the measured apparent Km for Zn2+ influx (240 μg l-1=3.7 μmol l-1) in identical conditions (Spry and Wood, 1989). Thus, we suggest that the initial response of the fish to Zn2+ is to compensate for the net Ca2+ loss by markedly increasing the affinity of the transporting sites. Once the Jmax for Ca2+ has been ameliorated, and therefore the Jin for Ca2+ has been restored, the system is tuned to limit the influx of Zn2+ rather than to improve the Ca2+ uptake further.
The measured changes in Zn2+ influx fit with the hypothesis that Ca2+ and Zn2+ share a common uptake route through the gills. After 1 day of Zn2+ exposure, the Zn2+ influx was not different from that of control fish (Fig. 5). According to the Ca2+ influx kinetics, the affinity of the carriers increased, but there was also a decrease in the number of available uptake sites, leaving the Zn2+ influx unchanged. During the following few days, the number of carriers was restored and the Zn2+ influx was, therefore, increased. This restoration of the number of uptake sites was followed by a decrease in affinity of the Ca2+ transporting system, which would explain the subsequent reduction in Zn2+ influx. A decreased Zn2+ influx is likely to be an important physiological adaptation to an elevated water Zn2+ concentration. In a previous study (Hogstrand et al. 1994), we reported that the branchial influx of Zn2+ was decreased in rainbow trout exposed to Zn2+ during a 2 month period. The present study provides experimental evidence that this reduction in Zn2+ influx is linked to a decreased affinity of a common Ca2+/Zn2+ transporting system, but final proof requires Zn2+ kinetic studies.
With the exception of day 15, the pattern of Zn accumulation in the gills, during the 24 h flux period, paralleled the Jin values for Zn2+ (Figs 5 and 6). The disagreement between the day 15 and day 29 results, as to the effect of Zn2+ exposure on Zn accumulation in the gills, prevents any firm conclusion determining whether the apical entry step or the basolateral transfer of Zn to the blood is altered. However, there was a scattered, but highly significant, linear relationship between the Zn2+ influx and the accumulation of Zn in gills (r=0.420, P<0.0001, N=160). We therefore speculate that the rate-limiting step for Zn2+ influx is the transport through the apical membrane and that the influx of Zn2+ may be regulated by changing the permeability of the apical membrane to Zn2+. The available evidence suggests that both Ca2+ and Cd2+ are transported through the gill epithelium by the chloride cells (Perry and Wood, 1985; Verbost et al. 1987, 1988, 1989; Perry and Flik, 1988; Flik et al. 1993; Perry et al. 1992; Marshall et al. 1992; McCormick et al. 1992). Both of these ions are thought to pass through the apical membrane of the chloride cells passively, through a voltage-independent Ca2+ channel, driven by their electrochemical gradients (Perry and Flik, 1988; Verbost et al. 1989). Furthermore, there is evidence that the influxes of Ca2+ and Cd2+ are subject to regulation by the hormone stanniocalcin at the level of the apical membrane (Lafeber et al. 1988; Verbost et al. 1989; Flik, 1990). In view of this information and the results from the present study, it seems quite possible that the branchial influx of Zn2+ can be regulated by a change in affinity of a common apical Ca2+/Zn2+ channel that is also permeable to Cd2+. More research is required to verify this hypothesis. Such experiments may include manipulation of the permeability of the apical membrane for Ca2+ and Zn2+ by lanthanum exposure and/or stanniocalcin treatment.
Both the present study and our previous investigation (Hogstrand et al. 1994) document temporal fluctuations in the kinetic variables for Ca2+ influx in untreated fish. Cyclic variations in branchial Ca2+ influx have been shown by Wagner et al. (1985, 1986), who also found a corresponding pattern of responsiveness to stanniocalcin. The effect of injected stanniocalcin was only found in the sequences of the cycle where Ca2+ influx was high (Wagner et al. 1986).
The Zn2+-exposed rainbow trout, in the present study, did not accumulate any Zn or lose any Ca on a whole-body basis (Table 1). During the later part of the experiment there was, however, a slight increase in the branchial Zn content that correlated with an induction of MT. At this point, the increased accumulation of Zn by the gills did not seem to impair the health of the fish, as judged by normal plasma [Ca], protein turnover rates and growth. The increased MT concentration of the gills presumably helped the fish to immobilize the potentially toxic Zn (Hogstrand and Haux, 1991). It is interesting to note that the whole-body content of Ca and the concentration of Zn in gill tissue increased in the control group during the experiment. Furthermore, the increasing branchial Zn content was paralleled by an elevated branchial accumulation of Zn in gill tissue on days 17 and 31 compared with the value on day -2 (Fig. 6). We speculate that these effects are associated with normal growth in juvenile rainbow trout.
The rainbow trout in the present study did not show an increased tolerance in the acute Zn2+ toxicity tests (Table 3), in contrast to our earlier study (Hogstrand et al. 1994). However, in that study the fish were larger (21 g versus 4 g), the experiment was carried out in a different season (winter versus late spring) and MT induction did not occur. McDonald and Wood (1993) introduced the concept of the ‘exposure window’ within which an increased tolerance to metals can be conferred. The concept is based on the observation that increased tolerance to metals is preceded by initial physical damage to the gill structure followed by a repair process. If the exposure level is too low to cause substantial morphological damage (e.g. oedema, inflammation, cell sloughing), no increased tolerance will develop. It is quite possible that the level of Zn2+ exposure (150 μg l-1=2.3 μmol l-1 as total Zn) chosen by Hogstrand et al. (1994) and also used here was borderline between the nutritional and toxicological range for Zn2+ and caused only minor damage. Spry et al. (1988) found that the same Zn2+ concentration in identical water conditions stimulated the growth of juvenile rainbow trout without any effects on whole-body Ca and Zn levels. Although the fish from the present study did not exhibit an increased tolerance to lethal concentrations of waterborne Zn2+ (4 mg l-1= 61 μmol l-1 as total Zn), they did show a physiological recovery in terms of restored plasma [Ca] (Fig. 1) and there was no loss in whole-body [Ca] (Table 1). Furthermore, when challenged with a sublethal concentration of Zn2+ (450 μg l-1= 6.9 μmol l-1 as total Zn) three times higher than the pre-exposure level, they showed a significantly greater ability to maintain plasma [Ca] than the controls (Table 2). The conclusion that can be drawn from these results is that the mechanisms for toxicity at acutely lethal concentrations of waterborne Zn2+ are different from the effects on the Ca2+ homeostasis that can be found when fish are exposed to waterborne Zn2+ at sublethal levels. Our conclusion is consistent with previous experiments which have established acute hypoxia as the cause of death during exposure of rainbow trout to Zn2+ at 40 mg l-1 (=610 μmol l-1 as total Zn) and 1.5 mg l-1 (=22 μmol l-1 as total Zn) (Skidmore, 1970; Spry and Wood, 1984), whereas hypocalcaemia and acid–base disturbances are the major causes of toxicity at lower concentrations (800 μg l-1=12 μmol l-1 as total Zn) of waterborne Zn2+ (Spry and Wood, 1984, 1985). The results underline the importance on conducting studies of potentially toxic metals at environmentally relevant concentrations.
On day 9 of the experiment, the protein synthesis rate of the gills was significantly higher in Zn2+-exposed fish than in controls (Fig. 8). We speculate that this enhanced protein synthesis rate reflected the repair mechanisms that were activated after the initial damage caused by Zn2+ exposure. The Ca2+ kinetics data support this explanation by indicating that there was a regeneration of Ca2+ carriers during this part of the experimental period. Furthermore, the protein degradation rate was elevated in Zn2+-exposed fish at the same time, reflecting the breakdown of damaged proteins. Associated with the faster protein turnover in the gills was, undoubtedly, an increased metabolic cost, which was likely to be the cause of the temporary attenuation of growth observed for Zn2+-exposed fish in the middle of the experimental period (Fig. 7). On day 18, there were no longer any differences in the rates of protein synthesis and degradation between the treatment groups. In accordance with this observation, the Zn2+-exposed fish were subsequently able to compensate for the lag in growth and, by the end of the experiment (day 29), fish from both groups were of similar masses. In a recent study, we found no increase in the fractional protein synthesis rates in gills, liver or whole body after 23 days of exposure to the Zn2+ concentration used here (Hogstrand et al. 1994). We speculated that no effects were seen because the analyses were performed late in the treatment, when the adaptation to Zn2+ had already been established. This corresponds well with the data obtained from the present study, where we show that adaptation to Zn2+ is initially associated with an elevated metabolic cost, but that the fish are able to compensate for their temporarily stalled growth rate within 1 month.
ACKNOWLEDGEMENTS
This study was supported by a DFO/NSERC Science Subvention Program Grant to C.M.W. and a PDF grant to C.H. from the Swedish Council for Forestry and Agricultural Research. The advice of Dr R. W. Wilson and skilful technical assistance of Mr R. S. Munger are gratefully appreciated.