Copper metabolism in actively growing rainbow trout (Oncorhynchus mykiss): interactions between dietary and waterborne copper uptake
1 McMaster University, Department of Biology, 1280 Main Street West, Hamilton, Ontario, Canada L8S 4K1 and
2 Department of Fisheries and Oceans, West Vancouver Laboratory, 4160 Marine Drive, British Columbia, Canada V7V 1N6
Present address: The August Krogh Institute, Zoophysiological Laboratory, University of Copenhagen, DK-2100 Copenhagen, Denmark
*Author for correspondence (e-mail: kamundcn{at}mcmaster.ca)
Accepted 8 October 2001
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Summary |
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Key words: Cu homeostasis, Cu deficiency, waterborne Cu uptake, dietary Cu uptake, gill, rainbow trout, Oncorhynchus mykiss.
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Introduction |
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Despite extensive studies (Harris, 1991; Linder and Hazegh-Azam, 1996
; Pena et al., 1999
), the exact mechanisms of Cu homeostasis in mammals are not well understood. Much less is known about Cu metabolism and regulation in fish, although in contaminated environments fish may take up Cu through both the gut and the gills (Dallinger et al., 1987
). Despite substantial literature pertaining to Cu uptake via either gills or gut (McDonald and Wood 1993
; Handy, 1996
), the interactions between the two routes of uptake are yet to be clearly determined. One study (Miller et al., 1993
) did examine this potential interaction in rainbow trout but started with the assumption that uptake from the water was zero at control (low) waterborne Cu levels of 79205 nmol l1, an assumption which is not substantiated by the present study. The assessment of Cu requirement in fish is much more complex than in mammals because of this potential for extra-intestinal Cu uptake via the gills, and the fact that Cu is ubiquitously present in the aquatic environment as a result of both natural and anthropogenic processes. While acknowledging a possible complication due to branchial Cu uptake, previous studies that have determined Cu requirements (Ogino and Yang, 1980
; Satoh et al., 1983
; Lorentzen et al., 1998
) failed to assess the potential contribution of waterborne Cu.
Although previous studies have independently assessed toxic effects (for reviews, see McDonald and Wood, 1993; Handy, 1996
) or nutritional requirements (Ogino and Yang, 1980
; Murai et al., 1981
; Satoh et al., 1983
; Lorentzen et al., 1998
), no study has simultaneously investigated Cu metabolism in states of experimental deficiency and sublethal loading in fish. In particular, the interaction of dietary and waterborne Cu uptake has yet to be unequivocally demonstrated, a finding that would allow the determination of the relative contributions of waterborne and dietary Cu in nutrition and toxicity.
This study was therefore conducted to investigate Cu metabolism during both Cu restriction and elevated levels of dietary Cu exposure in juvenile rainbow trout. Firstly, we set out to establish conditions under which Cu deficiency could be induced in fish and to determine whether they could obtain their Cu requirement from water. Secondly, we assessed the effects of Cu restriction and excess levels of dietary Cu exposure on growth and whole body and tissue Cu reserves. Thirdly, we used direct measurements of 64Cu fluxes to quantify waterborne Cu uptake in fish in which whole body Cu had been depleted or elevated, and were thereby able to separate quantitatively Cu uptake from diet and from water in order to determine their relative contributions. Finally, we assessed possible interactions between dietary and waterborne Cu uptake.
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Materials and methods |
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Cu-supplemented and Cu-deficient diets were prepared at the West Vancouver Laboratory, Department of Fisheries and Oceans, West Vancouver, British Columbia. The diet composition (Table 1) was based on known requirements for rainbow trout (NRC, 1993) and the only variable was the Cu content. This diet fulfilled the criteria necessary for diets intended for nutrient requirement studies (Baker, 1986
).
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Waterborne Cu uptake kinetics
The effect of the exposure conditions on waterborne Cu uptake kinetics by gills was assessed at week 7 (day 50) over a range of waterborne Cu concentrations. Each treatment was divided into five groups (N=9); each group was then exposed to waterborne 64Cu at a nominal total Cu concentration of either 31, 47, 79, 94 or 126 nmol l1. The radioisotope 64Cu (as CuNO3) was prepared at the McMaster University Nuclear Reactor. On the day of the experiment, 0.7 µCi l1 of 64Cu (specific activity 0.35 µCi µg1) was introduced into each experimental tank; the tanks had been pre-dosed with CuSO4·5H2O to bring the concentration to the nominal level. The radioisotope dosage administered added a total concentration of 3 nmol l1 (0.2 µg l1) Cu to the water, and therefore did not substantially elevate the water Cu concentration. The fish were then exposed to the 64Cu for 12 h under static water conditions. A 10 ml water sample was taken from each tank 15 min after introduction of 64Cu and again after 12 h. Over this period the water 64Cu activity and total Cu concentration changed by no more than 6.5 %.
Analysis
Cu concentrations in water, tissue, and food samples were determined by atomic absorption spectroscopy (AAS; Varian AA-1275 with GTA furnace atomizer) using a 10 µl injection volume and the operating conditions for Cu specified by the manufacturer. Certified Cu standards (National Research Council of Canada) run at the same time were within the specified range. Water samples were acidified (0.5 % nitric acid), while solid samples were weighed and digested overnight at 70°C with 6 volumes of 1 mol l1 nitric acid (Fisher Scientific, trace metal grade), and then centrifuged for 4 min at 13000 g. A subsample of the supernatant was diluted appropriately with 0.5 % nitric acid. For day 50, the tissues and water samples were first measured for 64Cu activity on a Canberra-Packard Minaxi Gamma counter with an on-board program for decay correction, and then analyzed as described above for determination of total Cu concentrations.
Calculations
Whole body total Cu concentration was calculated by dividing the sum of Cu contents (concentration multiplied by mass) of all the tissues plus the carcass by the sum of the masses of all the tissues plus carcass.
Whole body uptake of waterborne 64Cu was calculated by adding 64Cu activities (cts min1) in all tissues plus carcass. Fish masses were determined by summing up the masses of liver, gills, gut tissue (washed) and carcass for each fish. Whole body Cu uptake was then calculated from the formula
![]() | (1) |
where a is the 64Cu of fish (cts min1 g1), b is the 64Cu of water (cts min1 l1) and c is the total Cu concentration in the water (nmol l1). The uptake was then divided by the time of exposure (12 h) to convert it into a rate. The resulting values were rather small, hence they are reported as pmol g1 h1.
Specific growth rate (SGR) was calculated on a per tank basis for three growth periods of 2 or 3 weeks using the formula:
![]() | (2) |
Where m1=mass at beginning of growth period (g), m2=mass at end of growth period (g), t=duration of growth period in weeks).
Food conversion efficiency (FCE) was calculated on a per tank basis for growth periods 02, 24 and 47 weeks:
![]() | (3) |
To calculate true bioavailability of dietary Cu, we first estimated Cu uptake from water over 7 weeks by adjusting waterborne Cu uptake rates measured at the end of week 7 for size using the mean fish masses determined for weeks 02, 24 and 47 using the Cu uptake rate versus body mass relationship determined by Kamunde et al. (2001). It was assumed that all the Cu accruing from waterborne uptake was accumulated.
True bioavailability of dietary Cu (%), defined as the percentage retention of Cu ingested via diet after subtracting the accumulation that occurred by waterborne uptake, was then calculated as:
![]() | (4) |
where totCuf and totCu0 are whole body total Cu at the end and beginning of the experiment, respectively, Cuwater is the total Cu taken up from the water and Cudiet is the total Cu ingested with the diet over the experimental period. Visual observation during feeding showed that all the food was ingested. Thus, to calculate Cudiet, the total amount of food delivered (ration) and the Cu concentration of the food were used.
Relative contributions of dietary and waterborne Cu to the total body metal burden were calculated as:
![]() | (5) |
![]() | (6) |
The assumptions for this calculation were as for the bioavailability calculation (Equation 4).
Somatic indices for liver, gill and gut were calculated as:
![]() | (7) |
where x is wet mass of the organ or tissue of interest.
For gill the entire gill basket was used, whereas for the gut, gut contents and extraneous tissues such as fat were removed.
Statistical analysis
Data are presented as means ± S.E.M. (N). Effects of exposure conditions on growth, tissue Cu concentration and subsequent waterborne Cu uptake at each sampling point were assessed using a two-way analysis of variance (ANOVA) with time, diet and waterborne Cu concentrations as variables. Percentage data were subjected to arc-sin transformation prior to statistical testing. In all cases, significance was set at P<0.05. StudentNewmanKeuls pairwise multiple comparison procedure was used to make comparisons between measurements. One-way ANOVA and Bonferronis test were used to compare changes in FCE and SGR at P<0.05 and curve fitting for whole body Cu concentration patterns over time was done with Statistica 5.1 using individual data points by the Quasi-Newton estimation method.
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Results |
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Discussion |
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Whole body Cu status
Cu concentration data were expressed on a wet mass basis since a previous study (Shearer, 1984) on rainbow trout of varying body size showed that whole body wet mass concentrations are more useful for comparison of trace elements than dry mass concentrations. In fact, since there were no treatment-related or time-related effects on moisture content, the same trends would have been seen even if the data had been expressed on a dry mass basis.
Although an ideal biomarker of Cu status in mammals has yet to be identified (Milne, 1998), several indicators have been used by different authors to assess Cu nutritional status. These include growth, activities of cuproenzymes, and plasma Cu concentration (Baker, 1986
; Gatlin and Wilson, 1986
; Turnlund et al., 1997
, 1998
). Based on previous studies (Grosell et al., 1997
, 1998
, 2001
; Kamunde et al., 2001
), plasma Cu concentration cannot be used as a sensitive indicator of Cu status in fish since it is very tightly regulated during waterborne and dietary Cu exposure. In this study whole body and liver Cu concentrations were sensitive indicators of Cu exposure. Baker (1986
) pointed out that although growth data are in the long term the only defensible way to establish trace element requirement, the use of body stores also provides an important indicator in determining the nutrient requirement.
Whole body Cu concentration declined exponentially over time during deficiency, but increased linearly during exposure to high dietary Cu levels. Lauren and McDonald (1987) described a linear loss of whole body Cu after 28 days of exposure to high waterborne Cu levels. Although these authors used larger fish, there appear to be notable differences in the kinetics of elimination of abnormally high body Cu concentrations (depuration) (Lauren and McDonald, 1987
) and the decline of normal body Cu concentrations in the face of deficiency (as in the present study). For actively growing juvenile rainbow trout, simple growth dilution was evident and could account for most of the decline in whole body Cu concentration. Fish mass increased by approximately 250 %, while whole body Cu concentration declined by 60 % over the same period, almost exactly the percentage that would be expected by growth dilution. Furthermore, growth of all the organs and tissues sampled was well correlated with body mass, independent of treatment. It is notable that body mass accounted for 90, 96 and 99 % of the change in liver, gut and carcass mass, respectively (Table 3). Since these organs were the main Cu reservoirs, a change in Cu concentration in the whole body due to growth dilution would reflect the change seen in these tissues. Overall, the decline in whole body Cu concentration fitted a one-compartment model (simple negative exponential), and the increase in whole body concentration during dietary loading was linear, an indication that the latter is not a well-regulated phenomenon.
Interestingly, despite the decline in whole body Cu concentration in the deficient fish, all the groups had significantly higher Cu levels per fish at the end of exposure compared to the levels at the beginning of the experiment (Fig. 9). Total Cu content in fish on a high dietary Cu level and normal water increased 65-fold, whilst for fish on inadequate Cu via both routes, only a twofold increase occurred (Fig. 9). For the groups receiving normal Cu levels in the diet or water in combination or separately, the Cu content increased fivefold. This increase occurred in the absence of changes in whole body moisture content. Thus in all the treatment combinations the fish extracted Cu from the water and their diet, although the amount obtained by fish exposed to low diet and low water Cu levels was not adequate to meet normal growth requirements or the normal tissue concentration. Nonetheless, this observation illustrates that both the gill and gut Cu uptake mechanisms are highly efficient.
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The concentration of Cu in the liver strongly influenced whole body Cu content although the liver represented only 1.3-2 % of the body mass. At the beginning of the exposure approximately 20 % of the body Cu burden was in the liver. This proportion remained between 1030 % in all the groups except the group on a high dietary Cu level, which held 75 % of the body Cu in the liver by the end of the exposure. Chronic dietary Cu exposure is characterized by a continuous accumulation of Cu in the liver as seen in the present and previous studies (Handy, 1993; Kamunde et al., 2001
). Although we noted massive accumulation of Cu in the liver in this study, there was no indication of toxicity since the fish grew at the same rate as the controls.
Cu content of gut tissue was greatly elevated in fish exposed to high dietary Cu but appeared to level out over time, an indication that this tissue effectively regulates its internal Cu levels, as suggested in previous studies (Berntssen et al., 1999; Kamunde et al., 2001
). Furthermore, Cu build-up in gut tissue is diagnostic of dietary Cu exposure and does not occur during waterborne Cu exposure to any great extent (Kamunde et al., 2001
). A common trend with Cu uptake kinetics and accumulation in gut is that early in the exposure, a high proportion of the metal burden is held within the gut tissue but subsequently this is mobilized into other tissues. Later in the exposure, the gut tissue attains steady state despite continued exposure to elevated dietary Cu levels, suggesting that prolonged exposure stimulates clearance of Cu from the gut to other tissue, increases loss through faeces and mucosal exfoliation, or decreases absorption. Our data suggest stimulated Cu mobilization into other tissues, especially the liver, under these conditions.
During elevated levels of dietary Cu exposure in normal water, the gills accumulated significant amounts of Cu, in agreement with previous studies (Miller et al., 1993; Kamunde et al., 2001
), thus pointing to a potential role for the gills in Cu excretion. Although the changes in carcass Cu concentration during periods of Cu deficiency and exposure to elevated dietary Cu levels were small, the change in Cu content was enormous given the large mass that the carcass comprises. This compartment held the highest proportion of whole body Cu burden in all the groups except in the group receiving a high dietary Cu level, in which the liver was the dominant Cu reservoir.
Whole body waterborne Cu uptake
Copper uptake rates were measured after 7 weeks of continuous exposure to constant conditions of dietary and waterborne Cu, by which time any acclimation process would presumably be complete. Fish deprived of Cu in the water or diet together or separately had high uptake rates at the low waterborne Cu concentrations (<100 nmol l1 Cu), which increased dramatically above this concentration (Fig. 8). Two types of Cu binding sites, the high-affinity low-capacity binding sites, and the low-affinity high-capacity binding sites, have been recently described in trout gills (Taylor et al., 2000). These authors demonstrated saturation of the high-affinity low-capacity sites at <315 nmol l1 Cu, and recruitment of low-affinity, high-capacity sites above this concentration. In the present study, which measured transport rather than binding, saturation of the high-affinity sites appeared to occur at much lower water Cu concentrations. The generally higher uptake rate at a waterborne concentration of 126 nmol l1 may represent the point at which the low-affinity high-capacity sites start to be recruited. It appears that restriction of Cu in diet increases the capacity and affinity of both types of binding sites.
Uptake of waterborne Cu via gills has been studied mainly as it pertains to Cu toxicity (for a review, see McDonald and Wood, 1993), while a possible role for the gills in normal Cu metabolism has been largely disregarded. Gills play vital roles in gaseous exchange, acidbase balance, and ionoregulation; the present study suggests an additional, novel role of the gills in trace metal nutrition and homeostasis. We report for the first time that exposure of fish to conditions deficient in Cu causes an upregulation of branchial Cu uptake. Furthermore, there is reduced branchial uptake following pre-exposure to high dietary Cu (see also Kamunde et al., 2001
). Thus fish respond to different levels of dietary Cu by varying the rate of Cu absorption from water. This strategy may serve to minimize or prevent the development of Cu deficiency when intake is low and, conversely, Cu toxicity when intake is high, and indicate that Cu is under tight homeostatic control.
These observations possibly suggest the presence of a Cu transporter in the fish gills that responds to body Cu status. Mammalian studies have shown several specific P-type ATPases that serve for Cu transport, e.g. the Menkes and Wilsons proteins, and are involved in Cu homeostasis (Bingham et al., 1998; Roft and Hediger, 1999
). For fish, Campbell et al. (1999
) demonstrated vanadate-sensitive Cu transport (indicative of the involvement of a P-type ATPase) in perfused whole gills of rainbow trout, and Bury et al. (1999
) reported an ATP-dependent silver uptake by trout gill basolateral membrane vesicles. Silver can substitute for Cu in bacterial Cu-ATPase (Solioz and Odermatt, 1995
) and silver transport in rainbow trout gills could thus well be via a Cu-ATPase.
Interactions between dietary and waterborne Cu
Only a few studies have assessed the interaction between dietary and waterborne metal uptake in fish. Miller et al. (1993) argued that Cu assimilated from either route partitioned into functionally independent compartments in rainbow trout. Furthermore, using whole body Zn burden, Spry et al. (1988
) reported no interaction between dietary and waterborne Zn uptake in the same species. Both these studies based their conclusions, at least in part, however, on the assumption of zero uptake from their control water Cu (79205 nmol l1) and Zn (107 nmol l1) levels. The current data (Fig. 8) show that this is clearly not the case for Cu at least. Furthermore, these measurements revealed a marked interaction between dietary and waterborne Cu uptake geared toward maintaining Cu homeostasis during deficiency or excess Cu exposure.
We estimated the relative contribution of waterborne and dietary Cu uptake to the whole total body Cu load using measured waterborne Cu uptake rates, feeding rates and dietary Cu concentrations (see Materials and Methods). At low dietary Cu, water was clearly the main source of Cu, contributing 60 % of the total (Fig. 10A). With increasing dietary Cu, the contribution of dietary Cu increased whilst that of waterborne Cu decreased. In the group maintained on normal dietary and waterborne Cu, water contributed less than 10 % of the body Cu. At the highest dietary Cu concentration, diet was clearly the main source of Cu (99 %) and water contributed insignificant amounts to the total Cu burden. A previous study on relative contributions of waterborne and dietary Cu uptake to liver Cu concentration (Miller et al., 1993) showed increasing contribution of waterborne Cu uptake as waterborne Cu concentration increased.
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Murai et al. (1981) noted that the responsiveness of catfish to graded levels of dietary Cu was less pronounced than in most terrestrial animals and argued that Cu metabolism in catfish may have been affected by waterborne Cu. However, these authors did not provide any waterborne Cu uptake to support this insight. The present study not only provides this missing link (waterborne Cu uptake data) but ascribes to the gills a key role in normal Cu metabolism in fish. Branchial uptake contributed approximately 60 % of the body Cu load during deficiency, but diet was the preferred source of Cu under normal dietary and waterborne conditions, contributing more than 90 % of the body burden. These findings coupled with recent reports of branchial Cu excretion (Grosell et al., 2001
; Kamunde et al., 2001
) persuasively underline a key role of the gills in Cu homeostasis in fish and provide evidence of the gill as an organ of nutritional regulation.
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Acknowledgments |
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