P O S S I B L E M E C H A N I S M S OF A L U M I N U M TOXICITY IN A

D I L U T E , ACIDIC E N V I R O N M E N T TO F I N G E R L I N G S AND

OLDER LIFE STAGES OF S A L M O N I D S *

C H R I S T I N E M. N E V I L L E

Ministry of the Environment, Water Resources Branch, 125 Resources Road, Rexdale, Ontario M9W 5L1, Canada

and

P E T E R G. C. C A M P B E L L

Universitd du Qudbec, INRS-Eau, C.P. 7500, Sainte-Foy, Qudbec GIV 4C7, Canada

(Received May 9, 1988; revised October 17, 1988)

Abstract. The respiratory, acid-base, and ionoregulatory responses of juvenile rainbow trout (Salmo gairdneri) were monitored during exposure of the fish in the laboratory to inorganic A1 (2.8 rtM) over the pH range 4.0 to 6.5. Responses to A1 were most severe at pH 6.1 and 4.5, mortality being primarily due to asphyxia at pH 6.1 and to electrolyte loss at pH 4.5. Competition between the H+-ion and A1 for binding at the gill surface is offered as an explanation for the decreased toxicity of A1 at pH 4.0, one which is compatible with the free-ion toxicity model that has been developed for other metals. The physiologically distinct response of S. gairdneri to A1 at pH 6.1 is less amenable to unambiguous interpretation. If a mixed ligand hydroxo-At complex is incorporated in the free-ion model, and if it is assumed that the two A1 species, [A1-L-gill] and [HO-A1-L-giU], provoke distinct toxicological responses, then a bimodal toxicological response to A1 is indeed predicted. An alternative explanation of the apparent toxic action of A1 at pH 6.1, i.e., at pH values close to that of minimum AI solubility, is the precipitation of solid AI(OH)3 at the gill surface, i.e., a 'physical' effect rather than a biochemical one.

1. Introduction

Declining fish populations in soft-water aquatic environments exposed to acid precipi- tation have been reported in Europe and eastern N. America (Sevaldrud et al., 1980; Harvey et al., 1981; Haines, 1981; Haines and Baker, 1986). Aluminum, mobilized from surrounding soils and bedrock by the acid deposition, has been implicated as the toxic agent in short-term exposures leading to fish kills (Cronan and Schofield, 1979; Schofield and Trojnar, 1980; Dickson, 1983; Skogheim etal . , 1984; Haines, 1986).

However, the toxicological and physiological responses of juvenile and adult fish to acid and A1 exposure are varied, complex and confusing. They depend primarily on the pH and the AI, Ca, and dissolved organic C concentrations of the ambient water, but are also affected by other ion concentrations as well as temperature and exposure period. Moreover, sensitivity to acid and A1, and also to the ameliorating effects of Ca, varies among species as well as life stages. Ingersoll et al. (1985) have shown that in brook trout (Salvelinusfontinalis) , the least sensitive salmonid to acid (Leivestad et al.

* Correspondence concerning the physiological responses of fish to aluminum should be addressed to CMN; enquiries regarding the free-ion model of aluminum toxicity should be sent to PGCC.

Water, Air, and Soil Pollution 42:311-327, 1988. © 1988 Kluwer Academic Publishers. Printed in the Netherlands.

312 CHRISTINE M. NEVILLE AND PETER G. C. CAMPBELL

1976; Rosseland and Skogheim, 1984a), sensitivity to acid and A1 is reduced by increasing Ca concentrations in the case of yolk sac and swim-up fry but not for adults. In contrast, both life stages are affected in brown trout (Salmo trutta) (Muniz and Leivestad, 1980a, b; Brown, 1983; Dalziel et al., 1986). Literature reports suggest that at pH values in the 5.9 to 6.5 range the A1 either has no effect (Muniz and Leivestad, 1980a; Rosseland and Skogheim, 1984b; Fivelstad and Leivestad, 1984; Ingersoll et aL, 1985; Leivestad et aI., 1985; Neville, 1985a; Sadler and Lynam, 1987) or can be lethal (Grahn, 1980; Schofield and Trojnar, 1980; Dickson, 1983; Neville, 1985a). In more acidic environments the most severe effect, compared to acid exposure alone, has been found to occur at pH 5.0 with concentrations of A1 ~ 7.4 gM (Muniz and Leivestad, 1980a; Schofield and Trojnar, 1980; Haines, 1981; Fivelstad and Leivestad, 1984), but at pH 4.5 (Neville, 1985a) to 5.5 (Sadler and Lynam, 1987) with lower A1 concentrations (~ 2.8 gM). However at pH 4.0 A1 decreases the toxicity of H + (Muniz and Leivestad, 1980a; Schofield and Trojnar, 1980; Neville, 1985a).

To predict biological responses to A1 accurately in the field, it is clearly important to know the mechanism(s) of A1 toxicity. Both asphyxiation and impaired ionoregulation have been proposed as the major cause of mortality (Haines, 1981; Neville, 1985a). In either case, to elicit a biological response from the target organism, the A1 must obviously interact with a biological (gill) membrane. For other metals such as Cd, Cu, or Zn, this interaction with gill surface has been represented in terms of the formation of M-L-cell surface complexes (Pagenkopf, 1983),

M z+ + L-cell ~ M z+ - L-cell, (1)

where M = metal, z = cationic charge, L-cell = a cellular ligand, and K = the thermo- dynamic stability constant. The biological response is assumed to be proportional to the concentration of surface complexes, [M z+ - L-cell]

biological response oc [M z+ - L-cell] = K [L-cell] [M z+ ]. (2)

By analogy with this approach, which has proved successful in explaining the toxicity of metals to aquatic micro-organisms (Nelson et al., 1981; Petersen, 1982; Morel and Morel-Laurens, 1983; Sunda and Huntsman, 1983; Morel, 1984) and to a somewhat lesser extent to fish (Pagenkopf, 1983), one might expect [A13 + ] to be the best predictor of A1 toxicity. Indeed, complexation by organic ligands (e.g., citrate - Driscoll et al., 1980; humic acid - Van Coillie et al., 1983; Neville, 1985a) does reduce AI toxicity, and it has been suggested that only labile inorganic species of A1 are toxic (Driscoll et al., 1980; Baker and Schofield, 1980).

The functional groups present at the gill surface, and involved in A1 binding, will also interact with other cations, notably H + itself. Depending on the pK a values of the functional groups involved, changes in pH might thus affect the degree of protonation of the A1 binding sites

H + + L-cell ~ H + - L-cell. (3)

In such a case, either the H + ion will protect the functional groups from toxic metal

POSSIBLE MECHANISMS OF ALUMINUM TOXICITY 313

concentrations or vice versa, depending on their relative concentrations and binding site affinities. Both would be in competition with Ca.

In the present study, using data from experiments on the physiological responses of rainbow trout to inorganic A1 in soft water at different pH levels (Neville, 1985a), we test the adequacy of this simplified model of metal toxicity and propose a modified model better adapted to A1. The proposed mechanism is based on the respiratory, acid-base, and ionoregulatory responses of juvenile rainbow trout exposed to inorganic A1 over the pH range 4.0 to 6.5. The A1, Ca, and other major ion concentrations used in the exposure medium were similar to those found in the Muskoka-Haliburton district of Ontario during snowmelt. Experimental procedures and methods of exposure were as outlined elsewhere (Neville, 1985a); only a brief summary will be given here.

2. Materials and Methods

Juvenile rainbow trout (Salmo gairdneri, 197 to 469 g), obtained from a commercial fish hatchery and maintained in the Ministry of the Environment laboratories at Rexdale, Ontario, were acclimated to soft water (pH = 6.1 to 6.3, [Ca] = 0.05 to 0.125 mM) prepared by reverse osmosis from dechlorinated Toronto tap water, at 10 + 1 °C for 10 d. After cannulation of the dorsal aorta, they were placed in individual, clear plexiglass respirometers of 2 L capacity in a darkened room and allowed a 24-hr recovery period at pH 6.1 to 6.3 before being exposed to the test medium. Each

experiment consisted of 2 groups of 6 fish exposed to the same pH value. Aluminum was added to 1 group to give a final concentration of 2.8 gM. The second 'acid only' group served as a control for the effects of A1 in the 'acid + Al' group. The pH values tested were 6.5, 6.1, 5.5, 5.0, 4.5, and 4.0. All were repeated at least once except pH 6.5 and pH 5.5. To identify effects due to acid exposure alone, fish exposed to pH 6.5, 6.1, and 5.5 without A1 were used as controls for the experiments run at higher acidity (pH 5.0, 4.5, and 4.0).

A flow-through exposure system was used in most experiments. This was later

changed to a recirculation system when organic-A1 effects were investigated (runs 3, 5, and 7), since only a limited supply of natural humic water was available. In the flow-through system, pH adjusted water was divided equally between the two groups. Aluminum sulphate ( A 1 2 ( 8 0 4 ) 3 • 18 H20 ) was added to the head tank for the acid + A1 group from a 28 ~tM stock solution (prepared every 48 hr and stored in a glass container) using a peristaltic pump operating at a rate of 4 mL min - 1. The concentration of A1 added did not significantly alter the pH of the water. The solution was well mixed before flowing at a rate of approximately 3.2 L rain- 1 into flow splitters and then to the individual respirometers at 500 mL min- 1. Excess solution was used partly for chemical analysis and the rest was discarded. Molecular replacement time (99~o) of the respirometers was approximately 20 min. No significant differences were observed between A1 concentrations measured at the inflow and outflow of the respirometers (2.8 + 0.4 gM, measured as total extractable A1 - see below for details of A1 measure- ments).

314 CHRISTINE M. NEVILLE AND PETER G. C. CAMPBELL

The short water retention time in the respirometers was designed to ensure that total A1 was essentially equal to inorganic A1 since organic A1, formed from A1 ions combining with mucous and excretory products, would be minimal. Indeed, in a preliminary experiment, labile A1 concentrations proved to be essentially equal to total A1 concentra- tions. The equilibrium speciation of inorganic A1 was calculated with the MINEQL chemical equilibrium model (Westall et al., 1976), after revision of certain of the original thermodynamic data. The following stability constants were employed: AIOH 2÷ (log K = 9.01); AI(OH)~- (17.87); AI(OH)4 (33.84); A1F 2÷ (6.98); A1F~ (12.6); A1F ° (16.65); A1SO~ (3.2); AI(SO4) ~- (5.1); AI(OH)3(s ) (32.65). The MINEQL programme was run in either of two modes: allowing precipitation of AI(OH)3(s ) (as micro- crystalline gibbsite) or suppressing this reaction.

The recirculation system, with 90 ~o replacement of the 100 L exposure medium every 24 hr, was used for the acid compared to acid + A1 experiment at pH 5.5, and for additional duplicate experiments at pH 5.0 and 4.5 comparing the effect of exposure to 2.8 gM inorganic A1 with and without the presence of organic A1. Each exposure medium was prepared daily. For the acid + inorganic A1 groups, A1 was added to pH-adjusted soft water using a 0.5 mM stock A1 sulphate solution. For the groups exposed to acid + inorganic A1 + organic A1, water from an acidic stream in Muskoka, Ontario, (DOC = 15-20 mg L - l ) was diluted with soft water to give a dissolved organic acid concentration of 8 to 10 mg L-1, and the pH adjusted with H2SO 4. Sufficient A1 was then added from a 1.25 mM stock A1 sulphate solution to leave an inorganic A1 concentration of approximately 2.8 ~tM after A1 complexation by the organic acids. After preparation, the media were aerated well for 10 min to ensure adequate mixing and degassing, then pH readjusted if required, before being used in the exposure system. The pH was monitored in the flow splitter and readjusted as required throughout the exposure period. Both stock solutions used in these experiments were prepared at the beginning of each 6-d experiment and stored in glass containers. Inorganic A1 concentrations calculated at the beginning and end of the successive 24-hr exposure periods showed a decrease of approximately 0.4 gM. These losses were calculated as the difference between measured total and organic AI concentrations, and presumably were due to adsorption on living and inert surfaces and to complexation by fish excretory products.

Ventilation parameters (ventilation rate, amplitude, cough rate and 02 uptake) and degree of activity were measured during the first 1 to 2 hr of exposure to the test medium, and then daily throughout the experiment. Arterial blood samples for measuring acid- base and blood gas parameters (pH, total carbonate - for calculating Pco2 - haemoglobin, 0 2 content, 0 2 capacity - for calculating O 2 saturation - and lactate) and electrolytes (plasma chloride, Na, K, Ca, and Mg) were taken on the last day. The fish were then taken out of the respirometers and stunned before gill samples were taken for electron microscopy, epaxial muscle samples were taken for determination of water content and electrolyte levels, and gill, liver, kidney, gonad, and intestine samples were taken for measurement of A1 concentrations.

Aluminum speciation in the exposure media was measured by the Laboratory Services

POSSIBLE MECHANISMS OF ALUMINUM TOXICITY 315

and Applied Research Branch of the Ontario Ministry of the Environment. Labile A1 concentrations were determined by rapid extraction with an oxine/methylisobutyl ketone reagent (Barnes, 1975; LaZerte, 1984). Aluminium concentrations in the extracts were determined by graphite furnace atomic absorption spectrophotometry. For samples from the recirculation system, the chelation-extraction procedure was performed before and after the sample was treated with a cation exchange resin to remove the inorganic exchangeable A1 (modified from Driscoll, 1984). Labile inorganic A1 was calculated from the difference between these two measurements (total labile A1 - organic labile A1).

Aluminum concentrations in fish tissue were measured after acid digestion in a mixture of concentrated HCIO4 (1.5 mL) and HNO 3 (7 mL). Tissue samples were dried at 105 °C for 48 hr, weighed, heated with the acid mixture at 55 °C overnight (16 hr) and than at 95 ° C for 24 hr. Aluminum concentrations were measured in the digested samples by flame atomic absorption spectrophotometry. Aqueous samples of known A1 concentration and spiked tissue samples were included with each batch of samples. Spiked tissue samples were prepared by dividing one sample between two test tubes and adding a known amount of A1 to one of these tubes before the digestion procedure.

Complementary water quality parameters, including the major ions (Ca, Mg, Na, K, C1, SO 4, CO3) and several trace metals (Cd, Cr, Cu, Zn), were routinely determined by the Laboratory Sciences Branch, Ontario Ministry of the Environment.

3. Results

Gill A1 concentrations, as determined on the last day of exposure to 2.8 gM inorganic A1, were highest at pH 6.1 and 6.5, moderately high in the experiments run at pH 5.5 to 4.5, and quite low at pH 4.0.

Gill A1 concentration Exposure pH (~tmol. g- 1) n (d)

6.5 10.2 + 1.4 6 11 6.1 18.5 + 2.6 10 2-11 5.5 3.5 + 0.4 6 6 5.0 4.5 + 0.5 11 6-11 4.5 6.5 + 0.6 25 4-11 4.0 1.1 + 0.1 9 5-11 Control a 0.1 + 0.0015 17 6-11

a pH 6.5, 6.1, and 5.5 without A1.

Interpretation of this general trend in terms of gill-A1 interactions is not straight- forward, however, since (i) exposure times and ventilation rates and amplitudes (indicat- ing ventilation volume) varied among the various pH conditions and (ii) some of the A1 may have been cleared from the gill surface by coughing, the frequency of which also varied among the various pH conditions. Aluminum accumulation in the liver, kidney, gonad and intestine was not significantly different from control levels.

316 CHRISTINE M. NEVILLE AND PETER G. C. CAMPBELL

T E S T p H

6 . 5

I v NT,LAT,ONI 02 I N A • IRATE • AMPLITUD~ )I|IN ~RT~:IZlAI. BLO00 )

6.1

5.5

AZ~

~, 51GNIFICANT DIFFERENCE FROM THE ACID GROUP

• SZGNIFICANT DIFFERENCE FROM THE CONTROL

"~ C ~.US E OF D E/~,TH

~ClO P = PLASMA

~ C I D ÷ 2 , B ~ M Al inorg " M = MUSCLE

Fig. 1A. Fig, 1. Summary of the major effects on juvenile rainbow trout of 6 days exposure to acid (IN) or acid + 2.8 gM inorganic A1 ([]) (a) at pH 6.5, 6.1 and 5.5, (b) at pH 5.0, 4,5 and 4.0. Mean and standard error are compared to controls (100 ~o); control values correspond to the observations made on the fish exposed to pH 6.5, 6.1 and 5.5 in the absence of Al. Deviations above and below 100~o indicate an adverse response.

Data from Neville (1985a, b).

A summary of the major physiological responses is shown in Figures 1A and lB. (For responses of individual parameters , see Neville, 1985a.) The 100~/o level shows the

responses of control fish (pH 6.5, 6.1, and 5.5; absence of A1). Deviat ions above 100~ for ventilatory parameters , and below 100 ~o for the other parameters , indicate adverse effects; greater deviations indicate more severe effects. Responses to the inorganic A1 were mos t severe at p H 6.1 and 4.5, mortality being primarily due to asphyxia at p H 6.1

and to electrolyte loss at p H 4.5. The p H 5.5 and 5.0 A1 groups showed modera te responses and a transition between the major response mechanisms. No adverse

response to A1 was noted at the upper or lower limits to the studied p H range. At p H 6.5, apart f rom a slight increase in coughing and ventilation rate for two days during the exposure period, there was no significant difference in activity, ventilation, and physiological parameters from the controls, even though gill-A1 concentrat ions reached about half those at p H 6.1. At p H 4.0, the severity of the response to the H + ions was reduced in the presence of A1.

P O S S I B L E M E C H A N I S M S O F A L U M I N U M T O X I C I T Y 317

TEST pH

5 . 0

4 . 5

% I ~z;::~.2. ,or ,I,,. o~.~.:: ..... I

1 0 0 ~ - -

oe i; zx

H A ÷ ! C I - I SURVIVAL ( 60AYS EXPOSURt )

IkALX ~/s

AZX

4 . 0 ,oo~ i t- . . . . . -L- • • A A A A • A •

A SIGNIFICANT DIFFERENCE FROM THE ACID GROUP • SIGNIFICANT DIEFERENCE FROM THE CONTROL

T CAUSE OF OE,~,TH

ACID P = PLASMA

ACID + 2.8 /xM Alinorg ' M = MUSCLE

• DEATHS OCCURRED ON DAYS I end 2

0 DEATHS OCCURRED ON DAYS 4 , 5 and 6

e • O•

Fig. lB.

A diagram of the severity of the toxicological impact of 2.8 gM inorganic AI, as superimposed on the effect of the acid treatment alone, is shown in Figure 2.

W High i

Z ' o

I , I L

) . n r I--

I . u ~ , ,,,o, 0 Low

4 . 0 4.5 5.0 5.5 6 . 0 6.5

pH

Fig. 2. Diagram to show the toxicological impact of low concentrations of inorganic A1 (2.8 pM), as superimposed on that of inorganic acid alone: ( - - ) = overall response to A1 + acid; ( - - - ) = response to

acid alone; (---) = approximate response to inorganic A1.

318 CHRISTINE M. NEVILLE AND PETER G, C. CAMPBELL

4. Discussion

4.1. A P P A R E N T T O X I C A C T I O N O F L A B I L E I N O R G A N I C A L U M I N U M A T D I F F E R E N T pH L E V E L S I N L O W C A L C I U M E N V I R O N M E N T S

4.1.1. pH6.5-5.5

At pH 6.1 with 2.8 ~tM inorganic A1 in the flow-through system, there was no significant difference between inflow and outflow total A1 concentrations; therefore A1 was not lost from the exposure system except for that which adhered to the fish themselves (gill epithelium; skin). Neville (1985a) suggested that juvenile rainbow trout could clear the epithelium by coughing if the ambient A1 concentrations were low (~ 1.6 ~tM) or the exposure period short, but after 1 to 2 days exposure to moderate concentrations (~2.8 ~tM) at 10 °C the accumulation orAl became so great that 0 2 uptake across the gill epithelium was reduced despite increases in ventilation rate and amplitude. Arterial blood 02 saturation levels became so deplated (43 ~o - see Figure 1) that the fish died from asphyxia. Blood samples taken shortly before the fish became moribund showed that ionoregulation was not significantly impaired (moribund fish would have shown severe electrolyte losses). In the pH 6.5 and 5.5 groups with 2.8 gM inorganic A1 the amount of A1 that adhered to the gill surface was reduced. Higher A1 concentrations could probably have still asphyxiated the fish as outlined above.

4.1.2. pH < 5.5-4.5

At pH values less than 5.5 the response of the fish became further complicated by increases in permeability of the gill membranes. At pH 4.5 in soft water there appeared to be an increase in the passive influx of H + across the gill membranes into the fish (Neville, 1985a). With 2.8 gM inorganic A1 at the same pH, Neville (1985a) also found that the loss of Na and chloride ions from the fish was critically increased. By com- parison with known Na and chloride influx rates of resting, adult rainbow trout in soft water at pH 7.6 (McDonald et al., 1983), the losses of Na and chloride during the exposure period were too great to be attributed only to impaired active uptake sites, to breakdown of enzymes involved in the uptake mechanisms or to lack of 0 2 (since arterial blood 0 2 saturation levels were not significantly reduced). Hence, the loss of Na and chloride was unlikely to have been caused only by impaired active uptake mechanisms. Apparently in soft water (i.e., [Ca] less than ,,~ 0.1 raM), the combination of approximately 2.8 }aM inorganic Al and 30 gM H + (pH 4.5) was sufficient to increase gill membrane permeability to H +, Na, and chloride ions.

Thus, the mechanisms of toxic action of Al in soft water at pH values between 5.5 and 4.5 depend on the A1 concentration as well as the pH. As the pH decreases below pH 5.5 asphyxiation can only be caused by increasingly high concentrations of A1, whereas the concentration needed to increase membrane permeability decreases.

4.1.3. pH 4.0

At pH 4.0 with 2.8 gM inorganic A1 there appeared to be yet another change in the action of labile inorganic A1. The H + ions were then at such a high concentration

POSSIBLE MECHANISMS OF ALUMINUM TOXICITY 319

(~ 100 jaM) that they were more toxic than the A1 and, for a short time at least, the small amount of A1 that adhered to the gill tissue (one sixth of that in the pH 4.5 + A1 group) protected rather than impaired the functions of the gill epithelium. Evidence to support this interpretation is given by the severe ventilatory response of the pH 4.0 'acid-only' group of fish within 5 min of initial contact and the slightly less severe, slower response of the pH 4.0 + A1 group (Neville, 1985a). A ventilatory response can be initiated directly by low ambient or internal 0 2 tensions or indirectly by high ambient or internal CO2 tensions. However, the response cannot be initiated directly by H + ions themselves

(Neville, 1979a). Since the ambient O 2 and CO2 tensions remained at normal levels in the soft water exposure medium, and there was no other toxicant present in the acid-only group*, nor time for accumulation of metabolically produced H + within the fish, the ventilatory response must have been due to a secondary effect of the ambient H + ions. Most likely there was a rapid diffusive influx of H + through the gill epithelium into the blood plasma in the gill arterioles. Plasma bicarbonate would be converted to CO2 and the ventilatory response of the fish would be due to the resulting increase in CO 2 tension.

Working with juvenile rainbow trout, McDonald et al. (1983) have shown that initially there is a net influx of H + (or efflux of bicarbonate ions) in less than 1½ hr exposure to pH4.0 in decarbonated, hard water ([Ca] = 2.85mM) and soft water ([Ca] = 0.03 mM). Results from the Neville (1985a) study showed that the influx in soft water must take place in less than 5 rain (see above). In contrast, Neville (1979a) showed that there was no ventilatory response during 45 min exposure to pH 4.0 in decarbonated, moderately hard water ([Ca] = 1 mM). Thus, the rate of H + influx is reduced in the presence of higher concentrations of Ca, which decrease the permeability of membranes (McWilliams, 1983). Similarly, lower Na and chloride losses were seen in fish exposed to pH 4.0 in moderately hard water (Neville, 1979b) compared to the losses in soft water measured by McDonald et al. (1980) and indicated by the concentra- tions after only two days exposure in the pH 4.0 "acid-only" group (Neville, 1985a). It would appear that low concentrations of gill A1 at pH 4.0 (approximately 1.1 lamol g- 1 dry weight) may temporarily substitute for Ca and give some protection to the mem- branes by slightly decreasing membrane permeability, as suggested by Burrows (1977), even though high concentrations (6.5 lamol g- 1 dry wt.) at pH 4.5 have the opposite effect.

The low A1 concentrations in the gill tissues at pH 4.0 were not due to low ventilation volumes or to high cough rates since ventilation rate and amplitude were both high and cough rate was only slightly increased. Rather, they must have been a function of the greater ambient H + ion concentration.

4.2. RELATIONSHIPS BETWEEN THE TOXIC ACTION OF ALUMINUM AND ITS

SPECIATION IN SOLUTION

Given the observed bimodal physiological response of juvenile rainbow trout to A1 over the pH range 6.5 to 4, and the known sensitivity of A1 speciation to pH changes in this

* Measured concentrations of Cd ( < 0 . 2 g g L - t ) , Cu (6ggL-1) , and Zn (1 ggL 1) in the acid-only exposure media were welt below known toxic threshold values.

320 CHRISTINE M. NEVILLE AND PETER G. C. CAMPBELL

range, it is tempting to relate the toxic action of A1 to its speciation in solution (Fivelstad and Leivestad, 1984; Sadler and Lynam, 1987). Of obvious importance in this context is the solubility of aluminum in the various assay media:

p H Ca l cu l a t ed solubil i ty (IxM)

6.5 0.59

6.1 0.39

5.5 2.41

5.0, 4.5, 4.0 > 2 . 8 .

Equilibrium calculations indicate that for experiments in the pH range 6.5 to 5.5, the solutions were in fact oversaturated with respect to AI(OH)3(s), i.e., one would predict the formation of a solid precipitate (microcrystalline gibbsite). The precipitation of AI(OH)3(s) is notoriously slow, however, particularly at low degrees of oversaturation and in the absence of suitable nuclei (Hem, 1968; Turner, 1976). Indeed, in the flowthrough experiments, the total extractable A1 concentrations at the inlet and outlet of the exposure chambers were unchanged. In all experiments, extractable A1 concentra- tions were essentially equal to the nominal total A1 levels (2.8 gM), even at the higher pH values. This is clear evidence of the absence in the assay media of any appreciable particulate or colloidal AI(OH)3, or of any dissolved polynuclear species. Accordingly, the equilibrium distribution of inorganic A1 among its various hydroxo-, fluoro-, and sulfato-complexes has been calculated for a total A1 concentration of 2.8 gM, i.e., with the precipitation of AI(OH)3 suppressed (see Table I).

Examination of Table I reveals few surprises. As expected, at pH values > 6.1 the

T A B L E I

Equi l ib r ium spec ia t ion o f i no rgan i c a l u m i n i u m in the b i o a s s a y m e d i a

Rela t ive con t r i bu t i on ( % )

R u n p H C o n d i t i o n a A13 + A I O H 2 + A I ( 0 H ) ~ A I ( 0 H ) 4 A1F 2 + A1F2 + A 1 S 0 4 +

1 6.5 F T - - 9 8 9 - 1 -

2 6.1 F T - 4 34 54 4 3 -

3 5.5 R 4 12 26 3 38 15 -

4 5.0 F T 5 4 3 - 56 32 -

5 5.0 R 9 8 5 - 59 18 1

6 4.5 F T 33 9 2 - 50 3 3

7 4.5 R 8 2 - - 65 24 1

8 4.0 F T 21 2 - - 65 9 3 9 6.1 F T - 4 31 49 8 8 -

( low A1)

" Condi t ion : F T = f low- th rough sys tem; R = rec i r cu la t ing system.

POSSIBLE MECHANISMS OF ALUMINUM TOXICITY 321

predominant A1 species in the exposure media is predicted to be the aluminate ion, AI(OH)4 ; for runs at pH < 5.5 this species is negligible and the mono- and di-fluoro complexes assume considerable importance. There is an appreciable contribution from the free aquo ion, AP +, in this latter pH range, but its proportion never exceeds 33 To.

The total fluoride concentrations varied by a factor of about 2 among the nine different assays, as a result of varying initial concentrations in the reverse osmosis water used to make up the exposure media. Of particular interest is the fortuitous difference between the two runs performed at pH 4.5; in the flow-through system fluoride concentrations were markedly lower (1.8 ~tM) than in the recirculating system (4.2 pM), and the resulting A1 speciation was correspondingly different (Table I: compare runs 6 and 7). However, despite the fact that the calculated proportion of free A13 + + hydroxo-A1 complexes in run # 6 (44~o) was four times higher than in run # 7 (10~o), no notable differences were noted in the physiological responses exhibited by the test organisms (Neville, 1985a). Moreover, despite similar proportions of fluoro-A1 complexes, major physiological differences were seen at pH 5.0, 4.5, and 4.0.

It would thus appear that pH-induced changes in A1 speciation in the bulk solution, though surely a contributing factor, are insufficient in themselves to explain the bimodal response of S. gairdneri to AI over the pH range 6.5 to 4.0. A more promising approach is to couple these changes to the pH-induced changes that occur at the gill-solution interface (see below).

4.3. SIMULATIONS OF A1-Gill INTERACTIONS

Possible interactions of A1 with the gill surface have been studied with a chemical equilibrium model (MINEQL). By analogy with the approach pioneered by Pagenkopf (1983), we have treated the metal-binding sites at the gill surface as dissolved ligands present at low concentration, i.e. with A1 present in excess of the ligands. It should be emphasized that these 'simulations' remain essentially qualitative in nature. Their purpose was not to represent metal-gill surface interactions in all their intricacy, but rather to illustrate trends that occur with changing pH for (i) the simplest formulation of the free-ion model; (ii) the free-ion model adapted to include H + -gill surface inter- actions; and (iii) the free-ion model modified to include these latter interactions and the formation of mixed ligand hydroxo-A1 complexes [HO-AI-L-gill]. It should be noted that the free-ion model only considers metal-gill surface interactions and relates these to overall metal toxicity (Pagenkopf, 1983); it cannot be expected to predict the post- uptake fate of the metal or the physiological response of the organism to metal stress. Note too that the toxic effect of the H ÷-ion per se is ignored.

In addition to the solution equilibria involving A1 and its hydroxo- and fluoro- complexes (cf. Table I), we have considered one or more of the following reactions (charges on the L-gill species, and neutral water molecules bound to A1, have been neglected for simplicity).

H + + L-gill ~ H-L-gill K 4 , (4)

2 H + + L-gill ~ H2-L-gill K 5 , (5)

322 C H R I S T I N E M, N E V I L L E A N D P E T E R G. C. C A M P B E L L

A13 +

Ca 2+ + L-gill ~ Ca-L-gill K 6 , (6)

A13+ + L-gill ~ A1-L-gill K 7 , (7)

+ O H - + L-gill ~ HO-AI-L-giU K s . (8)

Values assigned to the various equilibrium constants (Table II) were chosen on the basis

of available experimental data (K 4, K 5, K 6 - Pagenkopf, 1983), by analogy with in vitro

A1 complexation reactions (K7 - Smith and Martell, 1976), and finally by iteration (Ks).

TABLE II

Summary of the equilibrium constants used in the simulations of Al-gill interactions ~

logK

Run K 4 K 5 g 6 K 7 K 8

1 - - 3 . 7 8 -

2 4.5 4.6 3.7 8 - 3 4.5 4.6 3.7 8 1,6

a In all runs [A1] = 2.8 gM; [Ca] = 90 taM; [F] = 4gM; [L-gill] = 0.1 gM; ionic strength = 1 mM; pH variable.

The free-ion model in its simplest formulation involves only reaction (7), the toxic response to A1 being directly proportional to the concentration of [A1-L-gill]; simu- lations with this model indicate a steadly increasing concentration of [A1-L-gill] as the

pH is decreased from 7 to 4, i.e., as the competition afforded by the hydrolysis of AI in solution (A13 + + n O H - = AI(OH)n ) decreases (Figure 3A). This gradual increase

in toxicity with decreasing pH does not conform to the toxicological observations

discussed previously, nor to those described in the literature and mentioned in the Introduction, which show maximum A1 toxicity at one or more intermediate pH values.

An obvious improvement to this initial formulation of the free-ion model would be

to take into account the interactions of the H + ion itself with the gill surface, i.e., reactions (4) and (5) (cf. Pagenkopf, 1983; Campbell and Stokes, 1985). I f the functional

groups involved in metal binding at the gill surface are assigned pKa values of ca. 4.5, then the calculated concentration of [A1-L-gill] reaches a maximum value at pH ~ 5.0 and decreases at higher or lower pH values (Figure 3B). By varying the importance of the H + -ion competition, i.e., by adjusting the pK~ values, one can obviously shift this maximum to higher or lower pH values. This prediction is clearly closer to the experi-

mental results than was that of the initial simulations. However, the model does not yet predict the bimodal response to A1 that is evident in the present study, i.e., the marked toxicity of A1 at pH 6.1.

The final simulations were performed for a system including the equilibria discussed above and, in addition, the formation of a mixed [HO-AI-L-giI1] complex, reaction

POSSIBLE MECHANISMS OF ALUMINUM TOXICITY 323

tOO

z IE :~ 50 _1

£ 0

Z

O IX) i00

50

la.I I '- ct) 0 J J (..9 I.I. I00 0 Z 0

50 er" 0

0 I~ 0 Q.

%

i I X ~ " ~ 5 6

% ® 0~•-0-0-0-0~ O/O/ O~'O\

\ ] \

I I

4 5 6 %

"-[AI-L- c Jilt ] ee @

/I// "t . e

\• \

[.o-Ai-L-~i,] o~,'~.

4 5 6

pH

Fig. 3. Equilibrium simulations of Al-gill interactions: (A) free-ion model; (B) free-ion model adapted to include H + -gill interactions; (C) free-ion model adapted to include H + -gill interactions and the formation

of a mixed ligand hydroxo-A1 complex. See Table II for a summary of the input data.

The existence of mixed hydroxo species, HO-A1-L, L ' , has been postulated for/n vitro

systems containing A1 and such ligands as salicylate, pyrocatechol and 1,2-dihydroxy- naphthalene-4-sulfonate (Marklund et al., 1986), and it seems reasonable to consider their possible existence at biological interfaces. Their formation can be represented as in reaction (8) or, perhaPs more realistically, as the loss of a H + from one of the water molecules in the hydration sphere of the [A1-L-gill] complex,

(H20)nAl-L-gill ~ ( H z O ) n _ 1HO-A1-L-gill + H + K 9 (9)

both representations being thermodynamically equivalent. If the equilibrium constant for this hydrolysis reaction is assigned a value of 10 - 6, simulations over the pH range 7 to 4 show both the expected bell-shaped curve for [A1-L-gill], as before, and a similar curve for [HO-A1-L-gill] displaced about 1.2 pH units higher (pH 6) (Figure 3C). If these two A1 species, [A1-L-gill] and [HO-A1-L-gill], were to have different toxicologi- cal effects at the gill surface, related perhaps to the different formal charges on the central

324 CHRISTINE M. NEVILLE AND PETER G. C. CAMPBELL

A1 cation, then a bimodal toxicological response to A1 would be predicted over the pH range of interest. However, if the formation of both species involves the displacement of Ca 2 + - which is reasonable to expect if they both involve the same ligand, L-gill - then they might both be expected to upset osmoregulation. At pH 6. I osmoregulation was almost normal.

An alternative, and perhaps more likely, explanation for the marked toxicity of A1 in the experiments run at pH 6.1, and the different physiological response at this pH, would be to invoke precipitation ofAl(OH)3(s ) at the gill surface (Schofield and Trojnar, 1980) with attendant effects on gas exchange, i.e., a 'physical' effect rather than a biochemical one*. Indeed, the maximum concentrations of A1 associated with the gills are noted at this pH, which coincides rather closely with the maximum degree of oversaturation in the bulk solution. As has been mentioned previously, there was no analytical evidence for the formation of colloidal or particulate hydroxy-Al species in the exposure medium, but this does not preclude the possibility that such precipitation may have occurred at the gill surface, given suitable sites for nucleation and an appreciable degree of over- saturation. Sadler and Lynam (1987) have suggested that transitional forms of A1, formed following rapid changes in pH, may be particularly toxic.

In this latter context, it would be helpful to know the effective pH close to the gill surface. The need for frequent addition of acid to maintain a constant pH when fish are exposed to an acidic medium (pH < ,~ 6.3) is evidence that the fish do undergo a net loss of base under such conditions. The loss is unlikely to be via the kidney since urine excretion is minimal at pH 4.3, and is slightly acidic (McDonald, 1983). It cannot be due to faecal contamination or increased mucous production as the fish are unfed during the exposure period and the loss of base occurs within a few hours of acid exposure, before there is any evidence of increased mucous. McDonald (1983) also calculated a net influx of H + -ions (or loss of HCO3- ) across the gills of acid-exposed rainbow trout. Therefore, in an acidic environment the pH at the gill epithelial surface in juvenile and adult fish is likely to be higher than in the ambient water. Consequently, A1 speciation close to the gill surface, and the degree of supersaturation with respect to AI(OH)3(s ), would also be different from that in the ambient water. Similarly, with regard to the severe physiological response to A1 at an ambient pH of 4.5 (Figure 2), it follows that the effective pH close to the gill surface may well have been somewhat higher; such a shift, i.e., to a slightly higher pH, would improve the agreement with the calculated pH response (Figure 3).

* This mechanism for A1 toxicity has been discounted by Fivelstad and Leivestad (1984), who successfully reared Atlantic Salmon parr at pH 6.5 in tanks fed by an acidic natural water that was neutralized but not filtered prior to use. Total A1 concentrations in the raw water supply varied from 2.2 to 14.8 laM and in separate experiments nearly 50~o of this A1 could be removed by filtering (0.05 gm) the water after neutralization. The authors concluded that it was 'highly unlikely' that this particulate AI was toxic to fish. It should be noted however that their reference water had a high humic acid content; the 'brown, metallohumic precipitate' formed on neutralization was clearly different from the uncoated inorganic colloids discussed here.

POSSIBLE MECHANISMS OF ALUMINUM TOXICITY 325

5. Conclusions

In the recent literature one finds frequent experimental observations that the toxicity of A1 to fish and other aquatic biota is greatest at 'intermediate' pH values, i.e. in the pH range 5.5 to 4.5 (Muniz and Leivestad, 1980a; Schofield and Trojnar, 1980; Haines, 1981; Helliwell et al., 1983; Fivelstad and Leivestad, 1984; Sadler and Lynam, 1987). There has been a tendency to interpret this response as an indication that one or more of the hydroxo-A1 complexes, AI(OH)n with n = 1 or 2, is the 'bioavailable form' or 'toxic species' of A1 (Helliwell et al., 1983; Fivelstad and Leivestad, 1984; Sadler and Lynam, 1987). Such an interpretation is clearly incompatible with current equilibrium models of metal ion toxicity (Morel, 1984). Since equilibrium among the various monomeric A1 species (A13+, A1OH 2+, AI(OH)~ . . . . ) is rapid (Baes and Mesmer, 1976; Plankey et al., 1986), and since they are all charged, hydrophilic species unlikely to pass biological membranes by simple diffusion, no one form can be considered more

or less available for equilibration with binding sites on the membrane. As demonstrated in the present paper, competition between the H + -ion and AI for binding at the gill surface is a more reasonable explanation for the decreased toxicity of A1 at low pH values, and one which is compatible with the free-ion toxicity model that has been developed for other metals.

The physiologically distinct response of Salmo gairdneri to A1 at pH 6.1 is less amenable to unambiguous interpretation. If a mixed ligand hydroxo-A1 complex is incorporated in the free-ion model, and if it is further assumed that the two A1 species, [A1-L-gill] and [HO-A1-L-gill], provoke distinct toxicological responses, then a bimodal toxicological response to A1 is indeed predicted. An alternative explanation of the apparent toxic action of A1 at pH 6.1, i.e., at pH values close to that of minimum A1 solubility, is the precipitation of solid AI(OH)3 at the gill surface, i.e., a 'physical' effect rather than a biochemical one.

Since the solubility of A1 trihydroxide is minimal at pH values close to 6.1, it might be argued that the combination of a toxic concentration of A1 (2.8 ~tM) and a pH of 6.1 is not environmentally relevant, i.e., should not occur in nature. However, precipi- tation reactions of A1 hydroxide in water are notoriously slow (Hem, 1968; Turner, 1976) and it is not uncommon to encounter such conditions in the field. An obvious example would be a (large) water body of pH ~ 6 that receives acidic, Al-rich inputs from small streams draining its watershed, as described by Wright and Skogheim (1983) and Neville (1985a). Similar conditions have been reported for a formerly acid lake that was subjected to annual liming and exhibited fish kills during the post-liming periods (Dickson, 1983).

Finally, the limitations of the free-ion model for metal toxicity should be stated clearly. Predictions of the metabolic fate of the metal and the detailed physiological response of the target organism are clearly beyond the capacity of the chemical equilibrium approach. In addition, the model limits the role of the proton to that of a competitive inhibitor of metal binding; acidification is considered only to affect metal binding at the gill surface, and the possible toxic effect of the H + -ion itself is ignored.

326 CHRISTINE M, NEVILLE AND PETER G. C. CAMPBELL

Acknowledgments

This s tudy was suppor ted by the Onta r io Min i s t ry o f the E n v i r o n m e n t ( C M N ) and in

par t by grants f rom the Na t u r a l Sciences and Engineer ing Resea rch Counc i l of C a n a d a

and f rom Fisheries and Oceans C a n a d a ( P G C C ) . The au thors wish to t h a n k M. R e n d a z

and A. Pa ren t for their help with the figures.

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