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Distribution, cycling and mean residence time of 226Ra, 210Pb and 210Po in the Tagus estuary
The Science of the Total Environment 196 (1997) 151-161
Distribution,
cycling and mean residence time of 226Ra, 210Pb and 210Poin the Tagus estuary Fernando
Direwoo
Gem1 do Ambiente,
Departamento
de Proteccao
P. Carvalho* e Seguranca
Radiologica,
EN IO, P-2685
Sacariem,
Portugal
Received 14 August 1996; revised 22 November 1996; accepted 29 November 1996
Abstract Resultsfor dissolvedand particulate “6Ra, ““Pb and 2’0Poin the Tagus river, estuary and coastalseasystemshow different distribution and chemicalbehaviour patterns for theseradionuclidesin the three aquatic environments. 22hRafrom riverborne particlesdissolvesin the estuary and contributes to increasedconcentrationsof dissolved226Ra in estuarinewater. In the estuary, dissolved2’oPb and 210Pofrom river dischargeand atmosphericdeposition are scavengedby suspendedmatter, which in turn becomesenriched in thesenuclidesin comparisonwith riverborne particles. As a result of these processes,the estuarine water flowing into the coastal sea contains enhanced concentrationsof dissolved“6Ra, but is depletedin dissolved“‘Pb and 2’oPo. Under averageriver flow conditions, massbalancecalculations for dissolved“‘PO and 2’0Pb in the estuary allowed their mean residencetimes to be estimatedas 18 and 30 days, respectively. Due to the rapid sorption of theseradionuclideson to settling particles, bottom sedimentsin the estuary representa sink for 210Pband “‘PO from both natural sourcesand industrial waste releases.Resultsalso suggestthat partial re-dissolutionof theseradionuclidesfrom bottom sedimentsand intertidal mudflatsis likely to occur in the mid- and low-estuaryzones.Nevertheless,box-model computationsindicate that the dischargeof “‘Pb and 2’oPointo the coastal seatakes place mainly with the transport of sediment,whereasthe dischargein the dissolvedfraction can only account for one third of the activities entering the estuary in the soluble phase. Implications of these results to the cycling of radionuclidesin phosphate waste releasesinto estuarine environmentsare discussed.0 1997Elsevier ScienceB.V. Kqmwds:
Estuaries;Radium; Polonium; Radioactive lead; Distribution coefficients;Phosphatewastes
1. Introduction * Present address: International Atomic Energy Agency, Marine Environment Laboratory. P.O. Box 800, MC 98012 Monaco Cedex. Principaute de Monaco.
Radium-226 ( Tljz = 1622 a), lead-2 10 (z-,/Z = 22.3 a), and polonium-210 (T,,,= 138.4 d), are
O048-9697;971$17.00 8 1997 Elsevier Science B.V. All rights reserved. PI/ SOO48-9697(96)05416-2
152
F. P. Cardho
I) The Science
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naturally occurring radionuclides of the uranium series and originate in the earth’s crust (Osmond and Ivanovich, 1992). Although ubiquitous in the environment, the concentration of these radionuclides in river-estuarine systems may vary widely due to the geology of the watershed and chemical weathering conditions. Furthermore, natural levels of these radionuelides in aquatic ecosystems may be increased by industrial waste releases. This is the case for non-nuclear industries, such as phosphate ore processing, whose wastes contain relatively high concentrations of 226~~, “‘Pb and 210Po (Koster et al., 1992; Carvalho, 1995a). The environmental cycling of these radionuclides, especiallv of ““PO, is of the utmost interest due to their high contribution to the radiation dose received by humans through the diet (Carvalho, 1995b). The fate of radionuclides in phosphate wastes discharged into estuaries and coastal zones has recently been investigated in several European countries (Germain et al., 1992; Koster et al., 1992; Perianez and Garcia-Leon, 1993; Carvalho, 1995a: Poole et al., 1995). However, a major difficulty encountered in predicting the fate of ‘26Ra, “‘Pb and “*PO from phosphate waste discharges is the paucity of data on the biogeochemistry of these radionuclides in estuaries. Therefore, a deeper understanding of the distribution and behaviour of these radionuelides is necessary for the interpretation of enhanced radioactivity levels as well as for the prediction of radionuclide fluxes and sinks in estuarine and coastal ecosystems. Here we present distribution data for 226Ra, ‘“‘Pb and “‘PO in the lower reach of the Tagus river and in the Tagus estuary, Portugal. Mass balances of “OPb and 2”)Po are calculated in order to determine the mean residence time of these radionuclides in the soluble phase and to estimate ““Pb and “‘PO retention in the estuary. This is achieved through the assessment of ““Pb and ““PO inputs, combined with inventories of dissolved zzhRa. “‘Pb and “(‘PO in the CL “f1Ll;il-C
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2. Methods 2.1. Study area
The river Tagus drains a catchment area of 86 000 km’, with its lower reaches flowing across a sedimentary flood-plain and ending in an extended fully-mixed type estuary (Fig. 1). The Tagus estuary covers an area of about 300 km2, and is composed of a 2-4 m deep and broad inner bay and a 40 m deep and narrow channel which connects to the sea. The shallow inner bay displays a complex bottom topography with tidal mud flat areas, sand islands and beds of dead oyster shells (Vale and Sundby, 1987). In the inner bay the tidal currents are weaker and allow the deposition of fine grain sediments (silt) at rates up to 3 cm a - ’ in some areas. The narrow sea entrance of the estuary has consolidated sandy gravel bottoms and, due to the higher velocity of tidal currents, little deposition of fine sediments occurs (Castanheiro, 1982). At the upper estuary, in the low salinity zone, the turbidity is usually the highest and attains 1.5 g 1- ’ . The main input of freshwater to the estuary is the discharge of the Tagus river, 400 m3 s - ‘. Other small rivers discharging into the estuary, like the Sorraia with 30 m3 ss’, provide little contribution to the fresh-
Fig. I. Catchment basin of the Tagus stations (RI R3) in the freshwater zone.
river
and
sampling
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Curcalito
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water inflow (7%) of Tagus discharge). Annually, the Tagus discharges about 4 x 10’ tons of suspended sediment into the estuary. Due to the small water depth and extended surface of the inner bay, tidal currents and winds intensively mix the estuarine water and re-suspend bottom sediments. The amount of sediment re-suspended and re-deposited during a fortnightly tidal cycle may attain the equivalent to the Tagus annual sediment discharge (Vale, 1986; Vale and Sundby, 1987). Furthermore, the estuary receives wastes from anthropogenie activities which carry contaminants into the system. Domestic sewage and major industrial waste outfalls (petrochemistry, fertilizers. smelters, etc.) are located on the south bank of the estuary, although several important point sources are located on the north bank (Barros, 1986). 2.2. Suwpling
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Fig. 2. Estuary of the River Tagus and sampling stations estuary (El -ElO) and adjacent coastal sea (Cl).
153
in the
und analvtical procedures
Water samples for 2’6Ra. ?“‘Pb and ““PO analyses were collected (December 1989) under average river flow conditions in the lower reach of the Tagus river, in the Tagus estuary and in the coastal sea. In the estuary, samples were collected in the mainstream channels in order to reflect the fresh-estuarine water flow better and to avoid, as much as possible, the influence of waste outfalls (Figs. 1 and 2). A total of 30&60 1 water samples were filtered in situ, using a portable generator and pumping-filtering systems, through pre-weighed 140 mm diameter, 0.45 pm pore size Sartorius membrane filters. Filtered water samples were acidified with HNO, to pH < 2 and a 209Po isotopic tracer and stable lead carrier (20 mg) were added immediately. In the laboratory, after addition of 224Ra to water samples, the analysis of dissolved radionuclides was made following co-precipitation with MnO, and sequential separations on anion and cation exchange resins (Bojanowski et al., 1983). Filters were oven dried at 80°C and weighed for determination of the suspended matter load. Samples of suspended matter, after addition of 209Po and 224Ra tracers
and Pb carrier, were dissolved and analyzed using the same procedure as for water samples. PO isotopes were plated onto a Ag disc for n-spectrometry (Carvalho, 1995a,b). ““Pb was precipitated as lead chromate, filtered and the chemical yield determined gravimetritally. Filters with the precipitate were stored for 6-12 months and, after a second addition of 2osPo tracer and dissolution of chromate, “OPb was determined through the measurement of “‘PO ingrowth from the radioactive lead decay. 226Ra was determined using an “4Ra isotopic tracer. The 224Ra tracer was separated from 228Th just before use and added to the samples in activity amounts significantly greater than the natural 224Ra levels. Following the electroplating of radium isotopes onto a stainless steel disc, immediate tl counting of 2’4Ra allowed the radiochemical yield to be determined. Measurement of 226Ra was performed on the same disc 1 month later to allow for ‘14Ra decay. Radioactivity measurements were made using silicon surface barrier detectors (400 mm’, R type, EG and G ORTEC) connected to a multichannel analyzer.
38”40’ w 39”50’ w
Concentrations
Ocean, -10”30’ Ocean, -10”50’
36
N
water water water water
-
~ -
(25)
(57) (62) (44) (24) (7)
1.26 k 0.06 0.56 + 0.03 0.52 k 0.03 2.56kO.06 0.84 & 0.04 0.96 + 0.09 0.43 * 0.02 0.56 + 0.06 0.19 f 0.01 0.96 k 0.04 0.90 * 0.04
2.3 + 0.5 1.8 * 0.5 0.80 & 0.3 0.11.-0.74
Dissolved
in percent
(particulate
and
open
(9)
(9)
(13) (33)
(8-J)
(41) (84) (57)
(60)
(74)
(28)
(41) (56) (50)
(“At)
ocean
concentration.
0.23 + 0.14
0.2 f 0.1
-
0.14 & 0.02 0.45 * 0.02
0.51 + 0.06 1.62 k 0.13 0.78 k 0.23 1.82kO.13 4.01 + 0.30 1.32 f 0.13 3.14 + 0.55 .-
0.14 * 0.02 2.22 * 0.30 0.04 f 0.01
Particulate
sea
plus dissolved)
2.28 +0.18 (n = 2) (‘%,) of total
0.97 + 0.04 1.97+0.10 1.23 f 0.10 0.18 & 0.06 0.81 k 0.04
1.60 k 0.05
-
(46) (2)
(3)
(“h)
1.20 + 0.24 (n = 4)
0.03 0.23 0.02 0.07 0.10 0.06 0.07 0.08 0.08 0.05
0.23 + 0.01 0.64 i 0.17 0.02 * 0.01
Particulate
coastal
21”Pb (Bq rn-‘)
rhtuary.
2.9 i 0.5 1.9kO.6 (n = 5) 2.3 k 0.6
+ f + + + k + & + +
m-‘)
river,
3.50+0.10 1.24 + 0.06 (n = 4) 1.20~0.10
1.22 8.21 0.58 2.50 3.85 2.24 2.45 2.70 2.68 1.77
7.9 * 0.2 0.75 * 0.02 0.88 * 0.02
phase are also given
9.3 8.4 39.0 26.0 16.5 25.8 37.0 15.4 20.8 5
5.3 42.7 1.0 -
(Bq
phases in the Tagus
Dissolved
“6Ra
and particulate
Suspended matter dry (g m-7
in dissolved
in the particulate
36
N
35.5 36
Coastal sea Cl Shelf. 38”40’
N
0.278 0.329 9.019 8.653 20.534 28.411 29.338 27.859 26.828 35.458 35.458
Estuary El E2 E3 E4 E5 E6 E7 E8 E9 El0 El0 rep
fresh fresh fresh fresh
Salinity tg 1-l)
c~mcrntrations
River RI (120 km) R? (70 km) R3 (20 km) Other (R2-R3, II = 9)
R;dionuclidr
* 0.09 + 0.08 * 0.10 +0.08 + 0.09 * 0.01 + 0.06 f 0.01 f 0.08 f 0.04 + 0.05
I .72 + 0.02 (n = 2)
2.6 + 0.6 1.3 +0.6 (n = 5) 1.8 kO.2
1.45 0.21 0.35 0.11 < 0.01 0.05 0.23 0.55 0.29 1.08 1.27
0.67 k 0.03 0.50 + 0.36 0.59 & 0.05 0.03-0.65
Dissolved
zloPo (Bq m--‘)
0.64 f 0.20
0.4 + 0.3
--
0.83 ) 0.08 1.71 f 0.06
0.92 f 0.09 2.15 k 0.21 4.60 + 0.51 1.84 f 0.21 0.94 f 0.10 1.67kO.17 2.92 + 0.70
0.34 f 0.03 2.90 + 0.34 0.05 2 0.01
Particulate
(27)
(18)
(4) (57)
(38) (92) (92) (92) (1f-v (98) (91)
(85) (8)
(34)
WI
F.P. Carvalho
1 The Science
Table 7 Radionuclide concentrations in suspended the Tagus river, estuary and open ocean. Sampling station Bq kg ’
Concentration Bq kg-’ (dry) 22hRa
matter
of the Total
and radionuclide
in suspended
matter
2 I up0
?“‘Pb
Environment
concentration
Concentration solved phase 2,0,,,,+6Ra
196 (1997)
ratios
ratios
151-161
in filtered
155
water
and suspended
matter
Concentration ticulate phase
ratios
210pb:2’bRa
‘“lPo:Z’Of,b
0.22 0.77 0.38 0.88
0.61 + 0.09 3.47 f 0.31 2.00* 1.11
2.4 + 0.4 1.3 kO.2 1.2 f 0.4
kO.08 k 0.38 k 0.06 * 0.03
0.32 * 0.04
1.8 i-O.3 1.7 2 0.2 5.9 k 1.9 1.0+0.1 0.2 $- 0.03 1.3*0.2 0.9 * 0.3 -~
in dis-
?,OpO:Z1Op,,
of
in par-
----___ River RI R2 R3 Other (R2 R3, n =9)
45 f 2 15 + 0.4 24 k 6
Estuar! El E2 E3 E4 E5 E6 El E8
172f5 25 If: 1 76+2 15+2 7&2 22 * 1
E9 E IO Ocean 38”40’ 39”40’
26 52 43 88
f + i *
3 I 3 35
55 k6 193 & 16 84 If 6 56 + 5 243 f 18 51*5 85k 15 81 k7 50*5 28,4
N- lO”30’ N- lO”50’
W W
20* 23k
10 14
65 * 6 68 * 8 48 & 5 98k44
99*10 328 k 118*13 70 f 57 + 65 + 79+
25 8 6 7 19
I .03 0.07 0.89 1.02 0.22 0.42 0.18 0.21
* 0.02 * 0.04 f 0.06 * 0.03 * 0.01 2 0.03 +O.Ol + 0.02
0.29 0.28 0.73 0.27
1.15 0.38 0.67 0.04
+ + * k
f 0.22 k 0.26 & 0.11
0.80 0.92 3.26 7.33 3.88
i+ t & *
0.24 0.08 0.36 1.42 0.71
166 * 17
0.07 * 0.002 0.54 * 0.03
1.52 f 0.28 1.12 + 0.06
5.9 + 1.0
40 * 30 64 + 20
1.9 & 0.3 1.9 * 0.2
0.78 i 0.56 0.75 + 0.28
2.0 f I .8 2.8 k 1.9
3. Results
3.1. 226R~, ‘l”Pb muttcv
0.3 + 0.6 2.4 k 0.7 0.9 & 0.3
and “‘PO in water and suspended
Concentrations of these radionuclides in samples from the Tagus river, estuary and coastal sea are shown in Tables 1 and 2. Concentrations of total (dissolved and particulate) 226Ra varied in the range of 0.90-8.2 Bq m - 3, depending on the load of suspended matter and location of the sampling zone. The relatively high concentration of dissolved 226Ra in the freshwater station Rl is due to weathering of secondary uranium deposits in soils of the Tagus watershed in the region of Nisa (Fig. 1). Downstream of Rl (stations R2El), water samples showed concentrations of dissolved 226Ra around 1 Bq mp3, which are typical of the river water discharge into the estuary. The relatively high concentration measured at the up-
per estuary, station E2, was possibly influenced by waste from a cement factory and drainage of agricultural fields discharged at the narrow entrance of the estuary near Vila Franca. Throughout the estuary, concentrations of dissolved 226Ra increased seaward reaching a maximum of 3.85 Bq m - 3. These concentrations remain near constant in the river plume in coastal sea, dropping off rapidly to 1.2 Bq m ~ ’ in more distant continental shelf sea water and in the open ocean (Table 1; Fig. 3). Concentrations of 226Ra in suspended matter were also higher in the upper estuary. It is interesting to note that ‘*‘Ra concentrations in suspended particles decrease in a seaward direction, which corresponds to dissolution/desorption of particulate ‘16Ra with increased salinity (Table 2). Concentrations of total *lOPb ranged from 0.8 Bq m ~ 3 in freshwater to 4.8 Bq m - 3 in the estuary. As with 226Ra, these concentrations de-
F.P.
ti
0
I
I
10
20
Carualho
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I,,,
30
Salinity
(g L-l]
Fig. 3. Concentrations of dissolved 226Ra. Area between lines indicates concentration trends throughout the Tagus river-estuary-coastal sea system.
pend on the load of suspended matter. On average, concentrations of dissolved z’“Pb in river water were higher than in the estuary and were lower than concentrations of dissolved *lOPb in the coastal sea (Table 1; Fig. 4). Concentrations of ?‘OPb in estuarine particles were consistently higher than concentrations measured in riverborne and marine suspended matter (Table 2), confirming the enhanced scavenging of dissolved Z’OPb taking place in the estuary. Concentrations of total “‘PO varied from 0.6 to 5.0 Bq m ’ and, in the majority of the samples, a very high percentage of total *“PO was associated
I O.O’ L
I 0
10
P I d-----J 20
30
Salinity l:tg. 4. Concentrations ““PO copen circles) \V~lClll
in
of dissolved the Tayus
(g L-l) ““Pb (solid squares) river-estuary-coastal
and sea
Enrironment
196 (1997)
151- 161
with particles (Table 1). Concentrations of dissolved 2’oPo were typically about 0.50 Bq m ’ in river water and, thus, generally higher than concentrations of dissolved “‘PO in the maximum turbidity zone of the estuary. In the upper and mid estuary (stations E22E6), ‘l”Po in solution further decreased to very low concentrations. < 0.01 Bq m - ‘, to increase again towards the sea from station E&E10 (Table 1; Fig. 4). Similar to “‘Pb, the concentration of “‘PO in suspended matter was generally higher in the upper estuary than both in riverborne and marine particles (Table 2). corresponding to enhanced particle sorption of 2’“Po occurring in that zone. More informative are the activity concentration ratios of these radionuclides in the soluble and particulate phases (Table 2). ‘10Pb:2’6Ra concentration ratios in the soluble phase of freshwater and low salinity water (upper estuary) were consistently higher than concentration ratios in mid and lower estuary zones, but gradually increased in the coastal sea. This is simultaneously due to enrichment of “‘Pb in estuarine suspended matter and to dissolution/desorption of particulate IZ6Ra in the estuary (Table 2). 2’oPo:“oPb ratios in the soluble phase decreased from freshwater (0.73) to the mid estuary ( < 0.01) where *“PO was almost entirely removed from solution. From the mid-estuary seaward, an increased Po:Pb ratio was observed, attaining values higher than unity at the mouth of the estuary and stabilizing at about 0.7 in the coastal sea. Since in the estuary no desorption of “‘PO from suspended matter was observed, the increased concentration of ““‘PO in the soluble phase is likely to originate in “OPo re-dissolution from bottom sediments and intertidal mudflats. The increased 2’oPo:“oPb ratios simultaneously observed in suspended matter at the mouth of the estuary and at coastal sea. sampled with the low tide outflowing water, further supports this interpretation. A common feature of the three radionuclides investigated is that total concentrations in water are chiefly regulated by the load of suspended matter. Furthermore, concentrations measured in suspended matter in the estuary are roughly identical to concentrations measured in the fine fraction ( < 63 jlrn) of bottom sediments suggesting
F.P.
Cardho
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that resuspension of bottom sediment is the main mechanism controlling the distribution of these radionuclides in the estuary (Carvalho, 1995a). Results from analyses of bottom sediments lead also to the identification of areas with higher than average ““Pb and ““PO background levels. Sediments in a very small area around Barreiro Peninsula, Lavradio, in the south bank, contain radionuclide levels enhanced by industrial waste releases. An old phosphate fertilizer plant during decades discharged phosphate wastes (process water and phosphogypsum) into self contained gypsum ponds on the river bank (Lavradio). After decantation of the gypsum, the overlying water generally containing little gypsum, much less radioactivity and with a pH near neutrality, was released into the estuary causing enhancement of radionuclide levels in local sediments. The total activity of radionuclides released is difficult to estimate but, since no evidence was obtained for technologically-enhanced radioactivity levels elsewhere in the estuary (Carvalho, 1995a) we conclude that it is a negligible input. Therefore, the measurements made in water samples collected in mainstream channels (this study) mainly reflect the distribution of naturally-occurring nuclides. However, an exception to this might be the concentrations of 2’6Ra in stations El -E2, which could be enhanced by point source discharges. The Distribution Coefficient (I&) of radionuelides between suspended solids and the dissolved phase (& = (Bq kg.-- ’ dry suspended matter)/(Bq kg-’ filtered water)) were computed using the concentrations given in Tables 1 and 2. Kd for “‘Ra in freshwater is (1.7 + 0.6) x lo4 and of the same order of magnitude in the upper estuary, but decreases to about 6 x lo3 in the mid estuary. Kd for ““Pb in freshwater is (3 + 1) x lo4 and (1.5 + 0.4) x lo5 in the estuary. K,, for 21”Po is (1.0 & 0.3) x 1Oj in freshwater and (9 t 5) x lo5 in brackish water. The I& values of both “‘PO and ““Pb in oceanic water seem to decrease to (l4) x 104. High Kd values for “‘Pb and “‘PO are a result of the strong adsorption of these metals onto suspended particles in aquatic systems, comprising of freshwater, estuarine and marine environments. Moreover, these values reflect the increased scavenging of dissolved “‘Pb and ‘l”Po
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196 (1997)
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157
in the estuary and the effect of particle concentration on the Kd values. 3.2. Preliminary Tugus estuary
balance oj’ “‘Pb
rind -“‘PO in the
Based on the radioanalytical results, we can estimate the apparent mean residence time of dissolved *l”Pb and “OPo in the estuary as well as the export of the soluble fraction of both radionuelides into the coastal sea. Since the geomorphology of the estuary (a broad inner bay with two narrow entrances) and the characteristics of a fully-mixed estuary (vertical mixing of water in the inner bay) meet the basic requirements of box-models, the use of this approach seems acceptable to make a mass balance of the rddionuelides. For this, we assume that measured concentrations of radionuclides (Tables 1 and 2) are representative of average annual concentrations and that the estuary may be represented by a single box-model with steady state conditions holding for at least several months. We start by considering the hydrology of the estuary. The flux of water through the estuary depends upon river discharges, 1.30 x 10’” m7 a -I, and direct rainfall into the estuary, 700 1 m-* a-‘. Taking into consideration the surface area of the estuary, direct rainfall contributes with 1.40 x lo* m3 a-‘, which only accounts for 1.1% of the annual river discharge. For simplification, urban sewage and industrial liquid effluents released into the estuary are considered equivalent in volume to evaporation. The average volume of water in the estuary was estimated at 2.0 x 10” m3. Assuming that this volume is constant, the outflow of water is identical to the inflow, i.e. 1.30 x 1O’Om3 a - ’ Therefore, the mean residence time of water in the estuary, R,, is 0.15 a (55 days) and the rate constant of water mass outflow, 2, = R, ‘, is 6.6 a-’ under average water flow conditions. The balance of dissolved “‘PO in the estuary depends upon discharges of dissolved “‘PO into the estuary, “‘PO produced through radioactive decay of dissolved *‘OPb, 210Po radioactive decay, 210Po removal from solution and “‘PO exported with the water discharged to the coastal sea. The annual supply of 2’oPo dissolved in river water
158
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(average concentration 0.58 f 0.08 Bq m ~ ‘) discharged into the estuary is 7.5 x IO9 Bq a - ’ contributing 3.7 Bq rnpi per year to the estuarine water. Atmospheric deposition of “OPo in the area is 8.3 2 2.8 Bq m ~~’ per year (unpublished), and it is a significant contribution that we consider as entirely soluble. In the estuary, the ingrowth of ““PO through radioactive decay of dissolved “‘Pb, was calculated from the inventory of 210Pb (based on the average “‘Pb concentrations in stations E2-E9), i.e. 1.6 x lo9 Bq. In the same way. the inventory of dissolved 2’oPo in the estuary, based on the average of concentrations determined in E2 through to E9, is estimated at 4.4 x 10’ Bq. Therefore, the mass balance equation for dissolved ?‘OPo in the estuary is 2,+APh + I, + I,, = &,APo + &A,,
+ &4,,
(1)
where, i,,, is the radioactive decay constant of ““PO, 1.83 a ‘, A,,, is the concentration of dissolved “OPb in the estuary, 0.84 Bq m ~ 3qA,, is the concentration of dissolved “‘PO in the estuary, 0.22 Bq m ‘, Z, is the annual input of dissolved 2’oPo in the river discharge, 3.7 Bq m ’ per year, Z, is the direct annual atmospheric deposition of 210Po to the estuary and diluted in the estuary volume. 0.83 Bq m 3 per year, 2, the scavenging rate constant for dissolved ““PO onto suspended matter and sediment particles, a ~ ‘, and i,,,, is the constant for water mass outflow into the coastal sea, 6.6 a ~ ‘. Substituting the values of the data into Eq. (1) and resolving for 2s gives a result of 20 a--’ and the mean residence time of “‘PO in solution relative to adsorption on sediment particles, is estimated at 0.05 a or 18 days. The other mechanism contributing to the removal of dissolved “‘PO from the estuary is the outflow of water into the sea. Taking into consideration that the inventory of dissolved “‘PO in the estuary is 4.4 x 10” Bq, the ““PO exported with a rate of i., = 6.6 a -- ’ is 2.9 x 10’ Bq a ~~‘. This export corresponds to about 30% of the annual delivery of dissolved “‘PO to the estuary. Therefore. most of dissolved s’“Po (about 70%) entering the estuary is removed by sorption onto sedimenting particles and trapped, at least temporarily. in the estuary. The concentration of dissolved “‘PO remaining with the water outflou will reduce to about 0.2 Bq m -’
Emironnzent
196 (1997)
151-161
(without considering mixing with sea water) which is in agreement with the very low concentrations measured at stations E5 and E6 (Table 1). An identical mass balance equation may be written for dissolved “‘Pb with /1p’y4’.& + z, + z, = &,A,,
+ /i,Apb + i,,A,,
(2)
where, /i,, is the radioactive decay constant of 2’0Pb, 31.1 x 10W3 a-‘, ;Ipb A,, corresponds to the 2’oPb ingrowth from dissolved 226Ra in the estuary, and the other variables have for “‘Pb meanings similar to those given above for 2”‘Po. Based on the average “OPb concentration in filtered river water, 1.6 ) 0.6 Bq m ~ 3, the annual input of dissolved 2’oPb by river discharge is 2.1 x 10” Bq a ~ ‘, contributing to (Z,) 10 Bq m ’ per year in the estuarine water. The direct deposition of atmospheric 2’0Pb, 56 -t 12 Bq m em2per year, accounts to (ZJ 5.6 Bq rnp3 per year in the water of the estuary, corresponding to 56% of the dissolved 2’0Pb discharged by the Tagus river. The contribution of dissolved 226Ra through radioactive decay to total dissolved 2’0Pb is minimal, 0.6%, when compared with other “OPb sources (Z, and ZJ. Solving Eq. (2) in order to A,, 1, is equal to 12 a ~ ’ and the mean residence time of soluble 2’0Pb in the water of the estuary is 0.08 a or 30 days. The flux of dissolved 2’0Pb delivered to the coastal sea is computed at 1.l x 10” Bq a - ’ and the average concentration of dissolved 210Pb in the water outflow is 0.85 Bq m -‘. This discharge of soluble z’“Pb into the sea accounts for 35% of the annual input of soluble 210Pb received by the estuary. Therefore, most of the 2’0Pb (65%) entering the estuary in the soluble phase is likely sorbed onto suspended particles and trapped in sediment accumulation zones of the estuary, or discharged bound to suspended matter into the coastal sea.
4. Discussion and conclusions The distribution of dissolved 226Ra concentrations indicates a partial dissolution of 226Ra in riverborne particles and, eventually, in industrial wastes discharged into the estuary. This is further confirmed by the decrease of 226Ra concentrations in suspended matter with increased salinity. More-
F.P.
Carvalho
/ The Science
of’ the Total
over, results show that the outflow of estuarine water is a source of dissolved 2’6Ra to coastal sea water. Similar trends for Ra dissolution in estuarine environments were reported before (Moore, 1981; Elsinger and Moore, 1983; Helz et al., 1985/ 86). In contrast with 226Ra, dissolved *‘OPb and *“PO entering the estuary are removed from solution in the freshwater/salt water mixing zone, probably by a combination of co-precipitation with humic substances, Fe-Mn hydroxides and by adsorption on to suspended particles (Sholkovitz et al., 1978). This is clearly shown by “‘Pb and 2’0Po concentrations in suspended matter which are higher in this mixing zone (Table 1; Carvalho, 1995a). Furthermore, results indicate that no noticeable desorption of these radionuelides occurs during the transport of suspended matter across the estuary. The mean estuarine residence times calculated for dissolved *“Pb and *“‘PO relative to sorption onto particles, 30 days and 18 days respectively, are in agreement with the few values reported for lakes, estuaries and coastal zones (Santschi et al., 1979; Benoit and Hemond, 1987). For instance, mean residence times for dissolved *“PO of l-5 days in winter and up to 2 months in the early summer have been reported for Narrangasett Bay, USA, whereas dissolved *“Pb displayed about 35 days in the same environment (Santschi et al., 1979). Our measurements and results from box-model computations indicate that direct deposition of atmospheric “‘Pb represents a significant contribution (36%‘) to the inventory of dissolved 2’0Pb in the estuary. On the other hand, the deposition of atmospheric *‘OPb over the much wider catchment basin of the Tagus River does not result in a significant input to the estuary, which implies an efficient immobilization of *“Pb and *“PO in the soils of the watershed. Furthermore. little “OPb and “‘PO in estuarine water is discharged in the soluble phase to the coastal sea. Instead, these radionuclides are mainly adsorbed by suspended matter and seem to accumulate in bottom sediments in the estuary. Previous results on *“Pb in sediment cores from the upper estuary indicated an excess of 2’0Pb relative to 226Ra above the average inventory supported by the flux of atmospheric *“Pb
Environment
196 (1997)
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159
(Carvalho, 1995a). This provides additional verification on the enhanced scavenging of “‘Pb and 210Po in the maximum turbidity zone. However, sediment cores sampled in the mid estuary displayed lower than expected “‘Pb inventories, suggesting remobilization of deposited 2’oPb-2’oPo back into solution. Increased concentrations and increased 2’0Po:2’0Pb ratios in filtered water and suspended matter were measured in samples collected at the mouth of the estuary (Tables 1 and 2) confirming remobilization of both radionuelides and especially of *“PO. Although not investigated in estuarine environments, it is known that “‘Pb and larger amounts of *“PO are released from sediments when anoxic conditions develop (Benoit and Hemond, 1990). Whether these re-dissolved 2’oPb-2’oPo totally leave the estuary or are partly reabsorbed on to suspended particles and trapped in the estuarine circulation, can not be concluded from our results. In the light of results obtained from other estuaries, it seems possible that particles of marine origin enter the estuary, contribute to the scavenging of dissolved radionuelides and finally accumulate in the low salinity zone (Olsen et al., 1989). Furthermore, there is evidence that with average river flow the discharge of suspended matter from the Tagus estuary is minimal and particles are retained in the estuarine circulation (Vale, 1986). This allows for sediment accumulation inside the estuary and, therefore, for the build up of *‘OPb and *“PO reservoirs in sediments. The focusing of *“Pb deposition in upper estuary sediments was observed also in other estuaries such as the Mira estuary (unpublished) and the Tamar estuary (Clifton et al., 1995). These radionuclide reservoirs may intermittently be reduced by exceptionally large floods. In the Tagus estuary these floods may occur two to three times per century and carry out to the sea as much sediment in a few days as transported with average river flow during one year (Vale, 1986). Still in these circumstances the sudden transport of *lOPb and “‘PO to the coastal sea will be made in the particulate phase. The following general conclusions can be drawn on the implications of these results to the geochemical cycling of Ra, Pb and PO in phosphate wastes released as aqueous slurry into estuarine environments:
1. 226Ra in particulate form will dissolve in the estuary and will be discharged into the coastal sea in the soluble phase. In the Tagus estuary the mean residence time of dissolved 276Ra will be controlled by the residence time of water averaging 55 days; 3L. since the residence times of “OPb and ‘*‘PO in the soluble phase are much shorter than the water residence time, these radionuciides will be rapidly adsorbed onto suspended matter. In the Tagus estuary about 2/3 of dissolved radionuclide inputs will be trapped by particles and may give rise to enhanced radioactivity levels in sediment accumulation zones; 3. most of the ‘“‘Pb and Z’“Po flowing through the estuary will be exported to the coastal sea bound to bottom sediments and suspended matter rather than in the soluble phase. In the Tagus estuary, probably after several cycles of adsorption-redissolution. only l/3 of dissolved ““Pb and “‘PO in waste discharges will be transferred to the sea in the soluble phase. One implication of the rapid sorption of ““PO and “‘Pb on to sediments could be a reduced bioavailability of these radionuclides to estuarine organisms. This would prevent a significant accumulation of “‘Pb and ““PO from industrial phosphate waste discharges by estuarine biota. Actually, no enhancement of ‘IrJPb and “‘PO levels could be detected in biota in the Tagus estuary. Furthermore, field studies carried out in the Seine estuary in zones receiving phosphate wastes released as aqueous slurry and dumped as dry phosphogypsum, indicate no noticeable enhancement of ““Pb and “OPo levels in biota (Germain et al.. 1992, 1995). Other studies carried out in estuaries and coastal areas receiving discharges of low pH phosphate wastes (acid stream and gypsum slurries directly from phosphoric acid production) have shown significant enhancement of radioactivity levels in the aquatic environment (Koster et al., 1992; Pennders et al.. 1992; Perianez and Garcia-Leon. 1993: Poole et al., 1995). In these areas, elevated concentrations of these radionuclides were observed in biota samples collected near the point of waste releases (Roll0 et al.. 1992; Swift et al., 1995). Enhanced bioaccumulation, in particular of “(‘PO, probably takes
place before neutralization of highly acidic wastes and radionuclide sorption on to particles have contributed to remove radionuclides from the soluble phase to sediment. Nevertheless, delayed accumulation of these radionuclides from sediments by benthic and filter-feeding organisms and subsequent radionuclide food-chain transfer, can not be entirely excluded with the present knowledge (Carvalho and Fowler, 1993, 1994) and justifies specific investigation.
Acknowledgements The excellent technical assistance of Mr J.M. Oliveira and Mrs G. Albert0 (DGA/DPSR) is gratefully acknowledged.
References Barros. M.C.. 1986. A case study of waste inputs in the Tagus Estuary. In: G. Kullenberg (Ed.), The Role of the Oceans as a Waste Disposal Option. D. Reidel, pp. 307-324. Benoit, G. and H.F. Hetnond, 1987. A biogeochemical mass balance of ““PO and ““Pb in an oligotrophic lake with seasonally anoxic hypolimnion. Geochimica et Cosmochimica Acta, 51: 1445-1456. Benoit, G. and H.F. Hemond, 1990. ““PO and *‘“Pb remobilization from lake sediments in relation to iron and manganese cycling. Environ. Sci. Technol., 24: 1224-1234. Bojanowski, R.. R. Fukai, S. Ballestra and H. Asari, 1983. Determination of natural radioactive elements in marine environmental materials by ion-exchange and cl-spectrometry. Proc. 4th Symp. Determination of Radionuclides in Environmental and Biological Materials, Laboratory of the Government Chemist, London, Paper No. 9. Carvalho, F.P., 1995a. ““Pb and “‘PO in sediments and suspended matter in the Tagus estuary. Local enhancement of natural levels by wastes from phosphate ore processing industry. Sci. Total Environ., 159: 201-214. Carvalho, F.P., 1995b. ““PO and ““Pb intake by the Portuguese population: the contribution of sea food in the dietary intake of ““PO and “(‘Pb. Health Phys., 69: 469-480. Carvalho, F.P. and S.W. Fowler, 1993. An experimental study on the bioaccumulation and turnover of polonium-210 and lead-210 in marine shrimp. Mar. Ecol. Prog. Ser.. 102: 125-133. Carvalho, F.P. and S.W. Fowler, 1994. A double-tracer technique to determine the relative importance of water and food as sources of polonium-210 to marine prawns and fish. Mar. Ecol. Prog. Ser., 103: 251 264.
F.P.
Catwalk
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Castanheiro, J.M.. 1982. Distribution, transport and sedimentation of suspended matter in the Tagus estuary. In: Estuarine Processes, An Application to the Tagus Estuary. Proc. UNESCO/IOCCNA Workshop held at Lisbon, Portugal, pp. 13-- 16 December. Clifton R.J., P.G. Watson, J.T. Davey and P.E. Frickers, 1995. A study of processes affecting the uptake of contaminants by intertidal sediments, using the radioactive tracers: ‘Be, “‘Cs and unsupported *“‘Pb. Estuarine, Coastal and Shelf Sci., 41: 4599474. Elsinger, R.J. and W.S. Moore, 1983. z24Ra, ‘18Ra, and zr6Ra in Winydh Bay and Delaware Bay. Earth Planet. Sci. Lett., 64: 430-436. Germain. P., G. Leclerc and S. Simon, 1992. Distribution of ““PO in MyMus edulis and Fucus vesiculosus along the channel coast of France; influence of industrial releases in the Seine river and estuary. Radiat. Prot. Dosim.. 45: 251-260. Germain, P.. G. Leclerc and S. Simon, 1995. Transfer of polonium-210 into Myths edulis (L.) and Fucus vesiculosus (L.) from the baie de Seine (Channel coast of France). Sci. Total Environ.. 164: 1099123. Helz, G.R., G.H. Setlock, A.Y. Cantillo and W.S. Moore, 1Y85/86. Processes controlling the regional distribution of 2”‘Pb, 226Ra and anthropogenic zinc in estuarine sediments. Earth Planet. Sci. Lett., 76: 23-34. Koster, H.W., P.A. Marwitz, G.W. Berger, A.W. van Weers, P. Hagel and J. Nieuwenhuize, 1992. *‘“PO, 2’oPb, 226Ra in aquatic ecosystems and polders, anthropogenic sources, distribution and enhanced radiation doses in the Netherlands. Radiat. Prot. Dosim., 45: 7155719. Moore, W.S.. 1981. Radium Isotopes in the Chesapeake Bay. Estuarine, Coastal and Shelf Sci., 12: 713-723. Olsen, CR., M. Thein, 1.L. Larsen, P.D. Lowry, P.J. Mulholland, N.H. Cutshall, J.T. Byrd and H.L. Windom, 1989. Plutonium, lead-210 and carbon isotopes in the Savannah estuary: riverbome versus marine sources, Environ. Sci. Technol., 23: 1475-1481.
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Osmond, J.K. and M. Ivanovich, 1992. Uranium-series mobilization and surface hydrology. In: M. Ivanovich and R.S. Harmon (Eds.), Uranium-series Disequilibrium. Applications to Earth, Marine and Environmental Sciences. Oxford Science Publications, Clarendon Press, pp. 2599289. Pennders, R.M.J., H.W. Koster and J.F. Lambrechts. 1992. Characteristics of L’“Po and “‘Pb in effluents from phosphate-producing industries: a first orientation. Radiat. Prot, Dosim., 43: 7377740. Perianez, R. and M. Garcia-Leon, 1993. Ra-Isotopes around a phosphate fertilizer complex in an estuarine system at the southwest of Spain. J. Radioanal. Nucl. Chem.. Articles, 172: 71--79. Poole, A.J., D.J. Allington, A.J. Baxter and A.K. Young, 1995. The natural radioactivity of phosphate ore and associated products discharging into the eastern Irish Sea from a phosphoric acid production plant. Sci. Total Environ., 1731174: 1377149. Rollo, S.F.N., W.C. Camplin, D.J. Allington and A.K. Young, 1992. Natural radionuclides in the UK marine environment. Radiat. Prot. Dosim., 45: 203-209. Santschi, P.H., Y.-H. Li and J. Bell, 1979. Natural radionuelides in the water of Narragansett Bay. Earth Planet. Sci. Lett., 45: 201-213. Sholkovitz, E.R., E.A. Boyle and N.B. Price, 1978. The removal of dissolved humic acids and iron during estuarine mixing. Earth Planet. Sci. Lett., 40: 130- 136. Swift, D.J.. D.L. Smith, D.J. Allington and K. Winpenny. 1995. A laboratory and field study of 2’“Po depuration by edible winkles (Littorina littorea L.) from the Cumbrian coast (north eastern Irish Sea). J. Environ. Radioact., 27: 13-33. Vale, C., 1986. Transport of particulate metals at different fluvial and tidal energies in the Tagus River estuary. Rapp. P.-V. Reun. Cons. Int. Explor. Mer, 186: 306-312. Vale, C. and B. Sundby, 1987. Suspended sediment fluctuations in the tagus estuary on semi-diurnal and fortnightly time scales. Estuarine. Coastal and Shelf Sci., 25: 4955508.
Distribution,
cycling and mean residence time of 226Ra, 210Pb and 210Poin the Tagus estuary Fernando
Direwoo
Gem1 do Ambiente,
Departamento
de Proteccao
P. Carvalho* e Seguranca
Radiologica,
EN IO, P-2685
Sacariem,
Portugal
Received 14 August 1996; revised 22 November 1996; accepted 29 November 1996
Abstract Resultsfor dissolvedand particulate “6Ra, ““Pb and 2’0Poin the Tagus river, estuary and coastalseasystemshow different distribution and chemicalbehaviour patterns for theseradionuclidesin the three aquatic environments. 22hRafrom riverborne particlesdissolvesin the estuary and contributes to increasedconcentrationsof dissolved226Ra in estuarinewater. In the estuary, dissolved2’oPb and 210Pofrom river dischargeand atmosphericdeposition are scavengedby suspendedmatter, which in turn becomesenriched in thesenuclidesin comparisonwith riverborne particles. As a result of these processes,the estuarine water flowing into the coastal sea contains enhanced concentrationsof dissolved“6Ra, but is depletedin dissolved“‘Pb and 2’oPo. Under averageriver flow conditions, massbalancecalculations for dissolved“‘PO and 2’0Pb in the estuary allowed their mean residencetimes to be estimatedas 18 and 30 days, respectively. Due to the rapid sorption of theseradionuclideson to settling particles, bottom sedimentsin the estuary representa sink for 210Pband “‘PO from both natural sourcesand industrial waste releases.Resultsalso suggestthat partial re-dissolutionof theseradionuclidesfrom bottom sedimentsand intertidal mudflatsis likely to occur in the mid- and low-estuaryzones.Nevertheless,box-model computationsindicate that the dischargeof “‘Pb and 2’oPointo the coastal seatakes place mainly with the transport of sediment,whereasthe dischargein the dissolvedfraction can only account for one third of the activities entering the estuary in the soluble phase. Implications of these results to the cycling of radionuclidesin phosphate waste releasesinto estuarine environmentsare discussed.0 1997Elsevier ScienceB.V. Kqmwds:
Estuaries;Radium; Polonium; Radioactive lead; Distribution coefficients;Phosphatewastes
1. Introduction * Present address: International Atomic Energy Agency, Marine Environment Laboratory. P.O. Box 800, MC 98012 Monaco Cedex. Principaute de Monaco.
Radium-226 ( Tljz = 1622 a), lead-2 10 (z-,/Z = 22.3 a), and polonium-210 (T,,,= 138.4 d), are
O048-9697;971$17.00 8 1997 Elsevier Science B.V. All rights reserved. PI/ SOO48-9697(96)05416-2
152
F. P. Cardho
I) The Science
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the Total
naturally occurring radionuclides of the uranium series and originate in the earth’s crust (Osmond and Ivanovich, 1992). Although ubiquitous in the environment, the concentration of these radionuclides in river-estuarine systems may vary widely due to the geology of the watershed and chemical weathering conditions. Furthermore, natural levels of these radionuelides in aquatic ecosystems may be increased by industrial waste releases. This is the case for non-nuclear industries, such as phosphate ore processing, whose wastes contain relatively high concentrations of 226~~, “‘Pb and 210Po (Koster et al., 1992; Carvalho, 1995a). The environmental cycling of these radionuclides, especiallv of ““PO, is of the utmost interest due to their high contribution to the radiation dose received by humans through the diet (Carvalho, 1995b). The fate of radionuclides in phosphate wastes discharged into estuaries and coastal zones has recently been investigated in several European countries (Germain et al., 1992; Koster et al., 1992; Perianez and Garcia-Leon, 1993; Carvalho, 1995a: Poole et al., 1995). However, a major difficulty encountered in predicting the fate of ‘26Ra, “‘Pb and “*PO from phosphate waste discharges is the paucity of data on the biogeochemistry of these radionuclides in estuaries. Therefore, a deeper understanding of the distribution and behaviour of these radionuelides is necessary for the interpretation of enhanced radioactivity levels as well as for the prediction of radionuclide fluxes and sinks in estuarine and coastal ecosystems. Here we present distribution data for 226Ra, ‘“‘Pb and “‘PO in the lower reach of the Tagus river and in the Tagus estuary, Portugal. Mass balances of “OPb and 2”)Po are calculated in order to determine the mean residence time of these radionuclides in the soluble phase and to estimate ““Pb and “‘PO retention in the estuary. This is achieved through the assessment of ““Pb and ““PO inputs, combined with inventories of dissolved zzhRa. “‘Pb and “(‘PO in the CL “f1Ll;il-C
Emironment
196 (1997)
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2. Methods 2.1. Study area
The river Tagus drains a catchment area of 86 000 km’, with its lower reaches flowing across a sedimentary flood-plain and ending in an extended fully-mixed type estuary (Fig. 1). The Tagus estuary covers an area of about 300 km2, and is composed of a 2-4 m deep and broad inner bay and a 40 m deep and narrow channel which connects to the sea. The shallow inner bay displays a complex bottom topography with tidal mud flat areas, sand islands and beds of dead oyster shells (Vale and Sundby, 1987). In the inner bay the tidal currents are weaker and allow the deposition of fine grain sediments (silt) at rates up to 3 cm a - ’ in some areas. The narrow sea entrance of the estuary has consolidated sandy gravel bottoms and, due to the higher velocity of tidal currents, little deposition of fine sediments occurs (Castanheiro, 1982). At the upper estuary, in the low salinity zone, the turbidity is usually the highest and attains 1.5 g 1- ’ . The main input of freshwater to the estuary is the discharge of the Tagus river, 400 m3 s - ‘. Other small rivers discharging into the estuary, like the Sorraia with 30 m3 ss’, provide little contribution to the fresh-
Fig. I. Catchment basin of the Tagus stations (RI R3) in the freshwater zone.
river
and
sampling
F.P.
Curcalito
:) The Science
of the Total
water inflow (7%) of Tagus discharge). Annually, the Tagus discharges about 4 x 10’ tons of suspended sediment into the estuary. Due to the small water depth and extended surface of the inner bay, tidal currents and winds intensively mix the estuarine water and re-suspend bottom sediments. The amount of sediment re-suspended and re-deposited during a fortnightly tidal cycle may attain the equivalent to the Tagus annual sediment discharge (Vale, 1986; Vale and Sundby, 1987). Furthermore, the estuary receives wastes from anthropogenie activities which carry contaminants into the system. Domestic sewage and major industrial waste outfalls (petrochemistry, fertilizers. smelters, etc.) are located on the south bank of the estuary, although several important point sources are located on the north bank (Barros, 1986). 2.2. Suwpling
Emvironnzent
196 (1997)
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Fig. 2. Estuary of the River Tagus and sampling stations estuary (El -ElO) and adjacent coastal sea (Cl).
153
in the
und analvtical procedures
Water samples for 2’6Ra. ?“‘Pb and ““PO analyses were collected (December 1989) under average river flow conditions in the lower reach of the Tagus river, in the Tagus estuary and in the coastal sea. In the estuary, samples were collected in the mainstream channels in order to reflect the fresh-estuarine water flow better and to avoid, as much as possible, the influence of waste outfalls (Figs. 1 and 2). A total of 30&60 1 water samples were filtered in situ, using a portable generator and pumping-filtering systems, through pre-weighed 140 mm diameter, 0.45 pm pore size Sartorius membrane filters. Filtered water samples were acidified with HNO, to pH < 2 and a 209Po isotopic tracer and stable lead carrier (20 mg) were added immediately. In the laboratory, after addition of 224Ra to water samples, the analysis of dissolved radionuclides was made following co-precipitation with MnO, and sequential separations on anion and cation exchange resins (Bojanowski et al., 1983). Filters were oven dried at 80°C and weighed for determination of the suspended matter load. Samples of suspended matter, after addition of 209Po and 224Ra tracers
and Pb carrier, were dissolved and analyzed using the same procedure as for water samples. PO isotopes were plated onto a Ag disc for n-spectrometry (Carvalho, 1995a,b). ““Pb was precipitated as lead chromate, filtered and the chemical yield determined gravimetritally. Filters with the precipitate were stored for 6-12 months and, after a second addition of 2osPo tracer and dissolution of chromate, “OPb was determined through the measurement of “‘PO ingrowth from the radioactive lead decay. 226Ra was determined using an “4Ra isotopic tracer. The 224Ra tracer was separated from 228Th just before use and added to the samples in activity amounts significantly greater than the natural 224Ra levels. Following the electroplating of radium isotopes onto a stainless steel disc, immediate tl counting of 2’4Ra allowed the radiochemical yield to be determined. Measurement of 226Ra was performed on the same disc 1 month later to allow for ‘14Ra decay. Radioactivity measurements were made using silicon surface barrier detectors (400 mm’, R type, EG and G ORTEC) connected to a multichannel analyzer.
38”40’ w 39”50’ w
Concentrations
Ocean, -10”30’ Ocean, -10”50’
36
N
water water water water
-
~ -
(25)
(57) (62) (44) (24) (7)
1.26 k 0.06 0.56 + 0.03 0.52 k 0.03 2.56kO.06 0.84 & 0.04 0.96 + 0.09 0.43 * 0.02 0.56 + 0.06 0.19 f 0.01 0.96 k 0.04 0.90 * 0.04
2.3 + 0.5 1.8 * 0.5 0.80 & 0.3 0.11.-0.74
Dissolved
in percent
(particulate
and
open
(9)
(9)
(13) (33)
(8-J)
(41) (84) (57)
(60)
(74)
(28)
(41) (56) (50)
(“At)
ocean
concentration.
0.23 + 0.14
0.2 f 0.1
-
0.14 & 0.02 0.45 * 0.02
0.51 + 0.06 1.62 k 0.13 0.78 k 0.23 1.82kO.13 4.01 + 0.30 1.32 f 0.13 3.14 + 0.55 .-
0.14 * 0.02 2.22 * 0.30 0.04 f 0.01
Particulate
sea
plus dissolved)
2.28 +0.18 (n = 2) (‘%,) of total
0.97 + 0.04 1.97+0.10 1.23 f 0.10 0.18 & 0.06 0.81 k 0.04
1.60 k 0.05
-
(46) (2)
(3)
(“h)
1.20 + 0.24 (n = 4)
0.03 0.23 0.02 0.07 0.10 0.06 0.07 0.08 0.08 0.05
0.23 + 0.01 0.64 i 0.17 0.02 * 0.01
Particulate
coastal
21”Pb (Bq rn-‘)
rhtuary.
2.9 i 0.5 1.9kO.6 (n = 5) 2.3 k 0.6
+ f + + + k + & + +
m-‘)
river,
3.50+0.10 1.24 + 0.06 (n = 4) 1.20~0.10
1.22 8.21 0.58 2.50 3.85 2.24 2.45 2.70 2.68 1.77
7.9 * 0.2 0.75 * 0.02 0.88 * 0.02
phase are also given
9.3 8.4 39.0 26.0 16.5 25.8 37.0 15.4 20.8 5
5.3 42.7 1.0 -
(Bq
phases in the Tagus
Dissolved
“6Ra
and particulate
Suspended matter dry (g m-7
in dissolved
in the particulate
36
N
35.5 36
Coastal sea Cl Shelf. 38”40’
N
0.278 0.329 9.019 8.653 20.534 28.411 29.338 27.859 26.828 35.458 35.458
Estuary El E2 E3 E4 E5 E6 E7 E8 E9 El0 El0 rep
fresh fresh fresh fresh
Salinity tg 1-l)
c~mcrntrations
River RI (120 km) R? (70 km) R3 (20 km) Other (R2-R3, II = 9)
R;dionuclidr
* 0.09 + 0.08 * 0.10 +0.08 + 0.09 * 0.01 + 0.06 f 0.01 f 0.08 f 0.04 + 0.05
I .72 + 0.02 (n = 2)
2.6 + 0.6 1.3 +0.6 (n = 5) 1.8 kO.2
1.45 0.21 0.35 0.11 < 0.01 0.05 0.23 0.55 0.29 1.08 1.27
0.67 k 0.03 0.50 + 0.36 0.59 & 0.05 0.03-0.65
Dissolved
zloPo (Bq m--‘)
0.64 f 0.20
0.4 + 0.3
--
0.83 ) 0.08 1.71 f 0.06
0.92 f 0.09 2.15 k 0.21 4.60 + 0.51 1.84 f 0.21 0.94 f 0.10 1.67kO.17 2.92 + 0.70
0.34 f 0.03 2.90 + 0.34 0.05 2 0.01
Particulate
(27)
(18)
(4) (57)
(38) (92) (92) (92) (1f-v (98) (91)
(85) (8)
(34)
WI
F.P. Carvalho
1 The Science
Table 7 Radionuclide concentrations in suspended the Tagus river, estuary and open ocean. Sampling station Bq kg ’
Concentration Bq kg-’ (dry) 22hRa
matter
of the Total
and radionuclide
in suspended
matter
2 I up0
?“‘Pb
Environment
concentration
Concentration solved phase 2,0,,,,+6Ra
196 (1997)
ratios
ratios
151-161
in filtered
155
water
and suspended
matter
Concentration ticulate phase
ratios
210pb:2’bRa
‘“lPo:Z’Of,b
0.22 0.77 0.38 0.88
0.61 + 0.09 3.47 f 0.31 2.00* 1.11
2.4 + 0.4 1.3 kO.2 1.2 f 0.4
kO.08 k 0.38 k 0.06 * 0.03
0.32 * 0.04
1.8 i-O.3 1.7 2 0.2 5.9 k 1.9 1.0+0.1 0.2 $- 0.03 1.3*0.2 0.9 * 0.3 -~
in dis-
?,OpO:Z1Op,,
of
in par-
----___ River RI R2 R3 Other (R2 R3, n =9)
45 f 2 15 + 0.4 24 k 6
Estuar! El E2 E3 E4 E5 E6 El E8
172f5 25 If: 1 76+2 15+2 7&2 22 * 1
E9 E IO Ocean 38”40’ 39”40’
26 52 43 88
f + i *
3 I 3 35
55 k6 193 & 16 84 If 6 56 + 5 243 f 18 51*5 85k 15 81 k7 50*5 28,4
N- lO”30’ N- lO”50’
W W
20* 23k
10 14
65 * 6 68 * 8 48 & 5 98k44
99*10 328 k 118*13 70 f 57 + 65 + 79+
25 8 6 7 19
I .03 0.07 0.89 1.02 0.22 0.42 0.18 0.21
* 0.02 * 0.04 f 0.06 * 0.03 * 0.01 2 0.03 +O.Ol + 0.02
0.29 0.28 0.73 0.27
1.15 0.38 0.67 0.04
+ + * k
f 0.22 k 0.26 & 0.11
0.80 0.92 3.26 7.33 3.88
i+ t & *
0.24 0.08 0.36 1.42 0.71
166 * 17
0.07 * 0.002 0.54 * 0.03
1.52 f 0.28 1.12 + 0.06
5.9 + 1.0
40 * 30 64 + 20
1.9 & 0.3 1.9 * 0.2
0.78 i 0.56 0.75 + 0.28
2.0 f I .8 2.8 k 1.9
3. Results
3.1. 226R~, ‘l”Pb muttcv
0.3 + 0.6 2.4 k 0.7 0.9 & 0.3
and “‘PO in water and suspended
Concentrations of these radionuclides in samples from the Tagus river, estuary and coastal sea are shown in Tables 1 and 2. Concentrations of total (dissolved and particulate) 226Ra varied in the range of 0.90-8.2 Bq m - 3, depending on the load of suspended matter and location of the sampling zone. The relatively high concentration of dissolved 226Ra in the freshwater station Rl is due to weathering of secondary uranium deposits in soils of the Tagus watershed in the region of Nisa (Fig. 1). Downstream of Rl (stations R2El), water samples showed concentrations of dissolved 226Ra around 1 Bq mp3, which are typical of the river water discharge into the estuary. The relatively high concentration measured at the up-
per estuary, station E2, was possibly influenced by waste from a cement factory and drainage of agricultural fields discharged at the narrow entrance of the estuary near Vila Franca. Throughout the estuary, concentrations of dissolved 226Ra increased seaward reaching a maximum of 3.85 Bq m - 3. These concentrations remain near constant in the river plume in coastal sea, dropping off rapidly to 1.2 Bq m ~ ’ in more distant continental shelf sea water and in the open ocean (Table 1; Fig. 3). Concentrations of 226Ra in suspended matter were also higher in the upper estuary. It is interesting to note that ‘*‘Ra concentrations in suspended particles decrease in a seaward direction, which corresponds to dissolution/desorption of particulate ‘16Ra with increased salinity (Table 2). Concentrations of total *lOPb ranged from 0.8 Bq m ~ 3 in freshwater to 4.8 Bq m - 3 in the estuary. As with 226Ra, these concentrations de-
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I,,,
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Salinity
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Fig. 3. Concentrations of dissolved 226Ra. Area between lines indicates concentration trends throughout the Tagus river-estuary-coastal sea system.
pend on the load of suspended matter. On average, concentrations of dissolved z’“Pb in river water were higher than in the estuary and were lower than concentrations of dissolved *lOPb in the coastal sea (Table 1; Fig. 4). Concentrations of ?‘OPb in estuarine particles were consistently higher than concentrations measured in riverborne and marine suspended matter (Table 2), confirming the enhanced scavenging of dissolved Z’OPb taking place in the estuary. Concentrations of total “‘PO varied from 0.6 to 5.0 Bq m ’ and, in the majority of the samples, a very high percentage of total *“PO was associated
I O.O’ L
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10
P I d-----J 20
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Salinity l:tg. 4. Concentrations ““PO copen circles) \V~lClll
in
of dissolved the Tayus
(g L-l) ““Pb (solid squares) river-estuary-coastal
and sea
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with particles (Table 1). Concentrations of dissolved 2’oPo were typically about 0.50 Bq m ’ in river water and, thus, generally higher than concentrations of dissolved “‘PO in the maximum turbidity zone of the estuary. In the upper and mid estuary (stations E22E6), ‘l”Po in solution further decreased to very low concentrations. < 0.01 Bq m - ‘, to increase again towards the sea from station E&E10 (Table 1; Fig. 4). Similar to “‘Pb, the concentration of “‘PO in suspended matter was generally higher in the upper estuary than both in riverborne and marine particles (Table 2). corresponding to enhanced particle sorption of 2’“Po occurring in that zone. More informative are the activity concentration ratios of these radionuclides in the soluble and particulate phases (Table 2). ‘10Pb:2’6Ra concentration ratios in the soluble phase of freshwater and low salinity water (upper estuary) were consistently higher than concentration ratios in mid and lower estuary zones, but gradually increased in the coastal sea. This is simultaneously due to enrichment of “‘Pb in estuarine suspended matter and to dissolution/desorption of particulate IZ6Ra in the estuary (Table 2). 2’oPo:“oPb ratios in the soluble phase decreased from freshwater (0.73) to the mid estuary ( < 0.01) where *“PO was almost entirely removed from solution. From the mid-estuary seaward, an increased Po:Pb ratio was observed, attaining values higher than unity at the mouth of the estuary and stabilizing at about 0.7 in the coastal sea. Since in the estuary no desorption of “‘PO from suspended matter was observed, the increased concentration of ““‘PO in the soluble phase is likely to originate in “OPo re-dissolution from bottom sediments and intertidal mudflats. The increased 2’oPo:“oPb ratios simultaneously observed in suspended matter at the mouth of the estuary and at coastal sea. sampled with the low tide outflowing water, further supports this interpretation. A common feature of the three radionuclides investigated is that total concentrations in water are chiefly regulated by the load of suspended matter. Furthermore, concentrations measured in suspended matter in the estuary are roughly identical to concentrations measured in the fine fraction ( < 63 jlrn) of bottom sediments suggesting
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that resuspension of bottom sediment is the main mechanism controlling the distribution of these radionuclides in the estuary (Carvalho, 1995a). Results from analyses of bottom sediments lead also to the identification of areas with higher than average ““Pb and ““PO background levels. Sediments in a very small area around Barreiro Peninsula, Lavradio, in the south bank, contain radionuclide levels enhanced by industrial waste releases. An old phosphate fertilizer plant during decades discharged phosphate wastes (process water and phosphogypsum) into self contained gypsum ponds on the river bank (Lavradio). After decantation of the gypsum, the overlying water generally containing little gypsum, much less radioactivity and with a pH near neutrality, was released into the estuary causing enhancement of radionuclide levels in local sediments. The total activity of radionuclides released is difficult to estimate but, since no evidence was obtained for technologically-enhanced radioactivity levels elsewhere in the estuary (Carvalho, 1995a) we conclude that it is a negligible input. Therefore, the measurements made in water samples collected in mainstream channels (this study) mainly reflect the distribution of naturally-occurring nuclides. However, an exception to this might be the concentrations of 2’6Ra in stations El -E2, which could be enhanced by point source discharges. The Distribution Coefficient (I&) of radionuelides between suspended solids and the dissolved phase (& = (Bq kg.-- ’ dry suspended matter)/(Bq kg-’ filtered water)) were computed using the concentrations given in Tables 1 and 2. Kd for “‘Ra in freshwater is (1.7 + 0.6) x lo4 and of the same order of magnitude in the upper estuary, but decreases to about 6 x lo3 in the mid estuary. Kd for ““Pb in freshwater is (3 + 1) x lo4 and (1.5 + 0.4) x lo5 in the estuary. K,, for 21”Po is (1.0 & 0.3) x 1Oj in freshwater and (9 t 5) x lo5 in brackish water. The I& values of both “‘PO and ““Pb in oceanic water seem to decrease to (l4) x 104. High Kd values for “‘Pb and “‘PO are a result of the strong adsorption of these metals onto suspended particles in aquatic systems, comprising of freshwater, estuarine and marine environments. Moreover, these values reflect the increased scavenging of dissolved “‘Pb and ‘l”Po
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in the estuary and the effect of particle concentration on the Kd values. 3.2. Preliminary Tugus estuary
balance oj’ “‘Pb
rind -“‘PO in the
Based on the radioanalytical results, we can estimate the apparent mean residence time of dissolved *l”Pb and “OPo in the estuary as well as the export of the soluble fraction of both radionuelides into the coastal sea. Since the geomorphology of the estuary (a broad inner bay with two narrow entrances) and the characteristics of a fully-mixed estuary (vertical mixing of water in the inner bay) meet the basic requirements of box-models, the use of this approach seems acceptable to make a mass balance of the rddionuelides. For this, we assume that measured concentrations of radionuclides (Tables 1 and 2) are representative of average annual concentrations and that the estuary may be represented by a single box-model with steady state conditions holding for at least several months. We start by considering the hydrology of the estuary. The flux of water through the estuary depends upon river discharges, 1.30 x 10’” m7 a -I, and direct rainfall into the estuary, 700 1 m-* a-‘. Taking into consideration the surface area of the estuary, direct rainfall contributes with 1.40 x lo* m3 a-‘, which only accounts for 1.1% of the annual river discharge. For simplification, urban sewage and industrial liquid effluents released into the estuary are considered equivalent in volume to evaporation. The average volume of water in the estuary was estimated at 2.0 x 10” m3. Assuming that this volume is constant, the outflow of water is identical to the inflow, i.e. 1.30 x 1O’Om3 a - ’ Therefore, the mean residence time of water in the estuary, R,, is 0.15 a (55 days) and the rate constant of water mass outflow, 2, = R, ‘, is 6.6 a-’ under average water flow conditions. The balance of dissolved “‘PO in the estuary depends upon discharges of dissolved “‘PO into the estuary, “‘PO produced through radioactive decay of dissolved *‘OPb, 210Po radioactive decay, 210Po removal from solution and “‘PO exported with the water discharged to the coastal sea. The annual supply of 2’oPo dissolved in river water
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(average concentration 0.58 f 0.08 Bq m ~ ‘) discharged into the estuary is 7.5 x IO9 Bq a - ’ contributing 3.7 Bq rnpi per year to the estuarine water. Atmospheric deposition of “OPo in the area is 8.3 2 2.8 Bq m ~~’ per year (unpublished), and it is a significant contribution that we consider as entirely soluble. In the estuary, the ingrowth of ““PO through radioactive decay of dissolved “‘Pb, was calculated from the inventory of 210Pb (based on the average “‘Pb concentrations in stations E2-E9), i.e. 1.6 x lo9 Bq. In the same way. the inventory of dissolved 2’oPo in the estuary, based on the average of concentrations determined in E2 through to E9, is estimated at 4.4 x 10’ Bq. Therefore, the mass balance equation for dissolved ?‘OPo in the estuary is 2,+APh + I, + I,, = &,APo + &A,,
+ &4,,
(1)
where, i,,, is the radioactive decay constant of ““PO, 1.83 a ‘, A,,, is the concentration of dissolved “OPb in the estuary, 0.84 Bq m ~ 3qA,, is the concentration of dissolved “‘PO in the estuary, 0.22 Bq m ‘, Z, is the annual input of dissolved 2’oPo in the river discharge, 3.7 Bq m ’ per year, Z, is the direct annual atmospheric deposition of 210Po to the estuary and diluted in the estuary volume. 0.83 Bq m 3 per year, 2, the scavenging rate constant for dissolved ““PO onto suspended matter and sediment particles, a ~ ‘, and i,,,, is the constant for water mass outflow into the coastal sea, 6.6 a ~ ‘. Substituting the values of the data into Eq. (1) and resolving for 2s gives a result of 20 a--’ and the mean residence time of “‘PO in solution relative to adsorption on sediment particles, is estimated at 0.05 a or 18 days. The other mechanism contributing to the removal of dissolved “‘PO from the estuary is the outflow of water into the sea. Taking into consideration that the inventory of dissolved “‘PO in the estuary is 4.4 x 10” Bq, the ““PO exported with a rate of i., = 6.6 a -- ’ is 2.9 x 10’ Bq a ~~‘. This export corresponds to about 30% of the annual delivery of dissolved “‘PO to the estuary. Therefore. most of dissolved s’“Po (about 70%) entering the estuary is removed by sorption onto sedimenting particles and trapped, at least temporarily. in the estuary. The concentration of dissolved “‘PO remaining with the water outflou will reduce to about 0.2 Bq m -’
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(without considering mixing with sea water) which is in agreement with the very low concentrations measured at stations E5 and E6 (Table 1). An identical mass balance equation may be written for dissolved “‘Pb with /1p’y4’.& + z, + z, = &,A,,
+ /i,Apb + i,,A,,
(2)
where, /i,, is the radioactive decay constant of 2’0Pb, 31.1 x 10W3 a-‘, ;Ipb A,, corresponds to the 2’oPb ingrowth from dissolved 226Ra in the estuary, and the other variables have for “‘Pb meanings similar to those given above for 2”‘Po. Based on the average “OPb concentration in filtered river water, 1.6 ) 0.6 Bq m ~ 3, the annual input of dissolved 2’oPb by river discharge is 2.1 x 10” Bq a ~ ‘, contributing to (Z,) 10 Bq m ’ per year in the estuarine water. The direct deposition of atmospheric 2’0Pb, 56 -t 12 Bq m em2per year, accounts to (ZJ 5.6 Bq rnp3 per year in the water of the estuary, corresponding to 56% of the dissolved 2’0Pb discharged by the Tagus river. The contribution of dissolved 226Ra through radioactive decay to total dissolved 2’0Pb is minimal, 0.6%, when compared with other “OPb sources (Z, and ZJ. Solving Eq. (2) in order to A,, 1, is equal to 12 a ~ ’ and the mean residence time of soluble 2’0Pb in the water of the estuary is 0.08 a or 30 days. The flux of dissolved 2’0Pb delivered to the coastal sea is computed at 1.l x 10” Bq a - ’ and the average concentration of dissolved 210Pb in the water outflow is 0.85 Bq m -‘. This discharge of soluble z’“Pb into the sea accounts for 35% of the annual input of soluble 210Pb received by the estuary. Therefore, most of the 2’0Pb (65%) entering the estuary in the soluble phase is likely sorbed onto suspended particles and trapped in sediment accumulation zones of the estuary, or discharged bound to suspended matter into the coastal sea.
4. Discussion and conclusions The distribution of dissolved 226Ra concentrations indicates a partial dissolution of 226Ra in riverborne particles and, eventually, in industrial wastes discharged into the estuary. This is further confirmed by the decrease of 226Ra concentrations in suspended matter with increased salinity. More-
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over, results show that the outflow of estuarine water is a source of dissolved 2’6Ra to coastal sea water. Similar trends for Ra dissolution in estuarine environments were reported before (Moore, 1981; Elsinger and Moore, 1983; Helz et al., 1985/ 86). In contrast with 226Ra, dissolved *‘OPb and *“PO entering the estuary are removed from solution in the freshwater/salt water mixing zone, probably by a combination of co-precipitation with humic substances, Fe-Mn hydroxides and by adsorption on to suspended particles (Sholkovitz et al., 1978). This is clearly shown by “‘Pb and 2’0Po concentrations in suspended matter which are higher in this mixing zone (Table 1; Carvalho, 1995a). Furthermore, results indicate that no noticeable desorption of these radionuelides occurs during the transport of suspended matter across the estuary. The mean estuarine residence times calculated for dissolved *“Pb and *“‘PO relative to sorption onto particles, 30 days and 18 days respectively, are in agreement with the few values reported for lakes, estuaries and coastal zones (Santschi et al., 1979; Benoit and Hemond, 1987). For instance, mean residence times for dissolved *“PO of l-5 days in winter and up to 2 months in the early summer have been reported for Narrangasett Bay, USA, whereas dissolved *“Pb displayed about 35 days in the same environment (Santschi et al., 1979). Our measurements and results from box-model computations indicate that direct deposition of atmospheric “‘Pb represents a significant contribution (36%‘) to the inventory of dissolved 2’0Pb in the estuary. On the other hand, the deposition of atmospheric *‘OPb over the much wider catchment basin of the Tagus River does not result in a significant input to the estuary, which implies an efficient immobilization of *“Pb and *“PO in the soils of the watershed. Furthermore. little “OPb and “‘PO in estuarine water is discharged in the soluble phase to the coastal sea. Instead, these radionuclides are mainly adsorbed by suspended matter and seem to accumulate in bottom sediments in the estuary. Previous results on *“Pb in sediment cores from the upper estuary indicated an excess of 2’0Pb relative to 226Ra above the average inventory supported by the flux of atmospheric *“Pb
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(Carvalho, 1995a). This provides additional verification on the enhanced scavenging of “‘Pb and 210Po in the maximum turbidity zone. However, sediment cores sampled in the mid estuary displayed lower than expected “‘Pb inventories, suggesting remobilization of deposited 2’oPb-2’oPo back into solution. Increased concentrations and increased 2’0Po:2’0Pb ratios in filtered water and suspended matter were measured in samples collected at the mouth of the estuary (Tables 1 and 2) confirming remobilization of both radionuelides and especially of *“PO. Although not investigated in estuarine environments, it is known that “‘Pb and larger amounts of *“PO are released from sediments when anoxic conditions develop (Benoit and Hemond, 1990). Whether these re-dissolved 2’oPb-2’oPo totally leave the estuary or are partly reabsorbed on to suspended particles and trapped in the estuarine circulation, can not be concluded from our results. In the light of results obtained from other estuaries, it seems possible that particles of marine origin enter the estuary, contribute to the scavenging of dissolved radionuelides and finally accumulate in the low salinity zone (Olsen et al., 1989). Furthermore, there is evidence that with average river flow the discharge of suspended matter from the Tagus estuary is minimal and particles are retained in the estuarine circulation (Vale, 1986). This allows for sediment accumulation inside the estuary and, therefore, for the build up of *‘OPb and *“PO reservoirs in sediments. The focusing of *“Pb deposition in upper estuary sediments was observed also in other estuaries such as the Mira estuary (unpublished) and the Tamar estuary (Clifton et al., 1995). These radionuclide reservoirs may intermittently be reduced by exceptionally large floods. In the Tagus estuary these floods may occur two to three times per century and carry out to the sea as much sediment in a few days as transported with average river flow during one year (Vale, 1986). Still in these circumstances the sudden transport of *lOPb and “‘PO to the coastal sea will be made in the particulate phase. The following general conclusions can be drawn on the implications of these results to the geochemical cycling of Ra, Pb and PO in phosphate wastes released as aqueous slurry into estuarine environments:
1. 226Ra in particulate form will dissolve in the estuary and will be discharged into the coastal sea in the soluble phase. In the Tagus estuary the mean residence time of dissolved 276Ra will be controlled by the residence time of water averaging 55 days; 3L. since the residence times of “OPb and ‘*‘PO in the soluble phase are much shorter than the water residence time, these radionuciides will be rapidly adsorbed onto suspended matter. In the Tagus estuary about 2/3 of dissolved radionuclide inputs will be trapped by particles and may give rise to enhanced radioactivity levels in sediment accumulation zones; 3. most of the ‘“‘Pb and Z’“Po flowing through the estuary will be exported to the coastal sea bound to bottom sediments and suspended matter rather than in the soluble phase. In the Tagus estuary, probably after several cycles of adsorption-redissolution. only l/3 of dissolved ““Pb and “‘PO in waste discharges will be transferred to the sea in the soluble phase. One implication of the rapid sorption of ““PO and “‘Pb on to sediments could be a reduced bioavailability of these radionuclides to estuarine organisms. This would prevent a significant accumulation of “‘Pb and ““PO from industrial phosphate waste discharges by estuarine biota. Actually, no enhancement of ‘IrJPb and “‘PO levels could be detected in biota in the Tagus estuary. Furthermore, field studies carried out in the Seine estuary in zones receiving phosphate wastes released as aqueous slurry and dumped as dry phosphogypsum, indicate no noticeable enhancement of ““Pb and “OPo levels in biota (Germain et al.. 1992, 1995). Other studies carried out in estuaries and coastal areas receiving discharges of low pH phosphate wastes (acid stream and gypsum slurries directly from phosphoric acid production) have shown significant enhancement of radioactivity levels in the aquatic environment (Koster et al., 1992; Pennders et al.. 1992; Perianez and Garcia-Leon. 1993: Poole et al., 1995). In these areas, elevated concentrations of these radionuclides were observed in biota samples collected near the point of waste releases (Roll0 et al.. 1992; Swift et al., 1995). Enhanced bioaccumulation, in particular of “(‘PO, probably takes
place before neutralization of highly acidic wastes and radionuclide sorption on to particles have contributed to remove radionuclides from the soluble phase to sediment. Nevertheless, delayed accumulation of these radionuclides from sediments by benthic and filter-feeding organisms and subsequent radionuclide food-chain transfer, can not be entirely excluded with the present knowledge (Carvalho and Fowler, 1993, 1994) and justifies specific investigation.
Acknowledgements The excellent technical assistance of Mr J.M. Oliveira and Mrs G. Albert0 (DGA/DPSR) is gratefully acknowledged.
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Osmond, J.K. and M. Ivanovich, 1992. Uranium-series mobilization and surface hydrology. In: M. Ivanovich and R.S. Harmon (Eds.), Uranium-series Disequilibrium. Applications to Earth, Marine and Environmental Sciences. Oxford Science Publications, Clarendon Press, pp. 2599289. Pennders, R.M.J., H.W. Koster and J.F. Lambrechts. 1992. Characteristics of L’“Po and “‘Pb in effluents from phosphate-producing industries: a first orientation. Radiat. Prot, Dosim., 43: 7377740. Perianez, R. and M. Garcia-Leon, 1993. Ra-Isotopes around a phosphate fertilizer complex in an estuarine system at the southwest of Spain. J. Radioanal. Nucl. Chem.. Articles, 172: 71--79. Poole, A.J., D.J. Allington, A.J. Baxter and A.K. Young, 1995. The natural radioactivity of phosphate ore and associated products discharging into the eastern Irish Sea from a phosphoric acid production plant. Sci. Total Environ., 1731174: 1377149. Rollo, S.F.N., W.C. Camplin, D.J. Allington and A.K. Young, 1992. Natural radionuclides in the UK marine environment. Radiat. Prot. Dosim., 45: 203-209. Santschi, P.H., Y.-H. Li and J. Bell, 1979. Natural radionuelides in the water of Narragansett Bay. Earth Planet. Sci. Lett., 45: 201-213. Sholkovitz, E.R., E.A. Boyle and N.B. Price, 1978. The removal of dissolved humic acids and iron during estuarine mixing. Earth Planet. Sci. Lett., 40: 130- 136. Swift, D.J.. D.L. Smith, D.J. Allington and K. Winpenny. 1995. A laboratory and field study of 2’“Po depuration by edible winkles (Littorina littorea L.) from the Cumbrian coast (north eastern Irish Sea). J. Environ. Radioact., 27: 13-33. Vale, C., 1986. Transport of particulate metals at different fluvial and tidal energies in the Tagus River estuary. Rapp. P.-V. Reun. Cons. Int. Explor. Mer, 186: 306-312. Vale, C. and B. Sundby, 1987. Suspended sediment fluctuations in the tagus estuary on semi-diurnal and fortnightly time scales. Estuarine. Coastal and Shelf Sci., 25: 4955508.