Previous Page Table of Contents Next Page


PART I
TECHNICAL AND REVIEW PAPERS (Contd.)

ESSAI D'ÉVALUATION DES RESSOURCES ICHTYOLOGIQUES ACTUELLES ET POTENTIELLES DANS LE BASSIN DE L'OURTHE (BASSIN DE LA MEUSE) EN BELGIQUE

J. C. Philippart1

Institut de Zoologie de l'Université de Liège, Service d'Ethologie Animale—Aquarium, Unité de Recherches Piscicoles, 22, quai Van Beneden, B-4020, Liege, Belgique

RÉSUMÉ

Ce rapport présente une estimation des ressources ichtyologiques actuelles (situation 1970–79) et potentielles du bassin de l'Ourthe (3 672 km2; zones à truite, ombre et barbeau) dans la perspective d'une évaluation économique de la pêche récréative et de l'incidence halieutique de la régression ou de l'altération des stocks. Les données de base sur les populations de poissons proviennent de nombreux inventaires quantitatifs par pêche à l'électricité. La faune ichtyologique du bassin de l'Ourthe comprend actuellement 32 espèces dont les populations représentent un stock de 127 tonnes pour l'ensemble du bassin: 80% de ce stock est du à S. trutta, T. thymallus, B. barbus, C. nasus, L.cephalus et L. leuciscus. La faune ichtyologique de l'Ourthe supérieure (barrage), de l'Amblève (pollution organique et chimique par une tannerie + aménagements hydro-électriques + barrages empêchant les mouvements des poissons) et de la Vesdre (pollution urbaine et industrielle aiguë depuis 1835) subissent des altérations qui se soldent par une perte de stock de 38,6 tonnes affectant principalement S. trutta, B. barbus, C. nasus et T. thymallus; le remplacement des espèces polluosensibles par des espèces polluorésistantes entraîne le statu quo du stock de L. cephalus et la légère augmentation du stock de L. leuciscus. Pour ces six espèces intéressantes pour la pêche, une perte de stock de 38,6 tonnes correspond approximativement à une perte annuelle de production de 16,1 tonnes.

ABSTRACT

This report is a first attempt to assess the effective (1970–79) and potential fish resources in the Ourthe River Basin (3672 km2; trout, grayling and barbel zones); it aims at providing an objective basis for an economical evaluation of recreative fishing and fishery incidences of the fish stock alterations caused by human activities. The data on the fish population were collected during numerous electro-fishing surveys. Thirty-two fish species were recorded from the Ourthe River Basin during the last decade. Their populations make up together a stock of 127 tons for the whole basin; 80% of this stock are contributed by S. trutta, T. thymallus, B. barbus, C. nasus, L. cephalus and L. leuciscus. The upper Ourthe (direct and indirect effects of a barrage), the Warche-Amblève river system (strong organic and chemical pollution caused by a tannery + hydro-electric impoundments + barriers to fish movements) and the Vesdre River (very strong urban and industrial pollution since 1835) are suffering major alterations of their fish fauna which result in a stock reduction of ± 38,6 tons. This stock reduction mainly concerns S. trutta, B. barbus, C. nasus and T. thymallus; these sensitive species are replaced by more resistant ones, the stock of which either remains relatively the same (L. cephalus) or slightly increases (L. leuciscus). When considering these six species only, the loss of annual production is 16,1 tons/year.

1 Chercheur qualifié du Fonds National belge de la Recherche Scientifique (F.N.R.S.).

INTRODUCTION

Depuis 1970, l'Unité de Recherches Piscicoles de l'Institute de Zoologie de l'Université de Liège étudie l'écologie, la dynamique et la production des populations de poissons dans les cours d'eau du bassin de la Meuse et spécialement du bassin de l'Ourthe (Philippart 1977a, b; 1979a,b,c; 1980a; Vranken en préparation). Des recherches concernent également l'incidence de la pollution de l'eau et des travaux sur les communautés de poissons (Philippart 1980b) ainsi que l'impact écologique et les aspects socio-économiques de la pêche récréative (Philippart 1979b; Gilon 1980).

Le présent rapport intègre ces résultats et propose des évaluations des ressources ichtyologiques actuelles et potentielles du bassin de l'Ourthe dans la perspective (CECPI 1977) d'une évaluation économique de ces ressources.

PRÉSENTATION DU BASSIN DE L'OURTHE

Le bassin de l'Ourthe (Fig. 1) couvre une superficie de 3 672 km2, soit environ 10% du bassin mosan. L'Ourthe prend naissance sur le plateau des Ardennes belges (altitude: 510 m) sous la forme de deux branches, l'Ourthe orientale (bassin versant: 326 km2; longueur: 45 km; débit moyen annuel en 1975: 1,6 m3/sec; pente moyenne générale: 5,2) et l'Ourthe occidentale (417 km2; 53 km; 3,2 m3/sec; 4,4) qui débouchent dans le lac du barrage de Nisramont (capacité: 3.106 m3; superficie: 47 ha) mis sous eau en 1958. De l'aval de ce barrage au confluent avec la Meuse à Liège, l'Ourthe proprement dite (135 km; 37,7 m3/sec; 1,5) reçoit deux affluents importants: l'Amblève (1 080 km2; 93 km; 13,0 m3/sec; 5,4) et la Vesdre (705 km2; 73 km; 7,9 m3/sec; 7,7) dans les bassins desquels existent quatre lacs artificiels de barrage (capacité et superficie totales: respectivement 71.106 m3 et 440 ha).

Fig. 1

Fig. 1. Situation du bassin de l'Ourthe, localisation des principales sources d'altération du milieu ( pollutions aigues; ■ barrage de Nisramont) et répartition des ressources ichtyologiques. Les évaluations des stocks se rapportent aux secteurs de rivière délimités par le signe ↕. Cercles: absence naturelle des poissons (pH acide); points noirs; élimination presque totale des poissons à cause de la pollution; pointillé: ichtyomasse des deux Salmonidae S. trutta et T. thymallus.

Les cours d'eau étudiés appartiennent aux zones à truite, à ombre et à barbeau selon Huet (1949) et représentent une gamme très variée de milieux au point de vue des caractéristiques physico-chimiques naturelles de l'eau.

La qualité des eaux dans l'entièreté du bassin est connue grâce à plusieurs études récentes (Descy et Empain 1979; Descy 1980; Empain et al 1980; Fabri et Leclercq 1977; Kaiser et al 1979) basées sur des analyses chimiques et biologiques. Mais ce rapport n'envisage que les modifications de la qualité de l'eau et du milieu physique (barrages et aménagements hydro-électriques divers) ayant une incidence majeure sur les ressources ichtyologiques et sur la pêche. Les zones critiques signalées sur la Fig. 1 et caractérisées au point de vue qualité des eaux par les données du Tableau 1 sont: l'amont et l'aval du barrage de Nisramont sur l'Ourthe supérieure, la Warche à l'aval du barrage de Robertville, une grande partie de l'Amblève et la totalité du cours de la Vesdre.

COLLECTE DES DONNÉES SUR LES POPULATIONS DE POISSONS

De 1976 à 1979, des sondages et inventaires démographiques par pêche à l'électricité furent effectués dans une centaine de stations choisies en fonction du type natural de cours d'eau (zone piscicole, largeur et débit, composition chimique de l'eau), de la nature et de l'intensité des altérations de la qualité de l'eau, du milieu physique (barrage, captages, chenalisation) et des populations de poissons (pêche). La plupart des inventaires démographiques quantitatifs consistent en dénombrements par la méthode des deux efforts de capture (Seber et Le Cren 1967; Laurent et Lamarque 1974). Dans les stations soumises à un seul effort de pêche, le peuplement réel est obtenu en appliquant un coefficient d'efficacité de pêche déterminé pour un milieu de même type et étudié de la même manière.

Pour l'Ourthe, une série d'informations quantitatives plus anciennes sont exploitées: un dénombrement par marquage-recapture effectué en 1973–74 dans un secteur de 10 km (Philippart 1977a,b; 1979a) et des inventaires (un effort de capture) réalisés par Micha (1971) en 1968–69 et par Huet et Timmermans (1966, 1976) en 1963–64.

Tableau 1. Qualité chimique et biologique des eaux dans quatre rivières du bassin de l'Ourthe où apparaissent des altérations majeures des communautés ichtyologiques. Sources: (a) Descy et Empain (1979); Descy (1980) (b) Empain et al. (1980); Mouvet (sous presse) (c) Kaiser et al. (1979).

Paramètres de pollution (moyennes pour la période indiquée)Période
(nombre d'analyses)
OurtheaAmblève zonesWarchebVesdrec
A
(n = 2)
B
(n = 2)
C
(n = 3)
(a) Oxygène dissous (% sat.)76–78(5)99821031146640
(minimum) (91)(34)(98)(97)(9)(16)
(a) pH76–78(5)6,97,07,17,57,27,4
(a) Ammoniac NH4 + (μN/l)76–78(5)71 1143322076622 374
(a) Phosphates PO4--- (μgP/l)76–78(5)46966060732 238
(a) Cadmium (μ Cd/l)78(2)(2,6)2,22,92,11,612,2
(a) Zinc (μg Zn/l)78(2)8,522334646253
(b) Chrome (μgCr+++/l)78(4)?2525618500?
(b) Chrome dans les Bryophytes 116 0251 87573812 600193
(mg Cr+++/kg sec)78(4)      
(a) Indice diatomique de Descy76–78(5)4,13,23,83,82,72,0
Indice biotique (invertébrés) 9–10----3–4 (c)
Pollution NulleOrganique + chrome + cadmiumOrganique + chrome + cadmiumOrganique + cadmium

a Ourthe en aval du barrage de Nisramont; les analyses chimiques et les indices diatomiques et biotiques révèlent une qualité d'eau satisfaisante pour les poissons mais il faut savoir que cette partie de la rivière a subi sporadiquement des “stress” qui ont eu de graves répercussions sur les communautés ichtyologiques: réduction du débit d'étiage due au captage d'eau potable et à la retenue d'eau en période de sécheresse (64,67,71,76); arrêt complet de l'écoulement de l'eau pendant 2–3 heures en 1977 pour faciliter la pose d'une canalisation en rivière (), importants lachers d'eau en mai-juin (reproduction) pour permettre des compétitions de kayak, eutrophisation du lac avec phénomènes de fleurs d'eau (ex 1976), pollution accidentelle à la chaux vive par la station d'épuration du barrage ().
b Avant le confluent avec l'Amblève.
c Zone de pollution maximale.

Dans chaque station de pêche, les caractéristiques écologiques du milieu sont déterminées afin d'établir les correspondances entre stations sur la base de critères morpho-dynamiques (pente, largeur, débit, sinuosité), physico-chimiques (température, pH, alcalinité, teneur en calcium, nitrates, phosphates) et biologiques (Indice biotique basé sur les Invertébrés benthiques, Indice diatomique de Descy et structure des communautés ichtyologiques). Ces informations constituent le matériel de base pour une étude statistique des facteurs influençant la structure (nombre et types d'espèces) et la biomasse absolue des populations et communautés de poissons.

ÉTAT ACTUEL DES RESSOURCES ICHTYOLOGIQUES

Composition de l'Ichtyofaune

D'après les renseignements disponsibles pour la période 1970 à 1979, l'ichtyofaune du bassin de l'Ourthe comprend 32 espèces; le Tableau 2 donne la liste de ces espèces; et la biomasse moyenne des principales populations en 1977–79.

Parmi les espèces introduites après ± 1850, celles notées xx sont acclimatées de manière certaine mais les autres (x) ne semblent se maintenir que par les rempoissonnements ou les introductions accidentelles (utilisation comme appât pour la pêche). Dans la catégorie des espèces disparues, il faut signaler Salmo salar (disparition vers 1930–40), Lotta lotta (1900?) et Lampetra fluviatilis (?). Une cartographie détaillée de la distribution géographique (présence-absence) et de la biomass absolue des populations de toutes les espèces de poissons du bassin de l'Ourthe est en préparation (Philippart 1979a; Vranken en préparation); la Fig. 2 montre les résultats obtenus avec S. trutta.

Stocks par Rivière et pour l'Ensemble du Bassin

Les données sur la biomasse des populations aux points d'échantillonnage ont été extrapolées à des cours d'eau entiers ou à des portions de cours d'eau bien différenciés et homogènes quant aux caractéristiques physiographiques (zones piscicoles), aux facteurs naturels de productivité (pH et teneur en calcium) et aux perturbations anthropiques.

Les résultats des évaluations sont illustrés par la Fig. 1 où les stocks (toutes espèces) par rivière et tronçon de rivière sont proportionnels aux surfaces des carrés. La sommation des stocks partiels indiqués sur cette figure donne, pour l'ensemble du bassin de l'Ourthe, un stock total de 127 tonnes qui se répartit comme suit entre les trois principaux sous-bassins: Amblève: 28,6 tonnes; Vesdre: 0,9 tonnes; Ourthe: 97,5 tonnes. La Fig. 3 montre la répartition par espèce du stock global.

Tableau 2. Biomasse des principales populations de poissons dans le bassin de l'Ourthe en 1977–1979 (69 stations) et liste des autres espèces présentes (période 71–79).

EspècesaNombre de stationsBiomasse (kg/ha)
MoyenneMaximale
Salmo trutta6371,2298
Thymallus thymallus4025,0242
Barbus barbus1054,2188
Chondrostoma nasus  855,2236
Leuciscus cephalus2731,5155
Leuciscus leuciscus1721,9171
Gobio gobio22  5,9  34
Rutilus rutilus1725,7204
Perca fluviatilis23  7,2  17
Esox lucius  7  6,1  18
Anguilla anguilla1917,0130

a Autres espèces: Salmo gairdneri (xx), Salvelinus fontinalis (x), Coregonus sp. (x?) Phoxinus phoxinus, Alburnoïdes bipunctatus, Alburnus alburnus, Scardinius erythrophthalmus, Abramis brama, Blicca bjoerkna, Tinca tinca, Cyprinus carpio, Carassius carassius, Carassius auratus, Gymnocephalus cernua, Stizostedion lucioperca (x?) Nemacheilus barbatulus, Cottus gobio, Gasterosteus aculeatus, Lepomis gibbosus (xx), Ictalurus melas ou nebulosus (x?) (Lampetra planeri, Cyclostme).

Fig. 2

Fig. 2. Biomasse absolue des populations de S. trutta dans le bassin de l'Ourthe (□: pH acide: aucune population permanente).

Fig. 3

Fig. 3. Composition par espèces du stock ichtyologique dans le bassin de l'Ourthe (lacs de barrage exclus). En blanc: stock actuel (1970–79): 127 tonnes dont 102 pour les six espèces principales: S. trutta, T. thymallus, B. barbus, C. nasus, L. cephalus, L. leuciscus; en noir: stock perdu: 39 tonnes pour la six espèces ci-dessus.

ALTÉRATIONS QUALITATIVES ET QUANTITATIVES DES COMMUNAUTÉS ICHTYOLOGIQUES

Ourthe Supérieure

Au cours des vingt dernières années, la faune ichtyologique de la zone à ombre (type inférieur) de l'Ourthe a subi de profondes modifications structurelles (Fig. 4) liées aux effets directs et indirects de la construction du barrage de Nisramont au confluent des deux Ourthe:

Pour l'ensemble du système Ourthe supérieure (133 ha), le bilan qualitatif et quantitatif de l'évolution du stock ichtyologique de 1964 à 1979 s'établit comme suit:

Warche-Amblève

La Warche, principal affluent de l'Amblève, reçoit les effluents (matières organiques + alcalis + chrome) d'une tannerie industrielle (Tableau 1); les poissons ont totalement disparu jusqu'au confluent avec l'Amblève et la perte de stock est évaluée (zone à truite-ombre) à 2,0 tonnes (1,9 t pour S. trutta et 0,1 t pour T. thymallus).

Fig. 4

Fig. 4. Évolution structurelle (biomasse relative des espèces) entre 1964 (recencements par Huet et Timmermans 1966) et 1979 des communautés ichtyologiques dans trois secteurs de l'Ourthe supérieure. Les stations 1 (Ourthe orientale) et 2 (Ourthe occidentale) sont situées à l'amont du barrage de Nisramont, la station 3, 11 km à l'aval.

Via la Warche, l'Amblève subit une pollution organique et chimique chronique dont les effets se combinent à ceux de divers aménagements hydro-électriques et de barrages physiques du cours. Trois zones sont reconnues (Fig. 5):

Zone A (42 ha)

Forte pollution organique et contamination par le chrome (Tableau 1): faune ichtyologique limitée à des espèces résistantes comme G. aculeatus (limite de tolérance à 15–18°C: 1,2 ppm Cr3+ d'après Erichsen Jones 1964, p. 69). Stock potentiel d'une zone à ombre du type supérieur (pente 4,4; largeur: 20–30 m) moyennement productive (15–20 mg Ca++/l) évalué à 7,6 t (S. trutta, T. thymallus et L. cephalus).

Zone B (67 ha)

Réapparition de S. trutta, T. thymallus, L. cephalus, G. gobio et R. rutilus mais ichtyomasse (30–50 kg/ha) nettement inférieure à la normale (180 kg/ha) à cause de la contamination par le chrome (>1 000 mg/kg de poids sec dans les Bryophytes d'après Empain et al 1980 et Mouvet sous presse), des pollutions organiques sporadiques et des fluctuations journalières du débit (conduite forcée d'une centrale électrique). Perte de stock estimée (zone à ombre) à 9,8 t.

Zone C (75 ha)

Pollutions organique et par le chrome fortement résorbées. Ichtyomasse quantitativement normale (>160 kg/ha) mais structurellement très déséquilibrée (remplacement des espèces polluo-sensibles B. barbus et C. nasus par les espèces polluo-résistantes L. cephalus et L. leuciscus) sauf à proximité du confluent avec l'Ourthe (station 7) où apparait une situation assez représentative du type de rivière qu'est la basse Amblève.

Fig. 5

Fig. 5. Profil longitudinal de la composition par espèces et de la biomasse absolue des communautés ichtyologiques dans l'Amblève. principales perturbations:

  1. pollution organique et chimique (Cr, Cd) aigue via la Warche
  2. cascade artificielle infranchissable par les poissons
  3. exutoire de la conduite forcée d'une centrale hydro-électrique (fluctuations anormales du débit)
  4. barrage de prise d'eau pour la conduite forcée d'une centrale hydro-électrique (réduction du débit normal dans un secteur de rivière de 7 km)
  5. exutoire de la conduite forcée
  6. barrage infranchissable pour les poissons jusqu'en 1978

Les lettres W (Warche), L (Lienne) et S (Salm) désignent les principaux affluents. Pour la station 7, l'histogramme des biomasses est à l'échelle ½.

Globalement, l'impact des altérations chimiques, hydrologiques et mécaniques dans le système Warche-Amblève se solde par une perte d'ichtyomasse de l'ordre de 19,5 t (Tableau 3).

Vesdre

La Vesdre comprend une partie amont (25 km) hostile pour les poissons (pH: 4,0–6,0) et une partie aval (48 km) gravement polluée (égoûts urbains, papeterie, lainerie, métallurgie des non-ferreux) depuis le début de l'époque industrielle (±1835). Dans cette zone polluée (Tableau 1), les sondages par pêche électrique révèlent que quelques espèces (N. barbatulus, P. phoxinus, L. cephalus et même S. trutta ou S. gairdneri) subsistent à l'aval des barragesdéversoirs, au confluent des ruisseaux non pollués et dans un tronçon en voie d'épuration; mais partout, les biomasses sont négligeables (40 kg/ha dans la zone épurée et rempoissonnée) et la fréquentation halieutique est faible.

En fonction des caractères physiographiques, de la zonation piscicole théorique (zone à ombre) et de la haute productivité potentielle (50 mg Ca++/l au confluent avec l'Ourthe) de la Vesdre inférieure, la perte en ichtyomasse est évaluée à 18,7 t (Tableau 3).

Bilan pour le Bassin de l'Ourthe

Dans l'ensemble du bassin de l'Ourthe, la perte d'ichtyomasse s'élève à 39 t, soit environ 30% du stock actuel (127 t) et 23% du stock potentiel (127 + 39 = 166 t). En tenant compte des remplacements d'espèces, le bilan global, en terme de perte de biomasse et de production annuelle, pour les six espèces dominantes s'établit comme indiqué dans le Tableau 3.

L'apport des lacs artificiels de barrage peut être évalué à 20–25 tonnes mais concerne des espèces (ex. R. rutilus, A. brama, P. fluviatilis) différentes de celles affectées par la réduction des stocks.

DISCUSSION ET CONCLUSIONS

Qualité des Évaluations

Bien que basées sur un échantillonnage très complet du bassin de l'Ourthe, les évaluations de stock présentées dans ce rapport sont perfectibles. Il s'agirait surtout d'étendre l'étude aux nombreux petits ruisseaux (moins de 3 m de largeur) qui n'ont pas été envisagés en première approche et d'affiner toutes les évaluations se rapportant aux milieux (parties profondes des rivières, lacs artificiels) où la pêche à l'électricité se révèle peu efficace (emploi de méthodes de capture et de dénombrement plus appropriées: pêche au moyen de filets et de pièges; dénombrements par marquage-recapture).

Dans l'Ourthe supérieure, une évaluation objective des modifications qualitatives et quantitatives de l'ichtyofaune est obtenue en comparant des inventaires démographiques à 15 ans d'intervalle. Mais dans l'Amblève et surtout la Vesdre où l'on ne dispose pas de références quantitatives sur la situation ancienne, les modifications de l'ichtyofaune sont mesurées en recourant à une méthode de similitude typologique, ce qui conduit inévitablement à un certain degré d'approximation dans les évaluations. Une analyse statistique des facteurs influençant la répartition et les caractères quantitatifs des populations de poissons devraient fournir prochainement des modèles simples (équations de régression multiple) permettant de prédire la densité et la biomasse “normales” des populations spécifiques ou de l'ichtyocénose dans tout milieu pour lesquels on connaît les paramètres physiographiques, physico-chimiques et biologiques déterminants.

Tableau 3. Bilan de la modification des stocks des principales espèces de poissons dans le bassin de l'Ourthe.

EspèceBiomasse (tonnes)Productiona
(tonnes/an)
OurtheAmblèveVesdreTotal
S. trutta--7,4-4,8-12,2-6,1
T. thymallus--1,6-1,7-3,3-1,7
B. barbus-3,4-6,5-4,2-14,1-4,7
C. nasus-2,3-4,1-2,7-9,1-3,5
L. cephalus+3,2+0,3-4,7-1,2-0,7
L. leuciscus+2,1-0,2-0,6+1,3(+0,6)
Total-0,4-19,5-18,7-38,6-16,1

a Rapport production/biomasse évalué à 0,5 pour les deux Salmonidae et repris de Philippart (1977b, 1979c) pour les autres espèces.

Conséquences Halieutiques

En admettant que les pêcheurs prélèvent annuellement 20% du stock (Philippart 1977b, 1979c), le potentiel et le déficit annuels d'exploitation halieutique du besoin de l'Ourthe sont respectivement de 25 t et 8 t. Ce déficit d'exploitation correspondrait à un manque à pêcher pour 3 200 personnes (prises moyennes: 2,5 kg/an, Philippart 1979b) dont les dépenses brutes annuelles s'élèveraient (Gilon 1980) à près de 30 millions de F.B. Mais ces premières évaluations devront être précisées par des études portant sur i) l'estimation exacte des prises par les pêcheurs dans les différents types de rivières et sur ii) les aspects économiques de la pêche (enquêtes globales + étude de sites).

Protection et Restauration des Communautés Ichtyologiques

Les mesures concrètes à prendre dans le bassin de l'Ourthe pour assurer la protection des stocks actuels et favoriser la restauration des stocks éliminés sont de trois types:

  1. mise en place d'une politique énergique d'épuration des eaux;

  2. aménagement de passes à poissons pour le franchissement d'obstacles (recolonisation); et

  3. transfert de population reproductrice ou, mieux, pisciculture de repeuplement pour les espèces sauvages (ex: B. barbus).

Au plan de la recherche, les efforts doivent porter sur i) l'évaluation quantitative complète et sytématique des stocks et leur surveillance continue, ii) la définition de critères de qualité d'eau et du milieu pour assurer le développement optimal des communautés ichtyologiques locales et iii) la mise au point d'outils mathématiques pour prévoir, qualitativement et quantitativement, les impacts sur ces communautés et populations de toute forme de modification (positive ou négative) future du bassin de l'Ourthe.

REMERCIEMENTS

Ces recherches ont été réalisée grâce aux autorisations de pêche à l'électricité accordées par l'Administration des Eaux et Forêts et avec l'appui financier du Fonds National belge de la Recherche Scientifique (F.N.R.S.), de la Commission provinciale de Liège du Fonds Piscicole et du Service d'Ethologie animale—Aquarium de l'Institut de Zoologie de l'Université de Liège.

RÉFÉRENCES BIBLIOGRAPHIQUE

CECPI (Commission Européenne Consultative pour les Pêches dans les Eaux Intérieures). 1977 Évaluation économique de la pêche sportive et commerciale. Rapport et documents techniques de la deuxième consultation européenne, Göteborg, Suède, 22–24 décembre 1975. EIFAC Tech. Pap., 26:186 p.

Descy, J.P. 1979 A new approach to water quality estimation using diatoms. Nova Hedwigia, 64:305–323.

Descy, J.P. 1980 Utilisation des algues benthiques comme indicateurs biologiques de la qualité des eaux courantes, pages 169–194 in P. Pesson, ed. La pollution des eaux continentales. Incidence sur les biocénoses aquatiques. Paris, Gauthier-Villars.

Descy, J.P., et A. Empain. 1979 Critères écologiques de la qualité des eaux courantes en Wallonie et établissement d'un réseau de surveillance. Laboratoire d'Hydrobiologie du Département de Botanique. Université de Liège, 17 p.

Empain, A., J. Lambinon, C. Mouvet et R. Kirchmann. 1980 Utilisation des bryophytes aquatiques et subaquatiques comme indicateurs biologiques de la qualité des eaux courantes, pages 195–223 in P. Pesson, ed. La pollution des eaux continentales. Incidence sur les biocénoses aquatiques. Paris, Gauthier-Villars.

Erichsen Jones, J.R. 1964 Fish and river pollution. London, Butterworths. 203 p.

Fabri, R., et L. Leclercq. 1977 Les ruisseaux de Haute Belgique. Caractéristiques physico-chimiques des eaux naturelles et polluées. Natura Mosana, 30(3):78–87.

Gilon, Ch. 1980 Enquête exploratoire sur la pêche sportive en Belgique. Étude sociologique des pêcheurs fédérés de la province de Liège et aperçu économique de la pratique de la pêche comme loisir. CECPI, Consultation technique sur la répartition des ressources ichtyologiques, Vichy, France, 20–24 avril 1980.

Huet, M. 1949 Aperçu des relations entre la pente et les populations piscicoles des eaux courantes. Revue Suisse d'Hydrologie, 11(3–4):332–351.

Huet, M. et J.A. Timmermans. 1966 La population piscicole de l'Ourthe (grosse rivière belge de la zone à ombre et du type supérieur de la zone à barbeau). Verh. Internat. Verein. Limnol., 16:1192–1203.

Huet, M., et J.A. Timmermans. 1976 Influence sur les populations de poissons des aménagements hydrauliques de petits cours d'eau assez rapides. Trav. Stat. Rech. Eaux et Forêts Groenendaal, série D, 46:1–27.

Kaiser, R., I. Jadot et L. Terroir. 1979 Bioindicateurs animaux, page 6 in J. Lambinon. La Vesdre aujourd'hui. Association Vesdre-Nature, Verviers: 8 p.

Laurent, M. et P. Lamarque. 1974 Utilisation de la méthode des captures successives (De Lury) pour l'évaluation des peuplements piscicoles. Ann. Hydrobiol., 5(2):121–132.

Micha, J.C., 1971 Étude des communautés piscicoles dans l'Ourthe liégeoise. Tribune du Cebedeau, 326:1–7.

Mouvet, C., Pollution de l'Amblève par les métaux lourds, en particulier le chrome: dosage dans Sous presse les eaux et les bryophytes aquatiques.

Philippart, J.C., 1977a Contribution à l'étude de l'écosystème “Rivière de la zone à barbeau supérieure”: Densité, biomasse et production des populations de poissons de l'Ourthe, pages 551–567 in P. Duvigneaud et P. Kestemont, eds. Productivité biologique en Belgique, Travaux de la Section belge du Programme Biologique International. Paris-Gembloux, Duculot. 617 p.

Philippart, J.C., 1977b Contribution à l'hydrobiologie de l'Ourthe. Dynamique des populations et production de quatre espèces de poissons cyprinidae: Barbus barbus (L.), Chondrostoma nasus (L.), Leuciscus cephalus (L.), Leuciscus leuciscus (L.). Thèse de Doctorat en Sciences Zoologiques, Université de Liège: 225 p.

Philippart, J.C., 1979a Évaluation de ressources piscicoles et halieutiques dans les rivières du bassin de la Meuse. Définition d'une méthode d'étude, pages 481–496 in L. Calembert, ed. Problématique et gestion des eaux intérieures. Liège, Editions Derouaux. 967 p.

Philippart, J.C., 1979b Introduction à l'étude des aspects écologiques et socio-économiques de la pêche sportive. Enquête sur la pêche récréative dans l'Ourthe à Hamoir. Bulletin de la Société géographique de Liège, 15:229–250.

Philippart, J.C., 1979c Sport fisheries, fish ecology and fishery research in the inland waters of Belgium with special reference to the River Meuse Basin (river Ourthe sub-basin), pages 32–42 in Proceedings of the 10th annual study courses of the Institute of Fisheries Management. Nottingham University, 18–20 September 1979. 304 p.

Philippart, J.C., 1979d Étude des populations de poissons dans trois ruisseaux oligotrophes du bassin de la Roer supérieure (Belgique). Bulletin de la Société des Sciences de Liège, 48(5–6):210–225.

Philippart, J.C., 1980a Démographie du hotu, Chondrostoma nasus (L.) (Pisces: Cyprinidae) dans l'Ourthe. Communication présentée à la séance du 15 décembre 1979 de la Société Royale Zoologique de Belgique (à paraître dans les Annales de la Société Royale Zoologique de Belgique).

Philippart, J.C., 1980b Incidence de la pollution organique et de l'eutrophisation sur la faune ichtyologique de la Semois. Annls Limnol., 6(1):77–89.

Seber, G.A.F., et E.D. Le Cren. 1967 Estimating populations parameters from catches large relative to the population. J. Anim. Ecol., 36:631–643.

Symoens, J.J. 1957 Les eaux douces de l'Ardenne et des régions voisines. Bull. Soc. R. Bot. Belg., 89:111–314.

Timmermans, J.A. 1974 Étude d'une population de truites (Salmo trutta fario L.) dans deaux cours d'eau de l'Ardenne belge. Trav. Stat. Rech. Eaux et Forêts Groenendaal, série D, No. 43:52 p.

Vranken, M., En prep. Cartographie de la répartition des poissons dans le bassin de l'Ourthe et les régions adjacentes.

POTENTIAL EFFECTS OF RADIOACTIVE RELEASES TO THE AQUATIC ENVIRONMENT1

T. R. Rice, D. W. Engel and F. A. Cross

Beaufort Laboratory, Southeast Fisheries Center, National Marine Fisheries Service, National Oceanic and Atmospheric Administration, Beaufort, North Carolina 28516 USA

ABSTRACT

The increasing amounts of radioactive waste presently being stored and the possibility for its accidental or controlled release to the aquatic environment emphasizes the need for assessing the potential dangers of radioactivity upon fishery organisms. In addition, there is a continuing possibility of radionuclides being released at any point in the fuel cycle from the mining of ore to the reprocessing of fuel elements after use in nuclear power reactors. Once radioactivity enters the environment, only physical decay can eliminate it. Radioactivity released to the aquatic ecosystem is taken up by both the biotic and abiotic components with most becoming associated with the abiotic portion. That portion which is accumulated by the biota is cycled through the food web and may be concentrated at any trophic level since organisms have different requirements or capacities to accumulate the different radionuclides. One of the most important concerns is how much radioactivity is required to affect adversely aquatic organisms. Both short- and long-term exposure, depending upon the level of radioactivity, can affect individual organisms as well as populations. Thus there is need for concern about the future handling of radioactive waste and the use of nuclear power due to their potential for impacting the environment. Decisions on radioactive waste disposal and the use of nuclear power should be based upon rationalism, not emotionalism.

RÉSUMÉ

Les quantités croissantes de déchets radioactifs que l'on entrepose actuellement en milieu aquatique et la possibilité que ces déchets y soient libérés accidentellement ou sous contrôle accentue la nécessité d'évaluer le danger potentiel de radioactivité pour les organismes marins. En outre, il y a risque permanent de rejet de radionuclides à n'importe quel stade du cycle, depuis l'extraction du minerai jusqu'au retraitement des déchets radioactifs. Une fois que la radioactivité a pénétré dans le milieu, seule la dégradation naturelle peut l'éliminer. Un certaine partie de la radioactivité libérée dans l'écosystème aquatique est absorbée par les composants biotiques mais la majeure part s'associe aux éléments abiotiques. La portion qui est accumulée par les organismes vivants parcourt toute la chaîne alimentaire et peut être concentrée à n'importe quel niveau trophique car les organismes n'ont pas les mêmes besoins ni les mêmes capacité d'accumuler les différents radionuclides. Un des points les plus importants à connaître est la dose de radioactivité qui est nuisible pour les organismes aquatiques. L'exposition, tant courte que de longue durée, selon le niveau de radioactivité, peut affecter les divers organismes ainsi que les populations. Ainsi, il faut se préoccuper pour l'avenir de la manutention des déchets radioactifs et de l'utilisation de l'énergie nucléaire étant donné les risques potentiels qu'ils représentent pour l'environnement. Les décisions concernent l'élimination des déchets radioactifs et l'utilisation de l'énergie nucléaire devraient être prises sur des bases rationnelles et non émotives.

1 Contribution number 80-44B, Southeast Fisheries Center, National Marine Fisheries Service, NOAA, Beaufort N. C. 28516.

INTRODUCTION

In recent years, man has added significant amounts of radioactivity to the environment, which could be potentially damaging to human beings and to the biota in freshwater and the oceans. Since many aquatic organisms are used as food by humans, and these organisms can accumulate large amounts of radioactive materials relative to their concentrations in the water, there is need to know the rates and levels of accumulation of radionuclides by the aquatic organisms. Further, since there is a possibility that radioactivity can be released in large quantities in limited areas of the aquatic environment as a result of accidents, there also is a need to know the amounts of radionuclides that can be returned in fishery products and the levels of radioactivity that may result in significant radiation effects on aquatic organisms.

The aquatic environment represents more than three-fourths of the earth's surface and has received a proportionate amount of fallout from nuclear weapons testing. Much of the fallout on land has leached from the soil, and has been carried through runoff to freshwater streams, and then to the oceans, while other radioactive materials have been released directly into the aquatic environment from nuclear reactors and radioactive waste outfalls. The oceans, which serve as a sink for many pollutants, may be destined to receive more radioactive wastes now stored on land. At the present time, the U.S. Government is considering the possible use of the seabed at great depths as a place to dispose of high-level radioactive wastes in specially designed containers.

Radioactivity is of concern because our senses cannot detect it and because radiation effects on man are being observed as a result of the atmospheric explosion of nuclear weapons during and following World War II. Also, once radionuclides are released into the aquatic environment, only the passage of time can reduce their radioactivity. Each radionuclide has its own rate of decay or half-life (time required for one half of that present to decay), which ranges from minutes to thousands of years. The detection of radioactivity in fish (Saiki et al. 1957) first made man aware that contaminants released into the oceans could return to him in seafood. Prior to this, the oceans had been considered so large that any unwanted material could be disposed of there safely. Because of the suspected dangers to man from such disposal, funding in the U.S. became available from the federal government for research on the cycling and effects of radiation so that radionuclides became the best understood and controlled of all the pollutants that man has produced.

The accumulation of radionuclides and their transfer through food chains have been followed in both laboratory and field studies (Cross et al. 1975). In the laboratory, it has been possible to determine the effects of food, feeding, temperature, salinity, pH, isotope dilution, and concentration of chemically similar elements upon the accumulation, assimilation and tissue distribution of radionuclides. Also, accumulation and food chain studies have been carried out in the aquatic environment where radionuclides have been released through nuclear weapons testing, reactor operation and radioactive waste release.

The effects of radiation upon aquatic organisms and upon humans from eating these organisms also has been intensively investigated (Foster et al. 1971; Templeton et al. 1971). Laboratory experiments at relatively high levels of radioactivity and field observations in areas where radioactive wastes have been discharged over long periods of time have been used in an effort to determine doses of radiation that are harmful to organisms. The restrictions that have been set by several nations on levels and rates of release of each radionuclide to the aquatic environment are based upon the radionuclides return to man from drinking water and food.

While this paper is concerned with the cycling of radionuclides and effects of radiation, it should be emphasized that it is not an exhaustive review but covers primarily those aspects of radioecology of relevance to fishery biologists who are not familiar with this discipline.

SOURCES OF MAN-MADE RADIONUCLIDES

There are a number of sources of man-made radionuclides (Table 1) that are derived either directly from nuclear fission or fusion, or indirectly through neutron bombardment of stable elements (induced) (Table 2). The largest amount entering the aquatic environment to date, by more than two orders of magnitude, has been from atmospheric fallout from weapons testing (Preston 1972). Probably the greatest future source of radioactivity to the aquatic environment will be from nuclear reactors, assuming that the world will not become involved in a nuclear war and that atmospheric nuclear testing does not resume to any significant level. Because of the volumes of water required to cool nuclear reactors, most have been built on lakes, rivers, estuaries and oceans. The majority have been designed to produce electrical power, nuclear fuel and explosives for weapons. The type and size of the reactor will determine the amount and type of waste that may be released into the aquatic environment (Blaylock and Wither-spoon 1978). Releases of radionuclides to the environment from a power reactor originate primarily from scheduled low-level emissions during normal operation, but there is always the possibility of accidental releases. With the exception of accidents, the annual releases of radioactivity have accounted for only small percentages of the legal permissible concentrations (Rice and Baptist 1974).

Table 1. Sources of radionuclides to the aquatic environment through man's activities (modified from Joseph et al. 1971).

SourceActivity
Nuclear fuel cycleMining, uranium processing, fabrication, spent fuel reprocessing
Nuclear reactorsElectric power generation, ships, satellite power, research, plutonium production, wastes
Nuclear explosivesMilitary and civilian applications
Encapsulated radioisotopes (power)Marine navigation aids, weather stations, artificial human organs
Encapsulated radioisotopes (radiation)Medical radiology, industrial radiography, research
RadionuclidesMedical uses and research activities

Table 2. Important man-made radionuclide in the aquatic environment and their half-lives (modified from Woodhead 1973).a

Hydrogen-312,3y
Carbon-145 730y
Manganese-54303d
Iron-552,6y
Cobalt-57270d
Cobalt-605,26y
Nickel-6392y
Zinc-65245d
Strontium-8952d
Strontium-9028,1y
Zirconium-9565,5d
Niobium-9535,0d
Ruthenium-10339,5d
Ruthenium-106368d
Silver-110 m255d
Antimony-1252,7y
Iodine-1318,05d
Cesium-13730,0y
Cerium-14133d
Cerium-144284d
Promethium-1472,62y
Europium-1551,81y
Plutonium-23923 309y

a d = day, y = year

Radioactive wastes can be disposed of as low-level liquid wastes or as high-level encapsulated wastes. Liquid wastes can be released on a routine basis from reactors or nuclear fuel reprocessing plants. These releases should fall within the International Commission on Radiological Protection (ICRP) guidelines for releases of radioactivity. Some of the most abundant liquid wastes may contain tritium (hydrogen-3), the fission products strontium-90, zirconium-niobium-95, cesium-137, cerium-144 and the induced radionuclides managanese-54, iron-55, cobalt-60 and zinc-65. High-level packaged wastes have been disposed of on the deep ocean seabed. Included in these wastes are radionuclides that have half-lives on the order of tens of thousands of years. Consideration is now being given to burying these wastes within the deep seabed sediments, which will reduce the probability of their release to the water column and their return to surface waters.

CYCLING

The biogeochemical processes that control the distribution of stable elements in the aquatic environment also control the distribution of radionuclides. Upon entering the aquatic environment, radionuclides can remain in solution or in suspension, precipitate and settle to the bottom, or be taken up by plants and animals. Certain processes interact to dilute and disperse these materials, while other processes simultaneously tend to concentrate them. Currents, turbulent diffusion, isotopic dilution and biological transport dilute and disperse radionuclides. Concentrating processes may be biological, chemical or physical. Radionuclides are concentrated biologically through uptake and assimilation by aquatic organisms and chemically and physically by adsorption, ion exchange, co-precipitation, flocculation, and sedimentation through the interaction of such biotic and abiotic processes. Radionuclides also are cycled through water, sediment, and biota, and each radionuclide tends to take a characteristic route and rate of movement through these components or reservoirs of the aquatic environment (Rice et al. 1965; Wolfe et al. 1973).

In freshwater systems and estuaries, chemical and biological processes tend to concentrate radionuclides. At the freshwater-saltwater boundary zone in estuaries, hydrological and physiochemical conditions can significantly influence the availability of radionuclides to the biota (Cross and Sunda 1978). The shallowness of estuarine and most freshwater habitats enhance the role of benthic communities in the exchange of radionuclides between sediments and water (Wolfe and Rice 1972). Such factors must be considered in estuaries, coastal waters and freshwater environments to a greater extent than in the open ocean. In the open ocean and deep lakes, however, thermal stratification, depth of water and circulation patterns make the sediments and benthic communities relatively unimportant in the cycling of radionuclides. In these systems, elements that are exchanged between the sediment and water may take thousands of years to reach the surface waters.

Sediments may accumulate radionuclides through the physical processes of exchange and adsorption (Duursma and Gross 1971). In effect, sediments and biota compete for radionuclides present in water. Although in some instances, sediments initially remove large quantities of radionuclides from the water and thus prevent their immediate uptake by the biota, this sediment-associated radioactivity later may affect many benthic organisms by exposing them to radiation (Woodhead 1973). Radionuclides also leach from the sediments back to the water and again become available for uptake by the biota. Even though radionuclides are associated with the sediment they may become available to the biota due to variation in the strength of the binding between the different radionuclides and the sediment particles. Loosely bound radionuclides on sediments can be stripped from particles of sediment and utilized by bottom-feeding organisms ingesting sediments (Luoma and Jenne 1976).

Aquatic plants and animals also play an integral part in the cycling of radionuclides (Fig. 1). They accumulate radionuclides by adsorption, absorption and ingestion. Conversely, radionuclides can be lost by desorption, excretion and decomposition. For example, even if an organism accumulates and retains radionuclides but dies, the radionuclides will be released back into the environment through organisms that decompose the dead organic material into its elemental components (Rice and Baptist 1974). In addition, radionuclides that are adsorbed or ingested but not assimilated by aquatic animals can be transported downward in the water column with fecal material (Osterberg et al. 1963) or in cast exoskeletons of pelagic crustacea (Fowler and Small 1967).

Fig. 1

Fig. 1. Processes involved in uptake and loss of radionuclides by marine biota (after Rice and Baptist 1974).

The extent to which fish can accumulate radionuclides depends upon their availability and the physical state of the radionuclides in the water. Radionuclides from food have been shown to be more available to fish than from the water (Jeffries and Hewett 1971; Pentreath 1973a, 1973b, 1973c). Several fission products—ruthenium-106, cerium-144 and zirconium-niobium-95—which are relatively insoluble, are poorly absorbed across the gut wall of fish (Pen-treath 1973d). Baptist and Hoss (1965) found that less than 1% of the cerium-praseodymium-144 ingested by croaker Micropogon undulatus was assimilated. The biologically significant induced radionuclides—managanese-54, iron-55, cobalt-60 and zinc-65—are assimilated much more readily across the gut wall than are the fission products discussed above (Osterberg et al. 1964; Cross et al. 1975). In addition, freshwater fish also have the capacity to accumulate relatively high levels of the radionuclides of strontium and cesium due to the relatively low levels of calcium and potassium in fresh water (Preston 1972). Cross et al. (1975) have reviewed the transfer of radionuclides through marine food webs leading to fish.

EFFECTS OF RADIOACTIVITY

While the accumulation of radioactivity by aquatic organisms is of importance because of their use as food by man, concern also has been expressed about the exposure of aquatic organisms to radioactivity from the water, sediments and from the radioactivity contained within the organisms themselves. To predict the possible effects of accidental releases of radioactivity on individual organisms and populations, it is necessary first to know if releases to date have had any impacts. Various approaches to the study of the effects of ionizing radiations on aquatic organisms have been used. For example, acute and chronic laboratory exposures, the study of animal populations in areas receiving radioactive waste and population dynamic models utilizing stressed populations (Ophel et al. 1976; Templeton et al. 1976; Blaylock and Trabalka 1978) have been used in an effort to reach a better understanding of radiation levels harmful to aquatic organisms.

In recent years, several review articles have summarized laboratory radiation experiments on aquatic organisms where acute radiation doses were used, and both lethal and nonlethal responses were examined (Polikarpov 1966; Rice and Wolfe 1971; Templeton et al. 1971; Chipman 1972; Rice and Baptist 1974; Ophel et al. 1976; Blaylock and Trabalka 1978). As Ophel et al. (1976) pointed out, most laboratory experiments involve acute exposures to radiation delivered over a short period of time. The data generated by such experiments show that aquatic organisms have LD50s (50% lethal dose), which range from 102 to 106 rads,2 and relate very poorly to environmental situations (Cross 1978; Wood-head et al. 1976) where the dose rates are low (0,6 to 3 400 micro rads/h) and the time periods are long (> year).

Attempts have been made to design chronic laboratory irradiation experiments so that exposures include a significant portion of the organism's life cycle. These experiments have used both sealed external cobalt-60 and cesium-137 sources and radionuclides released into the water (Blaylock and Trabalka 1978). Donaldson and Bonham (1964, 1970) irradiated coho and chinook salmon eggs (Oncorhynchus kisutch and O. tshawytscha) from fertilization through 80 days and looked for differences between the irradiated and control groups. No demonstrable effects were noted at a dose rate of 0,5 R/day,3 but at 10 R/day significant changes in sex ratios were observed, and at 20 R/day no females occurred. Other investigators have used fish, crabs, and snails in laboratory chronic irradiation experiments and have shown few adverse effects except at the highest dose rates used (Engel 1967; Cooley and Miller 1971; Kaufman and Beyers 1973).

In other attempts to design irradiation experiments that can be related to environmental radioactive contamination, some investigators have exposed different developmental stages of aquatic organisms to radionuclides dissolved in seawater. Developing eggs of several species of fish have been used extensively because eggs have been shown to be the most sensitive stage in the life cycle of fish. For example, Brown and Templeton (1964) did not observe any significant effect of protracted irradiation on the hatching of eggs or development of the plaice Pleuronectes platessa at dose rates ranging from 0,01 to 1 R/h or to total dose up to 500 R.

Some investigators have attempted to produce radiation damage in fish through the use of high concentrations of internally deposited radionuclides. For example, rainbow trout Salmo gairdneri showed no detrimental effects from body burdens of zinc-65 and phosphorus-32 from 100 to 10 000 times greater than the concentrations occurring in the Columbia River in 1965 (Foster and Soldat 1966). At higher levels, however, both zinc-65 and phosphorus-32 caused damage to the blood cell producing tissues in the trout (Nakatani 1966). A discussion of effects from internally deposited radionuclides and external radiation on aquatic organisms has been compiled by Templeton et al. (1971) and Ophel et al. (1976).

One of the most direct methods of obtaining an understanding of the impact of radioactive releases on natural populations is to study, intensively, environments where radioactivity has been released intentionally. By following changes in population size of resident organisms in such areas, it should be possible to determine the biological effects of accidental releases. Three such areas where intentional releases have been studied in considerable detail are the White Oak Lake at the Oak Ridge National Laboratory, Tennessee, the Columbia River downstream from the Hanford weapons production plant in Richland, Washington, and the Irish Sea in the vicinity of the Windscale reprocessing plant.

White Oak Lake was established as a radioactive waste settling basin for the production reactors at Oak Ridge, and as a result the organisms in the lake have been exposed to long-term chronic irradiation from 1943 to the present. Numerous investigations have been conducted on the fish and invertebrate populations of the lake. These investigations primarily have been concerned with the effects of ionizing radiations on the gene pools of aquatic populations.

Three natural populations of aquatic organisms studied intensively from 1963 to the present are the midge Chironomus tentans, the snail Physa heterostropha, and the mosquitofish Gambusia affinis. These species were exposed to low-level chronic irradiation for many generations. The Chironomus population was studied from 1960 through 1970 and was exposed to decreasing dose rates that ranged from 230 rad/year to 11 rad/year (Blaylock and Trabalka 1978). The irradiation caused an increase in the frequency of aberrations in the giant salivary chromosomes of the midge larvae and the frequency of aberrations was dose rate dependent (Blaylock 1966). When the dose rate decreased to 11 rad/year or less due to the decay of the radioactivity in the lake, the numbers of aberrations decreased and did not differ from the control population. Cooley and Nelson (1970) and Cooley (1973), using a natural population of the snail Physa heterostropha, demonstrated that a dose rate of 0,65 rad/day reduced egg capsule production when compared to an unir-radiated group. Egg production, however, was similar in both groups because there were more eggs per capsule in the irradiated population. The genetic variability of the mosquitofish Gambusia affinis also has been studied intensively (Blaylock 1969; Blaylock and Frank, in press). When irradiated and controlled populations of fish were compared, a larger brood size occurred in the irradiated population; there also was a significantly higher frequency of dead and abnormally formed embryos. It was suggested that the increased fecundity of the irradiated fish population compensated for the increased mortality caused by radiation. Trabalka and Allen (1977), however, demonstrated that laboratory-reared fish from the irradiated and controlled populations had the same fecundity but that the irradiated fish had many deleterious genes. At the same time the population in the field was thriving, which indicated that measurements of radiation damage such as genetic load may not be a valid index of population fitness.

The construction of the Hanford plutonium production reactors on the Columbia River resulted in the average release of 1 000 Ci/day4 of radioactivity directly into the River (Osterberg 1975). This plant used river water for once-through cooling, and the radionuclides released were primarily from the neutron activation of stable elements in river water, antifouling agents and corrosion products. The isotopes of greatest importance were phosphorous-32, chromium-51 and zinc-65. From the time the plant began operation in the mid-1940's until the last production reactor was shut down in 1971, relatively large quantities of radioactivity were released into the river. Large-scale research projects at universities and neighboring government institutions were organized to follow the cycling of the released radioactivity in the river, estuary and adjacent Pacific Ocean. The overall result was that no effects on the aquatic biota were detected from the radioactivity accumulated by various organisms of the food chains in the river, estuary or adjacent ocean (Osterberg 1975). Much valuable information on the partitioning of radionuclides in aquatic ecosystems and the coastal oceanographic processes of the river plume, however, came from these investigations.

Since about 1956 the Windscale reprocessing plant and reactors on the west coast of England have released low-level radioactive waste into the northeast Irish Sea, a productive fishing ground. During investigations conducted to determine whether released radioactivity would affect the resident population of plaice Pleuronectes platessa, Woodhead (1970) calculated the doses of radiation a developing plaice egg would receive from the sediment and from radioactivity adsorbed to the chorionic membrane. Laboratory experiments were conducted at ambient and higher dose levels to determine the effects on hatching and larval development. No significant damage occurred at a total integrated dose to the eggs of approximately 0,18 rad. In the environment the total dose rate to the eggs from released radionuclides was 9,32 × 10-2 micro rad/h, while the dose from the natural occurring radionuclide 40K was 6,96 × 10-1 micro rad/h is almost an order of magnitude greater. Templeton et al. (1976) concluded that the release of radioactivity to the environment at Windscale had not had any measurable effect upon the plaice population. These investigations have demonstrated that at higher levels of radioactivity in the natural environment, it was difficult to demonstrate any effects on the resident populations of organisms. Although no damage to population structure was detected, however, it does not eliminate the possibility that subtle changes in an individual's genome might have occurred. In nature, probably almost any mutation that weakens an animal will cause its death and therefore will result in the mutation being eliminated from the gene pool.

The failure to detect radiation-induced somatic changes at either individual or population levels in contaminated environments does not necessarily prove that no effects have occurred. Instead, it may reflect our lack of understanding of the natural variability within the system. It can be stated, however, that no catastrophic mortalities have occurred and that the subtle changes, which might be suspected, were not observed due to long-term fluctuations in the ecosystem.

In attempts to use fishery statistics and data, Zaystev and Polikarpov (Rice and Angelovic 1969) calculated the length of time required to reduce fish populations 50% if various percentages of the spawned eggs were killed by radiation. The calculations assumed constant levels of mortalities (i.e., 10% a year), constant levels of fishing and a set recruitment rate dependent upon spawning population size. Such assumptions are not valid, however, for many populations of marine fish because of natural and man-induced variability (Beverton and Holt 1957).

Possible effects of radiation at the population level were investigated by using knowledge of the population dynamics of selected commercial fish species by an International Atomic Energy Agency Panel in a report entitled “Effects of Ionizing Radiation on Aquatic Organisms and Ecosystems” (Templeton et al. 1976). The panel considered the role of density-dependent mortality in the stock-recruitment relationship in marine populations of both high and low fecundity (Fig. 2). Using commercially exploited fish stocks as an example, the authors concluded, “If mortality of eggs is being enhanced by the low levels of irradiation presently existing in the marine environment, then recruitment to the stocks of highly fecund marine species of fish is unlikely to be adversely affected unless those stocks are already at risk because of severe over-exploitation.” In other words, the result of mortalities of eggs caused by irradiation would decrease larval competition for food and space and, therefore, would increase the probability of survival for the remaining individuals.

Survival rates for highly fecund density-dependent fish stocks increase dramatically at low stock sizes (Fig. 3). The curve demonstrates the relationship between spawning stock size and survival for Atlantic menhaden Brevoortia tyrannus for the 1955–1970 year classes. Each point on the graph shows by year the estimated egg production and percent survival to age one. Similar population responses could occur, however, if some perturbation such as radiation was causing mortalities of eggs and larvae. This would ultimately reduce spawning stock size. Obviously, we would not expect the density-dependent relationship to compensate for radiation-induced mortalities in severely exploited fish stocks.

Evidence also exists that increased pressure on fish stocks by over-exploitation or other stresses may be compensated for by an increase in fecundity of surviving adults. Blaylock (1969), for example, reported that the mosquitofish Gambusia affinis increased fecundity relative to controls in the presence of chronic exposure to radioactivity in a freshwater environment.

Fig. 2

Fig. 2. An extreme example of a density-dependent relationship between spawning stock size and recruitment in highly fecund species (figure from Beverton and Holt 1957).

Fig. 3

Fig. 3. Relationship between spawning stock size and survival for Atlantic menhaden, Brevoortia tyrannus, for 1955–1970 year classes (W. Nelson, personal communication).

The inherent dynamics of highly fecund marine populations, therefore, could compensate for additional mortalities of young caused by contaminants such as radionuclides. This compensating mechanism is limited in its capacity to “protect” a species that experiences high mortalities in its early stages. The actual level of mortality that must occur to affect a population significantly will be highly variable and dependent on a number of additional factors such as predation, food supply, exploitation, etc.

2 Rad = absorbed dose of ionizing radiation from alpha, beta, and gamma rays.

3 R = roentgen, a unit of exposure dose for gamma rays or X-rays.

4 Ci = Curie, the basic unit measurement for radioactivity, equal to 3,7 × 1010 atomic disintegrations per second.

EXPOSURE OF MAN TO RADIOACTIVITY

Since it generally is accepted that any radiation dose above background levels to man will be accompanied by some risk of deleterious effects, releases of radioactivity to the aquatic environment must be kept to a minimum. To ensure that man's exposure to radiation will be minimal, restrictions have been placed upon release and disposal of radioactivity into the aquatic environment. These restrictions have been established through the efforts of the International Atomic Energy Agency and the International Commission on Radiation Protection.

In both the U.K. and the U.S., two different approaches have been taken to control routine releases of radioactivity into the aquatic environment. In the U.K. the critical pathway approach is used (Preston 1969), and in the U.S. the concept of maximum permissible concentration (MPC) is used (Code of Federal Regulations 1967). In the critical pathways approach, the pathways by which radionuclides can reach man are first identified and evaluated in relation to the level of radionuclides that can be transferred to man. The dose of radiation to man then is calculated for the “critical group” exposed. Radiation doses from any other pathway for that radionuclide cannot exceed those levels. An excellent example is consumption of ruthenium-106 contaminated seaweed Porphyra in laver-bread by a small coastal population near Wind-scale. This pathway sets the limits on ruthenium-106 discharges. Other examples of critical pathways to man in the U.K. are shown in Table 3. In the U.S. the concept of MPC is set forth in the “Code of Federal Regulations,” Title 10, Part 20. This limits, on a daily basis, the concentration of radionuclides in 2,2 liters of drinking water to that amount (contained in 2,2 liters of water per day), which will not exceed the maximum permissible radiation dose (ICRP 1959 and NCRP 1959).

Another approach is the use of specific activity for establishing permissible levels of environmental radioactivity. This was first recommended for radioactive waste disposal into the U.S. Pacific coastal waters (NAS-NRC 1962). The application of this approach is based mainly on two assumptions: (1) that a radionuclide introduced into the environment readily equilibrates with the non-radioactive element such that biological concentrating mechanisms will be unable to discriminate between different forms of the element, and (2) that the quantity of each stable element in each body organ is constant and does not fluctuate with intake of that element. Thus, the distribution of radionuclides will correspond to the distribution of the stable element.

In the calculations for radiation dose limits to man from incorporated radionuclides, the ICRP (1959) used the “standard man.” This model was based on a middle-European man with food habits of that geographical area. Tanaka et al. (1979) pointed out that such standards were not applicable to the Japanese because of differences in both body structure and culture. The primary difference is that the Japanese standard man has very different patterns of food consumption. Since “he” consumes many times more fishery products and seaweed than a European, Tanaka suggested that the standards should be designed to take into account the peoples and customs of the various cultures around the world. For example, a certain level of release of radioactivity may not affect a beef and potatoes food culture but could seriously impact a culture where fishery products are the primary protein source.

Table 3. Critical organisms and radionuclides in the pathways to man at four radioactive waste discharge locations in the U.K. (modified from Preston and Mitchell 1973).

SiteCritical organismsCritical radionuclide
WindscalePorphyra (seaweed)Ruthenium-106
 FishCesium-137
  Cesium-134
   
BradwellOysterSilver-110
  Zinc-65
   
DungenessFishCesium-137
  Cesium-134
   
Hinkley PointFish & shrimpCesium-137
  Cesium-134

CONCLUSIONS

The sudden and awesome beginning of the nuclear age during World War II and subsequent atmospheric nuclear testing gave rise to intensive international research on the environmental behavior of radionuclides and biological damage from radiation. Results of this research have been partially responsible for the establishment of extremely tight restrictions governing all uses and releases of radionuclides into the environment. As a result, peaceful uses of radioactivity have not interfered with commercial and recreational harvest of fisheries resources in either the freshwater or marine environment.

This has not been the case, however, with other pollutants. Excessive discharges of mercury and cadmium, for example, have resulted in adverse effects on man (Douglas-Wilson 1972; Kurland 1973). Discharges of these heavy metals have caused the closing of a number of rivers and lakes to recreational fishing in the U.S., Canada and Sweden. Even several years after environmental discharges of mercury had been sharply reduced or eliminated, certain species of recreational fish from several freshwater areas of the U.S., e.g., Holston River, Shenandoah River and Lake St. Clair, still cannot be eaten. In addition, excessive releases or uses of chlorinated hydrocarbons (e.g., DDT, PCB) have impacted the use of freshwater resources, such as salmon in Lake Michigan, and have been linked to the decline of two populations of birds—peregrine falcons and ospreys—along the east coast of the United States (Hickey 1969; Spitzer et al. 1978).

Since the aquatic environment, particularly the ocean, is the final repository for many of man's pollutants and also serves as one of the major sources of protein for the earth's growing population, extreme care must be exercised in the release and disposal of radioactivity. Constant vigil and continued research on the fate and effects of environmental releases of radioactivity must be continued to ensure that misuse of this potentially dangerous contaminant does not occur. In particular, research should be concentrated on obtaining a better understanding of the consequences of elevated levels of plutonium in the aquatic environment (National Academy of Sciences 1975) and developing techniques for damage assessment to fishery resources in the event of a catastrophic release of radioactivity from a nuclear reactor. At the present time, however, there are no known instances where radioactivity has had a deleterious effect on aquatic populations (Templeton et al. 1976; Cross 1978). With the growing number of nuclear power generating facilities and the accompanying wastes, such situations could occur if we do not maintain our vigilence.

ACKNOWLEDGEMENT

We express our appreciation to those members of the staff of the Beaufort Laboratory who provided critical review and editorial assistance in the preparation of this manuscript.

LITERATURE CITED

Baptist, J.P. and D.E. Hoss 1965 Accumulation and retention of radionuclides by marine fish, pages 14–19 in Annual report of the Bureau of Commercial Fisheries Radiobiological Laboratory, Beaufort, N.C., for the fiscal year ending June 30, 1963. U.S. Fish Wildl. Serv., Circ., 204.

Beverton, R.J.H. and S.J. Holt 1957 On the dynamics of exploited fish populations. Fish. Invest. Minist. Agric. Fish. Food (G.B.), Ser. II, 19. 533 p.

Blaylock, B.G. 1966 Chromosomal polymorphism in irradiated natural populations of Chironomus. Genetics, 53:131–136.

Blaylock, B.G. 1969 The fecundity of a Gambusia affinis affinis population exposed to chronic environmental radiation. Radiat. Res., 37:108–117.

Blaylock, B.G. and M.L. Frank. In press Effects of chronic low-level irradiation on Gambusia affinis. In N. Egami, ed. Symposium on effects of radiation on aquatic organisms. Tokyo, Japan Scientific Societies Press.

Blaylock, B.G. and J. R. Trabalka. 1978 Evaluating the effects of ionizing radiation on aquatic organisms, pages 103–152 in Advances in radiation biology, vol. 7. New York, Academic Press.

Blaylock, B.G. and J.P. Witherspoon. 1978 Evaluation of radionuclides released from the light water reactor nuclear fuel cycle to the aquatic environment, pages 851–865 in D.C. Adrino and I.L. Brisbine, eds. Environmental chemistry and cycling processes. Washington, D.C., U.S. Department of Energy.

Brown, V.M. and W.L. Templeton. 1964 Resistance of fish embryos to chronic irradiation. Nature (London), 203:1257–1259.

Chipman, W.A. 1972 Ionizing radiation, pages 1579–1657 in O. Kinne, ed. Marine ecology, vol. 1, pt. 3. New York, Wiley-Interscience.

Code of Federal Regulations, 1967 with supplements through 1970, Atomic Energy, Title 10. Standards for protection against radiation, pt. 20, p. 63–80b. Washington, D.C., U.S. Government Printing Office.

Cooley, J.L. 1973 Effects of chronic environmental radiation on a natural population of the aquatic snail Physa heterostropha. Radiat. Res., 54:130–140.

Cooley, J.L. and F.O. Miller. 1971 Effects of chronic irradiation on laboratory populations of the aquatic snail Physa heterostropha. Radiat. Res., 47:716–724.

Cooley, J.L. and D.J. Nelson. 1970 Effects of chronic irradiation and temperature on population of the aquatic snail Physa heterostropha. Oak Ridge National Laboratory, Oak Ridge, Tenn., ORNL-4612. 71p.

Cross, F.A. 1978 Impact of radioactivity on the marine environment, pages 63–72 in American-Soviet symposium on the biological effects of pollution on marine organisms, 1st, Gulf Breeze, Fla., 1976. Gulf Breeze, Fla., U.S. Environmental Protection Agency (EPA-600/9-78-007).

Cross, F.A., W.C. Renfro and E. Gilat. 1975 A review of methodology for studying the transfer of radionuclides in marine food chains, pages 185–210 in Design of radiotracer experiments in marine biological systems. Tech. Rep. Ser. Int. At. Energy Agency, 167.

Cross, F.A. and W.G. Sunda. 1978 Relationship between bioavailability of trace metals and geochemical processes in estuaries, pages 429–442 in M.L. Wiley, ed. Estuarine interactions. New York, Academic Press.

Donaldson, L.R. and K. Bonham. 1964 Effects of low-level chronic irradiation of chinook and coho salmon eggs and alevins. Trans. Am. Fish. Soc., 93:333–341.

Donaldson, L.R. and K. Bonham. 1970 Effects of chronic exposure of chinook salmon eggs and alevins to gamma irradiation. Trans. Am. Fish. Soc., 99:112–119.

Douglas-Wilson, I. 1972 Cadmium pollution and Itai-Itai disease. Lancet, 1972:382–383.

Duursma, E.K. and M.G. Gross. 1971 Marine sediments and radioactivity, pages 147–160 in Radioactivity in the marine environment. Washington, D.C., National Academy of Sciences.

Engel, D.W. 1967 Effect of single and continuous exposures of gamma radiation on the survival and growth of the blue crab, Callinectes sapidus. Radiat. Res., 32:685–691.

Foster, R.F., I.L. Ophel and A. Preston. 1971 Evaluation of human radiation exposure, pages 240–260 in Radioactivity in the marine environment. Washington, D.C., National Academy of Sciences.

Foster, R.F. and J.K. Soldat. 1966 Evaluation of exposure resulting from the disposal of radioactive wastes in the Columbia River, pages 683–696 in Disposal of radioactive wastes into seas, oceans and surface waters. Vienna, International Atomic Energy Agency.

Fowler, S.W. and S.L.F. Small. 1967 Moulting of Euphausia pacifica as a possible mechanism for vertical transport of zinc-65 in the sea. Int. J. Oceanol. Limnol., 1:237–245.

Hickey, J.J. 1969 Peregrine falcon populations, their biology and decline. Madison, University of Wisconsin Press. 618p.

International Commission on Radiological Protection. 1959 Report of Committe II on permissible dose of internal radiation. New York, Pergamon Press. ICRP Publ. 2. 233p.

Jeffries, D.F. and C.J. Hewett. 1971 The accumulation and excretion of radioactive caesium by the plaice (Pleuronectes platessa) and thornback ray (Raia clavata). J. Mar. Biol. Assoc. U.K., 51:411–422.

Joseph, A.B., P.F. Gustafson, I.R. Russell, E.A. Schuert, H.L. Volchok and A. Tamplin. 1971 Sources of radioactivity and their characteristics, pages 6–41 in Radioactivity in the marine environment. Washington, D.C., National Academy of Sciences.

Kaufman, G.A. and R.J. Beyers. 1973 Effects of chronic gamma irradiation on the fish. Oryzias latipes, pages 1119–1124 in D.J. Nelson, ed. Radionuclides in ecosystems, proceedings of the third national symposium on radioecology, Oak Ridge, Tenn., vol. 2. Oak Ridge, Tenn., U.S. Atomic Energy Commission (CONF-71050-P2).

Kuraland, L.T. 1973 The human health hazards of methylmercury, pages 283–297 in D.R. Buhler, ed. Mercury in the western environment. Corvallis, Ore., Continuing Education Publications.

Luoma, S.N. and E.A. Jenne. 1976 Factors affecting the availability of sediment-bound cadmium to the estuarine, deposit-feeding clam, Macoma balthica, pages 283–290 in C.E. Cushing, Jr., ed. Radioecology and energy resources. Stroudsburg, Pa., Dowden, Hutchinson & Ross.

Nakatani, R.E. 1966 Biological response of rainbow trout (Salmo gairdneri) ingesting zinc-65, pages 809–823 in Disposal of radioactive wastes into seas, oceans and surface waters. Vienna, International Atomic Energy Agency.

National Academy of Sciences. 1975 Assessing potential ocean pollutants. Washington, D.C., National Academy of Sciences. 438p.

National Academy of Sciences-National Research Council. 1962 Disposal of radioactive waste into Pacific coastal waters. NAS-NRC, Publ. 985. 87p.

National Committee on Radiation Protection. 1959 Maximum permissible body burdens and maximum permissible concentrations of radionuclides in air and in water for occupational exposure. U.S. Natl. Bur. Stand. Handb., 69. 95p.

Ophel, I.L., M. Hoppenheit, R. Ichikawa, A.G. Klimov, S. Kobayashi, Y. Nishiwaki and M. Saiki. 1976 Effects of ionizing radiation on aquatic organisms, pages 57–87 in Effects of ionizing radiation on aquatic organisms and ecosystems. Tech. Rep. Ser. Int. At. Energy Agency, 172.

Osterberg, C.L. 1975 Radiological impacts of releases from nuclear facilities into the aquatic environments, pages 25–35 in Impacts of nuclear releases into the aquatic environment. Vienna, International Atomic Energy Agency.

Osterberg, C., A.G. Carey, Jr. and H. Curl, Jr. 1963 Acceleration of sinking rates of radionuclides in the ocean. Nature (London), 200:1276–1277.

Osterberg, C., W.G. Pearcy and H. Curl, Jr. 1964 Radioactivity and its relationship to ocean food chains. J. Mar. Res., 22:2–12.

Pentreath, R.J. 1973a The accumulation and retention of 59Fe and 58Co by the plaice, Pleuronectes platessa L. J. Exp. Mar. Biol. Ecol., 12:315–326.

Pentreath, R.J. 1973b The accumulation and retention of 65Zn and 54Mn by the plaice, Pleuronectes platessa L. J. Exp. Mar. Biol. Ecol., 12:1–18.

Pentreath, R.J. 1973c The accumulation from sea water of 65Zn, 54Mn, 58Co and 59Fe by the thornback ray, Raja clavata L. J. Exp. Mar. Biol. Ecol., 12:327–334.

Pentreath, R.J. 1973d The roles of food and water in the accumulation of radionuclides by marine teleost and elasmobranch fish, pages 421–436 in Radioactive contamination of the marine environment. Vienna, International Atomic Energy Agency.

Polikarpov, G.G. 1966 Radioecology of aquatic organisms. New York, Reinhold. 314p.

Preston, A. 1969 Aquatic monitoring programmes, pages 309–324 in Environmental contamination by radioactive materials. Vienna, International Atomic Energy Agency.

Preston, A. 1972 Artificial radioactivity in freshwater and estuarine systems. Proc. R. Soc. Lond. B. Biol. Sci., 180:421–436.

Preston, A. and N.T. Mitchell. 1973 Evaluation of public radiation exposure from the controlled marine disposal of radioactive waste (with special reference to the United Kingdom), pages 575–593 in Radioactive contamination of the marine environment. Vienna, International Atomic Energy Agency.

Rice, T.R. and J.W. Angelovic. 1969 Radioactivity in the sea: effects on fisheries, pages 574–578 in F.E. Firth, ed. Encyclopedia of marine resources. New York, Van Nostrand Reinhold Co.

Rice, T.R. and J.P. Baptist. 1974 Ecologic effects of radioactive emissions from nuclear power plants, pages 373–439 in L.A. Sagan, ed. Human and ecologic effects of nuclear power plants. Springfield, Ill., Charles C. Thomas.

Rice, T.R., J.P. Baptist and T.J. Price. 1965 Accumulation of mixed fission products by marine organisms, pages 263–286 in E.A. Pearson, ed. Advances in water pollution research, vol. 3. New York, Pergamon Press.

Rice, T.R. and D.A. Wolfe. 1971 Radioactivity—chemical and biological aspects, pages 325–379 in D.W. Hood, ed. Impingement of man on the oceans. New York, Wiley.

Saiki, M., K. Shirai, S. Ohno and T. Mori. 1957 Studies on the radioelements in the contaminating radioactive fish. II. On skipjacks caught at the Pacific Ocean in 1956 (Part I). Bull. Jpn. Soc. Sci. Fish., 22:645–650.

Spitzer, P.R., R.W. Risebrough, W. Walker, R. Hernandez, A. Poole, D. Puleston and I.C.T. Nisbet. 1978 Productivity of ospreys in Connecticut-Long Island increases as DDE residues decline. Science, 202:333–335.

Tanaka, G., H. Kawamura, E. Nomura, H. Yamaguchi and Y. Nakahara. 1979 Studies of reference Japanese man, page 285 in Abstracts of papers, 6th International Congress of Radiation Research, Tokyo, 1979. Tokyo, Japanese Association for Radiation Research.

Templeton, W.L., M. Bernhard, B.G. Blaylock, C. Fisher, M.J. Holden, A.G. Klimov, P. Metalli, R. Mukherjee, O. Ravera, L. Sztanyik and F. Vab Hoeck. 1976 Effects of ionizing radiation on aquatic populations and ecosystems, pages 89–100 in Effects of ionizing radiation on aquatic populations and ecosystems. Tech. Rep. Ser. Int. At. Energy Agency, 172.

Templeton, W.L., R.E. Nakatani and E.E. Held. 1971 Radiation effects, pages 223–239 in Radioactivity in the marine environment. Washington, D.C., National Academy of Sciences.

Trabalka, J.R. and C.P. Allen. 1977 Aspects of fitness of a mosquitofish Gambusia affinis population exposed to chronic low-level environmental radiation. Radiat. Res., 70:198–211.

Wolfe, D.A., F.A. Cross and C.D. Jennings. 1973 The flux of Mn, Fe, and Zn in an estuarine ecosystem, pages 159–175 in Radioactive contamination of the marine environment. Vienna, International Atomic Energy Agency.

Wolfe, D.A. and T.R. Rice. 1972 Cycling of elements in estuaries. U.S. Natl. Mar. Fish. Serv. Fish. Bull., 70:959–972.

Woodhead, D.S. 1970 The assessment of radiation dose to developing fish embryos due to the accumulation of radioactivity by the egg. Radiat. Res., 43:582–597.

Woodhead, D.S. 1973 Levels of radioactivity in the marine environment and the dose commitment to marine organisms, pages 499–525 in Radioactive contamination of the marine environment. Vienna, International Atomic Energy Agency.

Woodhead, D.S., M. Bezzegh-Galantai, S.W. Fowler, A. Frantz, M. Ijuin, J.P. Oliver, J. Sas-Hubicki and E. Wanderer. 1976 Concentrations of radionuclides in aquatic environments and the resultant radiation dose rates received by aquatic organisms, pages 5–54 in Effects of ionizing radiation on aquatic organisms and ecosystems. Tech. Rep. Ser. Int. At. Energy Agency, 172.


Previous Page Top of Page Next Page