Nordic Radioecology: The Transfer of Radionuclides Through Nordic Ecosystems to Man

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NORDIC RADIOECOLOGY THE TRANSFER OF RADIONUCLIDES THROUGH NORDIC ECOSYSTEMS TO MAN

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Studies in Environmental Science 62

NORDIC RADIOECOLOGY THE TRANSFER OF RADIONUCLIDES THROUGH NORDIC ECOSYSTEMS TO MAN Edited by

H. Dahlgaard R i s National ~ Laboratory Roskilde, Denmark

ELSEVIER Amsterdam

- Lausanne - New York - Oxford - Shannon - Tokyo 1994

ELSEVIER SCIENCE B.V Sara Burgerhartstraat 25 P.O. B o x 21 1,1000 AE Amsterdam,The Netherlands

ISBN1 0-444-8 16 17-8

0 1994 Elsevier Science B.V. All rights reserved.

No part of this publication may be reproduced, stored in a retrieval system or transmitted in any form or by any means, electronic, mechanical, photocopying, recording or otherwise, withoutthe prior written permission of the publisher, Elsevier Science B.V., Copyright & Permissions Department, P.O. B o x 521,1000 AM Amsterdam, The Netherlands. Special regulations for readers in the USA - This publication has been registered with the Copyright Clearance Center Inc. (CCC), Salem, Massachusetts. Information can be obtained from theCCCaboutconditionsunderwhichphotocopiesofpartsofthispublication may bemadeinthe USA. All other copyright questions, including photocopying outside of the USA, should be referred to the copyright owner, Elsevier Science B.V., unless otherwise specified. No responsibility is assumed by the publisher for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions or ideas contained in the material herein. This book is printed on acid-free paper. Printed in The Netherlands

V

PREFACE

The present book is the final milestone in the radioecology programme, RAD, carried out from 1990 to 1993 under the Nordic Committee for Nuclear Safety Research, NKS. This work was done in parallel to three other NKS programmes: Reactor safety (SIK), Waste and decommissioning

(KAN), and Emergency preparedness (BER). The NKS was established in 1966 and was financed by the Nordic Council of Ministers from 1977 to 1989. It is now a joint Nordic committee financed by the Danish Emergency Management Agency, the Finnish Ministry of Trade and Industry, Iceland's National Institute of Radiation Protection, the Norwegian Radiation Protection Authority, and the Swedish Nuclear Power Inspectorate. The NKS is further co-sponsored by a number of Finnish and Swedish companies working in the field of civil nuclear energy and protection of the population. The preparation of this book involved much painstaking effort by the authors, the participants in the working groups and the four project leaders, Manuela Notter, Per Strand, Aino Rantavaara and Elis Holm. I would like here to express my gratitude for their contribution. The guidance and inspiration given by the RAD reference group is furthermore acknowledged. Finally, it should be mentioned that there would have been no Nordic collaboration on Nuclear Safety without the energetic, persistent, diplomatic and occasionally maddening efforts of our travelling "ambassador", Franz Marcus, executive secretary of the NKS from 1976 to 1994.

Henning Dahlgaard Co-ordinator of the RAD programme

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vii

CONTENTS

PREFACE CONTRIBUTORS AND PARTICIPANTS

Chapter 1

V XI

NORDIC RADIOECOLOGY 1990 -1993

1

1.1

The aims and justification of Nordic radioecology. H. Dahlgaard

3

1.2

General summary and conclusions. H. Dahlgaard, M. Notter, J. Brittain, P. Strand, A. Rantavaara and E. Holm

Chapter 2 2.1

AQUATIC ECOSYSTEMS

23

The characterization of radiocaesium transport and retention in Nordic lakes. H.E. Bjernstad, J.E. Brittain, R. SaxBn and B. Sundblad

2.3

21

Introduction to aquatic ecosystems. M. Notter, J. Brittain and

U. Bergstrom 2.2

7

29

The distribution and characterization of 137Csin lake sediments. A. Broberg

45

...

Vlll

2.4

Transport of 137Csin large Finnish drainage basins. R. Saxtn

2.5

The role of lake-specific abiotic and biotic factors for the transfer of radiocaesium fallout to fish. T. Anderson and M. Meili

2.6

105

Polonium-210 and radiocaesium in muscle tissue of fish from different Nordic marine areas. E. Holm

2.9

93

Radiocaesium in algae from Nordic coastal waters. L. Carlson and P. Snoeijs

2.8

79

Models for predicting radiocaesium levels in lake water and fish.

U. Bergstrom, B. Sundblad and S. Nordlinder 2.7

63

119

Radiocaesium as ecological tracer in aquatic systems - a review. M. Meili

127

AGRICULTURAL ECOSYSTEMS

141

3.1

Introduction to radioecology of the agricultural ecosystem. P. Strand

143

3.2

Direct contamination - seasonality. A. Aarkrog

149

3.3

Influence of physico-chemical forms on transfer.

Chapter 3

D.H. Oughton and B. Salbu

1 65

3.4

Contamination of annual crops. M. Strandberg

185

3.5

Transfer of 137Csto cows’ milk in the Nordic countries. H.S. Hansen and LAndersson

3.6

197

Radiocaesium transfer to grazing sheep in Nordic environments. K. Hove, H. Lijnsjo et al.

211

ix 3.7

Dynamic model for the transfer of 137Csthrough the soil-grass-lamb foodchain. S.P. Nielsen

3.8

229

Studies on countermeasures after radioactive depositions in Nordic agriculture. K. RosCn

239

FOREST AND ALPINE ECOSYSTEMS

26 1

4.1

Introduction to terrestrial seminatural ecosystems. A. Rantavaara

263

4.2

The transfer of radiocaesium from soil to plants and fungi

Chapter 4

in seminatural ecosystems. R.A. Olsen

265

4.3

Radiocaesium in game animals in the Nordic countries. K.J. Johanson

287

4.4

Pathways of fallout radiocaesium via reindeer to man. E. Gaare and H. Staaland

4.5

Chapter 5 5.1

5.2

303

The distribution of radioactive caesium in boreal forest ecosystems. R. Bergman

335

METHODOLOGY, QUALITY ASSURANCE AND DOSES

381

Introduction to intercalibration / analytical quality control and doses. E. Holm

383

Intercomparison of large stationary air samplers. I. Vintersved

3 85

X

5.3

Intercalibration of whole-body counting systems. T. Rahola, R. Falk and M. Tillander

407

5.4

Intercalibration of gamma-spectrometric equipment. E. Holm

425

5.5

Doses from the Chernobyl accident to the Nordic populations via diet intake. A. Aarkrog

5.6

433

Internal radiation doses to the Nordic population based on whole-body counting. M. Suomela and T. Rahola

457

DEFINITIONS, TERMS AND UNITS

473

INDEX

477

SPECIES INDEX

481

xi CONTRIBUTORS AND PARTICIPANTS

Hannele Aaltonen, STUK, P.O.Box 14, FIN 00881 Helsinki Asker Aarkrog, ECO-Riss, Postboks 49, DK 4000 Roskilde Magne Alpsten, Institut for Radiofysik, Sahlgrenska Sjukhuset, S 41345 Goteborg Inger Andersson, Lantbruksuniversitetet, Box 59, S 23053 Alnarp Tord Andersson, Naturgeografisk avd., Umel Universitet, S 90187 Umel Ronny Bergmann, FOA-4, S 90182 Umel Ulla Bergstrom, Studsvik Eco & Safety, S 61182 Nykoping Torolf Bertelsen, Statens Strllevern, Postboks 55, N 1345 0sterh Helge E. Bjernstad, Agricultural University of Norway, N 1432 AS-NLH Inggard Blakar, Agricultural University of Norway, N 1432 AS-NLH John Brittain, Oslo Universitet, Sars Gate 1, N 0562 Oslo Anders Broberg, Uppsala Universitet, Box 557, S 75122 Uppsala Lena Carbon, Avd. for Marinekologi, Box 124, S 22100 Lund Gordon Christensen, IFE, Postboks 40, N 2007 Kjeller Olof Eriksson, Lantbruksuniversitetet, Box 703 1, S 75007 Uppsala Ake Eriksson, Lantbruksuniversitetet, Box 7031, S 75007 Uppsala Sverker Evans, Statens Naturvbdsverk, Box 1302, S 17125 Solna Rolf Falk, Swedish Radiation Protection Institute, Box 60204, S 10401 Stockholm Torbjorn Forseth, Institut for Naturforskning, Tungasletta 2, N 7004 Trondheim Lars Foyen, Havforskningsinstituttet,Box 1870, N 5024 Bergen

Torstein Garmo, Agricultural University of Norway, N 1432 AS-NLH Eldar Gaare, Norwegian Institute for Nature Research, Tungasletta 2, N 7005 Trondheim Eva Hllkansson, Institut for Radiofysik, Sahlgrenska Sjukhuset, S 41345 Goteborg

Lars EUkansson, Uppsala Universitet, Viistra Agatan 24, S 75220 Uppsala Hanne S. Hansen, Agricultural University of Norway, N 1432 AS-NLH

Lars Egil Haugen, Agricultural University of Norway, N 1432 AS-NLH Knut Hove, Agricultural University of Norway, N 1432 AS-NLH

Erkki nus, STUK, P.O.Box 14, FIN 00881 Helsinki Kki Indridason, Geislavarnir rikisins, Laugavegur 118d, Is 150 Reykjavik

xii Tim0 Jaakkola, Radiokemiska institutionen, Pb 5, FIN 00014 Helsingfors Universitet Hans Pauli Joensen, Academia Faroensis, Noatun, FR 100 Torshavn Karl J. Johanson, Lantbruksuniversitetet, Box 7031, S 75007 Uppsala Bernt Jones, Lantbruksuniversitetet, Box 7038, S 75007 Uppsala Pekka Kansanen, Helsingin kaupungin ymp., Helsinginkatv. 24, FIN 00530 Helsinki Riitta Korhonen, VlT/YDI, Pb 208, FIN 02151 Espoo Vappu Kossila, Lantbrukets forskningscentral, FIN 31600 Jokioinen Andrew Liken, Agricultural University of Norway, N 1432 AS-NLH Hans Liinsjo, Lantbruksuniversitetet, Box 7031, S 75007 Uppsala Sigurdur Magnusson, Geislavarnir rikisins, Laugavegur 118d, Is 150 Reykjavik Soren Mattsson, Inst. for Radiofysik, Malmo Almanna Sjukhus, S 21401Malmo Marcus Meili, Uppsala Universitet, Box 557, S 75122 Uppsala Georg NeumaM, Lantbruksuniversitetet, Box 7031, S 75007 Uppsala Sven P. Nielsen, ECO-Riss, Postboks 49, DK 4000 Roskilde Sture Nordlinder, Studsvik Eco & Safety, S 61 182 Nykoping Tuire Nygren, Vilt- och Fiskeriforskningsinstitutet, Tutkimuslaitos, FIN 82950 Kuikkalampi Elisabet D. Olafsdijttir, Geislavarnir rikisins, Laugavegur 118d, Is 150 Reykjavik Rolf A. Olsen, Agricultural University of Norway, N 1432 AS-NLH Deborah H. Oughton, Agricultural University of Norway, N 1432 AS-NLH Olli Paakkola, Torpantie 1 B, FIN 01650 Vanda Arja Paasikallio, Lantbrukets forskningscentral, FIN 3 1600 Jokioinen Sigurdur E. Piilsson, Geislavarnir rikisins, Laugavegur 118d, Is 150 Reykjavik Tua Rahola, STUK, P.O.Box 14, FIN 00881 Helsinki Hannu Raitio, Skogforskningsinstitutet,FIN 39700 Parkano Kristina Rissanen, STUK, Louhikkotie 28, FIN 96500 Rovaniemi Klas Rosbn, Lantbruksuniversitetet, Box 7031, S 75007 Uppsala Brit Salbu, Agricultural University of Norway, N 1432 AS-NLH Chr. Samuekson, Institutionen f. Radiofysik, Lasarettet, S 22185 Lund Ritva Saxbn, STUK, P.O.Box 14, FIN 00881 Helsinki Tone Selnaes, IFE, Postboks 40, N 2007 Kjeller Pauli Snoeijs, Uppsala Universitet, Box 559, S 75122 Uppsala Riitta Sormunen-Christian, Lantbrukets forskningscentral, FIN 3 1600 Jokioinen Hans Staaland, Agricultural University of Norway, N 1432 AS-NLH Eiliv Steinnes, Universitetet, AVH, N 7055 Dragsvoll Morten Strandberg, ECO-Riss, Postboks 49, DK 4000 Roskilde

xiii Bjorn Sundblad, Studsvik Eco & Safety, S 61182 Nykoping Matti Suomela, STUK, P.O.Box 14, FIN 00881 Helsinki J6hann Thorsson, Agricultural Research Institute, Is 112 Reykjavik Michael Tillander, Helsinki Universitet, Radiokemiska inst., FIN 00014 Helsinki Ole Ugedal, Finmark Distrikth0yskole, Follumsvei, N 9500 Alta Finn Ugletveit, Statens Strilevern, Postboks 55, N 1345 0sterh Trygvi Vestergaard, Academia Faeroensis, Noatun, FR 100 Torshavn Ingemar Vintersved, Forsvarets Forskningsanstalt, S 17290 Sundbyberg

PROJECT LEADERS Elis Holm, Institutionen f. Radiofysik, Lasarettet, S 22185 Lund Manuela Notter, Statens NaturvArdsverk, Box 1302, S 17125 Solna Per Strand, Statens Strhlevern, Postboks 55, N 1345 0steris Aino Rantavaara, STUK, P.O.Box 14, FIN 00881 Helsinki

REFERENCE GROUP Asker Aarkrog, Rise National Laboratory, Postboks 49, DK 4000 Roskilde Henning Dahlgaard, Riss National Laboratory, Postboks 49, DK 4000 Roskilde (Co-ordinator) Sigurdur Magnusson, Geislavarnir rikisins, Laugavegur 118d. Is 150 Reykjavik Franz Marcus, NKS, Postboks 49, DK 4000 Roskilde Judith Melin, SSI, Box 60204, S 10401 Stockholm Eiiiv Steinnes, Universitetet, AVH, N 7055 Dragsvoll Matti Suomela, STUK, P.O.Box 14, FIN 00881 Helsinki Seppo Vuori, VTT/YDI, Pb 208, FIN 02151 Espoo Erik-Anders Westerlund, Statens StrAlevern, Postboks 55, N 1345 0sterh (Chairman)

CO-ORDINATOR Henning Dahlgaard, Rise National Laboratory, Postboks 49, DK 4000 Roskilde

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Chapter 1 NORDIC RADIOECOLOGY 1990 - 1993

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3 1.1. THE AIMS AND JUSTIFICATION OF NORDIC RADIOECOLOGY

HEN"G DAHLGAARD Risar National Laboratory, DK-4000 Roskilde, Denmark.

SUMMARY A description is given of the goals and background of the RAD programme described in this book. The overall scientific aim of the Nordic Radioecology programme was to perform a quantitative, comparative study of the pathways of Chernobyl-derived radiocaesium, in particular, through different Nordic ecosystems. Furthermore, the programme was to help a new generation of radioecologists become acquainted with different Nordic ecosystems and to foster Nordic contacts. The relevance of a radioecology programme for nuclear accident preparedness is furthermore stressed.

BACKGROUND The word RADIOECOLOGY came into being in the 1950's when it became evident that man-made radionuclides produced in atmospheric nuclear weapons tests had been spread globally and were transferred through various ecosystems to man. From the very beginning the scientific study of radioecology was developed by scientists with an interest in ecology and genetics. However, physicists, analytical chemists and engineers played an essential role because accurate measurements of the low levels of the relevant radionuclides - e.g. %k,'37Csand 239Pu- found

in the environment, required the elaborate analytical procedures and advanced electronic equipment that were gradually developed during the 1960's - the "Golden Age" of radioecology. At most institutions radioecology became a branch of health physics ultimately aiming at studying and reducing the radiation dose to man. Attempts were made at several institutions to incorporate the field in general ecology and to utilize the radionuclides as global-scale tracers for, e.g., studies of atmospheric pollutant transport and trace element turnover. However interest in radioecology dwindled with the declining activity from atmospheric fallout, and by the mid-1980's work in radioecology had been reduced to a minimum, or was even non-existent in several countries. Furthermore the integrity of radioecologists and health physicists had been challenged by "environmentalist" groups fighting the peaceful utilization of nuclear energy on a non-scientific basis. Several institutions thus reduced funding to radioecology to serve political ends. When the accident at the Chernobyl nuclear power station happened in April 1986

4

radioecology was reinvented throughout Europe and surviving centres of study were given an economic boost. At several places ecologists of different backgrounds introduced new and fruitful concepts, using the Chernobyl radiocaesium for more than just radiation protection studies. The Nordic countries, Denmark, Finland, Iceland, Norway and Sweden, have a long, historic tradition of cultural and scientific collaboration. This has also applied to radioecology, where the Nordic Committee for Nuclear Safety Research (NKS), financed by the Nordic Council of Ministers, included this subject in their programmes from 1977 to 1985. At the beginning of 1986

- a few months before the Chernobyl accident - general radioecology was removed from this collaboration', and from 1990 the NKS financing was transferred from the Nordic Council of Ministers to the national authorities responsible for nuclear safety and radiation protection in the different countries. The Nordic radioecology programme RAD, which is the subject of the present book, was run under the auspices of the new NKS from 1990 to 1993. Via the NKS, the RAD programme has had funding of around 6 million Danish kroner (- 1 million US $). As the contents of the present book will show, this is only a minor part of the total costs of the work described here. However, without the catalytic support provided by the NKS much of the present work would not have taken place, and efforts in different Nordic countries would not have been coordinated. Plans for the Nordic Radioecology programme 1990-1993were described in the Scandinavian languages in a publication issued by the Nordic Council of Ministers (NKS, 1989).

THE NORDIC RADIOECOLOGY PROGRAMME The RAD programme consists of four projects. As the largest doses to man immediately after the Chernobyl accident were derived from the consumption of terrestrial products and freshwater fish, the programme included 2 projects on terrestrial radioecology: RAD-3, Agricultural ecosystems (project leader: Per Strand) and RAD-4, Forest and alpine ecosystems (project leader: Aino Rantavaara), and one on aquatic radioecology: RAD-2, Aquatic ecosystems (project leader: Manuela Notter) that mainly dealt with Nordic lakes. Finally, RAD-1 included training, methodology, quality assurance and doses to the Nordic population (project leader: Elis Holm). Results from the four projects are presented in detail in chapters 2-5, and are summed up in the following chapter 1.2.

~

I: The AKTU program 1985 - 1989 did, however, include environmental radioactivity after the Chernobyl accident (Tveten, editor).

5

AIMS AND JUSTIFICATION After the Chernobyl accident it became clear that the transfer of radionuclides via food to man

could result in significant internal radiation doses to the Nordic population after nuclear accidents. In the long term the most significant internal doses from Chernobyl were expected to be related to the contamination of specially sensitive Nordic environments leading to a high transfer of radiocaesium to man. It was considered important for the authorities to have access to up-to-date knowledge of the spreading and turnover of radionuclides in different Nordic ecological systems in order to be able to decide on the relevant countermeasures. Furthermore, knowledge of the

contamination levels of agricultural products was necessary to assure exports and avoid unnecessary loss of resources. There is an immense variation within the Nordic countries not only in the distribution of the Chernobyl deposition, but also in the transfer of radiocaesium to man. The contamination of a highly productive agricultural area is expected to give relatively small individual doses to a large population during a short period, whereas the contaminationof the lichen carpets utilized as wintergrazing for reindeer, or of the abundant oligotrophic lakes, will give a larger individual dose to a small population for many years. The overall scientific aim of the Nordic Radioecology programme was to perform a quantitative comparative study of the pathways of selected radionuclides through different Nordic ecosystems. Moreover the programme aims at helping a new generation of radioecologists to become acquainted with different Nordic ecosystems and to foster Nordic contacts. The RAD programme has aimed at obtaining the widest possible coverage, i.e. the inclusion of as many Nordic radioecological centres as possible. This is not cost-effective with respect to research results, but it does promote Nordic radioecological contacts. As a consequence, the programme is to a large extent based on nationally-funded programmes. A general goal for the entire programme

- and a justification for the funding of the

programme by the nuclear safety authorities - is its benefits in respect of preparedness for nuclear accidents. On first thoughts this goal may seem remote from a scientific field programme on the cycling of caesium in the environment. However, one benefit of keeping radioecological centres alive is that the necessary measuring equipment is ready for use, and that competent staff are available to take suitable samples and carry out reliable radionuclide analyses the very day an accident happens. In addition, knowledge of the pathways of radionuclides through ecosystems to man will be available. A nuclear preparedness plan without working scientific projects is like an airforce without trained fighter pilots. Maybe the most important justification of such programmes is not the production of final reports, but rather the less definable benefits such as inspiration and collaboration based on the

6

close personal relations among individual scientists from different Nordic countries and institutions having common interests. A further aspect of the personal contact between Nordic radioecologists and radiation protection officials is that it will facilitate information exchange between the different countries in any future nuclear emergency.

REFERENCES

NKS (1989). Plan for Nordisk Kjernesikkerhetsprogram 1990-1993. Nordisk Md, Nordisk Ministerriid, NU 19895 (in the Scandinavian languages). Tveten, U. (editor). Environmental consequences of releases from nuclear accidents. Final report of the NKA project AKTU-200. IFE, P.O.Box 40, N - 2007 Kjeller, 1990. 261 pp.

7

1.2. GENERAL SUMMARY AND CONCLUSIONS

HENNING DAHLGAARD', MANUELA NOTTER', JOHN E. BRITTAIN3,PER STRAND4, AINO RANTAVAARA' AND ELIS HOLM6 'Riss National Laboratory, DK - 4000 Roskilde, Denmark. 2Swedish Environmental Protection Agency, S - 171 85 Solna, Sweden. 3FreshwaterEcology and Inland Fisheries Laboratory (LFI), University of Oslo, Sars gate 1, 0562 Oslo, Norway. 4Norwegian Radiation Protection Authority, P.O.Box 55, N - 1340 0sterA.9, Norway. 5Finnish Centre for Radiation and Nuclear Safety, P.O.Box 14, FIN - 00881 Helsinki, Finland. 6Departmentof Radiation Physics, Lund University, Sweden.

INTRODUCTION On Monday, 28th April, 1986, most Nordic radioecologists and health physicists realized the area was being contaminated by debris from a serious nuclear accident. The cloud from Chernobyl had already reached the Nordic countries on Sunday, 27th April, and contamination was to continue during May. Figure 1.2.1 shows the resulting ground deposition of 137Csin kBq m-2 in the Nordic countries Denmark, Finland, Norway and Sweden. Off the map, the Chernobyl contamination on Iceland and Greenland was very low, whereas the deposition on the Faroe Islands was 0.6-4.5 kBq 1 3 7 m-2 ~ ~

The Nordic post-Chernobyl radioecology programme, RAD, consisted of four projects. The main radionuclides chosen for study were the two radiocaesium nuclides, 137Csand 134Cs,because they appeared to be the most important contributors to doses to man after the Chernobyl accident, and because they are relatively simple to measure. However, a few results for %rand 210Powere also reported. The present chapter is intended to give an overview of the results from the RAD programme. RAD-1 (project leader Elis Holm) had a multiple purpose: methodology, training, quality assurance and doses. Initially, a major task was to conduct a two-week post-graduate training course in various aspects of radioecology. The course included 20 lectures by various Nordic radioecologists. These are published elsewhere (Holm, editor). An exchange programme permitting, preferentially, young scientists to stay for one or two weeks at another Nordic laboratory, e.g. to adopt a new radiochemical method, was also conducted by RAD-1. Three

8

separate programmes on quality assurance were carried out. Of these, the intercomparison of nine large, stationary air samplers and the intercalibration of 20 Nordic whole-body counting systems are especially remarkable. Finally, RAD-1 was responsible for dose assessments based partly on the results produced in the three other RAD projects. The results from RAD-1 are given in chapter 5 and in Holm (editor).

RAD-2: Aquatic ecosystems (project leader: Manuela Notter) mainly concerned Nordic lakes, as the major problems in aquatic environments after the Chernobyl accident appeared in freshwater systems. However, two minor projects were run in the marine environment. The results from RAD-2 are described in detail in chapter 2. RAD-3: Agricultural ecosystems (project leader: Per Strand) focused on various aspects of Nordic agriculture in relation to nuclear contamination: annual crops, cows’ milk, grazing sheep and on countermeasures. RAD-3 also included a study of physico-chemical forms and a model study. The results are given in chapter 3. Finally RAD-4: Forest and alpine ecosystems (project leader: Aino Rantavaara) concerned the natural terrestrial environment which, like the freshwater environment, appeared to surprise the authorities with high and variable radionuclide levels after the Chernobyl accident. RAD-4 studied radiocaesium transfer from soil to plants and fungi, game animals, the reindeer foodchain and boreal forests in general. The results are reported in chapter 4. AQUATIC ECOSYSTEMS With respect to Nordic aquatic ecosystems, the main exposure pathway of 137Csto man after the Chernobyl accident has been through the consumption of freshwater fish. Caesium accumulates in fish muscle due to its chemical similarity to potassium and the accumulation of 137Csis of particular importance in the Nordic countries where ionic concentrations in freshwaters are generally low. Chapter 2 identifies the

important parameters determining radionuclide

concentrations in fish, thereby permitting the development and assessment of potential remedial measures. Since the Chernobyl accident in 1986, there has been an intensive research effort in the Nordic countries aimed at obtaining reliable input data for prediction models and determining the important driving forces and parameters for such models. Lakes received radionuclides from Chernobyl fallout via two sources: direct fallout on the lake surface and leakage from the catchment. Chapter 2.2 describes fractionation techniques used in a study of the input of radiocaesium to three widely different Nordic lakes, Hillesjon in Sweden,

!&re Heimdalsvatn in Norway and Saarisjawi in Finland. Using hydrological data, the degree of retention of 137Csin these three lake systems was estimated. Transport of 137Csin plant material (Coarse Particulate Organic Material, CPOM) is considerable in Nordic lakes. Through its rapid

Figure 1.2.1. Ground deposition of 137Cs,kBq m-*,in Denmark, Finland, Norway and Sweden resulting from the Chernobyl accident.

10

assimilation into the invertebrate foodchain, it is potentially a major source of 137Csfor lake ecosystems. CPOM transport is higher in mountain and forest lakes than in lowland lakes in agricultural areas. However, in all lakes almost all such plant material is retained in the lake. The Nordic lakes studied differed in the concentration of 137Csin the various molecular weight fractions

in the water phase. Free ions may easily cross biological membranes and the low molecular weight fraction is assumed to have a high degree of bioavailability. However, both organic and inorganic substances in the water phase may affect the biological uptake of a given element. In fact, the low molecular weight fraction showed no retention in the three study lakes and was exported downstream. In contrast, half the colloidal (pseudocolloidal) fraction was retained during passage through both &re Heimdalsvatn and Saarisjarvi. In Hillesjon, ten times more 137Csflowed out sediments. than flowed in, due to resuspension of 137Cs-ri~h Although some of the radiocaesium from Chernobyl has been transported out of lakes because of the high flows associated with the spring snowmelt at the time of deposition, most of it still remains in lake sediments. Chapter 2.3 describes a study of the distribution, physicochemical forms and concentration of radiocaesium in lake sediments. In 1987, 137Cswas to a large extent bound to chemically labile fractions, but it has subsequently been transformed to less available fractions, thus reducing the tendency for resuspension. The horizontal distribution of 137Csin the sediments is affected by the shape of the lake basin, steep-sloping bottoms tending to focus the radiocaesium towards the deeper parts. The degree of bioturbation, diffusion and the rate of sedimentation determine the vertical distribution of 137Csin lake sediments. A strong tendency for resuspension was found in shallow lakes. Although this may transport 137Cs to deeper areas where it is less available, it also increases its availability to the biota, delaying recovery in shallow lakes. The importance of leakage from catchment areas has been studied on a large scale in Finland, where the whole country has been divided into seven different catchments, each with its own characteristics with regard to fallout, soil type and topography (chapter 2.4). However, during the first year after the fallout the activity concentrations in lake waters and fish could be estimated using simple relationships to the deposition. In subsequent years catchment characteristics played an increasing role, leading to differences between lakes in the different catchment areas. For example, a high incidence of bogs prolonged the decrease of 137Cs in lake waters and in fish, whereas a predominance of clay soils reduced the transfer to aquatic systems. A number of lake-specific factors, both abiotic and biotic, have been put forward as

determining the concentration of radiocaesium in fish. Chapter 2.5 describes a major study encompassing a large number of Swedish lakes, and assesses the importance of a wide range of such factors. The maximum activity concentration in fish was reached within three years in most

lakes and normally in the order small perch - trout and charr - larger perch - pike, a sequence reflecting their trophic level. However, the transfer to fish varied by up to an order of magnitude between lakes. Variation in the expected transfer to pike can be explained by differences in the theoretical residence time of 137Cs,determined from the mean hydraulic residence time and the scavenging capacity of the lakes, which in turn is well indicated by the concentration of base cations in lake waters. The model assessment in chapter 2.6 is based on three Nordic lakes for which extensive data are available, both in terms of the radiocaesium inventory and in terms of ecosystem characteristics. This allows an evaluation of the precision of the model predictions and an assessment of the parameters contributing to their uncertainty. The latter is particularly important in the long term when factors other than the primary load become important in determining radiocaesium concentrations in lake water and in fish. The compartment model gave satisfactory predictions for concentrations in fish and lake waters during the first five years after Chernobyl. However, the results were sensitive to appropriate parameter values such as the K, and the biological half-life in fish. Uncertainty analyses demonstrated that leakage from the drainage area is important for mountain lakes, while resuspension is of significance in lowland lakes. As indicated by the model uncertainty analyses in chapter 2.6 and the sediment studies in

chapter 2.3, the behaviour of Chernobyl caesium is now entering a new phase as different processes, insignificant in the short term, begin to increase in importance. It is therefore essential that the research effort initiated after the Chernobyl accident is maintained. This is necessary in order to understand the long-term consequences of fallout from Chernobyl and other similar events, especially in systems with long half-lives. It will also provide a different set of dynamics, which will increase our knowledge and experience, thus forming a broader base for prediction and remedial measures should there be future and perhaps even more serious nuclear accidents.

As mentioned in chapter 1.1, the main emphasis in the aquatic radioecology programme was put on fresh-water radioecology. However, chapters 2.7 and 2.8 deal with marine and brackish water environments. Chapter 2.7 describes a project where the brown alga Fucus vesiculosus was used to monitor the level of radiocaesium in the coastal waters of all the Nordic countries, Denmark, Finland, Iceland, Norway and Sweden, in 1991. The Chernobyl fallout pattern appeared clearly with highest concentrations in the southern Bothnian Sea. Fucus vesiculosus occurs along most Nordic coasts except in the northern parts of the Baltic Sea, where it becomes scarce because of the low salinity. Epilithic diatom communities proved useful as an alternative bioindicator for radiocaesium in these waters. Whereas the main work in the present programme was centred on radiocaesium, chapter 2.8 reports concentrations of the natural a-emitting radionuclide 2'oPoas well as radiocaesium in fish

12 muscle from different Nordic marine areas, the Baltic Sea, the Norwegian Sea and Icelandic waters. A dose assessment after the Chernobyl accident showed that the population received similar doses from *loPoand radiocaesium via fish caught in the Baltic, whereas from other locations the dose from 210Powas the most important from the marine environment.

In addition to the importance of radiocaesium in the aquatic foodchain in terms of dose to man, fallout from Chernobyl has enormous potential as an ecological tracer. Chernobyl caesium has been and will indeed continue to be used as a tracer to monitor and elucidate basic ecological processes, as reviewed in chapter 2.9.

AGRICULTURAL ECOSYSTEMS Nordic agriculture is highly variable because of differences in climate, latitude, altitude and soil types. It includes a wide spectrum of farming, ranging from highly intensive grain, meat and dairy centres in Denmark and part of Sweden, southern Finland and south-east Norway, to free-range goat and sheep grazing in natural environments in Iceland and the Norwegian mountains. Direct contamination of agricultural plants immediately after a nuclear accident is the fastest and most direct route to the human foodchain. Chapter 3.2 deals with the direct contamination of agricultural products including secondary direct deposition, i.e. rain splash and resuspension. The chapter focuses on seasonality, i.e. the varying response to contamination of crops according to the time of year when contamination occurs. The effect of seasonality is largest for short-lived radionuclides (such as

I3lI)

and for elements that mainly enter the foodchain by direct

contamination (e.g. 137Cs).As a result of seasonality, the transfer of radiocaesium to man from the Chernobyl accident was higher in southern than in northern Europe normalized to the same deposition density. The effects of the physico-chemical forms of the deposited radionuclides on transfer and mobility in the environment are dealt with in chapter 3.3. The activity levels of radionuclides (Bq m-2) deposited in the Nordic countries showed considerable variation, even within a single m2. Activities in vegetation and transfer factors also show variations between sites, within sites, with time and between the different radionuclides. In 1989 studies on the mobility of radionuclides

(137Csand %Sr) in Norwegian soil-plant systems indicated that the fraction of radionuclides deposited as fuel particles was not having any significant effect on the transfer of 137Csor %Sr. Apparently the lability of 137Csand

depends more heavily on the physical and chemical

properties of the soil and on the chemical properties of the element, than on the fallout speciation. Hence, the particle form of deposition from Chernobyl is not expected to be important for future transfer of radionuclides in the Nordic countries. In contrast, studies on soils collected from the

30 km zone around Chernobyl suggest that the lability (or rather "non-lability") of wSr is largely

13 determined by the fraction associated with fuel particles. Studies on Norwegian soils suggest that both transfer factors and mobility factors are needed for a full understanding of the processes involved and for future predictions of radionuclides in the other parts of the ecosystem.

In chapter 3.4 special emphasis is laid on annual crops as a vector for the transfer of radiocaesium to man. Barley, potato, cabbage, carrot and pea are used as examples. After a nuclear accident, a common trend is that contamination levels in annual crops decrease rapidly from the first to the second year. Thereafter the rate of decrease is more variable and it seems that long ecological half-lives are possible in some agricultural ecosystems.The uptake of radiocaesium from soil through roots to edible parts of annual crops is generally very low in Scandinavian agricultural ecosystems, except on peaty organic or sandy soils that are often used for other purposes such as livestock or forage production. The most important pathway for the transfer of radiocaesium from annual crops to man is through direct contamination, because of the low uptake from soil. Therefore the season of the year is the most important factor determining the transfer to man after a nuclear accident, as mentioned above and in chapter 3.2. On the Faroe Islands the uptake is generally between one or two orders of magnitude higher than in the other Nordic countries. The high content of organic matter and sand may be part of the explanation. An effective half-life for radiocaesium content in barley of between 5 and 10 years seems reasonable on common Nordic arable land soil types in the first years after an accident. In potatoes a similar value of 6 years was calculated for Denmark. Following the Chernobyl nuclear accident in 1986 several studies were made in Denmark, the Faroe Islands, Finland, Iceland, Norway and Sweden on the transfer of 137Csfrom feed to cows’ milk. The present review (chapter 3.5) shows that the transfer of 137Csto cows’ milk related to ground deposition was highest in the Faroe Islands, Iceland and Norway and lowest in Denmark, Finland and Sweden. The effective ecological half-life for Chernobyl I3’Cs ranged from 1-2 years for all the Nordic countries and was 18.4 years for global 137Csfallout in Iceland. Radiocaesium transfer in the soil-herbage-lamb foodchain was assessed in a four-year trial conducted in sheep production locations of the Nordic countries (chapter 3.6). Radiocaesium contamination of the topsoil ranged from 3 to 30 kI3q m-’ and was predominantly of Chernobyl origin in Finland, Norway, and Sweden, whereas in Iceland 137Cswas primarily of nuclear weapons test origin, and in Denmark and the Faroe Islands contamination was derived from both sources. Soil-to-herbage radiocaesium transfer factors were high on the organic and acidic soils of the Faroe Islands, Iceland, Norway, and Sweden, averaging 18-82 Bq 137Cskg-I herbage on a soil deposition of 1 kBq 137Csm-’, and much lower on the sandy soils of Denmark and clay soils in Finland (0.4-0.8). Herbage-to-lamb concentration factors were generally more homogeneous,

indicating that the absorption of radiocaesium from herbage was similar in each of the countries.

14 A I3'Cs deposition of 1 kBq m-' soil gave rise to much lower meat radiocaesium concentrations

at the sites in Denmark, the Faroe Islands, and Finland (0.5-3.0 Bq kg-I) than in Iceland, Norway, and Sweden (20-47 Bq kg-'). It is concluded that among the Nordic countries the soil-herbage-lamb pathway is clearly of greatest importance in Iceland and Norway, intermediate in the Faroe Islands, and of comparatively lesser importance in Denmark and Sweden. The data were further utilized in a dynamic radioecological model describing the transfer of radiocaesium through the soil-grass-

lamb foodchain (chapter 3.7). Finally, chapter 3.8 reviews experiments on countermeasures after radioactive deposition in Nordic agricultural systems carried out since the sixties. Experiments have mainly concerned two strategies: ploughing and fertilization. It was found that efficient placement below root depth can be achieved by means of two-layer ploughs and by deep-ploughing equipment. However, soil type and moisture conditions in the soil during ploughing will influence the quality of the work. Loose, sandy soils and heavy clays are more difficult to handle than other soil types. On soils with low clay content such as sandy soils and peat soils, fertilization with up to 200 kg potassium per hectare can efficiently reduce caesium uptake by both grass and arable crops. These soils have low potassium reserves and need new potassium dressings during crop rotation. Heavy clays generally need no extra potassium dressings to reduce crop uptake of caesium. FOREST AND ALPINE ECOSYSTEMS

There is an area of overlap between the agricultural and the natural ecosystems in the Nordic countries. Some of the results described under the agricultural ecosystems (chapter 3) relate to the utilization of more or less natural ecosystems, e.g. sheep production in part, whereas reindeer herding is treated in chapter 4 alongside forest ecosystems and game animals. In the early sixties during the major atmospheric nuclear tests, the transfer of radiocaesium in the lichen - reindeer man foodchain was a major radioecological factor in Scandinavia. It was therefore more of a political difficulty than a scientific puzzle when, after Chernobyl, the natural ecosystems gave rise to relatively high individual doses. However, the actual transfer of radiocaesium through natural terrestrial ecosystems, and in particular the role of fungi in this transfer, gave new results. Chapter 4.2 deals with the transfer of radiocaesium from soil to plants and especially to fungi in seminatural ecosystems. The radiocaesium concentration in fungal fruit bodies is often more than

50 times higher than in plants growing at the same location, and whereas the radiocaesium content in higher plants has decreased since 1988, in fungi it has tended to be stable or even increasing. Comparisons with measurements of old global fallout radiocaesium make it possible to predict that the content of Chernobyl radiocaesium in fungi will be high for many years in several Nordic ecosystems. This has implications for the radiocaesium content of wild as well as domestic animals

15

grazing in seminatural and forest ecosystems. Furthermore chapter 4.2 reports on studies of horizontal and vertical redistribution of Chernobyl radiocaesium after deposition. In the mostly acid seminatural and forest soils in the Nordic countries, practically no vertical transport of radiocaesium has occurred. More than 90% is still bound in the top 3-4 cm organic layer. In areas covered with snow during the deposition,

a horizontal redistribution took place during snowmelt giving rise to much higher variation in the area content than in nearby sites not covered in snow during deposition. This may in part explain the patchiness mentioned elsewhere, e.g. in chapter 3.3. One of the main pathways for the transfer of radiocaesium from natural ecosystems to man is via game animals (chapter 4.3). Roe deer consume large quantities of fungi in autumn, resulting in a high and very variable content of radiocaesium. Normally, the radiocaesium concentration in

roe deer peaks in August to October. The transfer per kg of moose is lower and not as variable, partly because of the smaller consumption of fungi. However because of the importance of this supply of meat in Sweden, Norway and Finland, the transfer of radiocaesium to man via moose is much higher than that via roe deer. There has been no significant decrease in the radiocaesium content of moose or roe deer after Chernobyl, implying that the effective ecological half-lives for the forest ecosystems are very long. It is suggested that the physical half-life of 137Csand 134Cs may be the best estimate. As mentioned above, the lichen - reindeer - man foodchain was studied in Scandinavia in the early days of radioecology, and the Chernobyl accident put new life into these studies (chapter

4.4). The reason for the importance of reindeer as a vector for radiocaesium is its choice of food, which consists of 70-80% lichen in winter and 10-20% in summer. Coupled with the short biological half-life of caesium in reindeer, 10-20 days, this leads to a strong seasonal variation of radiocaesium in reindeer meat: a late winter high that is about five times higher than the late summer low. In contrast to results from the game animals above, an effective ecological half-life of radiocaesium in reindeer meat after Chernobyl could be estimated to 3-4 years. For the lichen species serving as winter forage, effective ecological half-lives of 5-7 years on ridges and 6-11 years in more sheltered habitats were observed. Finally, chapter 4.5 reviews the distribution of radiocaesium in boreal forest ecosystems based on Chernobyl as well as global fallout results. The review thus focuses on data of relevance for both the early and the later phases after nuclear fallout over forest areas. In boreal forests the humus layer usually retains a major fraction of the deposited radiocaesium even decades after deposition. This feature, as well as a persistent high availability in important foodchains, may explain the long effective ecological half-lives, approaching the physical half-life of the radionuclides, observed for radiocaesium in forest ecosystems. This is in contrast to the intensive

agricultural ecosystems (chapter 3 ) and even to the reindeer ecosystem (chapter 4.4), where a significant decrease in concentrations with time is observed.

METHODOLOGY, QUALITY ASSURANCE AND TRAINING

For all environmental measurements, quality assurance of the analyzed values is of central importance. In the present programme, the concept of quality assurance has included qualitypromoting activities such as the exchange of analytical methodology, short exchange programmes for scientists wanting to acquire knowledge of an analytical method from one of the other Nordic laboratories, and a two-week postgraduate training course including 20 lectures on several aspects of radioecology from sampling and radiochemistry to statistical analysis. The course included a series of practical laboratory exercises. The 20 lectures are being published in book form (Holm, editor). In the field of radioecology, international intercomparisons of low-level radionuclide

concentrations, measured in thoroughly homogenized samples, are organized routinely by the International Atomic Energy Agency (IAEA) in Vienna and Monaco. Under the present Nordic programme, most of the old-established laboratories were already participants in the IAEA intercomparisons, and it was decided to urge the remaining laboratories to join. However, two types of equipment of central importance for the surveillance of nuclear fallout, and for dose assessment, are normally not quality-assured on an international scale: large stationary air samplers and whole-body counters. The reports in chapters 5.2 and 5.3 are therefore internationally unique. The intercomparison of large stationary air samplers (chapter 5.2) was performed by circulating two high-volume air samplers between the nine participating laboratories and operating them for two - six months parallel with the local air sampler. The intercomparison included several types of filter material, including glass fibre as well as organic filter media. During part of the test period (1990-1993), air concentrations of 137Cswere too low for high-quality measurements. The natural radionuclide 7Be was therefore used as the main basis for the comparisons showing a difference of up to 15% when using one type of glass-fibre filters and no significant difference using another type of glass fibre. This indicates that the quality of the data on radionuclides in air from the Nordic countries is surprisingly good. Whole-body counting is used for the determination of X- and y-emitting radionuclides in the human body. Its use includes the surveillance of selected groups of the general public and of radiation workers for dosimetric purposes. The intercalibration of 20 Nordic whole-body counting systems (chapter 5.3) was performed by circulating a modular phantom system filled with calibrated solutions of radiocaesium. The modular phantom could simulate all varieties of wholebody geometries in use. The observed quotient between measured and expected activity was 0.9 -

17

1.1 for most systems, i.e. *lo%. This is better than previously expected. Finally, two sets of homogenized samples intended for y-spectrometric analysis were distributed as a supplement to the above-mentioned IAEA sources. The results from 26 laboratories given in chapter 5.4 are generally satisfactory, although there were a few unexplained outliers. INTERNAL DOSES TO THE NORDIC POPULATION One of the aims of the RAD programme was to produce a good data background for the estimation of doses to the Nordic population after the Chernobyl accident. Furthermore this was a good basis

on which to make better predictions of population doses after any future nuclear contamination of various Nordic environments. Two main approaches were used for the dose estimates: food intake (chapter 5.5) and whole-body counting (chapter 5.6). The individual mean doses from radiocaesium intake with diet since the Chernobyl accident in 1986 were determined for Denmark, Finland, Iceland, Norway and Sweden (chapter 5.5). The estimates were obtained by two methods. The first used consumption data, i.e. information on the amounts of food eaten by an average individual in each of the five countries. The other method applied food production in the Nordic countries, ignoring the export and import of food but taking into account the amounts actually eaten. The consumption method gave an individual mean dose commitment of 1.3 mSv and the production method gave 1.0 mSv. In comparison the external mean dose, i.e. the dose received from penetrating radiation emitted by radionuclides outside the body, was 0.8 mSv for the Nordic countries. Figure 1.2.2 shows the relative intake of 137Csfrom different diet groups in % since the Chernobyl accident by an average person in Denmark, Finland, Norway and Sweden. The study emphasizes the importance of wild produce for the internal doses from radiocaesium. More than 50% of the total 137Csintake with the Nordic diet came from natural and seminatural ecosystems. In this context it is unfortunate that information on the consumption of and radiocaesium concentration in wild produce is relatively scarce. It is believed that the dose based on consumption data is an overestimate because of the lack of reliable information especially on wild produce, both with regard to amounts actually eaten and because the exact effective half-lives are not known. Nordic critical groups with high consumptions of fungi, wild berries, reindeer, freshwater fish, elk, lamb and goat products may receive dose commitments from dietary intake that are 1-2 orders of magnitude higher than those of the general population. Such groups are found in Norway, Sweden and Finland, in particular among the Lapp population. It should, however, be kept in mind that remedial measures introduced in the Nordic countries after Chernobyl significantly reduced the exposure of these population groups. After the Chernobyl accident whole-body measurements on selected population groups were performed in Denmark, Finland, Norway and Sweden. Chapter 5.6 presents the mean internal

18 Table 1.2.1. A comparison between the Nordic countries of radioecological sensitivities in total diet for Chernobyl 137Cs. Country

Population,

Area,

Sensitivity,

millions

109 m2

Bq kg-'

Denmark

5.1

43

4.4

Finland

5.0

338

13

Iceland

0.25

103

Norway

4.2

324

33

Sweden

8.4

450

20

Faroes

0.04

1.4

19

c

23

1259

18

* yr / kBq m-2

effective doses caused by '34Csand 13'Cs originating from the Chernobyl accident calculated on the basis of these measurements. The dose estimates above, based on dietary intake, were higher than the present estimates based on whole-body measurements ranging from a factor 1.2 for Denmark and up to a factor 8 for Sweden. One possible explanation suggested in chapter 5.6 could be that the biological half-life of radiocaesium in the Nordic countries is shorter than the internationally accepted values used in the calculation based on the food consumption data. If so, the whole-body content and the estimated dose would be lower than reported in Chapter 5.5. Other explanations could be that the selected whole-body groups were not representative enough, poor representativeness of the radionuclide concentration in samples used to estimate the radiocaesium content of the diet, or limited knowledge of the amounts of wild produce actually consumed. These last explanations might further explain the large discrepancy found in Sweden, where the contamination level was extremely variable resulting in almost unattainable representativeness, and the better correlation in Denmark, where fallout was lower and much more homogeneously distributed. The introduction of the term radioecological sensitivity reveals that, on average, the Chernobyl-derived radiocaesium concentration in a diet produced in Norway would be 7 times higher than that of a diet produced in Denmark for the same ground surface deposition (Table 1.2.1). The radioecological sensitivity for 137Csin diet is defined as the infinite time-integrated

19

Figure 1.2.2. Relative intake by an average person in Denmark (DK), Finland (SF), Norway (NO) and Sweden (SW) of 13'Cs from different diet groups in % since Chernobyl.

20 concentration of 137Csin the diet arising from a given deposition, Bq kg-' * yr / kBq m-* (Aarkrog, 1979). Table 1.2.1 also shows that, on average, a unit deposition in Finland would result in 3 times higher, and in Sweden and the Faroe Island 5 times higher diet concentrations than in Denmark. However, as food production in Denmark is much greater than in the other Nordic countries, contarnination in Denmark might give rise to a larger population dose if no countermeasures were introduced. By comparing the radioecological sensitivity for Chernobyl 137Cs in a diet produced in Denmark with comparable values found earlier for global fallout (Aarkrog, 1979), it is seen that the transfer of global fallout was transferred 2.5 times more efficiently to man than the Chernobyl debris*. The primary reason for this is seasonality (chapter 3.2), which resulted in lower 137Csconcentrations in the production of especially grain and milk during the first year after the Chernobyl accident than seen for similar depositions of global fallout.

REFERENCES Aarkrog, A. (1979). Environmental Studies on Radioecological Sensitivity and Variability with Special Emphasis on the Fallout Nuclides ? S r and I3'Cs. Rise-R-437. Holm, E. (editor). Radioecology. Lecture Notes in Environmental Radioactivity. World Scientific Publishing Co., Singapore. (1994, in press). NKS (1991). Radioecology in Nordic Limnic Systems - Present Knowledge and Future Prospects. SNV Report 3949.

* For total Danish diet 1963 - 1976, the radioecological sensitivity was 4.2 Bq 137Cs(g K)-' per kBq 137Csm-2 or 1 1 Bq 137Cskg-I per kBq '37Csm-2 (Aarkrog, 1979).

Chapter 2 AQUATIC ECOSYSTEMS

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23

2.1. INTRODUCTION TO AQUATIC ECOSYSTEMS

MANUELA NOTTER', JOHN E. BRITTAIN' & ULLA BERGSTROM' 'Swedish Environmental Protection Agency, 171 85 Solna, Sweden 'Freshwater Ecology and Inland Fisheries Laboratory (LFI), University of Oslo, Sars gate 1,0562 Oslo, Norway. 3Studsvik Eco and Safety, 61 1 82 Nykoping, Sweden.

SUMMARY This paper summarizes the background, objectives and major results of the NKS programme on aquatic radioecology and serves as an introduction to the more detailed research papers. The programme included both marine and freshwater studies. INTRODUCTION The NKS RAD-2 programme on aquatic radioecology continues a long Nordic tradition in cooperative work concerning the behaviour of radionuclides in aquatic ecosystems. In a previous Nordic project (Nilsson et al., 1981) the environmental status with regard to radioactive pollution in the seas surrounding the Nordic countries was studied using the seaweed, Fucus vesiculosus.

Fucus samples were analyzed for their content of radionuclides and distribution patterns and turnover times were obtained. More recently there has been a need to verify previous models and compare the behaviour of Chernobyl caesium with earlier results. From studies of fallout in the 1960's (Kolehmainen et al., 1966; 1967; 1986; Hasanen et al., 1963; 1967; 1968) it was known that predatory fish in oligotrophic lakes reach high concentration levels of caesium. It was also known that different fish species reach varying caesium levels depending on feeding habits (Hannertz, 1966; 1968). The Chernobyl accident took place four years prior to the start of the present programme. Oligotrophic lakes predominate in northern Scandinavia and fish from these lakes rapidly reached high concentrations of caesium in areas with high fallout rates. There was a considerable interest from the authorities for models to predict caesium concentrations in fish as the consumption of freshwater fish is the major source of the dose to the Nordic populations received via the aquatic food web. Model development and validation were also given high priority internationally. Several international studies were initiated to create and verify radioecological fish models.

24

OBJECTnTES The three main objectives of the RAD-2 project were to: -

collect data for developing and evaluating models for the prediction of caesium concentration in fish for different types of Nordic lakes,

-

earlier studies of the concentration of radionuclides in the bladderwrack Fucus

vesiculosus and to compare the uptake in Fucus with the accumulation rates in other algae, -

secure data for a relevant calculation of the dose to the Nordic population from the aquatic environment and to compare the dose contributed by Chernobyl with the dose received by radiation from natural sources, Numerous participants from all the Nordic countries have worked on the programme,

although in most cases RAD-2 has only given limited financial support. However, it has made it possible for Nordic scientists in the field of aquatic radioecology to meet in small groups to discuss mutual problems and to co-operate. Six seminars/workshops were held under the auspices of the programme. RAD-2 has had a total budget of Dkr 1.2 million, but the participants and their institutions have contributed substantially both in terms of funding and in personal involvement. Their joint efforts have also permitted the presentation of ongoing research projects outside RAD2, thereby contributing to the success of this work. BACKGROUND AND MAIN RESULTS

Carlsson et al. (1994) report the results from the efforts that were put into repeated Fucus investigations in 1991 in order to provide a picture of caesium distribution in the Nordic sea basins after Chernobyl. The accumulation rates and biological half-lives in Fucus are compared with those of other algal species, particularly benthic diatoms. A summary of the results was presented at the Nordic Radioecology Seminar in June 1992 (Carlsson et al., 1992). Resources were also directed towards assessing the radiation dose that can be received by the population through fish consumption. Several radionuclides were measured in herring, cod, perch and char. Fish also contain certain amounts of natural radionuclides, including 2'%,

which

will contribute to the dose. As very few data are available, this programme has encouraged analyses providing improved dose calculations for 2"%'oin fish from Nordic waters (Holm et al., 1994). In addition to the importance of radiocaesium in the aquatic food chain in terms of dose to man, fallout from Chernobyl has an enormous potential as an ecological tracer. Radionuclides

25

in general, and certainly Chemobyl caesium, have been and will indeed continue to be used as tracers to monitor and elucidate basic ecological processes. Meili (1994) provides a review of such studies. One of the main concerns after the Chernobyl accident was the concentration of I3’Cs in the aquatic food chain and particularly in freshwater fish. In lakes the main exposure pathway of I3’Cs to man is through the consumption of freshwater fish. Highest priority and considerable RAD-2 resources were given to studies of the behaviour and bioavailibility of caesium in freshwater systems. The main part of this chaper gives the results of these studies. Largely through co-ordinating of results from ongoing work in the Nordic countries, it was possible to study the influence of lake morphology and hydrology on caesium concentrations in fish and also within the relevant food webs. It was possible to elucidate the major factors determining concentrations in freshwater fish and in freshwater ecosystems in general, thereby contributing to dose assessment studies. The identification of the important parameters determining radionuclide concentrations in fish also permits the development and assessment of potential remedial measures in aquatic ecosystems. As a result of processes associated with the last Ice Age, lakes are a typical feature of the landscape in the Nordic countries. This is especially striking in Finland, although there is also a high incidence of lakes both in Norway and Sweden. In the Nordic countries, freshwater fishing is therefore widespread, both as a leisure activity and a commercial undertaking. Sports fishing is also an integral part of the tourism associated with the unspoilt countryside and pristine environments typical of the Nordic countries. In many areas freshwater fish also form an important part of people’s diet and there are several traditional methods of preparation. Caesium accumulates in fish muscle because of its chemical similarity to potassium. This accumulation is most pronounced in freshwater and is of particular importance in the Nordic countries where ionic concentrations in freshwaters are generally low. However, Nordic lakes differ widely in many other characteristics. For instance there are wide differences between lowland, coastal lakes and high altitude, mountain fresh waters in terms of, for example, temperature and fish species. Winter ice cover is also a feature of importance for many lakes, especially as much of the Nordic countries was still covered in ice and snow at the time of the Chernobyl accident. The environmental impact of radionuclide releases from nuclear installations can be predicted using assessment models. However, many of the models were developed and tested on the basis of the fallout from nuclear weapons testing in the 1950s and 1960s, or from laboratory experiments. In contrast, fallout from Chernobyl constituted a single mdionuclide pulse which entered natural, agricultural and urban ecosystems at the end of April 1986. The fallout was also

26 in physical and chemical forms differing from those of the weapons testing fallout because of its quite different origin. Thus, the Chernobyl accident provided a unique opportunity to test and validate radioecological models for point release. Since 1986 there has been an intensive research effort in the Nordic countries aimed at obtaining reliable input data for prediction models and determining the important driving processes and parameters for such models. This research has been funded by the national research councils, research institutions and universities. The research results presented in this chapter on Nordic lakes were supported by various sources. The Nordic Nuclear Safety Research Committee has supported certain projects and contributed to the collation and presentation of the results (NKS, 1991; Dahlgaard, 1994). Lakes received radionuclides from Chernobyl fallout via two sources: direct fallout on the lake surface, the primary load, and by leakage from the catchment, the secondary load. In the first instance the primary load is of major importance, but in the long term inputs from the catchment can be of importance in determining radiocaesium concentrations in fish. Bjarnstad et al. (1994), using fractionation techniques, studied the input of radiocaesium to three very different Nordic lakes, one each in Sweden, Norway and Finland. Using hydrological data, they also estimated the degree of retention of '37Csin these three lake systems, both in terms of total concentrations and in terms of the different sue fractions from plant material to low molecular weight species. This is a useful approach in explaining the transport, behaviour and biological uptake of radionuclides. Some preliminary results were given at the Nordic Radioecology Seminar in 1992 (Bjomstad et al., 1992). Although some of the radiocaesium from Chernobyl has been transported out of lakes because of the high flows associated with the spring snowmelt at the time of deposition, most of it still remains in lake sediments. The distribution, physico-chemical forms and concentration of radiocaesium in lake sediments are thus potentially of major importance in determining the Iong-

term fate of Chernobyl caesium in our lakes (Broberg, 1994). The importance of leakage from catchment areas has been studied on a much larger scale in Finland, where the whole country has been divided into seven different catchments, each with its own characteristics with regard to fallout, soil type and topography (Saxen, 1994). A number of factors, both abiotic and biotic, have been put forward as determining the concentration of radiocaesium in fish. In a major study, encompassing a large number of Swedish lakes, Anderson and Meili (1994) assessed the importance of a wide range of such factors. Such studies are essential in evaluating the appropriate model compartments. An assessment of whole-lake models is also included in this chapter (Bergstom & Sundblad, 1994). This is based on three Nordic lakes for which extensive data are available, both in terms

27 of the radiocaesium inventory and in terms of ecosystem characteristics. This enables an evaluation to be made of the precision of the model predictions and an assessment of the parameters contributing to their uncertainty. The latter is particularly important in the long term when factors other than the primary load become important in determing radiocaesium concentrations in lake water and in fish.

FUTURE RESEARCH The behaviour of Chernobyl caesium is now entering a new phase as different processes, insignificant in the short term, begin to increase in importance. It is therefore of considerable importance that the research effort initiated after the Chernobyl accident is maintained. This is necessary in order to understand the long-term consequences of fallout from Chernobyl and other similar events. It will also provide a different set of dynamics which will increase our knowledge and experience, thus forming a broader base for prediction and remedial measures in the case of future and perhaps more serious nuclear contamination. REFERENCES Anderson, T. and M. Meili. 1994. The role of lake-specific factors for the transfer of radiocaesium fallout to fish. In Dahlgaard, H. (ed.).Nordic Radioecology. Elsevier, Amsterdam. Carlsson, L., E. Ilus, G. Christensen, H. Dahlgaard and E. Holm. 1992. Radionuklidinnehillet i Fucus vesiculosus langs de nordiska kusterna sommaren 1991. Nordic Radioecology Seminar, Torshavn, 1992. Carlsson, L. and P. Snoeijs. 1994. Radiocaesium in algae from Nordic coastal waters. In Dahlgaard, H. (ed.). Nordic Radioecology. Elsevier, Amsterdam. Bergstrom, U. and B. Sundblad. 1994. Whole-lake models. In Dahlgaard, H. (ed.). Nordic Radioecology. Elsevier, Amsterdam. Bjornstad, H. E., J.E. Brittain, R. SaxCn, B. Sundblad and B. Salbu. 1992. Karakt;irisering av radionuklidtillforsel till Nordiska insjoar. Nordic Radioecology Seminar, Torshavn, 1992. Bj~rnstad,H. E., J.E. Brittain, R. SaxCn and B. Sundblad. 1994. The characterization of radiocaesium transport and retention in Nordic lakes. In Dahlgaard, H. (ed.) "Nordic Radioecology". Elsevier, Amsterdam. Broberg, A. 1994. The distribution and characterization of '37Csin lake sediments. In Dahlgaard, H. (ed.). Nordic Radioecology. Elsevier, Amsterdam. Dahlgaard, H. 1994. (ed.). Nordic Radioecology. Elsevier, Amsterdam. Hannertz, L. 1966. Fallout 13'Cs in fish and plankton from Lake Malar and the Baltic. Acra Radiologica, Suppl. 254:22-28. Hannertz, L. 1968. The role of feeding habits in the accumulation of fallout '"Cs in fish. Rep. Inst. Freshw. Res. Drottningholm 48: 112-119. Holm, E. and G. 1994. Christensen. Po-210 in muscle tissue of marine fish from different Nordic areas. In Dahlgaard, H. (ed.).Nordic Radioecology. Elsevier, Amsterdam. Hasanen, E. and J.K. Miettinen. 1963. Caesium-137 content of fresh-water fish in Finland. Nature. 2OO(49 10): 1018-1019.

28 Hasanen, E., S . Kolehmainen and J.K. Miettinen. 1967. Biological half-time of 137Csin three species of fresh-water fish: perch, roach, and rainbow trout. p. 921-924. In Radiological Concentration Processes. Eds: B. Aberg & F.P. Hungate. Proc. Int. Symp., Stockholm, April 1966. Pergamon Press, Oxford. Hasanen, E., S . Kolehmainen and J.K. Miettinen. 1968. Biological half-time of 137Csand "Na in different fish species and their temperature dependence. p. 401-406. In W. S. Snyder (Ed.). Proc. 1st Int. Congr. Radiolog. Protect. Vol. I . Pergamon Press, New York. Kolehmainen, S . , E. Hasanen and J.K. Miettinen. 1966. 13'Cs levels in fish of different limnological types of lakes in Finland during 1963. Health Physics. 12:917-922. Kolehmainen, S . , E. HZsiinen and J.K. Miettinen. 1967. 137Csin fish, plankton and plants in Finnish lakes during 1964-65. p. 913-919. In Radiological Concentration Processes. a s : B. Aberg & F.P. Hungate. Proc. Int. Symp., Stockholm, April 1966. Pergamon Press, Oxford. Kolehmainen, S . , E. Hasanen and J.K. Miettinen. 1968. '37Csin the plants, plankton and fish of the Finnish lakes and factors affecting its accumulation. p. 407-415. In W. S. Snyder (Ed.). Proc. 1st Int. Congr. Radiolog. Protect. Vol. 1. Pergamon Press, New York. Meili, M. 1994. Fallout caesium as an ecological tracer. In Dahlgaard, H. (ed.) Nordic Radioecology. Elsevier, Amsterdam. Nilsson, M., Dahlgaard, H., Edgren, M., Holm, E., Mattsson, S . and M. Notter. 1981. Radionuclides in Fucus from inter-ScandinavianWaters. IAEA-SM 248/ 107, pp 501-5 13. International Atomic Energy Agency, Vienna. NKS. 1991. Radioecology in Nordic Limnic systems - present knowledge and future prospects. SNV report 3949. Saxen, R. 1994. Transport of 137Csin large Finnish drainage basins. In Dahlgaard, H. (ed.) "Nordic Radioecology". Elsevier, Amsterdam.

29

2.2. THE

CHARACTERIZATION

OF

RADIOCAESIUM

TRANSPORT AND

RETENTIONIN NORDIC LAKES

HELGE E. BJPIRNSTAD', JOHN E. BRITTAIN*,RITVA SAXEN3 & BJORN SUNDBLAD' 'Laboratory of Analytical Chemistry, Agricultural University of Norway, P.O. Box 5026, N-1432

As, Norway. 'Freshwater Ecology and Inland Fisheries Laboratory (LFI),University of Oslo, Sarsgt. 1, 0562

Oslo, Norway. 3Finnish Centre for Radiation and Nuclear Safety, P.O. Box 268, 00101 Helsinki, Finland. 4Studsvik Ecology & Safety, 61 1 82 Nykoping, Sweden.

SUMMARY

Fractionation studies of radiocaesium have been carried out in three Nordic lakes, 0vre Heimdalsvatn in Norway, Hillesjon in Sweden and Saarisjarvi in Finland. These lakes differ markedly in several aspects and provide insight into the factors determining radionuclide transport in a range of lake ecosystems. Transport of I3'Cs in plant material (Coarse Particulate Organic Matter, CPOM) was about 17 times greater to 0vre Heimdalsvatn than Saarisjarvi, although over 99 % of the inflow CPOM was retained in both lakes. Inflows to Hillesjon were an order of magnitude lower than to Saarisjhi and the net retention was only 71 X, on account of the outflow of autochthonous production, largely water lily fragments. With regard to the water phase, the lakes differed in the activity of I3'Cs in the various molecular weight fractions. This was a function of catchment processes, resuspension and biological activity in the lakes. In 0vre Heimdalsvatn and Saarisjarvi 45 % of the '37Csin the water phase was retained in the lake, while in Hillesjon ten times more '37Csflowed out than flowed in, due to resuspension of '37Cs-ri~hsediments.

INTRODUCTION Fallout from the Chernobyl accident reached Finland, Sweden and Norway at the end of April 1986. Among the areas of high deposition ( > 70 kBq m-') were localities in central southern Finland near Lammi, around the city of Gavle in Sweden and in the Jotunheimen mountains of central southern Norway (NKS, 1991). Lakes in these areas have been the subject of several radioecological studies and thus formed a natural basis for the characterization of radionuclide inputs to Nordic lakes. Previous studies of the Norwegian subalpine lake, 0 v r e Heimdalsvatn, have shown the

30 importance of inputs from the catchment for lake radionuclide dynamics (Brittain et al., 1992; Salbu et al., 1992). Size distribution patterns elucidated by fractionation techniques and lake budget calculations have demonstrated the significance of transport forms for the degree of retention in the lake system. On account of differences in the biological, chemical and physical characteristics of lakes and their catchments, transport form and mechanisms are likely to differ among freshwater systems. In order to identify transport mechanisms, the waters and the plant material transported by them have been fractionated with respect to particle size. Based on the input-output budget, the fraction of radionuclides retained in the lake system can be estimated. Run-off during the spring snowmelt is an important pathway for radionuclide transport (Salbu et al., 1992). Therefore, during the spring snowmelt period of 1991 comparable investigations were carried out in 0vre Heimdalsvatn in Norway, Hillesjon in Sweden and Saarisjarvi in Finland.

SITE DESCRIPTIONS 0vre Heimdalsvatn, Norway The subalpine lake, 0vre Heimdalsvatn, is situated on the eastern edge of the Jotunheimen mountains in central southern Norway (Table 2.2.1, Fig. 2.2.1). The highest point of the catchment is 1843 m a.s.1. Vegetation ranges from subalpine birch forest with areas of mountain pasture to high alpine vegetation above 1600 m. The lake is poor in electrolytes and wind exposed. The average renewal period for the lake varies considerably between a few days at the peak of the spring spate and over 400 days during winter (Vik, 1978). The lake is ice-covered from midOctober until the beginning of June. The input of terrestrial plant (allochthonous) material from the catchment is of major importance as a source of organic matter for the lake (Larsson et al., 1978). The Concentration of I3’Cs in lake waters was 5.5 kBq m-3in June 1986just after ice break. The concentration fell to about 250 Bq m-3 by the spring of 1989.

Hillesjon, Sweden The lake, Hillesjon, is situated north of the town of Gavle about 5 km from the eastern coast of central Sweden (Table 2.2.1, Fig. 2.2.2). Over 80% of the catchment is covered by forest; the remainder is agricultural land and marshes. During summer large areas of the lake become covered with aquatic macrophytes. Hillesjon is eutrophic, with a primary production of approximately 100 g C m-*y-’. The lake sediments have an organic content of about 35 %. The lake is ice-covered

between December and April/May. The initial peak concentration of I3’Cs in lake waters was approximately 6.5 kBq m3. This had declined to about 1 kBq

by 1990 although winter values

32 were generally lower.

TABLE 2.2.1 - Selected physical, chemical and biological parameters of the investigated lakes. Heimdalsvatn

Hillesjon

Saarisjarvi

61" 25' N

60" 45' N

8" 50' E

17" 1 2 ' E

25" 7 ' E

1090

10

125

Catchment area kmz

23.6

19

7.9

Lake area km2

0.78

1.6

0.12

Catchment/ lake area

30

12

66

Max. depth m

13

3

Mean depth m

4.7

1.7

Mean renewal period -days

63

130

c. 110

I3'Cs deposition kBq ni2

130

100

35-70

Latitude Longitude Altitude

m a.s.1.

Trophic status

Oligotrophic

Eutrophic

61" 1 1 "

Mesotrophic

PH

6.8

7.3

6.2

Conductivity mSm-'

1.3

40

6.1

P pg 1"

2

11

29

Ca mg 1.'

1.7

1.o

7.2

K mg 1.'

0.4

3.0

Saarisjarvi, Finland Saarisjhi is situated in the municipality of Lammi, Finland. About 75% of the catchment is forest, 15% bogs and marshes and 10% farm pasture (Table 2.2.1). The catchment contains few lakes, and none occur on the major inflow river studied, Joutsjoki (Fig. 2.2.3).

33

Hillesjon

Figure 2.2.2. Location and catchment of Hillesjon, Sweden

4 Figure 2.2.3. Location and catchment of Saarisjarvi, Finland.

o

500m

34 The initial lake water concentration of '"Cs has not been measured, but a concentration of

4 . 6 kJ3q m-3 I3'Cs was measured in the nearby lake, Is0 Valkjarvi, in June 1987 (SaxBn, 1990). Chemical and radioecological data for nearby lakes are given in SaxBn (1988), Arvola et al. (1990) and Rask (1991).

SAMPLING AND FRACTIONATION TECHNIQUES Waters from the lakes, their inflows and outlets were collected during the spring of 1991. Material was collected from Hillesjon during the period 25 April-8 May, from Saarisjarvi 4 to 7 May and 0vre Heimdalsvatn from 24 May to 3 June. The stream and river waters were fractionated with respect to particle size: Coarse

particulate organic material (CPOM) was collected in drift traps suspended in the current. The traps consisted of oblong nets with an opening 5 x 25 cm and a mesh size 0.9 mm (Larsson & Tangen 1975, Aunan 1986). Discharge was measured directly, either using a current meter over a known profile or the salt dilution method (Hongve 1987). The macromolecularfraction

(m,

the pseudocolloidal fractions (CF1, CF2) and low molecular weight fraction (LJMF) were obtained by a tangial flow ultrafiltration unit (Millipore XX4202K50; Millipore, Bedford, Ma., U.S.A.). The fractions were produced using three different ultrafidtration membranes, with the levels of O.1lm (Millipore VVLP), 10 kDa (Millipore PTGC) and 1 kDa (Novesett NS001005, Filtron, Mass., U.S.A.). The fractionation was not performed sequentially, but on aliquots of the total sample. The standardization of ultrafiltration membranes is usually carried out using globular proteins, or dextrans. The membranes used were specified according to globular proteins. As the components in natural water seldom have the spherical structure of globular proteins and differ in atomic composition (e.g. Si, Al, Fe) compared to organic calibration components (e.g. C, H, N), we prefer metric units. 10 kDa and 1 kDa correspond approximately to a Nominal Molecular Diameter (NMD) of 1.5nm and 1.2nm, respectively (Amicon publ. 426V, Amicon, Ma., U.S.A.).

The HMF fraction corresponds to a NMD of

> 100nm.

Total and fractionated samples (251) were collected and after adding camers (20 mg Cs and

30 nig Y per sample) and preservatives (2 ml HN03/1 sample) they were stored at 4OC in polyethylene containers until analysis and weighed accurately. After analysis the different fractions were calculated according to the following equation:

T = HMF

+ CF, + CF, + LMF

where T = the total concentration of I3'Cs which can be normalized to 100%.

HMF describes particles with NMD > I OOnm, CF, components in the macromolecular range with

a lOOnrn
CF,

components in

the molecular range with

a

35 1S n m < NMD < 1.2nm and LMF describes components in the ionic range with a NMD < 1.2nm.

In the mass balance budget calculations for the lakes a state of hydrological equilibrium is assumed. This is a reasonable assumption as renewal periods are short and at a minimum during the spring.

SAMPLE PREPARATION Samples of CPOM were dried in the laboratory at 7OoCand weighed accurately. The predominant plant materials in each sample were sorted and weighed separately after drying. 251 water samples were evaporated to 11 and transferred to a Marinelli beaker (11) prior to gamma-spectrometry .

Gamma-spectrometry The evaporated water samples, total and filtered water, were analyzed with respect to ‘37Csusing a Canberra Ge detector (20% efficiency and 2 2 keV resolution at 1332 keV) interfacing a PC equipped with the spectrum AT software manufactured by Canberra (Conneticut, U.S.A.). Plant material samples (CPOM) were measured using a Minaxi 5000 Auto-Gamma with through-hole NaI scintillation detector (Packard International S.A., Illinois, U.S.A.). Measurements were carried out at the Isotope Laboratory of the Agricultural University of Norway. The counting errors were in the order of 1 % for the plant material and 10% for the water samples. The limit for quantitative determination of radioactivity (L,) was in accordance with Currie (1968).

RESULTS AND DISCUSSION The lakes studied represent a wide range in physical, chemical and biological characteristics. ‘37Cs deposition was of the same order of magnitude, although the lowest deposition was around the Finnish lake, about half that recorded for the Norwegian site. Fallout in the Swedish lake catchment was intermediate. The composition and I3’Cs concentration of the plant material (CPOM) transported into the Norwegian lake, 0 v r e Heimdalsvatn, by the inflowing streams has been dealt with in earlier publications (Brittain et al., 1992; Salbu et al., 1992). The major component, both in terms of the weight transported per 24h and its contribution to the L37Csload, was leaves of dwarf willow,

m. Although the highest

I3’Cs concentrations were recorded in lichens and mosses, their

occurrence and weight was very low compared to

m.There were also differences in I3’Cs

concentrations in the same plant material from different streams. This probably reflects local differences in deposition, soil type, vegetation cover and drainage (Hilton et al., 1992; Haugen et

al., in press).

36 TABLE 2.2.2 - Concentration of 137Csin plant material transported by the stream Hille syd, the proportion by weight of various components and their relative contribution to the 137Csload. Bq kg-I

% wt

% ‘37Cs

contrib. Herbs

260 f 11

8.9

4.0

Grasdcereals

164 f 9

16.8

4.8

Aq. macrophytes

265 f 12

7.0

3.2

Woodltwigs

< 70

Unidentified

760

Mean

579

0.3

c0.1

67.0

88.0

Similar differences were seen in the plant material from the Swedish and Finnish streams. The two investigated inflow streams to Hillesjon differ markedly in their catchment characteristics. Hille syd drains arable agricultural land, while &gland runs largely through coniferous forest. This was in part reflected in the composition of the CPOM samples (Tables 2.2.2, 2.2.3). In general, 137Csconcentrations were similar for the same material. Soil management and cultivation has not caused any major reduction in the radiocaesium concentration of the plant material carried by Hille syd in relation to hgland. This is in line with studies of radioactivity losses from agricultural areas (RosBn, 1991). A large proportion of the material in the CPOM traps was unidentified plant material, which had comparatively high 137Csconcentrations. This fraction also contained particulate mineral material transported by the streams as a result of soil and sediment erosion. Thus, it is likely that much of the 137Csactivity still remains in the soils and the stream sediments in the lake catchment (cf. Haugen et al., in press). This is supported by annual studies of the loss of I3’Cs from the Hillesjon catchment which have shown a total loss of only 8 % of the deposition during the period 1986-1991 (Sundblad et al., 1991). The 137Csconcentration in plant material in the outflow of Hillesjon was much higher than in the inflow streams (Table 2.2.4). This was especially true of the water lily leaves )”(

and

the unidentified fraction, both of which had concentrations in excess of 2 kl3q kg-I. Resuspension of lake sediments is a feature of the shallow, wind-exposed Hillesjon (Sundblad et al., 1989), and sediment material had most likely adhered to the leaves of water lilies. The unidentified fraction probably consisted largely of water lily leaves as the leaf fraction had a low occurrence compared to the stems, which are more resistant to physical breakdown than the leaves. In contrast to

37 lake, Saarisjarvi (Tables 2.2.5, 2.2.6). Apart from coniferous twigs in the inflow, which had a very high value, 137Csconcentrations were in the range 300-600 Bq kg-'.

TABLE 2.2.3 - Concentration of 137Csin plant material transported by the stream Angland, the proportion by weight of various components and their relative contribution to the 13'Cs load. (* approximate values). Bq kg-'

% wt

% 137cs

contrib. Equisetum

455 f 17

10.8

8.3

Aq. rnacrophytes

368 f 26

6.0

3.7

Grass

196

6.3

2.1

Deciduous leaves*

466

1.7

1.3

Woodltw igs

277 f 8

2.6

1.2

Bark

176 f 8

2.6

0.8

Pine needles*

129

2.9

0.6

Mossllichens*

690

Unidentified

720 f 53

Mean

627

* 11

0.2

0.2

67.0

81.7

TABLE 2.2.4. Concentration of '37Csin plant material transported by the outflow of Hillesjon, the proportion by weight of various components and their relative contribution to the I3'Cs load. Bq kg-'

% wt

% 137CS

contrib. Nuphar - stems

394 _+ 12

45.8

21.0

Nuphar - leaves

2363 f 109

1.6

4.5

297 _+ 27

20.7

7.2

4.3

<0.4

Phragmites Mvriophvlluin Unidentified Mean

<70 2079 f 1 13 857

27.6

67.0

38 TABLE 2.2.5 - Concentration of 137Csin plant material transported by the inflow of Saarisjarvi, the proportion by weight of various components and their relative contribution to the 137Csload. (* approximate values). Deciduous leaves, birch bark and pine rootslbark were below detection limit. Bq kg-‘

% wt

%

13’CS

contrib. Coniferous twigs

6561 f 154

*

2.9

43.9

Coniferous needles

519

37

11.8

14.1

Deciduous twigs

396 f 9

5.6

5.1

Grass

302 f 11

2.8

2.0

Unidentified*

205

76.9

34.8

Mean

433

TABLE 2.2.6 - Concentration of I3’Cs in plant material transported by the outflow of Saarisjarvi, the proportion by weight of various components and their relative contribution to the 137Csload. Bq kg-’

% wt

% ‘37cs

contrib. Equisetum

555 f 14

81.0

Unidentified

640 f 56

83.1

Mean

592

16.9

19.0

The total transport of CPOM 137Cs,a function of discharge and the concentration of 137Cs in the constituent plant materials, varied widely among the investigated streams (Table 2.2.7). Transport was clearly greatest in the streams flowing into 0vre Heimdalsvatn, although there was a major transport of 137Csin Joutsjoki, the inflow of Saarisjarvi. In fact, the CPOM concentration recorded in Joutsjoki, which drains a forested area, was the highest recorded in this study. This

is explained by the lack of lakes in the catchment of Saarisjarvi’s major inflow. In catchments with several lakes much of the CPOM will sediment in upstream lakes. Apart from Hillesjon, the transport into the lakes was considerably higher than the outflow transport. This means that 137Cs accumulates in these lakes, either entering the aquatic food chain or deposited as sediments, where it is exposed to microorganisms and can subsequently enter the food chain. Thus the CPOM

39 TABLE 2.2.7. Transport of coarse particulate organic matter (CPOM) and associated 137Csin Nordic lake inflows 0 and outflowing streams (0)during spring 1991. Mean discharge

I sec-l

CPOM

I3'Cs

g dw 24hl

Bq kg-'

Transport Bq 24h-I ~~~

Heimdalsvatn

285

4050

2703

10950

29

4005

6052

24240

1370

13

8881

120

Angland (I)

20

I02

627

64

Hille syd (I)

21

83

579

48

200

186

857

160

Joutsjoki (I)

160

12250

433

5300

outflow

270

77

592

45

Brurskardbk. (I) Lektorbekken (I) Hingrala (0) HiUesion

outflow Saarisiarvi

represents a potential reservoir for bioaccumulation of radionuclides such as caesium. Although the transport of 137Csin CPOM is a factor

less than that transported in the water phase, CPOM

is rapidly taken up in lake food chains and thus represents an important source for uptake. Hillesjon is an exception to the pattern of high CPOM concentrations in inflowing streams compared to the outflow, with sediment resuspension resulting in a higher CPOM 137Cs concentration in the outflow than in the inflowing streams, although the total input of CPOM from all streams was greater than the output.

The distribution of '37Cs in the various size fractions of the water phase in 0vre Heimdalsvatn has been reported earlier (Brittain et al., 1992; Salbu et al., 1992). About 80% of the 137Cstransported into 0vre Heimdalsvatn in the water phase is associated with the high molecular weight fraction, while this is reduced to about 50% in the lake outflow, due to the sedimentation of larger particles (Fig. 2.2.5.). Most of the '37Csin the water phase in the inflow stream of Hillesjon, Angland, was in the low molecular weight fraction (Fig. 2.2.6). This was also the case with outflow, although there was a clear increase in the high molecular weight fraction. This may be explained by the resuspension of sediment material with high "'Cs

concentrations. It has been shown in several

40

Figure 2.2.4. Transport of I3’Cs in plant material (CPOM)in and out of three Nordic lakes. Note logarithmic scale.

Figure 2.2.5.13’Cs budget for the water phase in 0vre Heimdalsvatn, Norway, spring 1989. Modified from Salbu et al. (1992).

41

Fig. 2.2.6. 137Csbudget for the water phase in Hillesjon, Sweden, spring 1991.

Fig. 2.2.7. 137Csbudget for the water phase in Saarisjki, Finland, spring 1991.

42 cases that much of the primary load of Chernobyl fallout caesium is now located in lake sediments (Sundblad et al., 1991; Blakar et al., 1992). The '37Csin the water phase transported by the inflow of Saarisjmi is more evenly dispersed among the fractions (Fig. 2.2.7). In the outflow, however, it appears that much of the 137Csassociated with the high molecular weight fraction has sedimented in the lake as 90% of the 137Csin the outflow stream was in the low molecular weight fraction. The low molecular weight fraction (LMF) has the potential to play an important role in radiocaesium mobility in freshwater systems and should be taken into account in model simulations of such ecosystems. However, the two colloidal fractions, CF, and CF,, may be the more "bioavailable" fractions in the water phase, as they are removed during passage through the lake (Fig. 2.3.6). Under the conditions studied the low molecular weight fraction (LMF) shows little or no retention and is even produced in Hillesjon. Thus the LMF-fraction appears to play a minor role in terms of bioavailability in the lakes studied. In both 0vre Heimdalsvatn and Saarisjarvi there is a net loss of 137Csfrom the water phase. In fact, in both lakes only 45% of the inflowing 137Csin the water phase is retained in the lake, the remaining 55 % being transported downstream. On account of resuspension in Hillesjon the relationship between inflow and outflow is entirely different (Fig. 2.2.6). As with the other lakes there is a net loss of 137Cs,but the similarity ends there. In Hillesjon over ten times more 137Cs

flows out of the lake than flows in. Annual estimates of 137Cslosses in relation to initial deposition for Hillesjon and its catchment have shown a total catchment loss of only 8% during the period 1986-1991 compared to a total loss from the lake of 30% (Sundblad et al., 1991). This clearly demonstrates the role of sediment resuspension and high macrophyte production, and indicates that Hillesjon and similar lakes are potentially important sources of '37Cs for downstream areas, including the coastal areas of the Baltic Sea (Evans, 1991). The fieldwork for this study was camed out during the spring snowmelt period, at a time when discharge and runoff are at a maximum. Care must therefore be taken in extrapolating the results to the whole year, although the differences between streams and lakes are likely to be similar during periods of low discharge. However, groundwater inputs, especially during the winter months, may affect 137Csconcentrations in lake inflows. Studies during the winter in 0vre Heimdalsvatn demonstrated a seasonal shift in the proportion of 137Csin the low molecular weight fraction (Brittain et al., 1992) and a similar shift is likely to occur in the other Nordic lakes. However, in contrast to "Sr, there was little difference in overall retention of 137Csbetween spring and winter.

43

CONCLUSIONS 1. The Nordic lakes, 0vre Heimdalsvatn, Hillesjon and Saarisjarvi differ markedly in several aspects and provide insight into the factors determining radionuclide transport in a range of lake ecosystems. 2. Transport of 137Csin plant material (Coarse Particulate Organic Matter, CPOM) was about 17 times greater to 0vre Heimdalsvatn than Saarisjhi, although over 99% of the inflow CPOM was retained in both lakes.

3 . CPOM inflows to Hillesjon were an order of magnitude lower than to Saarisjarvi and the net retention was only 71 %, on account of the outflow of autochthonous production. 4. The lakes differed in the activity of 137Csin the various molecular weight fractions in the water

phase. This was a function of catchment processes, resuspension and biological activity in the lakes.

5. In 0vre Heimdalsvatn and Saarisjarvi 45 % of the I3’Cs in the water phase was retained in the lake, while in Hillesjon ten times more 137Csflowed out than flowed in, due to resuspension of ‘37Cs-ri~h sediments.

ACKNOWLEDGEMENTS Support for the project has been given by the authors’ respective institutions. The studies in 0vre Heimdalen were financed by the Norwegian Agricultural Research Council’s Programme for Research on Radioactive Fallout. Anna N ~ r e nkindly assisted with the radioisotope analysis. We are also grateful for the assistance given by Dr. Jukka Ruuhijarvi and his staff at the Evo State Fisheries and Aquaculture Research Station during our fieldwork in Finland. Dr. Ruuhijarvi and Mr. Marko Jawinen kindly supplied chemical data for Saarisjarvi.

REFERENCES Arvola, L., T.-R. Metsala, A. Simila, and M. Rask. 1990. Seasonal fluctuations in the chemistry of small forest lakes in southern Finland with special reference to acidity. Aqua Fenn. 20: 7179. along inflow streams to 0vre Heimdalsvatn, Vigi, Aunan, K. 1986. Leaf production in &I& Oppland, and transport of allochthonous material to the lake. Thesis, University of Oslo. 67pp. Blakar, LA., D. Hongve, and 0. Njistad. 1992. Chernobyl cesium in the sediments of lake Hqsj~ien,Central Norway. J . Environ. Radioact. 17: 49-58. Brittain, J.E., H.E. Bj~rnstad,B. Salbu and D.H. Oughton. 1992. Winter transport of Chernobyl radionuclides from a montane catchment to an ice-covered lake. Analyst 117: 515-519. Cunie, L.A. 1968. Limits for quantitative detection and quantitative determination. Application to radiochemistry. Anal. Chem 40: 586-593. Evans, S . 1991. Impacts of the Chernobyl fallout in the Baltic Sea ecosystem. In: Moberg, L. (Editor). The Chernobyl Fallout in Sweden. Swedish Radiation Protection Institute, Stockholm, Sweden. pp. 109-127.

44 Haugen, L.E., H.E. Bj~mstadand J. E. Brittain. In press. Variation in deposition of radiocesium within a small watershed on natural pasture in the mountainous region of southern Norway. J Environ. Radioact. Hilton, J. ,F.R. Livers, P. Spezzano and D.R.T. Leonard. 1993. Retention of radioactive caesium by different soils in the catchment of a small lake. Sci. Tot. Environ. 129: 253-266. Hongve, D. 1987. A revised procedure for discharge measurement by means of the salt dilution method. Hydrol Proc. I : 267-270. Larsson, P., J.E. Brittain, L. Lien, A. Ldehammer and K. Tangen. 1978. The lake ecosystem of 0vre Heimdalsvatn. Holarct. Ecol. 1: 304-320. Larsson, P. and K. Tangen. 1975. The input and significance of particulate terrestrial organic carbon in a subalpine freshwater ecosystem, In: Ecological Studies, Analysis and Synthesis, 6, Wielgolaski, F.E. (editor). pp. 351-359. NKS. 1991. Radioecology in Nordic Limnic Systems - present knowledge and future prospects. Rep. Naturvhrdsverket 49. Rask, M. 1991. Is0 Vakjirvi research: an introduction to a multidisciplinary lake liming study. Finn. Fish. Res. 12: 25-34. RosCn, K. 1991. Effects of potassium fertilization on caesium transfer to grass, barley and vegetables after Chernobyl. In: Moberg, L. (editor) The Chernobyl Fallout in Sweden. Swedish Radiation Protection Institute, Stockholm, Sweden. pp. 305-322. Salbu, B., H.E. Bjumstad, and J.E. Bnttain. 1992. Fractionation of cesium isotopes and 90Srin snowmelt run-off and lake waters from a contaminated Norwegian mountain catchment. J . Radioanal. NucE. Chenz. 156: 7-20. Saxen, R. 1990. Radioactivity of surface waters and freshwater fish in Finland in 1987. Suppl. 3 Ann. Rep. STLJK-A77. 59 pp. Sundblad, B., S . Evans and U. Bergstrom. 1989. The turnover of Chernobyl fallout within two catchment areas - Hillesjon and Salgsjon - in the Gavle area, Sweden. Rep.Studsvik NP-89/5 1. 45 PP. Sundblad, B., U. Bergstrom and S . Evans. 1991. Long term transfer of fallout nuclides from the temestnal to the aquatic enviroment - evaluation of ecological models. In: Moberg, L. (editor) The Chemobyl Fallout in Sweden. Swedish Radiation Protection Institute, Stockholm, Sweden. pp. 207-238. Vik, R. (editor) 1978. The lake 0vre Heimdalsvatn - a subalpine freshwater ecosystem. Holarct. Ecol. 1: 81-320.

45

2.3. THE DISTRIBUTION AND CHARACTERIZATION OF 137Cs IN LAKE SEDIMENTS.

A. BROBERG Institute of Limnology, Uppsala University, Norbyvagen 20, S-752 36 Uppsala, Sweden.

SUMMARY

The 137Cs entering a lake is associated with particles and quite rapidly transferred to the sediments because of different mechanisms. The resulting horizontal and vertical distribution of 137Cs in lake sediment, together with physical and chemical associations, is very important for the bioavailability and circulation of *37Csin the ecosystem. In 1987 *37Cswas bound, to a great extent, to chemically labile fractions, but since then the isotope has, in some lakes, been transferred to more stable associations in the sediment. The highest concentration of 1 3 7 0 was found in the smallest particle fraction (< 3-4 pm), but this fraction constitutes less than 1 % of the sediment material and, hence, the largest amount of 1 3 7 0 was found in coarser fractions with a lower tendency to resuspension. In concave, flat lakes 137Cs was uniformly distributed horizontally in the sediment, whereas in convex lakes with steep sides 137'2s was focussed, for example through resuspension, on the deepest parts of the lake. The content of l37Cs in the uppermost sediment layer is crucial for determining the flow of Cs from the sediment to other parts of the ecosystem. A deep vertical distribution of deposited 137Cs was found in lakes or subareas in lakes with high rates of sediment accumulation in combination with strong mixing (bioturbation or resuspension).When mixing was missing as in the deep areas of certain lakes, the pulse of 137Cs was very quickly covered, at high sedimentation rates, by new material with less or no 137Cs.The combined effect of diffusion, mechanical mixing and bioturbation gives a stochastic depth distribution of 137Cs,and the pulse of Cs is mixed with a larger volume of sediment whereby the concentration in the exposed surface layer is lower. The depth of maximum Cs-activity in a sediment core is normally greater in deeper areas of a lake because of the increased sediment accumulation. Processes which can be important for the transport of 137Cs from sediment to water are diffusion, resuspension, bioturbation and biological uptake. The degree of resuspension was studied, by using sediment traps, in some shallow lakes in central Sweden. A very strong tendency to resuspension was found in these lakes with maximum depths between 4 and 11 m. Resuspension may transport 137Cs to deeper areas, which probably gives a faster burial of the isotope. On the other hand, the process will increase the availability of Cs to biota , which will delay the recovery of Cs-contaminated shallow lakes.

46

SEDIMENTATION PROCESSES, SEDIMENTATION RATES AND SEDIMENTING MATERIAL The radionuclide 137Cs has a strong particle affinity and the transfer from water column to sediments may be the result of several possible mechanisms (Kansanen et al., 1991); - sedimentation of insoluble detrital particles from nuclear fuel (aerosols) - adsorption or precipitation onto inorganic compounds, e.g. carbonates, clays or oxyhydroxides - sedimentation with humic matter - sedimentation with autochthonous organic matter either after assimilation or absorbtion on cell surfaces - direct uptake through assimilation by periphyton or other biota on surface sediments - direct adsorption on the surface sediments. According to Hesslein et al. (1980) direct binding to the sediment is of minor importance and the dominant sedimentation process for 137Csis coupled to particulate fractions (Santschi et al., 1988). The chemical properties of the water and the biological characteristics of the lake system, e.g. productivity, together with the physical and chemical state of the radionuclide itself will subsequently regulate the magnitude of each possible mechanism, while the settling velocity of the particle will regulate the removal rate of the radionuclides from the water mass. These sedimentation processes are relatively rapid and in 1990 more than 99 % of the radioactive Cs in a contaminated lake occurred in the sediment (Broberg & Andersson, 1991). The production of organic material in the lake and the inflow of material from outside are of vital importance for the fate of Cs within the lake ecosystem. High production will increase the sedimentation rate and dilute the Cs-containing sediment layer with uncontaminated particles or shorten the time for its burial below fresh sediment. Table 2.3.1 presents the distribution of 137Cs in two lakes during three different years. Both lakes have the same surface area (0.6 kmz) and are clearly humic but Flatsjon has higher levels of P and hence a higher production of organic material. The quotient between catchment area and lake area is 34 and 14 for Siggeforasjon and Flatsjon, respectively, indicating a very clear initial influence of the catchment area on Siggeforasjon. In 1990 a larger proportion of 137Cs was found in suspended matter in Flatsjon, owing to its shallowness (maximum depth 4 m) and high production, than was found in Siggeforasjon (maximum depth 11 m). During the first year after Chemobyl the amount of 137Cs deposited in lake catchment areas was tightly bound to the soil layer or transported away to the water systems and the annual inputs to lakes are today, in most cases, very low ( 0.01 - 0.2 % of original deposition). By using sediment traps the load of particle bound 137Cs in lakes was measured in seven lakes in J h t l a n d , Northern Sweden, during the period 1988-1990 (Hammaret al., 1991). The traps were placed about 1 m above the sediment for 50 - 100 days but the results can be used as an approximate measure of the annual load of 137Cs on the sediments. There are three main sources for the material in the traps, particles in inflowing waters, associations with autochthonous production and resuspended bottom sediment. The average amount (kl3q) of 137Cs in the traps was quite high and amounted to 0.25, 0.29 and 0.47 % of the deposition in the different catchment areas for 1988, 1989 and 1990,

41 respectively. The concentration of 137Cs(Bq g dw-1) in the traps was fairly constant and in some of the lakes higher than in surface sediments, indicating an inflow of 137Csfrom the catchment area (Robbins et al., 1990) or an accumulation in biological material.

Table 2.3.1 Main distribution of 137Csin lakes Siggeforasjon and L. Flatsjon 1986, 1988 and 1990 (kBq * 106 and %). 1986

%

1988

%

1990

%

weforasion Sediment Susp. matter Plankton Macrophytes

21.25 2.49 .06 .42

87.6 0.2 .26 .7

26.33 .70 .01 .10

96.9 2.6 .02 .36

30.45 .16 .o 1 .05

99.3 .5 .02 .15

Flats-ion Sediment Susp. matter Plankton Macrophytes

42.47 1.20 .02 .04

96.9 2.7 .05 .08

45.02

98.5 1.3 .02 .08

47.94 .35 .o 1 .02

99.1 .7 .02 .05

.58 .01 .03

In a regional study (Htikansson et al., 1989) the concentration of 137Cs in bottom traps in 1988 and 1989 was on average 15 % of the 1986 values. The variation was however very considerable (300-22 000 Bqkg dw). In the two lakes mentioned above ( Siggeforasjon and Flatsjon ) sedimentation has been followed since 1987 and the seasonal and annual variations in 137Cs loads can be deduced (Fig. 2.3.1 .). In the deeper Siggeforasjon the content of 137Cs in the traps is slowly decreasing, whereas little has changed in the shallow Flatsjon. Calculated sedimentation rates (using data from the traps) were, in both lakes, 5-10 times higher than expected for that type of lake, indicating that resuspension may be responsible for the delayed settling of 137Cs-containingparticles (see discussion below).

CHEMICAL AND PHYSICAL CHARACTERIZATION OF 137Cs IN THE SEDIMENTS In addition to the covering of 137Csby less contaminated material, the chemical and physical associations of I37Cs in the sediments are very important for its bioavailability and circulation within the ecosystem. Relatively large amounts of I37Cs move between water and sediment and are hence potentially available for uptake and transport in food chains.

48

Chemical fractions For chemical fractionation of the sediments a sequential extraction technique was used (Table 2.3.2). Sediments from four different lakes in GbtriMand/Uppland were analysed following the extraction scheme. Sediment cores were divided into 1 cm slices down to, in most cases, 6 cm sediment depth. The lakes represent a gradient with regard to deposition of Chernobyl 137Cs,bioproduction and potential for resuspension.

160

-

mp20 Q

a

\

$80

W

-

B

I-

2 40 -

Figure 2.3.1. Sedimentation of I37Cs in lakes Siggeforasjon and Flatsjon 1987 - 1992. Sedimenting material collected in traps close to the bottom. Values for sedimentation in Bq/day m*.

The most important associations are 137Cs bound by adsorption to inorganic and organic material, combined with clay, carbonates, hydroxides/oxides (Fe/Mn), sulphides and bound to autochthonous or allochthonous organic matter. For example, a high content of clay results in a 137Cs-fractionwhich is tightly bound to the sediment, and environmental changes have only small effects on the transport of 137Cs (Heit & Miller, 1987). A moderate reduction of pH will mainly increase the transport of adsorbed and carbonate-bound *37Csand decrease the fraction of 137Cs bound to organic complexes. In many stratified, productive lakes or in shallow lakes rich in organic matter there are periods with low concentrations of oxygen during summer and winter respectively. These low levels of oxygen may increase the mobility of 137Csassociated with hydroxides/oxides. A total lack of oxygen can, however, bind 137Csbecause of a transfer of the element to less soluble sulphides and a lower rate of anaerobic mineralization .

49 Table 2.3.2. Chemical fractionation of wet sediment with respect to 137Cs. Extraction solution

T ("C)

Time (h)

Sediment fraction

A. 1 M MgC12, pH=7 B. 1 M NaAc, pH=5

20 20 20

1 2 0.5

Adsorbed Cs Carbonate- bound Cs Cs bound to easily reducible compounds (especially Mn-oxides)

D. 1 M Hydroxyl Ammonium Chloride in 25 % HAc

20/90

130.5

E. 1 % HCI

20

24

F. 0.5 % NaOH

80

24

Cs bound to moderately reducible compounds, for example some Fe(II1)-oxides Cs bound to low molecular organic matter ( for example fulvic acids) Cs bound to high molecular organic matter ( for example humic acids) Cs content of the residual sediment portion

~

C. 0.1 M Hydroxyl Ammonium Chloride in 0.01 M HNO3

G. Residual fraction

~

In 1987 between 40 and 50 % of the 137Cs was found in the available fractions (Figure 2.3.2.). In the following year, 1988, less than 30 % of the total 137Csin Siggeforasjon sediment was found in the easily available fractions, whereas in Flatsjon there was little change (Figure 2.3.3.). In Siggeforasjon 137Cswas transferred to oxidized Fe (D) and especially the residual fraction (G). One main part of this residual fraction may be clay in which 137Csis very easily incorporated in the lattice of clay minerals (Squire & Middleton, 1966). This binding to clay is blocked by the action of organic material (Barber, 1966), which can be one of the reasons for the slower transfer of 137Cs in the very organic sediments of Flatsjon (compare Figures 2.3.2 and 2.3.3). Another difference between the two lake sediments was content of calcium, in Siggeforasjon 4.5 and in Flatsjon about 10 mg g dw-1. In sediment from Flatsjon the binding capacity of carbonates is probably very high and 137Csremained in this fraction. After another 4 years the speciation of '37Cs was about the same for the shallow sediments (4.5 m water depth) in Siggeforasjon (Figure 2.3.4.). In the deepest parts of the lake there was a shift of 137Cs from low molecular weight organics to Fe-oxides. One of the other lakes investigated, RAksjon, which is quite similar to Siggeforasjon regarding trophic level and sediment characteristics, showed a similar distribution of 137Cs (Figure 2.3.4 C).

50

Figure 2.3.2. Distribution of 137Csbetween various chemical fractions (A-G, see table 2.3.2) in 3 sediment layers (0-1, 1-2 and 2-3 cm sediment depth). A: Lake Siggeforasjon; 30 June 1987; sample from 6.5 m depth (max 11 m) B : Lake Flatsjon; 17 Aug. 1987; sample from 3.4 m water depth (max4 m)

51

Figure 2.3.3. Distribution of 137Cs between various chemical fractions (A-G, see table 2.3.2) in 3 sediment layers (0-1, 1-2 and 2-3 cm sediment depth). A: Lake Siggeforasjon; 28 March 1988; taken from 3.5 m depth (max 11 m) B : Lake Flatsjon; 10 Feb. 1989; sample from 3.6 m water depth (rnax 4 m)

53

Figure 2.3.4. Distribution of 137Csbetween various chemical fractions (A-G, see table 2.3.2) in 3 sediment layers (0-1, 1-2 and 2-3 cm sediment depth). A: Lake Siggeforasjon; 4 June 1992; sample from 4.5 m water depth (max 11 m) B: Lake Siggeforasjon; 4 June 1992; sample from 10.5 m depth (max 11 m) C: Lake Rksjon; 4 June 1992; sample from 9.1 m water depth (max 10 m)

There were generally no differences in the distribution pattern of l37Cs for the uppermost 4-5 cm of sediment. This was expected as the uppermost sediment layers in these types of lake are more or less homogenized by different forces, which will be discussed later. Sequential extractions for 137Cs have been performed on soil samples from Norway (Riise et al., 1990) showing that terrestrial 137Cs was distributed in the same way as in lake sediments, that is, the main part was present in the stable fractions. Similar extractions were reported for sediments from some larger lakes in J h t l a n d (Hammar et al., 1991). In these sediments there was no significant change in 137Cs-distributionduring the investigation period 1986 - 1989. The sediments had a rather low content of organic matter but a high proportion of inorganic mineral matter, obviously with a high capacity to bind added 137Cs.

Physical fractions To clarify the tendency of the carrier particles for I37Cs to be resuspended, the grain size distribution was measured for a number of sediments. The different grain sizes were analysed with regard to *37Cs,C, N, P, Fe, Mn and Ca. This type of fractionation was carried out for wet sediment (0-2 cm layer) from Ortrasket in Northern Sweden and from Siggeforasjon, Flatsjon and Riksjon in Central Sweden. Some of the results from the study are summarized in Table 2.3.3. In all sediments the concentration of 137Cswas highest in the smaller fractions probably owing to their high surface/volume ratio, which promotes adsorption processes. This part of 137Cs in the sediment is quite easily transferred to the water by resuspension and will be available to organisms in the water.

54

However, in all sediments, except for the deeper parts of Ortrasket, the larger grain sizes dominated. Thus 137Cs is to a large degree bound to the bottom strata in the lakes. This was especially true for Siggeforasjon where particles more than 35 pm clearly dominated. In Ortriisket a relatively large proportion of 137Cs was found in the smallest fraction probably associated with lowweight organic matter with a high tendency to resuspension. However, the lake is deep and the degree of resuspension in the deeper parts, according to the vertical distribution of 137Csin the sediments, is quite low.

Table 2.3.3. Distribution of grain size, I37Cs (Bq/g dw), and C (mg/g dw) in some freshwater sediments (0-2 cm). ~~

Grain size % in fract.

137~s

csin fract. csin fract.

(w-4

Lake Flatsion (sample 3.5 m depth, max 4.0 m) < 3.8 0.5 29.28

0.15 24.58 9.5 3.8 - 11 2.33 11 -23 18 22.22 4.00 23 - 36 9 17.65 1.59 > 36 63 13.21 8.32 Lake RBksion (sample 9 m depth, max 10 m) <3 0.5 28.47 0.14 3-8 6.5 25.05 1.63 8 - 23 5.0 22.22 1.11 23 - 3 5 65.2 15.67 10.22 135 22.8 12.86 2.93 Lake Si!xeforasjon (sample 6.5 m depth, max 11 m) < 3.8 0.5 11.91 0.06 0.5 10.74 3.8 - 8 0.05 8 - 19 18 9.52 1.71 8.65 19-35 21 1.82 7.0 1 > 35 60 4.20 Lake Ortrasket (sample 45 m depth, max 60 m) < 2.6 10 5.04 0.50 20 4.26 2.6 - 13 0.85 13- 19 20 3.21 0.64 32 2.52 19 - 49 0.81 > 49 18 2.48 0.45

c

(% )

0.9 14.2 24.4 9.7 50.8

119 181 205 223 203

0.9 10.2 6.9 63.8 18.3

129 149 175 167 161

0.8 0.6 21.8 23.2 53.5

74 88 95 93 92

15.4 26.2 19.7 24.9 13.8

140 68 65 68 64

55

DISTRIBUTION OF

137Cs

WITHIN THE SEDIMENTS

The distribution of deposited 137Cswithin the sediments is vital for the future effect of this pool of 137Cs on the lake ecosystem. Distribution includes both the spatial distribution (horizontal and vertical) and the association of l37Cs with different particulate fractions. The associations have already been discussed and the following section will focus on horizontal and vertical distributions and their regulating factors. The horizontal distribution of 137Csin lake sediments is mainly affected by the shape of the basin, which gives various patterns for resuspension and redistribution of the sediment owing to water movements and wave actions. A steeply sloping basin will result in a focusing of 137Cs towards the deeper parts of the lake, whereas the horizontal distribution is more uniform in concave lakes (Figure 2.3.5.). In lakes with vast areas of shallow water, relatively large amounts of 137Cs can be bound to the sediment surface even on exposed sites (Kansanen et al., 1991), partly by physicalkhemical adsorption and partly by binding to the benthic algal community or to macrophytes. This 137Cs in the shallow areas in a lake is, to a greater degree than that in the deeper areas, exposed to the water and therefore very important for the circulation of 137Csin the lake. In lakes with steeply sloping sides this 137Cs may, for example through resuspension, be transported towards the deeper parts of the lake (Kansanen et al., 1991). This transport of 137Csis clearly shown by comparing the distribution in Loppesjon in 1988 (see Figure 2.3.5. B) and 1992 (Figure 2.3.6. A). The deposition of 137Cs in 1986 was 20 kBq/m* and in 1988 the same level was found in most parts of the lake. The focusing factor (activity at a certain locatiodaverage activity in the lake (Eadie & Robbins, 1987)) was very high (1.8) at the deepest point in 1992, which clearly shows that I37Cs was redistributed and transported to accumulation areas in the lake. In the shallow, southern part of the lake the activity was close to the average, indicating that the sediment may be affected by resuspension and bioturbation but it sediments again at the same location . The same distribution pattern was obtained for Riksjon (Figure 2.3.6. B). The depth variation was less pronounced than in Loppesjon but there was still a focusing transport towards the deeper parts, by a focusing factor of 1.6. Conditions for sedimentation vary strongly in different parts of lakes, which results in large variations in 137Cs-contentof the sediment, but also variations in sediment structure and properties. In some lakes there is an increase in sediment radioactivity with increasing water depth (see above), whereas in lakes with complicated bottom topography there may not be any obvious pattern (Hongve et al., 1990). A pattern of horizontal (and vertical) divergence in 137Cs-concentrationis seen in lakes with large inlets and/or strong seasonal fluctuations in water flow. For these reasons, the mechanisms which regulate the exchange of 137Csbetween sediment and water are clearly different in different parts of such lakes. For example, 137Cs can be associated with different particulate fractions or the sediment structure may result in strong variations in resuspension between different areas. The content of 137Cs in the uppermost sediment layer is crucial for determining the flow of 137Cs from the sediment to other parts of the ecosystem. Shortly after the accident added 137Cswas deposited in surface sediment, but after that period, in most lakes, there was a very extensive vertical

56

Figure 2.3.5. Bathymetric maps of lakes Bottentjiirn (A) and Loppesjon (B) (Hudiksvall, central Sweden) showing the variation of area-specific total activities of 137Cs (kBq/mZ) in the sediment two years after the Chernobyl nuclear accident. Contour depths in m (from Meili et al., 1989).

57 redistribution of 137Cs.A deep vertical distribution of 137Cs is possible in lakes with high rates of sediment accumulation in combination with strong mixing processes (e.g. bioturbation, resuspension). When mixing processes are missing as in the deep areas of certain lakes, especially at low oxygen concentration and consequently little or no bioturbation, the pulse of 137Cs will be covered by new material without or with a lower content of 137Cs(Loppesjon, at max depth, 1992, Figure 2.3.7. A). The rate of the covering process depends on the sedimentation rate and on the transport of material to the lake. Lack of mixing presupposes that I37Cs is very strongly bound to particles that give little or no diffusion. If diffusion processes dictate redistribution, 137Cswill be transported both to the water and deeper down into the sediment. Diffusion depends on the concentration gradient between water and sediment and high rates can be maintained through turbulent conditions in the water. The density of the sediment affects the diffusive transport of 137Csto deeper sediment strata, and as the sediment is compacted the diffusion will decrease (Figure 2.3.7. A). The sediments can also be mixed down to a certain, constant depth, which gives uniform distribution of 137Cs within this layer. As new sediment, containing less 137Cs,is mixed with the original, the content of 137Cs decreases continuously. The combined effect of diffusion, mechanical mixing and bioturbation gives a stochastic depth distribution of 137Cs and the pulse of 137Cs will be mixed with a larger volume of sediment (Figure 2.3.7. B) whereby the concentration in the exposed surface layer will be lower. The depth of maximal 137Cs-activityis mainly dependent on sediment accumulation and there is normally a clear difference between deep and shallow areas within a lake (Figure 2.3.7. B,C) (Wass, 1992).

TOTAL INVENTORY OF 137Cs IN LAKE SEDIMENTS The fallout after the Chernobyl accident in April 1986 was deposited in lake sediments to various degrees. The north of Sweden was still covered in snow and much of the fallout was flushed away in spring floods, although new 137Cs entered the lakes from the drainage area (e.g. Malmgren & Jansson, 1991). In Halsingland, central Sweden, the content of l37Cs in small lakes with short turnover time was less than the fallout in the area (Konitzer, 1992). The short turnover times probably decreased the possibilities of 137Cs being bound during the period of snowmelt and high flow, and as no 137Cs-containingmaterial was added from the catchment area the resulting total inventory was low. Around many lakes in central Sweden there was no snow cover and hence no high water flow, resulting in 137Cs inventories in lake sediments very close to the area fallout (e.g. Broberg & Anderson, 1991). The total amount of 1 3 7 0 in most lakes has not changed since 1986. Some of the small lakes with short turnover times in Hgsingland had the same inventory in 1992 (Meili et al., 1989) and the same was reported for lakes in GastrikIandRippland (Broberg & Anderson, 1991). The content of 137Cs decreased in surface sediments (0-2 cm) in some lakes in Jtirntland, northern Sweden (Hammar et al., 1991). This was, however, just an effect of sediment redistribution and burial of 137Cs- contaminated layers, which were discussed earlier.

B kBq/m*

-

0

100m

1t

0

20

0

30

-

0

Figure 2.3.6. Bathymetric maps of lakes Loppesjiin (A) (Hudiksvall, central Sweden) and RPIrsjGn (B}(IJppland, cenlral Swedcn} showing the variation of area-specific total activities of 137Cs (kBqlm2) in the sediment six years after the Chernobyl nuclear accident. Contour depths in m ((A) from Konitzer, IW? and (B) cronom Wass, 1991).

500 m

59

Figure 2.3.7. Vertical distribution of 13’Cs in sediment from lakes bppesjon (A) (Hudiksvall, central Sweden) and Mksjon (B and C) (Uppland, central Sweden). Values in Bq/g dw. A Sediment core from maximal depth (14 m) in Loppesjbn, B Sediment core from maximal depth (10.1 m) in Mksjon and C: Sediment core from 5.5 m depth in Mksjon. (A from Konitzer, 1992 and B and C from Wass, 1992).

60

TRANSPORT OF 137Cs FROM SEDIMENT The processes which can be important for the transport of 137Csfrom sediment to water are: 1. diffusion

2. resuspension 3. bioturbation by benthic animals and fish 4. uptake by bacteria, benthic fauna, benthic algae and macrophytes.

The diffusion rate is dependent on the concentration gradient between water and sediment. If this gradient is maintained at a high level by, for example, turbulent conditions in lake water and a constant release of 137Cs to the pore water, diffusion can remain at a high rate and will add 137Cs to the water phase. The concentration of 137Cs in pore water and hence the diffusion is, according to Heit & Miller (1987), influenced by the presence of clay and organic matter, oxygen content and pH. They found that a low content of clay may result in mobilization of 137Csand this mobilization was also affected by organic ligands, which increase the binding capacity in combination with clay. Under anoxic conditions the production of N&+ in the sediment may mobilize I37Cs from clay minerals to pore water through ion exchange (Pardue et al., 1989 and Comans et al., 1989). The level of resusuension is dependent on lake-specific properties such as depth, bottom configuration and wind exposure, but also on sediment structure and its physical properties (e. g. density). In some shallow lakes such as Hillesjon (Sundblad et al., 1990) resuspension takes place over the entire lake. By this process 137Cs is transported out of the lake (Bjornstad et al., 1994). In many lakes, especially in deeper lakes such as Paijanne (Kansanen et al., 1991), resuspension results in a transport of a part of the 137Cs bound in shallow areas towards deeper areas. This redistribution of 137Csto deeper areas may result in a more rapid burial of 137Cs in undisturbed sediment, which was discussed in relation to the horizontal distribution of 137Cs. On the other hand, the 137Cs-contaminated sediment may meet anoxic conditions, which at first can increase the diffusion rate (Pardue et al., 1989). Resuspension also means that the level of 137Cs is maintained at a high level in organisms in shallow lakes, partly owing to a direct uptake of resuspended 137Cs and partly owing to an increased supply of oxygen to the particles, which increases mineralization and the bioavailability of sediment-bound 137Cs. In allochthonous organic matter the microbial activity and subsequently the rate of mineralization is lower than in material produced within the lake. For that reason 137Cs associated to this fraction is quite stable. However, this type of sediment has a high content of dissolved organic material, capable of binding 1 3 7 0 , in pore water, which may result in a release of 137Cs from the sediment on disturbances of the sediment surface (e.g. resuspension). The degree of resuspension was studied in some 137Cs-contaminatedlakes in central Sweden and selected results are presented in table 2.3.4. Both lakes are rather low-productive forest lakes with a maximum depth of about 10 m, but Riksjon is larger than Siggeforasjon, 1.2 and 0.73 km2 respectively. In both lakes the degree of resuspension was highest during the circulation period in May, but during summer with higher production and stratified conditions the resuspended material

61

Table 2.3.4. Lakes Siggeforasjon and Rksjon :Sedimentation rates (mg dw/day cm2) and content of 137Cs(Bq/g dw), C and N (mg/g dw) in sediment traps at various depths. Depth (m)

Sedimentation

137Cs

C

N

weeforasion (May 1992) 2.5

0.15

5.17

216

29

5

0.17

5.91

163

16

8

0.23

5.85

145

13

2.5

0.17

4.17

193

30

Sigeeforasion (July 1992) 5

0.17

5.43

167

23

8

0.18

6.18

152

21

2.5

0.11

4.00

358

40

5

0.09

5.80

285

25

8

0.09

4.49

27 1

22

2.5

0.12

1.79

35 1

54

5

0.13

2.01

34 1

49

8

0.13

4.02

325

43

RBksion (May 1992)

RBksion (July 1992)

sedimented. This settling was slow in Siggeforasjon and the I37Cs-content of the traps remained high. Bioturbation can have an effect similar to that of resuspension. When the sediment is mixed by the activity of benthic fauna or fish, the magnitude of the transport to or from the sediment often increases. In most cases oxygen is added to the sediment, which gives increased mineralization and a redistribution or increased outflow of I37Cs. The activity also results in a breakdown of the chemical or physical barrier to transport which is often found at the sediment surface (Hkansson & Jansson, 1983), and the flow of 137Cs from pore water may increase. Another very important effect of bioturbation is that 137Cs deposited on the sediment surface is distributed in a much larger volume of sediment, which will lower the concentration of 137Cs in the uppermost sediment layer. The uDtake of 137Cs in biota affects the transport from the sediment. As mentioned earlier, benthic algae and macrophytes in the littoral zone bound a fraction of deposited I37C.s during 1986 (Kansanen et al., 1991) and this binding gave a longer exposure of a large fraction of the pulse in shallow waters. This 137Cs was transported into the food chains by herbivores, although it was also released to the water on decomposition during the autumn of 1986 and in the subsequent years. Another transport mechanism via the biota is the uptake of sediment-bound 137Csby rooted vegetation and its excretion into the water or release into the water during decomposition.

62

REFERENCES

Barber, D. A. 1966. Influence of soil organic matter on the entry of cesium-137 into plants. Nature 204 pp. 1326. Bjornstad, H. E. ; Brittain, J. E. ; SaxBn, R. & Sundblad, B. 1994. The Characterization of Radiocaesium Transport and Retention in Nordic Lakes. (This volume chapter 2.3.) Broberg, A. & Andersson, E. 1991. Distribution and circulation of Cs-137 in lake ecosystems. In "The Chernobyl Fallout in Sweden", Ed. L. Moberg, The Swedish Radiation Protection Institute, pp. 151-175. Comans, R. N. ; Middelburg, J. J. ; Zonderhuis, J. ; Woittiez, R. W. ; DeLange, G. J. ; Das, H. A.& Van Der Weijden, C. H. 1989. Mobilization of radiocaesium in pore water of lake sediments. Nature 339 pp.367-369. Eadie, B.J. & Robbins, J. A. 1987. The Role of Particulate Matter in the Movements of Contaminants in Great Lakes. In "Sources and Fates of Aquatic Pollutants", Eds. R. A. Hites & S. J. Eisenreich, American Chemical Society, Washington D. C., Advances in Chemistry Series 216 pp. 3 19-364. Hammar, J. ; Notter, M. & Neumann, G. 1991. Cesium in Arctic Char lakes - Effects of the Chernobyl Accident. Information from Institute of Freshwater Research of the Swedish National Board of Fisheries 3 (in Swedish with English summary). Heit, M. & Miller, K. M. 1987. Cesium-I37 sediment depth profiles and inventories in Adirondack Lake sediments. Biogeochemistry 3 pp. 243-265. Hesslein, R. H.; Broecker, W. S. & Schindler, D. W. 1980. Fates of metal radiotracers added to a whole lake: sediment-water interactions. Can. J. Fish. Aquat.,,Sci. 37 pp, 378-386. Hongve, D. ; Blakar, I. & Brittain, J. E. 1990. Sediment studies in Ovre Heimdalsvatn. Inf. Statens Fagtjeneste for Lantbruket 28 pp. 173- 174. HAkansson, L. ; Andersson, T. & Nilsson, A. 1989. Caesium-137 in perch in Swedish lakes after Chernobyl - present situation, relationships and trends. Envir. Poll. 58 pp. 195-212. HAkansson, L. & Jansson, M. 1983. Principles of lake sedimentology. Springer Verlag, Berlin Heidelberg, 316 pp. Kansanen, P. H. ; Jaakola, T. ; Kulmala, S. & Suutarinen, R. 1991. Sedimentation and distribution of gamma-emitting radionuclides in bottom sediments of southern Lake Piiijiinne, Finland, after the Chernobyl accident. Hydrobiologia 222 pp. 121- 140. Konitzer, K. 1992. The importance of sediment transport for the distribution and availability of Cs137 in a forest lake in middle Sweden. Report, Institute of Limnology, Uppsala University, Uppsala (in Swedish). Malmgren, L. & Jansson, M. 1991. Contamination of freshwater and estaurine environments in northern Sweden by I37Cs from Chernobyl accident. Report, Department of Physical Geography, University of UmeB, Sweden. Meili, M. ; Rudebeck, A. ; Brewer, A. & Howard, J. 1989. Cs-137 in Swedish forest lake sediments, 2 and 3 years after Chernobyl. In "The Radioecology of Natural and Artificial Radionuclides", Ed. W. Feldt, Verlag TUV Rheinland GmbH, Koln, Germany pp. 306-3 11. Pardue, J. H. ; DeLaune, R. D. ; Patrick, W. H., Jr. & Whitcomb, J. H. 1989. Effect of redox potential on fixation of 137Cs in lake sediment. Health Physics 57 (5)pp.781-789. Riise, G. ; Bjornstad, H. E. ; Lien, H. E. ; Oughton, D. H. & Salbu, B. 1990. A study on radionuclide association with soil components using a sequential extraction procedure. J. Radioanal. Nucl. Chem. 142 pp. 531-538. Robbins, J. A. ; Murdoch, A. & Oliver, B. G. 1990. Transport and storage of 137Cs and *1oPb in sediments of Lake St. Clair. Can. J. Fish. Aquat. Sci. 47 pp. 572-587. Santschi, P. H. ; Bollhalder, S . ; Farrenkothen, K. ; Lueck, A. ; Zingg, S. & Sturm, M. 1988. Chemobyl radionuclides in the environment: Tracers for the tight coupling of atmospheric, terrestrial and aquatic geochemical processes. Environ. Sci. Technol. 22 pp. 510-5 16. Squire, H. M. & Middleton, L. S. 1966. Behavior of Cs-137 in soil and pastures. Pastures. Radiat. Bot 6 pp. 49. Sundblad, B. ; Evans, S. & Lampe, S. 1990. Radioecological observations in 1989-90 within the catchments of Lake Hillesjon and Salgsjon. Studsvik Report StudsvikRVS -90/145, Studsvik Nuclear, Nykoping, Sweden. Wass, E. 1992. Distribution of Cs-137 in sediment from Lake RAksjon, a forest lake in middle Sweden, six years after the Chernobyl accident. Report, Institute of Limnology, Uppsala University, Uppsala (in Swedish)

63

2.4. TRANSPORT OF I3’Cs IN LARGE FINNISH DRAINAGE BASINS

RITVA SAXEN

Finnish Centre for Radiation and Nuclear Safety, Laboratory for Foodchain Research, P.O.Box 268, 00101 Helsinki, Finland

SUMMARY Area-specific transfer parameters are useful in predicting internal radiation doses incurred via the consumption of fish or the drinking of fresh water, as well as for the estimation of countermeasures which might be needed for an accidental release of radionuclides. The results obtained from the countrywide studies of the Finnish Centre for Radiation and Nuclear Safety, carried out in the aquatic environment after 1986, were used to compare the transfer of 137Csin seven large drainage areas in Finland. During the first six months after the accident at Chernobyl, the correlation between area concentrations of ‘37Cs in water and area deposition of 137Cswas linear, as also the correlation between the median contents of 137Csin fish in the drainage basins concerned and the area deposition of 137Cs. In subsequent years, temporal changes of 137Csin water were more profoundly affected by the character of the catchment area. Though diffcrences in the amount of runoff were taken into account, differences in transfer of I3’Cs from deposition to water still existed between the drainage areas. Annual transfer factors from deposition to water were highest in the drainage basin characterized by large bog areas and lowest in the drainage basin characterized by large clay areas. In coastal drainage areas with a low percentage of lakes, transfer to water was somewhat lower than in inland drainage basins with higher lake percentages. The decrease of 137Csin watcr had at least two components in all the drainage areas studied. At first the concentrations decreased rapidly with an effective ecological halflife of 0.15 - 0.36 years, except in the drainage area with abundant bogs in the catchment where the decrease was remarkably slower with a halflife of 0.94 years. The slower components of the decrease began at different times in different areas. The halflives of the next phase varied from 1.2 to 2.8 years. The transfer of 137Csfrom deposition to fish was highest in the second or third year after the accident in the different drainage areas. The annual maximum transfer from deposition to predatory fish in the area where it was highest was about 0.4 m%g, whereas in the other areas it was about 0.2 m2/kg. The observcd ecological half-life of 13’Cs in predatory fish after the beginning of the decline was 0.7 years in the drainage area differing most from the others, whereas in the other areas it varied from 3 to 5 years. The 137Csdeposited in the drainage areas in 1986 was almost totally retained in the areas and only a little, 1 - 2%, was removed with the rivers flowing from the drainage areas to the sea for five years after the deposition.

64

INTRODUCTION Comparisons were made of the transfer of 137Csfrom Chernobyl deposition to water and fish in seven large drainage basins in Finland. The study was based on the results of the long-term surveillance programme carried out in the aquatic environment by the Finnish Centre for Radiation and Nuclear Safety (SaxCn & Aaltonen, 1987, SaxCn & Rantavaara, 1987, SaxCn, 1990, S a x h & Koskelainen, 1992). Lake-specific parameters were determined already during the fallout period and after the Chemobyl accident in the Nordic countries, but area-specific transfer parameters were not determined in the fallout period and even now are only established in smaller areas (Hammar et al., 1991, Andersson et al., 1990, HBkanson et al., 1990, Salbu et al., 1992). Area-specific values for transfer parameters are needed for several purposes; e.g. for predicting radiation doses after a deposition situation in different areas, and for estimation of the countermeasures that might be needed to keep radiation doses to people as low as possible, according to radiation protection principles.

In order to make areal comparisons of the transfer and behaviour of radionuclides, it is necessary to have representative experimental data from the area. Besides the consideration of radioecological factors in sampling, an adequate number of samples from each area is required.

The two paths by which radionuclides reach watersheds are direct deposition to the water and runoff from the surrounding drainage area. The longer the time lapse after deposition, the more decisive will be the character of the drainage basin for the transfer and transport of radionuclides. The existence of differences between areas regarding the long-term behaviour of radionuclides in water and fish will be apparent when considering a representative set of results for each area after they have been related to the average deposition.

MATERIAL AND METHODS Material for this study was obtained from the countrywide programmes for monitoring surface water and fish carried out after 1986 by the Finnish Centre for Radiation and Nuclear Safety (Arvela et al., 1990, Saxin & Aaltonen, 1987, Saxin & Rantavaara, 1987, SaxCn, 1990, Sax& & Koskelainen, 1992). Figure 2.4.1 shows the drainage basins studied and the sampling stations for surface water. Water samples were taken at four different stages of the hydrological cycle: in March, when the lakes are frozen, in May, just after the ice melts and during the spring water turnover period, in August, before the autumn turnover of water, and in October, after the autumn water turnover period and before the ice cover. Sampling of fish was focussed on the summer period, which is the most important fishing season. The number of fish samples analysed annually in the drainage

65

the most important fishing season. The number of fish samples analysed annually in the drainage areas studied is given in Table 2.4.1. The average municipal deposition values used in calculations were obtained from Arvela et al., 1990.

The drainage basins studied and surface-water sampling stations.

Fig. 2.4.1.

TABLE 2.4.1. Numbers of fish samples analysed annually in the drainage areas studied. Drainage area

1 2 3

4 5 6

7

Year 1986

1987

1988

1989

1990

1991

21 7 55 151 217 18 44

32 52 96 577 482 157 57

42 10 100 253 359 32 93

41 15 103 352 303 12 64

31 13 88 222 298 6 13

34 10 90 201 317 4 10

66

Description of the drainage areas Drainage areas 1, 2 and 6 mainly have rivers discharging into the Gulf of Finland, Archipelago Sea or the Gulf of Bothnia, and only a few larger lakes. Extensive agriculture is typical of these areas. Drainage area 3 discharges its waters via the river Vuoksi into lake Ladoga. The large area of forest and small arable area is typical here.

Only a small portion of the lower parts of drainage area 4 is rich in clay, the greater part being till. A typical feature of this drainage area is the scarcity of bogs. About 80% of the area of land is forest and less than 20% arable.

Fig. 2.4.2. Clay and silt areas in Finland.

Fig. 2.4.3. Regions where over 30% of the land area is bog.

A characteristic feature of the southcrn parts of drainage area 5 is the high proportion of clay in the land areas (Suomi Finland Uusi yleiskarttakirja, 1977) (Fig. 2.4.2). The amount of suspended solids in the river Kokemaenjoki, flowing from this area to the sea, is much higher than in the outlet rivers of the other basin areas studied. Bogs are only found in the northern parts of the catchment area. Forest accounts for > 70% of the land area and arable land for 25% (Forest Res. Lnst., 1980).

In drainage area 7 there is much land consisting of more than 50% bog and almost all the remaining area includes 30-50% bogs. There are virtually no clay areas (Suomi Finland Uusi yleiskarttakirja 1977) (Figs. 2.4.2 and 2.4.3). Forest accounts for about 80% and arable land for 40%.

67

TABLE 2.4.2. Figures characterizing the drainage basins. A = area (km?, MQ = mean discharge (m3/s), R = runoff (l/s4un2), L% = lake percentage

Drainage basin

A km2

11075 7135 61560 37235 27100 33215 22572

m3/s

110 80 550 290 210 370 250

R

9.4 8.1 8.9 7.8 7.9 8.2 11.1

L %

4.6 2.6 19.9 17.2 10.6 3.1 11.3

Calculations The average deposition D in the seven large drainage basins was calculated using the average

municipal deposition values of 137Csfor 1986 (Arvela et al., 1990) and weighting them with the areas of the municipalities within the drainage basin concerned. The values for 1986 were corrected annually for radioactive decay to obtain the values for subsequent years. Average area concentrations of 137Csin water, C,, were calculated as an arithmetic mean of the results from different sampling points in the drainage basin. Average area concentrations of K in water were calculated by means of gamma-spectrometric measurements of 40Kin the same water samples. Calculations were made of annual average concentrations of 137Cswith variation and medians of 137Cswith quartiles, C,, in three fish types having different nutrition (predators, non-predators and intermediate), as well as in all fish samples, for the seven drainage areas. Effective environmental halflives of *37Cs, Tin,,rf, in water and in fish in different drainage = ln2/k, where yo is the activity basins were calculated using the equations: y = yo x e-kl and Tllzefr

concentration of 137Csat the beginning of a time period, and y is the respective concentration after a time period of t and k is a constant. The time periods were determined by the experimental data.

The transfer of radionuclides in aquatic food chains is often described by transfer factors. Transfer from deposition to water is given by the transfer factor TF, = C, / D, where

C, = the

annual average concentration of 137Csin water (Bq/m3) and D = the average deposition of 137Csin the drainage area (Bq/rn?, corrected annually for the physical decay. The transfer factor from deposition to fish is given by: TF, = C, / D, where C, = the annual average concentration of 137Cs in fish (Bqkg wet weight) in the drainage area and D = the average deposition of 137Csin the same

68 area (Bq/m?, corrected annually for physical decay. Values for TF, and TFf were calculated annually for the drainage areas 1-7. Annual averages, not medians, for

c, were used in the

calculation of TF+, because this gives more a conservative estimate than the use of medians. Calculations were made of the amounts of 137Csremoved during the five years after the Chernobyl accident from the three drainage areas (4, 5 and 7) with the rivers discharging from these into the Baltic Sea. The concentrations of 137Csat the mouths of the rivers (Saxtn & Aaltonen, 1987, Saxen, 1990, Saxin & Koskelainen, 1992), the average flow rate values of the rivers (Atlas of Finland, 1986) and the total amounts of I3’Cs deposited in the drainage area were used in the calculations. The following formula was used: %Removed = 100 x (MQ (m3/y) 2 C, (Bq/m3))/ (Di (Bq/m2> x A (mp), where C, is the annual average concentration of 137Csin the river water, discharging from the drainage basin to the Baltic, in year i, and Di the original total deposition of 13’Cs in the drainage basin, corrected for the physical decay to the year i after 1986. In 1986 (i=l), when concentrations of 137Csin water decreased rapidly, shorter time periods were used in the calculations.

RESULTS IJ7Csdeposited in different drainage basins

The average deposition of 137Cswas highest (34 kBq/m2) in drainage basin 5 and lowest (3.8 q/m? lowest (3.8 kBq/m2) in drainage basin 7 in 1986 (Table 2.4.3.).

TABLE 2.4.3. The average depositions of

Drainage basin

1 2 3 4 5 6 7

137Csin the seven drainage basins studied.

137Cs,kBq/m2 1986

14.1 13.0 6.7 29.9 34.1 16.2 3.8

Transfer of 137Csfrom deposition to water Temporal changes of 137Csin water in different drainage basins after the Chernobyl accident are

69 given in Fig. 2.4.4. The correlation between average concentrations of 137Csin water during six months after the accident and area deposition of 137Cswas linear (Fig. 2.4.5). The 137Cscontents in water also seemed to depend on the K content of the water (Fig. 2.4.6).

BQ/KG

BQ/KG , - OO . O l

0.10

0.01

0.01

JANE6JAN87 JANBE JAN89 JAN90 JAN91JAN92 JAN93 DRAINAGE BASIN A7533 m 4 w5 m7

Fig. 2.4.4.

JAN86 JAN87 JAN88 JAN89 JAN90 JAN91 JAN92 JAN93 DRAINAGE BASIN M A 1

m2 w6

137Csin surface water in different drainage basins in Finland in 1986-1992.

cw

3000

2500

0

10

20

30

40

D kBq/m2 Fig. 2.4.5.

Correlation of I3’Cs in water (Bq/m3) and 137Csdeposited QBq/mZ> in different drainage areas in the first six months after the fallout. C, = 45.5 x D

- 59.2 (Bq/m3), 8=0.9599, p<0.0002

70

Csl37 Cw/D

A 0.01 0

A

A

a

o.ooo~I, 1

Fig. 2.4.6.

,

,

2

,

,

3

,

I

I

4

5

cs137 Cw/O O.lOol

0.0 10

0.010

1987

1388

1989

DRAINAGE BASIN a 3 m4

Fig. 2.4.7.

I

~

6

II 7

Dependence of the area 137Cscontent in water on the respective content of K in water, for 1988 - 1990.

Cs137 cw/o ,7001.o

1986

,

1990

1991

w5 w7

1

1986

1987

1988

DRAINAGEBASIN

1989

1990

1991

M IEEe2 w 6

Annual values a f 137Csin water, normalized with deposition, in seven large drainage areas in Finland in 1986-1991. The areas are shown in Fig. 2.4.1.

The transfer factors for I3’Cs from deposition to water were highest in drainage area 7, next highest in 3 and lowest in 1 and 2 (Fig. 2.4.7). Because runoff is one of the main factors affecting long-term transfer of radionuclidcs from catchment to water, the annual transfer factors, CJD, were normalized with the average values of runoff (R) in each area. The transfer of 137Cs was still most effective in drainage basin 7. For the ‘inland‘drainage areas, it was least in 5. For the coastal

71

drainage basins, area 6 had highest and 1 and 2 somewhat lower but almost equal values of

CJDR

for 137Cs.In coastal drainage areas (1, 2 and 6) transfer of 137Csfrom deposition to water was somewhat less than in 'inland' areas (3, 4, 5 and 7). The correlation of the annual contents of

lf7CS

in water in different drainage areas, normalized with deposition and runoff, with lake percentages (L%) in the areas was quite good during 1987-1991 when drainage area 7 was excluded

(4 =

0.80-0.95), and not so good when drainage area 7 was included (2 = 0.22-0.64) (Table 2.4.4.).

TABLE 2.4.4. Values for slopes (a), constants (b) and correlation coefficients (I"> of the lines between 137Csin water of the different drainage areas, normalized with deposition and runoff, and the lake percentages of the drainage areas, when 1) all areas are included and 2) drainage area 7 is excluded.

1986 1987 1988 1989 1990 1991

18 5.1 3.6 2.6 2.1 1.4

32 8.5 4.3 3.3 2.3 3.0

0.803 0.418 0.539 0.635 0.571 0.466

17 4.1 3.1 2.4 1.8 1.1

31 6.6 3.3 2.8 1.8 2.5

0.813 0.844 0.936 0.915 0.948 0.797

Effective environmental halflife of I3'Cs in surface water The decrease of 137Csin water in all areas has at least two components (Fig. 2.4.4). In most areas the 137Cscontents decreased to one half of the maximum values in about 2 months. In area 7, the decrease in 137Cscontent was clearly slower, one half of the maximum values remained about one year after the fallout. The component of the slower decrease of 137Csin water began at different times in different drainage areas, usually already in the year following the fallout, although in area 7 it did not take place until year three after the accident. The effective environmental halflives of the second phase of decrease were then 1.9 0.6 years. In some areas the experimental data refers to the third component in the decrease of 137Csin water. In area 6 a slight increase in 137Cs contents after 1991 reflects the increasing effect of runoff from the catchment area (Fig. 2.4.4, Table 2.4.5.).

12

TABLE 2.4.5. Effective environmental halflives of

I3’Cs

basins. The drainage basins are given in Fig. 2.4.1.

1 2 3 4 5 6 7

0.19 0.17 0.18 0.21

0.15 0.36 0.94

86 86 86 86 86 86-87 86-88

1.4 2.8 1.9 1.5 1.2 1.9 2.8

[y] in the water of large Finnish drainage

87-89 87-92 87-89 87-89 87-89 87-90 89-92

5.7

89-92

4.6 9.7 2.3

90-92 90-92 90-92 91-92

-

-1)

-

-

-

The contents of 137Cshad increased slightly in 1991-92, therefore it is impossible to calculate Tln,crf for the decrease.

Transfer of

from deposition to fish

The quarterly averages with variation of the content of

I3’Ck

in fish in the seven drainage areas are

given in Fig. 2.4.8. For six months after the deposition, the correlation between area medians of ‘37(3sin

fish and the area deposition was linear (Fig. 2.4.9).

The transfer maximum was reached in 1988 in drainage area 7, whereas in all other areas it was 1987. The transfer of 137Csfrom deposition to fish during the year of maximum transfer was highest in area 7, next highest in area 3 and lowest in the coastal drainage areas 1, 2 and 6 (Table 2.4.6).

Effective environmental halflife of I3’Cs in fish The halflives of 137Csobserved in predator, non-predator and intermediate fish varied from 1.5 to 4.8 years in areas 3, 4 and 5, but was remarkably shorter in area 7. The halflife in perch was

almost the same as that in the fish of the intermediate group because perch is the dominant species in this group, whereas that in pike was almost the same as that in all the predators (Table 2.4.7.).

The effective environmental halflives in fish were longest in drainage area 5, about 5 years, and shortest in drainage area 7, about 0.8 years. A larger uncertainty caused by the smaller number of samples is included in the estimated halflives of

137(=sin

fish in drainage areas 1, 2 and 6. These

areas are of minor importance for freshwater fishing in Finland.

73

1OOOOD

10000

looo# 100 10

If 1 /li .Ij

l . ,

100006~

,

,

(

1 1 1:11: f i4 ? iy i". i: DRAINAGE BASIN 4

10000

10

I

1-

,

DRAINAGE BASIN 5

,

,

I

,

I

I

1 10000

.

- 1 1986

Cf B q / k g 100000

1988

o

1990

l1

1992

1986

1988

1990

1992

DRAINAGE BASIN 7 A NON-PREDATOR

10000

0 INTERMEDIATE 0 PREDATOR

1

i 1986

Fig. 2.4.8.

1988

1990

1992

Quarterly averages of I3'Cs in three fish types, non-predator, intermediate and predator, with variation in seven large drainage areas after the accident at Chernobyl. The areas are shown in Fig. 2.4.1.

74

TABLE 2.4.6. Transfer factors TF, (m2/kg) from deposition to three different fish types (A=predator, B=non-predator, C=intermediate, All=all samples) in seven large Finnish drainage basins. The drainage basins are shown in Fig. 2.4.1.

Drainage basin

Fish type

1986

1987

1988

1989

1990

1991

1

A B C All

0.012 0.021 0.027 0.022

0.23 0.035 0.10 0.10

0.027 0.0080 0.022 0.019

0.020 0.0053 0.012 0.012

0.043 0.0068 0.011 0.012

0.013 0.0047 0.0034 0.0055

2

A

0.036 0.032 0.043 0.038

0.12 0.077 0.14 0.12

0.082 0.021 0.12 0.074

0.045 0.0092 0.032 0.030

0.033 0.0058 0.051 0.035

0.041 0.0094 0.048 0.038

0.17 0.13 0.27 0.20

0.12 0.057 0.079 0.085

0.10 0.031 0.11 0.084

0.11 0.027 0.072 0.066

0.073 0.023

All

0.016 0.030 0.026 0.026

4

A B C All

0.032 0.037 0.061 0.046

0.082 0.031 0.082 0.065

0.12 0.031 0.082 0.077

0.10 0.019 0.059 0.060

0.069 0.020 0.048 0.045

0.066 0.018 0.043 0.042

5

A B C All

0.055 0.054 0.098 0.073

0.090 0.033 0.12 0.086

0.12 0.023 0.11 0.093

0.084 0.019 0.084 0.070

0.070 0.012 0.087 0.067

0.086 0.0099 0.067 0.062

6

A B C All

0.011 0.020 0.064 0.042

0.086 0.034 0.093 0.072

0.063 0.022 0.059 0.050

0.067 0.021 0.039 0.042

0.10 0.013 0.023 0.075

0.061 0.022 0.032 0.037

7

A B C All

0.025 0.025 0.059 0.041

0.23 0.12 0.15 0.16

0.37 0.088 0.22 0.24

0.20 0.053 0.16 0.16

0.028 0.0097 0.021 0.020

0.022 0.0080 0.024 0.019

B C All 3

A B C

0.060

0.054

75

3000. 2500. 2000 1

20

10

0

40

30

0 kBq/m2 Fig. 2.4.9.

Correlation of activity concentrations of 137Csin fish (medians with quartiles)(l3q/kg fresh weight) and deposited 137Cs(kBq/m2) in different drainage areas in the first half year after the fallout.

G = 62.6 x

D

- 193 (Bqkg wet weight), ?=0.9348, p<0.0006

TABLE 2.4.7. Effective environmental halflife of 137Csin different fish types (non-predator, intermediate, predator, and in perch and pike) after reaching m a values in seven drainage basins. Drainage basin 1 2 3 4

5 6 7

Non-predator

TIRcff [YI Intermediate

1 2 1.8 2.9 3.9 2 0.91

1 1 1.5 3.4 2.1 2 0.84 ~~~

~~

Predator

T1acff

Perch

[YI

Pike

1 2 2.9 3.3 4.8

1.5 3.2 5.0

4.8 3.0 5.2

0.72

0.81

0.75

~~

Amounts of 137Csretained in the drainage areas The amounts of 137Csfrom the Chernobyl fallout deposited on the drainage areas were almost

totally retained in these areas. Only a minor part, 1.7, 1.1 and 1.3%of the total amount deposited was removed by the end of 1990 with the rivers flowing to the Baltic Sea from areas 4, 5 and 7, respectively.

76

DISCUSSION Though the transfer of radiocaesium deposited in the aquatic environment is a complex process depending on numerous factors, the use of the area-specific parameters obtained here provide a basis for rapid prediction of the levels of I3’Cs in water and in fish in different drainage areas after the deposition of 137Cs.Immediately after deposition, the season, the chemical form of the radionuclides, as well as other features of the deposition, will mainly determine the behaviour of radionuclides in the aquatic environment. In the first six months after a short release of ‘”Cs during spring, activity concentrations of 137Csin water and in fish in southern and central Finland can be estimated by: C, = 45.5 x D - 59.2 Bq/m3 and C, = 62.6 x D

- 193 Bqhg, respectively (D

must be given here in kBq/m2). After estimating the activity concentrations of 137Csin surface water and in fish, the countermeasures that may be needed in the fallout situation can be planned.

In later years temporal changes in the 137Cscontent of water have been still more profoundly affected by the character of the catchment area, and activity concentrations in surface water can be estimated by means of area-specific TF,s. The maximum variation in the annual values for

TFws,normalized with runoff, was a factor of 5 betwecn the seven drainage areas in 1987 and 1988. Extensive agriculture in drainage areas 1 and 2, and hence fairly high contents of nutrients in surface waters, may explain why transfer from deposition to surface water was lowest in these areas. Similarities in the behaviour of 13’Cs in water in areas 1, 2 and 6 (Fig. 2.4.7) are at least partly explained by the fact that fresh water in these areas is mainly derived from rivers discharging into the sea, while the number of central lakes is small. This suggests that water residence times might be relatively short in these areas. The large number of bogs in drainage area 7 may explain, at least partly, the higher transfer

of 137Csfrom deposition to water and slower decrease of 137Csin water compared to other areas. The higher transfer from deposition to water may be due to that the amount of 137Csleached from the drainage area and arriving to the lakes with runoff is also higher there. This agrees well with the assumption (Bergman R., 1994) that peat bogs constitute the main source for loss of 137Csfrom catchments in the boreal zones. The 13’Cs activity levels in Devoke Water, Cumbria (UK), also remained high for much longer than was expected after the Chemobyl accident. This was shown to result from the release of 137Csfrom the catchment to the lake. Among different soil types, fibrous peat soils were noticed to lose caesium rapidly. Rankers and amorphous peats lost smaller amounts of 137Csand podzols the least (Hilton et al., 1993). Due to the bogs in the catchment area, the content of humic substances in water is relatively high. This is also shown by the relatively high colour values of lake waters in the area. Humic substances, in turn, may retain caesium in water, possibly as colloids; otherwise caesium would be bound to suspended solids and sink to the bottom. In a lake-specific study (Ilus et al., 1993) the ratio of I3’Cs in the water of a lake in

drainage area 7 to 137Csin deposition was also clearly higher than the corresponding ratios in the other drainage areas. Relatively large clay and silt areas, especially in drainage basins 1, 2 and 5, explain the lower transfer of 137Csto water in the coastal drainage areas 1 and 2 compared to that in area 6, and the lower values of the 'inland' drainage area 5 compared to those in the other 'inland' areas

3, 4 and 7. Clay particles bind caesium tightly, and ion exchange on the clay minerals is the prevailing sorption mechanism for caesium (Ritchie et al., 1990). Therefore, the runoff of 137Cs from catchment areas rich in clay is minimal and only takes place with suspended clay particles, which in turn easily sink to the bottom sediment of the lake. According to the results obtained, '"Cs was transferred to fish most effectively in drainage area 7, where the transfer factors for predator, intermediate and non-predator were 0.37, 0.22 and 0.088 m2/kg in the year of maximum transfer, which was 1988 in this drainage area. This is

supported by a Swedish study (Hlkanson et al., 1990) in which one of the characteristics of the drainage basin best explaining the increased transport of radiocaesium to pike, was the large number of mires in the catchment area. Though transfer to fish was most effective in drainage area 7 after the fallout, the contents in fish also decreased most rapidly of all, and thus the halflives of

137Csin fish were shortest here (Table 2.4.7.). Maybe the chemical form of the 137Csdeposited in this area has gradually changed. The more rapid decrease of 137Csin fish in this drainage area could mean that there are diffcrences in the uptake and metabolism of different chemical forms of 137Csby fish. It seems evident that the chemical form of a radionuclide much affects its transfer in the aquatic environment. Though the variation of radiocaesium concentrations in water and especially in fish is large, and though there are a number of factors affecting the transfer of radionuclides deposited in the aquatic environment, differences did appear in the behaviour of 137Csin the large drainage areas. They may at least partly be explained by differences in soil type, land use and number of lakes in the area, as was the case in this report, but further analyses of the results are, however, needed to verify in greater detail the factors causing areal differences in the long-term behaviour of radiocaesium.

78 REFERENCES Andersson T, Forsgren G, HAkanson L, Malmgren L, Nilsson A (1990). Radioaktivt cesium i fisk i svenska sjoar efter TjernobyLSSI-rapport 90-04, Statens strilskyddsinstitut, Stockholm, Sverige. Arvela H, Markkanen M. and Lemmela H. (1990). Mobile survey of environmental gamma radiation and fall-out levels in Finland after the Chernobyl accident. Radiation Protection Dosimetry 32, 3, 177-184. Atlas of Finland, Folio 132 (1986). Water. National Board of Survey, Geographical Society of Finland. Bergman R. (1994). The distribution of radioactive caesium in boreal forest ecosystems. Chapter 4.5, this book. Forest Research Institute, (1980). Yearbook of forest statistics. 1980. Helsinki. Hammar J, Notter M. and Neumann G (1991). Radioaktivt cesium i rodingsjoar - effekter av Tjernobylkatastrofen. Information frh sotvattenslaboratoriet, Drottningholm 3, 1-152, Sverige. Hilton J, Livens F.R, Spezzano P. and Leonard D.R.P (1993). Retention of radioactive caesium by different soils in the catchment of a small lake. The Science of the Total Environment 129, 253-266. Hakanson L, Kvamas H, Andersson T, Neumann G, Notter M. (1990). Cesium i gadda efter Tjernobyl - dynamisk och ekometrisk modellering. SSI-rapport 90-09, Statens stralskyddsinstitut, Stockholm, Sverige. Ilus E, Puhakainen M. and SaxCn R. (1993). Gamma-emitting radionuclides in the bottom sediments of some Finnish lakes. Report STUK-A112. Finnish Centre for Radiation and Nuclear Safety, Helsinki, Finland. Ritchie JC, McHenry JR.(1990). Application of radioactive fallout cesium-137 for measuring soil erosion and sediment accumulation rates and patterns: A Review. Journal of Environmental Quality 19,215-233. Salbu B, Bjornstad H.E. and Brittain J.E. (1992). Fractionation of cesium isotopes and '%r in snow melt run-off and lake waters from a contaminated Norwegian mountain catchment. Journal of Radioanalytical and Nuclear Chemistry, Articles 156, 1, 7-20. S a x h R. (1990). Radioactivity of surface water and freshwater fish in Finland in 1987. Report STUK-A77, Supplement 3 to Annual Report STUK-A74. Helsinki: Finnish Centre for Radiation and Nuclear Safety. Saxdn R. and Aaltonen H. (1987). Radioactivity of surface water in Finland after the Chernobyl accident in 1986. Report STUK-A60, Supplement 5 to Annual Report STUK-ASS. Helsinki: Finnish Centre for Radiation and Nuclear Safety. SaxCn R. and Koskelainen U. (1992). Radioactivity of surface water and freshwater fish in Finland in 1988-1990. Report STUK-A94, Supplement 6 to Annual Report STUK-A89. Helsinki: Finnish Centre for Radiation and Nuclear Safety. Saxin R. and Rantavaara A. (1987). Radioactivity of freshwater fish in Finland after the Chernobyl accident in 1986. Report STUK-A61. Supplement 6 to Annual Report STUK-ASS. Helsinki: Finnish Centre for Radiation and Nuclear Safety. Suomi Finland, Uusi yleiskarttakija, Maanmittaushallituksen karttapaino, 1977.

79

2.5. THE ROLE OF LAKE-SPECIFIC ABIOTIC AND BIOTIC FACTORS FOR THE TRANSFER OF RADIOCAESIUM FALLOUT TO FISH

TORD ANDERSSONl & MARKUS MEILI2 1Dept. of Physical Geography, University of UmeB, 901 87 Umel, Sweden 2Inst. of Earth Sciences, Uppsala University, 752 36 Uppsala, Sweden

SUMMARY Lake-specific factors are used as a tool to explain the variation in transfer of 137Cs to fish at a certain level of fallout, and to describe and evaluate the temporal development of this transfer. The effect of biotic interrelationships and environmental variables influencing the tmphic structure of the lakes is briefly reviewed and discussed. The study is based on a large amount of existing Nordic post-Chernobyl data on 137Cs in different species of freshwater fish. The maximum transfer was reached within the first three years for all species in most lakes and normally in the order, small perch - trout and charr - larger perch - pike, a sequence that seems to reflect the trophic level each species occupies. Thus, the fish-eating pike is the species with the most extended temporal development and is also the species with the highest values of total time-integrated transfer. The transfer to fish differed over an order of magnitude between lakes, and lakes with a high total transfer to small perch also show a high total transfer to pike. Inter-lake differences in this respect could not be explained by the difference in growth rate of pike between lakes. Information is provided about the ranges for the total transfer to some common fish species in Nordic waters at different ranges of the theoretical residence time of 137Cs in lake waters (T&. Tcsis determined from the mean hydraulic residence time and the scavenging capacity of the lakes. The amount and nature of scavenging agents (possibly clay minerals) were well indicated, but not determined, by the natural concentration of base cations in lake water.

INTRODUCTION After the Chernobyl accident, high levels of radiocaesium became a major environmental problem in Sweden. 137Cs has been the major concern in most studies on the fate and effect of the Chernobyl fallout in Nordic lakes, partly due to its long physical half-life, partly because it is readily accumulated by organisms because of its chemical similarity to potassium, which is a major component in cell metabolism. In freshwater fish, which are an important factor in the Scandinavian life-style, high levels of radiocaesium occurred in the flesh and this led to various governmental actions. An estimated 14000 Swedish lakes contained fish with concentrations of 137Cs above 1500

Bq kg-1 wet weight during the autumn of 1987 (Hkanson et al., 1992). In this study, lake-specific abiotic and biotic factors are used as a tool to explain the variation in transfer of 137Cs to fish at a certain level of fallout, and to describe and evaluate the temporal development of this transfer. The study is based on a large amount of existing Nordic post-Chernobyl data on 137Cs in different species of freshwater fish. Much of the data on perch and pike originates from a comprehensive study initiated in 1986 on

80 "Liming-Mercury-Cesium", which was originally intended to evaluate different measures to decrease the mercury content in fish (HAkanson et al., 1990). However, in April 1986 the Chernobyl accident occurred and 41 lakes included in the project were affected by a major fallout of 137Cs in the range of 3 to 70 kBq m-2 . High levels of 137Cs in small perch from these lakes were recorded shortly after the fallout (Andersson et al., 1990; HAkanson et al., 1992), and the ongoing sampling schedule and planned remedial measures intended for mercury seemed well fitted to be supplemented with analyses of 137Cs and remedial treatments also directed towards reducing the levels of 137Cs in fish. Results from the project up to 1989 have been presented earlier, dealing, for example, with the relationships between drainage area characteristics and lake water quality (Nilsson and Hflkanson 1992), and effects of remedial measures on the content of radiocaesium in fish (Andersson et al., 1991; Hflkanson and Andersson 1992). In the present paper, data reported earlier are supplemented by new, unpublished data from the period 1990 to 1992 and by data on perch and pike (Andersson et al., 1990; Broberg and Andersson 1991; SaxCn 1992; Sundblad et al., 1991) and brown trout and Arctic charr (Hammar et al., 1991). Drainage lakes and their surroundings are complex systems where many chemical, physical and biological factors interact and potentially influence the distribution of different elements and their ecological impact. This paper concentrates on factors and processes which in the Nordic countries have turned out to be empirically significant, and which may affect the transport and fluxes of radiocaesium in the lakes. Beside the influence of abiotic factors, the effect of biotic interrelationships on the transfer and environmental variables influencing the trophic structure of the lakes are discussed. The basic transport route to fish is from atmospheric deposition to lake water with a certain retention within the catchment and the lake itself, and finally an uptake primarily via the food web. 137Cs is rather strongly adsorbed onto suspended matter, which implies that the flux of particles within lakes must be accounted for in lake budgets (Stumm and Morgan 1981). Radiocaesium has been widely used for sediment dating (Jaakola et al., 1983; Pennington et al., 1973) and more generally as a tracer of particle transport to and within aquatic environments (Campbell et al., 1982; Santschi et al., 1988; Walling 1989). Several papers have shown that great attention must be. paid to environmental factors influencing the distribution and behaviour of radiocaesium in lakes, such as the size fractionation of radiocaesium-carrying agents (Salbu et al., 1992) and the influence on the sorption of 137Cs to particles of the content and type of clay minerals (Cremers et al., 1988; Evans et al., 1983; Heit and Miller 1987), the organic content (Longmore et al., 1983) and the concentration of competing cations such as

m+or K+ (Comans et al., 1989; Evans et al., 1983).

As regards the radiocaesium content in fish, several water-quality variables have been reported to correlate significantly negatively, i.e., ionic strength and intercorrelated parameters such as hardness and potassium concentration (Andersson et al., 1990; HAkanson et al., 1992). It is also well documented that the relationship between the concentration of radiocaesium in organisms and in ambient water should be inversely proportional to the potassium concentration in lake water. Lake-specific parameters describing hydrology and lake morphology (e.g., the hydraulic residence time and lake depth) have also been identified and confirmed as correlates in several works

81 (Andersson et al., 1991; Hammar et al., 1991; HAkanson et al., 1989; HAkanson et al., 1992). The interpretation of these abiotic factors has mostly been focused on the retention of 137Cs in the lakes, but it has also been suggested that they could indicate inter-lake differences in bioproduction and biotic interrelationships, as the magnitude of the transfer to a given species or size group of fish appears to be a function of diet and feeding rate (Meili 1991; Forseth et al. 1992). The remedial measures tested in the "Liming-Mercury-Caesium" project included the addition of lime and potash. This made it possible to study if and how the distribution and retention of 137Cs were changed by the altered chemical conditions, and to what degree the retention was related to water chemistry as compared to lake characteristics being more or less unaffected by the remedial measures, such as hydraulic residence time, lake morphology and fundamental particle properties. The results of that study (Andersson, 1993) are used here to evaluate the role of the physical and chemical lake properties in the time-dependent transfer to fish. The Chernobyl fallout coincided with the snowmelt and a high runoff in the study region, which gave rise to a marked pulse of 137Cs traceable in different lake compartments and in different biota. This situation provided an occasion to quantify the decline and the time needed to reach conditions closer to steady state, and the results of this study might give some insight into these matters too.

MATERIAL AND METHODS As the results and data used in this work originate from several sources and projects, we will just give a brief description of the most important sources of error in the study. A more detailed description of the analytical methods and sampling routines as well as information concerning the temporal and spatial representativity and statistical variability of the data is given elsewhere (Andersson 1993; Andersson et al., 1991; Hammar et al., 1991; HAkanson and Andersson 1992; HAkanson et al., 1992; Hfianson et al., 1990). The overall strategy was to assign to each lake a representative mean value for each variable during a defined period of time. Annual fish samples, usually 10 to 15 individuals of each species, were used to give each lake a value of 137Cs for a specific period. The intra-lake variability for the same species and size was very high during 1986 when, for example, the coefficent of variation (CV) for small perch had an average of 0.60. The individual variation has gradually decreased and the possible error (10.95)in each yearly lake mean value is now mostly <20 %. Fallout on the lakes and their catchments was generally calculated from fallout maps based on aerial surveys of soil radiation, which were performed in 1986 at regular flight path distances of usually 5 to 10 km and calibrated against soil sample measurements. Considering that wide local variations occur, that numerous soil and vegetation factors could affect the measured radiation and that certain interpolation errors are inevitable, it seems obvious that these fallout values give a rather crude estimate of true loading, and the potential error in each lake mean value is probably higher than the calculated error for the content in fish. The relatively large possible errors in both radiocaesium content in fish and estimated fallout must be accounted for when dealing with transfer coefficients. Certain outlier lakes may be explained

82 by an erroneous estimate of, for instance, fallout, but because of the large number of lakes investigated it is unlikely that the generality of the results is significantly affected.

RESULTS AND DISCUSSION The effect of radioactive contamination is a function of the input of isotopes to ecosystems, given as a rate or as an accumulated amount. In the case of an atmospheric input of short duration as from the Chernobyl accident, concentrations and burdens of radioisotopes in different ecosystem compartments are most suitably normalized by means of the total areal deposition during April and May 1986 and expressed as transfer coefficients or factors (e.g. Bq kg-1 wet weight in fish per Bq m-2 deposited = m2 kg-1). Other commonly used ratios are concentration factors with respect to the activity in sediments or in ambient water (e.g. Bq kg-1 wet weight in fish per Bq m-3 in the ambient water). Transfer and concentration factors vary significantly between organisms, ecosystems and situations, and they become an intricate function of time when ecological turnover times exceed the duration of the input. Consequently, the transfer needs to be defined in relation to the time elapsed since contamination. In addition, the transfer from abiotic to biotic compartments is determined by a large number of physical, chemical, physiological and ecological processes that alter in importance as isotopes are redistributed within the ecosystem. As a result of the pulse-like type of emission, the transfer of Chernobyl caesium to fish showed strong temporal variation. At any time (t) the content of radiocaesium in fish within a lake (Cs-fish(t)) can be related to a load parameter, for example the fallout (Cssoil) as: TC(t)=Cs-fish(t)/Cssil (m2 kg-1)

UI

The magnitude of this lake specific ratio or transfer coefficent (TC) was largely dependant on the actual fish species and time after fallout. This is illustrated in Figure 2.5.1 where the variation of the transfer coefficient with time is shown for small perch (Percafluviufilis),pike @sox lucius), brown trout (Sulmo b u m ) and Arctic charr (Salvelinus alpinus) as compiled from data from a large number of Nordic lakes. The most striking feature of Figure 2.5.1 is perhaps the wide range of transfer coefficients in different lakes for the same species and size category at a particular time. This large inter-lake variation, of course, gives rise to a wide variety of possible lake specific curves for a certain species. The different temporal development of the transfer in different lakes is illustrated with curves for pike in lakes Lovsjon and Hamstasjon and for trout in Storsjouten and Grundvattnet respectively. Another measure of the spread around the mean TC(t) is given (Figure 2.5.1 A) by the yearly quartile values for small perch and pike in the most well-documented population of lakes (n=4 1), Another conclusion which can be derived from Figure 2.5.1 is that the maximum transfer was reached within the first three years for all species in most lakes and normally in the order, small perch - trout and cham - larger perch - pike. This sequence seems to reflect the trophic level each species occupies. Thus, the fish-eating pike is the species with the most extended temporal development and

83 is also the species with the highest values of total time-integrated transfer (F, Table 2.5.1). In this context it should also be noted that the types of lakes occupied by pike and perch on the one hand and trout and charr on the other are normally quite different.

A.

Jan-86

Jan-88

Jan-90

Jan-92

Date

1

B. r A

ir

E

0,1

." - --

I

,

A LI

P

\

0,Ol

0,001 Jan-86

I

I

Jan-88

Jan-90

Jan-92

Date

Figure 2.5.1. Ranges (within arrows), inter-quartile ranges (vertical bars) and mean values (lines) for the time-dependent transfer coefficient TC(t) between the Chernobyl fallout and the lake mean content of 137Cs in fish, for (A) small perch, 4 0 g (unfilled symbols) and pike (filed) and, (B) Arctic charr (unfilled, dashed line) and brown trout (filled, full line), based on data from 230 Nordic lakes. Examples of lakes with a low and high total transfer to pike (A) and to trout (B), respectively. After the maximum value on the transfer (TC(*o))was reached, the decline can be fairly well described by an exponentially decreasing function: Cs-fish(t) = Cs-fish(t=O)*e-kt, from which the apparent or ecological half-life (TE) for each lake can be derived: TE = In 2 / k. This makes it possible to extrapolate future transfer coefficients and to estimate the total, time-integrated value of

84 the total expected transfer given by the transfer factor F F = F(t0) + TC(a)*TE/ln 2 where F(t0) is a linear approximation of the time-integrated transfer up until the maximum transfer coefficient was reached. Planktivores, like small perch, reached their maximum transfer coefficient within a few months after the fallout and consequently F(tO) becomes small compared to the total transfer for these fish categories. From Eqn. 2 it is also obvious that the magnitude of F is determined both by the magnitude of the initial transfer and the rate of decrease as expressed in TE. Table 2.5.1. Mean, maximum and minimum values of the the total expected transfer (F, m2 kg-1 yr> of Chernobyl fallout to pike, small perch ( 4 0 g), brown trout and Arctic charr. Fpi(6):F,i and Fpe(3):Fpegives the fraction (in %) of F transferred after 6 years for pike and 3 years for small perch.

MGXl

Mm Max

Fpike

Fpi(6):Fpi

Fperch

Fpe(3):Fpe

FTrout

FCharr

0.95 0.11 5.0

74 50 85

0.50 0.04 1.34

84 36 99

0.55 0.13

0.46 0.18 0.90

1.o

Table 2.5.1 gives the expected total transfer (mean and ranges) for some different common fish species in Nordic lakes, and also the transfer after 3 years (F3) and 6 years (F6), respectively, in relation to the total expected transfer F. Annual and seasonal fluctuations and an increase of TE with time due to a future increased impact of factors controlling the secondary load (such as resuspension (Broberg and Andersson 1991; Hkkanson and Andersson, 1992) are possible. However, in small perch (which in this data set show a decreasing concentration of radiocaesium for the longest time, > 6 years), there is a tendency for an increase of TE during the last 3 years compared to the values (0.6
Turnover of radiocaesium in fish Several important controlling factors for radiocaesium turnover in fish were identified. Generally, both intake and excretion of radiocaesium vary with season and with the life stage of the fish. The turnover is rapid in summer, with both high intake and rapid excretion, and slower during winter with lower intake and slower excretion (Forseth et al., 1991; Meili, 1991). The dependence of uptake and elimination on temperature in a given type of organism may, however, differ significantly. Furthermore, turnover is faster in smaller than in larger fish, as small fish have a higher rate of metabolism. Food selection may also vary with season and fish size, thus influencing the intake of radiocaesium. Uptake of caesium from contaminated food is the major source of radiocaesium in fish, and

85

intake from the water is of negligible significance to the body burden in natural freshwater systems. The radiocaesium intake from food is thus determined by food selection, radiocaesium levels in the prey, and amount of food consumed (see above). On the other hand, excretion is probably determined by the metabolic rate (Carlsson, 1978; Evans, 1988) and may be species-specific. Several studies have shown that the elimination rate, similar to the metabolic rate, depends on both fish size and temperature (Ugedal et al., 1991 and references therein; Evans, 1989). In Nordic lakes, temperature varies with depth and time, with maximum temperatures in epilimnetic waters during summer. Excretion may therefore be linked to the habitat of the fish which, in some species or size groups, varies with respect to temperature. The same applies to the feeding rate of fish, but not necessarily in the same manner (Forseth et al., 1991). The biological half-life of radiocaesium was estimated experimentally in brown trout (Ugedal et al., 1991) and in one size class of roach (Evans, 1989). The former study confirmed the strong dependence on temperature, to a lesser degree there was also an effect of fish size. The latter study showed no effects of potassium addition to the water (not to the food) on the elimination rate of radiocaesium in roach. There exists a very clear inter-lake relationship between the total transfer to perch and to pike (Figure 2.5.2), i.e., lakes with a high total transfer to small perch also show a high total transfer to pike. The amount of appropriate data concerning trout and charr is not sufficient to make any similar comparisons of the total transfer between different lakes, but the high correlation during autumn 1987 (r2= 0.68, n = 12 lakes, Hiikanson et al., 1992) suggests that there exists a similar, but somewhat weaker, relationship as between pike and perch.

A

0.1

I 1

Fpike (mz kg-1 yr)

Figure 2.5.2. Relationship between the total expected transfer of Chernobyl fallout (F) to pike and small perch in 44 lakes, Fpike(6) = the time-integrated transfer to pike after 6 years. An attempt was made to relate the inter-lake variability in the activity ratio Cs-pike:Cs-perch to

the growth rate of pike expressed as mean age of 1-kg pike (Figure 2.5.3 A). However, the

86 inter-lake differences could not be explained by the difference in pike growth rate between the lakes. The total transfer to pike after 6 years (Fpike(6))did not show any clear relation to the mean age of pike, even if one might see a tendency to a faster transfer in lakes where the pike has a faster growth rate (Figure 2.5.3 B). Thus, other factors must be considered when explaining the inter-lake differences in transfer from fallout to fish.

A.

4

4,5

5 5,5 6 Mean age of pike (years)

6,5

7

B.

4

4,5

5

535

6

7

Mean age of pike (years)

Figure 2.5.3. The growth rate of pike expressed as mean age of I-kgpike compared to (A) the ratio of the concentration of 137Cs in pike and perch (Cs-pi:Cs-pe) during 1987, 1988 and 1992, and (B) compared to the time-integrated transfer to pike after 6 years, i.e. until 1992 (Fpike (6). m2 kg-1 yr-1).

The transfer to fish in relation to abiotic factors Earlier studies in Swedish lakes (Anderson et al., 1990; H&anson et al., 1992) have shown that the factors of highest statistical significance for the initial transfer of radiocaesium to fish were (sign

87 of correlation within brackets) water retention time (+) and any of the more or less intercorrelated parameters hardness (-), potassium concentration (-), humic content (+) and ionic strength (-). How are these parameters related to the total transfer factor (F), and what causal relationships could these correlations reflect? Andersson (1993) showed that the lake mean distribution coefficient (Kd = mean activity in settling particles / mean activity in lake water) for 137Cs varied between 1 * 104 and 7 * 104 cm3 g-1 in the Swedish lakes (n=15) most studied, a range which covers the values reported for Lake Ziirich (Santschi et al., 1990) and Lake Paijbne (Kansanen et al., 1991) and also most of the modelled, strongly time-dependent distribution coefficients in the epilimnion of Lake Constance during 1986 (Robbins et al., 1992). The inter-lake variation of Kd,Cs was quite strongly correlated to the natural (i.e. before liming operations) concentration of major base cations expressed as hardness (r2=0.58, n=15) and intercorrelated parameters like pH and alkalinity. h , c swas also well correlated negatively to the carbon content in settling material. The sedimentation rate of radiocaesium as expressed by the sedimentation coefficient (Ksed,Cs [d-I] = &,Cs

*

particle settling flux / lake mean depth) was well correlated to the natural

concentration of major base cations (r24.81, n=15) and intercorrelated parameters such as pH, alkalinity and conductivity (Andersson 1993). The higher scavenging capacity in lakes with higher concentration of major base cations was due to both higher particle sedimentation rates and higher & values in these lakes. However, water chemistry was probably not causal in this respect, despite the high correlation. Liming operations and potash addition caused a highly significant increase in mean values of hardness (and intercorrelated parameters) in lakes with initially low concentrations of major base cations (Hfkanson & Andersson, 1992), but this markedly increased concentration of major base cations in most lakes during 1988 and 1989 did not notably affect the mean distribution and sedimentation coefficients of 137Cs in the lakes (Andersson, 1993). The lack of effect of the remedial measms on the distribution and sedimentation coefficients of radiocaesium, suggests that a likely causal factor would rather be the amount and nature of scavenging agents, which in these lakes were well correlated to the natural concentration of base cations in the water. This is supported by the fairly good relationship between Ksed.Cs and the fraction of particulate inorganic matter in setding particles (PIM), which was very similar before and after remedies, respectively. The high correlation of Ksd,cs to natural water hardness might reflect that the concentration of major base cations in this calcite-poor region is positively correlated to the content of clay minerals in soils and sediments. Lakes with higher sedimentation coefficients generally also had a higher bioproduction as expressed by the concentration of total phosphorous, but the correlation of K,j,cs to total-P in water, and the ratio of C:N in settling matter, was much weaker compared to the correlation with hardness. In a model for scavenging of 137Cs in the epilimnion of Lake Constance (Robbins et al., 1992), the best fit to observed activities was obtained using a substitute (particulate aluminium) for clay minerals as the "reactive phase" of total suspended matter, while the affinity to calcite was found to be negligble. Cremers et al. (1988) and earlier works have shown that the division of

88 radiocaesium between solid and liquid phase in soils is regulated by a small number of highly selective ion-exchange sites, located at the frayed edges of micaceous clay minerals (illites). These observations are all consistent with the observed depletion of nuclear weapon radiocaesium in sediment inventories compared to the cumulative atmospheric deposition in North American softwater lakes (Heit and Miller, 1987), where low levels of binding clay minerals were suggested as a prime cause. Because of the mentioned reasons, however, this interpretation of the correlation between haxiness or ionic strength to scavenging capacity seems to be restricted to lakes within the zone of boreal forests and the relationship is different in geological areas dominated by calcite or dolomite weathering. The sedimentation coefficient Kse&csis inversely related to the residence time (Tsed, Eadie and Robbins, 1987) of 137Cs in the water body with respect to the removal by particle settling (Tsed =l/Ksed). The theoretical residence t h e of 137Cs in lake water (TcS)could then be obtained from:

where Tw= mean hydraulic residence time, Figure 2.5.4 shows the relationship between the mean hydraulic residence time and the theoretical residence time for 137Cs (TcS) within the lake water columns for different values of the sedimentation coefficient

(Ksed),

along with empirical values and modelled values based on natural

mean water hardness. It should be noted that there is a considerable variation between lakes in their theoretical retention of 137Cs. This variation is directly connected to the hydraulic residence time of the lakes, but also to the factors described earlier that influence the specific sedimentation rates of 137cs.

The theoretical residence time of 137Cs was determined in 15 lakes using sediment traps. As water hardness was found to be a good indicator of the scavenging capacity of 137Cs from lake water, the mean concentration of Ca+Mg in lake waters was used to calculate the removal rate due to sedimentation. In this context, it is important to note that the empirical values are based on data of gross sedimentation from sediment traps without accounting for any potential resuspension effects. The apparent residence time is almost always significantly longer than Tcs, as a result of resuspension of or possibly diffusion from Cs-contaminated sediments, and/or of input from the catchments. However, during the summer of 1986, after the large initial direct deposition and catchment-derived input that occurred within the f i s t month after Chernobyl, the lake water pools were comparatively large compared to the potential loading from resuspension and to the catchment input which declined very rapidly (Bergman et al., 1991; Santschi et al., 1988). Thus, it is likely that Tcs provides a fairly realistic description of the shape and magnitude of the pulse of radiocaesium activity through the lake water columns during the first important phase.

89

0

1

0.001 0.13

1.4

0.01

0.17 0.21

2.2 3.0

0.05

0.55

6.5

0.005

P

I

10

1000

100

Tw (days)

Figure 2.5.4. The relationship between the mean hydraulic residence time (Tw) and the theoretical mean residence time of 137Cs (TcS) for different values on the sedimentation coefficient (Ksed), along with empirical values based on sedimentation traps (n=15 lakes, unfilled diamonds) and modelled values (short lines) based on the natural mean hardness of lake water. This suggestion is supported by Figure 2.5.5, which shows the relationship between the theoretical residence time of 137Cs in lake water (Tcs) and the total transfer to pike. Three lakes with high transfer to fish despite low Tcs can be identified. Two of these are very shallow lakes (mean depths around 1 m), and a high degree of windwave-induced resuspension is likely. The third lake is situated closely downstream (< 500 m) of a much larger lake, and should in this context be regarded as a part of the larger lake with its much higher Tcs value. The relationship given in Figure 2.5.5 is therefore valid for lakes without major secondary inputs of 137Cs to lake waters from either major resuspension activity (or possibly diffusion) in sediments, or from major temporary traps within the catchment (lakes or bogs). From the rather high frequency of shallow lakes in the studied data set one would expect more outliers from the general pattern which does not account for resuspension or sediment-mediated uptake. This lack of outliers indicates that the biological availability of sediment-bound radiocaesium is generally rather low, even though the effect of resuspension or sediment-mediated uptake must be considered in certain lakes. Table 2.5.2. Normal ranges for the total expected transfer (F)of Chemobyl fallout to pike, small perch and brown trout in Nordic lakes at different theoretical residence times of 137Cs in lake water (Tcs).

Tcs (days)

Fpike (m2 kg-1 yr)

Fperch (m2kg-1 yr)

Feout (m2 kg-1 yr)

<30 30-50 50- 100 100-365 365-

<0.50 0.30-0.70 0.60- 1.2 0.80- 1.5 1.2-3.3

<0.25 0.20-0.40 0.35-0.60 0.50-1 .O 0.80-1.5

ca 0.20 ca 0.30 ca 0.50 ca 0.70 ca 1.0

90

Table 2.5.2 provides information about the ranges for the total transfer to some common fish species in Nordic waters at different ranges of the theoretical residence time of 137Cs (TcS). By combining the information in Figure 2.5.4 and Table 2.5.2 it is possible to obtain a fairly accurate estimation of the total transfer of Chernobyl caesium to, e.g., pike in a certain lake, based only on a knowledge of the hydraulic residence time and water hardness. -

10

c * 7 rn

*

N

€

v

1

-Bn L

0,1

1

10

'CC!

:2 c o

TCs (days)

Figure 2.5.5. The relationship between the theoretical mean residence time of 137Cs in lake water v c s ) and the expected total transfer to pike (Fpike) in 45 Nordic lakes, transfer values marked with unfilled squares are probably due to major secondary inputs (resuspension or upstream input, see text).

CONCLUSIONS The maximum transfer of Chernobyl caesium to fish was reached within the f i s t three years for all species and normally in the order; small perch - trout and charr - larger perch - pike, a sequence that seems to reflect the trophic level each species occupies. Thus, pike is the species with the highest total time-integrated transfer and the largest remaining fraction of the total transfer. The transfer from fallout to fish differed over an order of magnitude between lakes, and lakes with a high total transfer to small perch also show a high total transfer to pike. Inter-lake differences in the transfer from fallout to fish could not be explained by differences in pike growth rate. A better predicton was obtained from the mean hydraulic residence time and the scavenging capacity of the lakes, where the amount and nature of scavenging agents (possibly clay minerals) in this calcite-poor region were well indicated by the natural concentration of base cations in lake water. It seems likely that these, easily monitored, abiotic factors which are causally linked to the residence time of 137Cs in lake waters will be good predictors also of the transfer to fish after possible, future fallout events. ACKNOWLEDGEMENTS We are grateful to all people involved in sampling and data collection and Lars Sonesten for age determination of fish. The overall study "Liming-Mercury-Cesium" was financially supported by the

91 National Swedish Environmental Protection Agency ( S N V ) and the National Swedish Institute of Radiation Protection (SSI).

REFERENCES Andersson, T. (1993). Influence of abiotic lake characteristics on the distribution of 137Cs and Hg within lakes - implications for content in fish. In: Mercury and radiocaesium in Swedish lakes, Ph.D Thesis., GERUM 18, UmeA University, Sweden. Anderson, T. , G. Forsgren, L. Hdkanson, L. Malmgren, and 8. Nilsson (1990). Radioaktivt caesium i fisk i svenska sjoar efter Tjernobyl (with English summary) Swedish Radiation Protection Institute, SSI Rapport 90-04,41 pp. Andersson, T., L. H&kanson, H. Kvarnas, and 8. Nilsson (1991). Atgiirder mot hoga halter av radioaktivt caesium i fisk - Slutrapport f r b caesiumdelen av projektet Kalkning - kvicksilver caesium. (with English summary) Swedish Radiation Protection Institute, SSI Rapport 91-07, 114 PP. Bergman, R., T. NylCn, T. Palo and K. Lidstrom (1991). The behaviour of radioactive caesium in a boreal forest ecosystem. - In: The Chernobyl Fallout in Sweden (Ed. L. Moberg) Swedish Radiation Protection Institute, Stockholm, pp. 425-456. Broberg, A. and E. Andersson (1991). Distribution and circulation of Cs-137 in lake ecosystems. In: The Chernobyl fallout in Sweden, (Ed. L. Moberg) Swedish Radiation Protection Institute, Stockholm, pp. 151-176. Campbell, B. L., R. J. Loughran, and G. L. Elliott (1982). Caesium-137 as an indicator of geomorphic processes in a drainage basin system. J . Australian Geographic Studies 20: 49-64. Carlsson, S. and K. Liden (1978). 137Cs and potassium in fish and littoral plants from a humus-rich oligotrophic lake 1961-76. Oikos 30:126-132. Comans, R. N. J., J. J. Middelburg, J. Zonderhuis, J. R. W. Woittiez, G. J. De Lange, H. A. Das, and C. H. Van Der Weijden (1989). Mobilization of radiocaesium in pore water of lake sediments. Nature 339: 367-369. Cremers, A., A. Elsen, P. De Preter, and A. Maes (1988). Quantitative analysis of radiocaesium retention in soils. Nature 335: 247-249. Eadie, B. J. and J. A. Robbins (1987). The role of particulate matter in the movement of contaminants in the Great Lakes. In: Sources and fates of aquatic pollutants. Washington D.C.: American Chemical Society. pp.319-364. Evans, D. W., J. J. Alberts, and R. A. Clark 111 (1983). Reversible ion-exchange fixation of caesium- 137 leading to mobilization from reservoir sediments. Geochim. Cosmochim. Acta 47: 1041-1049. Evans, S. (1989). Biological half-time of Cs-137 in fish exposed to the Chemobyl fallout. Clearance of Cs-137 in roach exposed to various potassium concentrations in the water. Studsvik report NP89/74, Studsvik Nuclar, Nykoping, Sweden. Forseth, T., 0. Ugedal, B. Jonsson, A. Langeland, and 0. Njilstad (1991). Radiocaesium turnover in Arctic cham (Salvelinus alpinus) and brown trout (Salmo trutta) in a Norwegian lake. - J . Appl. E d . 28: 1053- 1067. Forseth, T., B. Jonsson, R. Nzumann, and 0.Ugedal (1992). Radioisotope method for estimating food consumption by brown trout (Salmo trutta). -Can. J . Fish. Aquat. Sci. 49:1328-1335. Hammar, J., M. Notter, and G. Neumann (1991). Radioaktivt cesium i rodingsjoar - effekter av Tjernobylkatastrofen. (with Enghlish summary) Swedish Freshwater Laboratory, Drottningholm. Information nr 3, 152 pp. Heit, M. and K. M. Miller (1987). Cesium-137 sediment depth profiles and inventories in Adirondack Lake sediments. Biogeochemistry. 3: 243-265. H&kanson,L. and T. Andersson (1992). Remedial measures against radioactive Caesium in Swedish lake fish after Chernobyl. Aquatic Sciences 54 (2): 141-164. HAkanson, L., T. Andersson, and 8. Nilsson (1989). Caesium- 137 in perch in Swedish lakes after Chemobyl - Present situation. relationships and trends. Environmental Pollution 58: 195-212. Hilkanson, L., T. Andersson, and 8. Nilsson (1992). Radioactive caesium in fish in Swedish lakes 1986-1988 - General pattern related to fallout and lake characteristics. J. Environ. Radioactivity 15: 207-229. Hdkanson, L., H. Borg, and R. Uhrberg (1990). Reliability of analyses of Hg, Fe, Ca, K, P, pH, alkalinity, conductivity, hardness and colour from lakes. Znt.Rev. ges. Hydrobiol. 75: 79-94. Jaakola, T., K. Tolonen, P. Huutunen, and S. Leskinen (1983). The use of fallout 137Cs and 239.24oPu for dating of lake sediments. Hydrobiologia 103: 15-19.

92 Kansanen, P. H., T. Jaakola, S Kulmala, and R. Suutarinen (1991). Sedimentation and distribution of gamma-emitting radionuclides in bottom sediments of southern Lake PtiijiijSinne,Finland, after the Chernobyl accident. Hydrobiologia. 2 2 2 121-140. Longmore, M. E., B. M. OLeary, and C. W. Rose (1983). Caesium-137 profiles in the sediments of a partial meromictic lake on Great Sandy Island, Queensland, Australia. Hydrobiologia. 103: 2127. Meili, M. (1991). In situ assessment of trophic levels and transfer rates in aquatic food webs, using chronic (Hg) and pulsed (Chernobyl 137Cs) environmental contaminants. - Verh. Internat. Verein. Limnol. 24:2970-2975. Meili, M. (1991). The importance of feeding rate for the accumulation of radioactive caesium in fish after the Chernobyl accident. In: The Chernobyl fallout in Sweden (Ed. L. Moberg), Swedish Radiation Protection Institute, 1991, 177-182. Nilsson, A. and L. H&anson (1992). Relationships between drainage area characteristics and lake water quality. Environmental Geology and Water Sciences. 19: 75-81. Pennington, W., R. S . Cambray, and E. M. Fisher (1973). Observation on lake sediments using Fallout 137Cs as a Tracer. Nature 242: 324-326. Robbins, J. A., G. Lindner, W. Pfeiffer, J. Kleiner, H. H. Stabel, and P. Frenzel (1992). Epilimnetic scavenging of Chernobyl radionuclides in Lake Constance. Geochim. Cosmochim. Acta 56: 2339-2361. Salbu, B., H. E. Bjornstad, and J. E. Brittain (1992). Fractionation of Cesium isotopes and 9oSr in snowmelt run-off and lake waters from a contaminated Norwegian mountain catchment. J . Radioanal. Nucl. Chem. 156 (7): Santschi, P. H., S . Bollhalder, K. Farrenkothen, A. Lueck. S . Zingg, and M. Sturm (1988). Chernobyl radionuclides in the environment: Tracers for the tight coupling of atmospheric, terrestrial and aquatic geochemical processes. Environ. Sci. Technol. 22: 510-516. Santschi, P. H., S . Bollhalder, S . Zingg, A. Luck, and K. Farrenkothen (1990). The self-cleaning capacity of surface water after radioactive fallout. Evidence from European water after Chernobyl, 1986-1988. Environ. Sci. Technol. 24: 519-527. Sax&, R.(1992) 137Cs i vatten och fiskar i Finland under 1986-1991. Presented at VI Nordic Seminary in Radioecology, Torshavn. Stumm, W. and J. J. Morgan (1981). Aquatic Chemistry. John Wiley & Sons. Sundblad, B., U. BergstrGm, and S. Evans (1991). Long term transfer of radioactive fallout nuclides from the terrestrial to the aquatic environment. Evaluation of ecological models. In: The Chernobyl fallout in Sweden (Ed. L. Moberg) Swedish Radiation Protection Institute, 207-238 Ugedal, O., B. Jonsson, 0. Njistad and R. Nreumann (1992). Effects of temperature and body size on radiocaesium retention in brown trout (Salmo trutta). -Freshwat. Biol. 28:165-171. Walling, D. E. (1989). Some applications of Caesium-137 measurements in sediment budget investigations. J . War. Res. 8 : 50-77.

93

2.6

MODELS FOR PREDICTING RADIOCAESIUM LEVELS IN LAKE WATER AND FISH

ULLA BERGSTROM, BJORN SUNDBLAD and STURE NORDLINDER Studsvik Eco & Safety AB, 5-611 82 Nykoping, Sweden. SUMMARY The most important pathway to man for 137Csentering fresh-water systems is via the consumption of fish. For pulse releases to fresh water it is therefore a major concern to predict future levels of radiocaesium in fish in order to identify cases where levels exceed required limits. The need to estimate the realibility of the models is, of course, important if one is to have confidence in such prognoses. For prediction purposes, the whole system needs to be taken into account, implying reservoirs for the lake waters, its biological system, sediment and drainage area. Such a model is presented in this paper and the calculated results are compared with the data observed for three Nordic lakes. In the uncertainty analysis, important processes that maintain increased levels of 137Csare identified and areas for model improvement recommended.

INTRODUCTION Every lake is unique, because the geological origin of a lake sets the limits for the morphometry or shape of its basin. In the Nordic countries, most lakes have been formed since the last ice age. During the post-glacial period, the lakes have changed as a result of different processes such as peat formation and eutrophication. There are a variety of physical, chemical and biological factors that have interacted to effect these changes. For 137Csreaching lakes directly or indirectly, the main exposure pathway to man is via the consumption of fish, for caesium accumulates in fish muscle because of its chemical similarity to potassium. This accumulation is most pronounced in freshwater and therefore of most radiological risk to individuals (not populations). For pulse releases to freshwater, it is important to identify cases in which the L37Cs levels in fish may exceed the authorities’ regulations, temporarily or during longer periods of time. Fishing in such lakes must then be prohibited until the concentration of 137Csmeets the required limits. For predicting future cesium levels, the whole system needs to be taken into account. In this context that means the lake waters, its biological system, sediments and drainage area. This modelling may also serve as a tool for identifying important processes and parameters maintaining increased concentrations of I3’Cs in fishes. This is especially true when the modelling is carried out with sensitivity and uncertainty analyses, in which the uncertainty in input parameter

94 values is generated through the model calculations, and correlation and regression methods are used to analyse the relations between input parameters and model responses. This type of modelling may also increase knowledge about ecosystem functioning, while using 137Csas an ecological tracer in the system (see Meili, this volume). Finally, as results from radioecological models may be used for decision making in connection with nuclear facilities, it is obviously of major importance to have confidence in the results. Radioecological models are mostly based upon compartment theory. It is therefore of scientific interest to take all possibilities of evaluating the compartment theory approach into account.

MODEL UNCERTAINTIES Because models have an inherent uncertainty, it is very important to know the magnitudes of this uncertainty in order to have confidence in the model results. This uncertainty arises from the different steps in the modelling such as conceptual modelling, consideration of processes and uncertainty in the input parameter values. When evaluating this uncertainty the most straightforward course is to use independent data sets for testing: blind testing. This implies that a model with parameter values is set up, run and compared with observed data, unknown to the modeller until the calculations are performed. This approach was taken in one of the scenarios in the f i s t BIOMOVS study, BIOMOVS 1 (Sundblad, editor, 1991). Additional testing is carried out within the IAEAKEC Co-Ordinated research programme VAMP (Validation of Model Predictions) (IAEA 1990), although not using the blind-test principle. MODEL STRUCTURE Modelling involves several steps and there is an inherent uncertainty coupled to each of them. Firstly, a conceptual model is designed based upon the system to be handled, in this case the lake with its drainage area. Of necessity this model a simplification of reality. The model includes compartments for the ingoing parts in the system. These should have similar properties and be in some context physically well-defined. Such ideal compartments may be the water above the thermocline which is well-fixed, zooplankton, benthos, fishes of a specific species and size. Considering the distribution of water depth in a lake, different zones are recognized. In shallow lakes, where the whole water mass is affected by external forces, one layer is usually sufficient to describe the lake, while deeper lakes demand at least two layers. However, the lake eventually has to be divided into several horizontal zones, compartments in modelling terms, if different lake basins are identified. Other areas may not be so ideal without having gradients of varying magnitude. Such areas may be divided into several compartments when simulating the system. One example of such a system is the sediments. The morphornetry of a lake provides a very useful means of describing a lake, and lake area, depth, volume, shoreline development are important parameters. There is a flow of 137Csbetween compartments because of physical, biological and chemical processes. In compartment modelling these are described by rate constants expressed as

95

turnover per unit time. These may be considered to be constant during the time studied, or vary such as seasonally. A typical example of the latter is the water residence time, which has a minimum in spring in connection with snow melting and the biological cycle. A typical example of a model structure is shown in Figure 2.6.1.

Fish 1

Fish 2

3 Drainage

outtlow

Deepsediment

-

Littoral sediment

Figure 2.6.1 Structure of general model for a lake eco system.

96

IMPORTANT PROCESSES The turnover of 137Cs within lake ecosystems is very complex. Many processes are involved in the transfer from the drainage area to the lake and within the lake itself. The most important processes are discussed below. Apart from the magnitude of the direct fallout upon the water surfaces, other factors influence the radiocaesium levels in water and fish. These are: season of fallout water retention time leakage of I37Cs from drainage area transfer to sediments turnover within sediments uptake into aquatic food-chains trophic level The characteristics of the lake and its drainage area play an important role for all these processes. 137Csmay enter a lake-ecosystem by either direct deposition on the lake surface or from the surrounding drainage basin (SaxCn, this volume). The chemical/physical form of the deposited 137Cs is important for the future mobility and bioavailability (Bjornstad et al., this volume). After deposition, I37Cs is exposed to different processes such as sorption, which affects its mobility (Anderson et al., this volume). The sorption depends, among other things, on soil type, organic matter content, soil water content, amount of competing elements and pH. A change in geochemical conditions such as acidification can cause increased mobility, i.e. a desorption process can take place. The sorption-desorption processes are usually simplified and only described by a distribution factor, Q. The resulting concentrations in lake ecosystems depend upon the time of year when the fallout occurs because of variations in water residence time, runoff and biological activity. The water residence time is determined by the inflow and outflow as well as the volume of the lake. The climate in the drainage basin determines the precipitation, evapotranspiration and discharge, all of which govern inflow and outflow. 137Csentering a lake sorbed on particles or in particulate form is subject to sedimentation, or may be transferred directly through the lake and leave the system. Parameters such as water depth, Q, suspended matter content and sedimentation rate were considered when modelling sedimentation. Within the sediments 137Csis affected by processes like resuspension, diffusion and bioturbation (Broberg, this volume). These may cause a reverse transfer back to the water as well as a migration downwards in the sediments. Resuspension can be of great importance in shallow lakes, where the wind action may affect the bottom sediments (Bjornstad et al., this volume and Sundblad et al., 1991). In general the modelling of biological uptake, in for example plankton and fish species, is described by bioaccumulation factors, if steady-state conditions prevail. If not, as for accidental contamination, the initial phase also demands a dynamic modelling of the uptake of 137Csto fish.

97 This can be done either by using rate constants based upon the bioaccumulation factor and the biological turnover time of 13’Cs within the species, or by using consumption values coupled with the fractional uptake. It is then important to consider the position of the species in the food chain. We have used this latter approach for the fish species and the former one for the plankton and the benthic amphipod. This is because of the fast turnover time of 137Cswithin these compartments compared to the fishes.

MODEL RESULTS, EXAMPLES FROM THREE NORDIC LAKES The turnover of 137Cs in three Nordic lakes was evaluated preliminarily against known observational data. The lakes are 0vre Heimdalsvatn in Norway, Hillesjon in Sweden and Is0 ValkjWi in Finland. The lakes cover a wide spectrum of conditions. Some of the most important characteristics from a modelling point of view are summarized in Table 2.6.1. Data on these lakes are also given in Bjarnstad et al. (this volume). Table 2.6.1 Lake characteristics Lake

Ovre Heimdalsvatn

Hillesjon

Is0 ValkjLvi

Alt m a s 1 Lake area (h2) ’ Max depth (m) Lake volume ( m3) Catchmentfllakearea Mean renewal period Stratification

1090 0.78 13 3.7E6 30 63 days winter strat 6.8 0.4 0.7 - 3.2 Oligo Oligo 25 - 30 0.3 60

10 1.6 3 2.7E6 12 130 days winter strat 7.3 3.0 36 - 57 Meso Eutro 100 5.0 600*

126 0.042 8 1.3E5

130 26 - 260

100

PH K (mgfl) Cond (mS/m) Humic status Trophic status Prim prod (g C/mz/y) Susp load (mg/l) Mass sedim rate (g/m’/y) Deposition (kBq/m2) - Mean - Ranges

* estimated value

4

3 years dimictic 5.1 0.4 1.7 Oligo Oligo 25

<1 70* 70

98 All lakes were covered in ice during the fallout. Normally ice covers the lakes for 40-60 % of the year. 0vre Heimdalsvatn has a very short water-residence time during the spring flood, when it is in the order of a few days. Hillesjijn also has a much shorter residence time at that time of the year, although some weeks are a more realistic period. Based upon this information a conceptual model (see Figure 2.6.1) was set up with compartments representing drainage area, water, sediment and types of biota. The last vary from lake to lake. For example, in the Swedish and Finnish lakes the trophic level from roach up to pike is represented. In the Norwegian mountain lake, trout is the dominating species. There are also compartments in the model for fish-food such as algae and plankton. The basic structure has different connections according to the lake to be modelled and the fish species present. Two compartments were used to describe the sediments, one for near-shore resuspension bottoms and another for the deeper accumulation bottoms. The rate constants in the model were based on the site specific information and biological parameters such as consumption values for the fish, fractional uptake and biological half-life of 137Csinside the fish. In addition the Kd-factor was used for obtaining the transfer rate from the water column to the sediments. The effect of seasonality was taken into account in the model by varying parameter values for the warm and cold season, respectively. Some of the most important parameter values with ranges are given in Table 2.6.2.

Table 2.6.2 Some important parameters, best estimate (b e) and ranges 0vre Heimdalsvatn be ranges

Hillesjon be ranges

Is0 Valkjtirvi be ranges

Bfzooplankton (kg d.w.P)

1.E4

3.E3- 3.E4 3.E3

1.E3 - 1.E4

1.E4

3.E3 - 3.E4

Bf gammarus (kg d.w./l)

2.5E4

8.E3- 8.E4 1.E4

3.E3- 3.E4

2.5E4

8.E3- 8.E4

Tin perch winter (month)

nc

nc

20

15 - 25

20

15 - 25

Ti12 perch summer n c (month)

nc

8

6 - 10

8

6 - 10

Ti12 trout winter (month)

15

10-20

nc

nc

nc

nc

Ti12 trout summer (month)

7

5-9

nc

nc

nc

nc

Kd ( l k g )

1.E4

(0.1 - 5)E4 1.E4

(0.1 - 5)E4

1.E4

0.1 - 5)E4

n c = notconsidered.

99 All rate constants were calculated within the code implying that the uncertainty in each parameter value is taken into account. For solving the differential equations, the BIOPATH code (Bergstrom et al., 1981) was used and the uncertainty analyses were carried out with the statistical error propagation package PRISM (Gardner at al., 1983), coupled to BIOPATH. The deposition onto the lakes and their drainage systems was assumed to be partly in particulate form with percentages varying according to the distance from Chemobyl (Korhonen, 1990). This fraction of the fallout was considered to be directly transferred to the lake sediments. The models were run for the lakes according to the site description and the deposition. Results were calculated twice a year from autumn 1986 until spring 1991. Emphasis was put on the uncertainty analyses and not on predicting peak values and the time of their occurrence. The calculated levels of I3Ts in the water and fish species were compared with the observed values. Some illustrative preliminary results are given below. Results for water and perch in Hillesjon are shown in Figs 2.6.2 and 2.6.3, respectively. The calculated results are given with 95 % confidence intervals according to Chebysev's formula (Boas, 1966). Results for perch in Is0 Valkjtirv and trout in @vre Heimdalsvatn are given in Figs 2.6.2 and 2.6.3 in a similar way. All predicted to observed ratios (P/O) for 137Csin the water for Hillesjon and Is0 Valkjtirvi are well within a factor of two, (Table 2.6.3) and the observed and predicted levels show about the same behaviour with time. The predicted values in this comparison are the arithmetic mean of the calculated distribution. For Heimdalsvatn there are few observations available about the levels in lake waters. However, when comparing results from that lake the predicted values are about a factor of 4 higher than the observed. One possible explanation of this overestimation is that the ice-cover was not considered. This ice-cover at the time of the fallout probably led to an immediate high transport of radiocaesium out of the lake by means of the spring flood. Another explanation is a too low transport of 13Ts to the sediments caused by a too low value of the &factor. For lake, I S ValkjPvi, ~ there is a slight tendency to underestimation.

Table 2.6.3 Predicted to observed rations (P/O) of 137Cs concentrations in lake water Season Autumn 1986 Spring 1987 Autumn 1987 Spring 1988 Autumn 1988 Spring 1989 Autumn 1989 Spring 1990 Autumn 1990 Spring 1991

0vre Heimdalsvatn

Hillesjon

Is0 Valkjavi

1.1

0.9 1.3 1.4 4.2 4.7

1.5 0.8 0.8 1.0 1.6

0.7 0.8 0.9 1.0 0.9 0.8

100

1El 0

Observed

1EO 0

1E-1 L

I

I

I

I

I

I

I

I

I

I

AS6 SS7 AS7 S S S ASS SS9 AS9 S90 A90 S91 Time (S = Spring, A = Autumn)

Figure 2.6.2 Predicted and observed concentrations of l37Cs in water from lake Hillesjon.

Bqkg fw 1E5

Observed

Q

1E4

1E3

I

I

I

I

I

I

I

I

I

I

AS6 SS7 AS7 S S S ASS SS9 AS9 S90 A90 S91 Time (S = Spring, A = Autumn)

Figure 2.6.3 Predicted and observed concentrations of 137Csin perch from lake Hillesjon

101

With respect to fish we have chosen to present results for one predator in each lake. Predicted and observed values are shown in Figs 2.6.3 to 2.6.5. For the Finnish oligotrophic lake, Fig 2.6.4, there is a tendency for the model to overestimate the levels in fish, while the opposite seems true for the other lakes. A plausible explanation for this is too high consumption values and too long biological turnover time of 137Cs in the fish, compare the dynamics. The observed values show a faster reduction than do the calculated. The overestimation in calculated fish-levels in the Norwegian lake is partly due to the overestimation of the levels in water. However, the values are not only overestimated but also show insufficient dynamic behaviour with too low a decrease of the levels with time. The model was rerun with modified parameter values, such as a much higher &-value, high removal of 137Csby the spring flood, and a faster biological turnover time of caesium in the trout. Higher &-values are in agreement with the findings of Anderson as well as Broberg (this volume).These results for trout are shown in Fig 2.6.5, model 2. In general the time behaviour has improved considerably, as well as the predicted levels of 137Csin water. Comparison with observed levels in water only showed discrepancies in the order of 20 %. The model was run on a monthly basis and the results for 0vre Heimdalsvatn would probably be improved if shorter period of times were used in the initial phase when the residence time of water is only a few days. The turnover of water is a very important "cleaning" process for lakes receiving Chemobyl fallout as this occurred during the spring (Hikansson, 1991). The dominating sources of uncertainty in the results vary between the lakes and as a function of time after deposition. For the Norwegian lake, the %-factor initially contributed most to the uncertainty, thereafter the leakage of *37Csfrom the drainage area to the lake is identified as the main contributor to the uncertainty in the I3?Cslevels in water. Also, initially the source term, that is deposited amount of activity, plays an important role in the initial calculations. After modifying some input values, the leakage from the drainage area becomes even more important as the importance of the &-factor has decreased considerably. This is because after a so-called breakpoint the influence of the &-factor decreases in relation to the other parameters used for obtaining the transport of 137Csto the sediments. This also reflects the high variation of deposition (Haugen et al., 1993) for Heimdalsvatn. The great importance of the leakage of 137Csfrom the drainage area is consistent with results from a model study in two subarctic salmonid ecosystems (Nordlinder at al., 1993) and agrees with the observations (BjBrnstad et al., this volume).When analysing the results for Hillesjon, resuspension seems to be the main process maintaining the 137Csconcentrations. This is in agreement with the observed results (see Bjernstad et al., this volume and Sundblad et al., 1991), where the loss of the nuclide from the lake is shown to be mainly from the amount deposited in the lake and not from leakage from the drainage area. For the Finnish lake, the speciation of the fallout 137Cscontributes significantly at the beginning, whereas the depletion of L37Csbcause of transfer to the sediments later becomes a dominating source of uncertainty. This could be expected to be in agreement with the long residence time of water in this lake.

102

Observed

0

1E4 @

Q

Q

I

I

c

t

1E3 I

I

I

I

I

I

I

0

I

0

I

AS6 S87 AS7 SSS ASS ,589 AS9 S90 A90 S91 Time (S = Spring, A = Autumn) Fig 2.6.4 Predicted and observed concentratins of 137Cs in perch from lake Is0 Valkjsivi.

Bqkg fw 1E5 Predicted Model 2

1E4

1E3 AS6 S87 A87 SSS AS8 SS9 AS9 S90 A90 S91 Time (S = Spring, A = Autumn) Fig 2.6.5 Predicted and observed concentrations of 137Cs in trout from lake Ovre Heirndalsvatn.

103 The results have greater similarity when comparing the dominant contributions to the uncertainty in the predators' levels of 137Cs. For all lakes the bioaccumulation factors for "fishfood" contribute to the uncertainty at all calculated time points. However, parameters contributing to the uncertainty in 137C.s concentrations in water may also contribute to the uncertainty in the 137Cslevels in fish. This is shown for example in 0 v r e Heimdalsvatn, where the initial deposition contributes to the uncertainty, even several years after deposition. In addition to the uncertainty analyses, a local sensitivity analysis, by varying all parameters similarly (10 % of the mean value was considered to be the standard deviation), was carried out for Hillesjon. Contrary to the uncertainty analyses, it showed that the parameter most sensitive to the results was water depth in the lake. For a full evaluation of the models, further analyses should be performed that include the amount deposited in the sediments.

CONCLUSIONS The calculations have shown that compartment models give satisfactory results for predictions of the concentrations in fish and water within five years after the deposition. However, the results are sensitive to appropriate parameter values such as & and biological turnover time in fish. More testing needs to be done for the long-term trends in the concentrations in fish species. According to the uncertainty analyses, the importance of processes varies with time and with the type of lake. Leakage from the drainage area is important for lakes at high altitudes and located in hilly terrain. For shallow lowland lakes, resuspension from the sediments appears to be the main contributor maintaining increased levels of 137Csin water. The evolution with time of the concentrations in fish needs to be addressed further in order to study the agreement between predicted and observed peak values and time of occurrence. Increased knowledge about the chemical speciation of 137Cscould also improve the accuracy of the model results.

ACKNOWLEDGEMENTS This study was partly supported by Swedish Nuclear Fuel and Waste Management. We thank also Anders Appelgren for his assistance with the graphical design.

I 04

REFERENCES BergstrBm, U., Edlund, O., Evans, S. and Rijjder, B. (1992). BIOPATH - A computer code for calculation of the turnover of nuclides in the biosphere and the resulting doses to man.

(STUDSVIK/NW-82/261)Studsvik Energiteknik AB. Boas, M. L. (1966). Mathematic methods in the Physical Sciences. John Wiley & sons, Inc, New York, London, Sydney. Gardner, R. H., Rojder, B. and Bergstrom, U. (1983). PRISM - A systematic method for determining the effect of parameter uncertainties on model predictions. (STUDSVIIVNW-83/55)Studsvik Energiteknik AB. Hilkansson, L. (1991). Radioactive caesium in fish in Swedish lakes after Chernobylgeographical distributions, trends, models and remedial measures. In: The Chernobyl fallout in Sweden (ed. L. Moberg), The Swedish Radiation Protection Institute, Stockholm, pp 239-281. IAEA 1990. International Atomic Energy Agency Division of Nuclear Fuel Cycle and Waste Management, Vamp, Aquatic working group Status Report, December 1990. Korhonen, R. (1990). Modelling transfer of 137Cs fallout in a large Finnish watercoarse. Health Physics 59,4, pp 443-454 Nordlinder, S, Bergstrom, U, Hammar, J and Notter, M. (1993). Modelling Turnover of Cs-137 in Two Subarctic Salmonid Ecosystems. Nordic J. Freshw. Res. 68, pp 21-33. Sundblad, B. (editor) (1991). Dynamics within lake ecosystem. (BIOMOVS TR 12). Swedish Radiation Protection Institute. Sundblad B, Bergstrom, U and Evans, S. (1991). Long term transfer of fallout nuclides from the terrestrial to the aquatic environment. Evaluation of ecological models. In:The Chernobyl fallout in Sweden. (ed L Moberg) The Swedish Radiation Protection Institute, pp 207-238.

105

2.7. RADIOCAESIUM IN ALGAE FROM NORDIC COASTAL WATERS

LENA CARLSON' & PAULI SNOELIS~' 'Department of Marine Ecology, Lund University, P.O. Box 124, S-221 00 Lund. *Departmentof Ecological Botany, Uppsala University, Villavagen 14, S-752 36 Uppsala. SUMMARY The results are presented from surveys of radiocaesium concentrations (I3'Cs and 134Cs)in the brown alga Fucus vesiculosus from Nordic coastal waters (Finland, Sweden, Norway, Denmark, Iceland) and in epilithic diatom communities from the Swedish part of the Gulf of Bothnia in 1991. The Baltic Sea was significantly contaminated by the Chernobyl accident, and radiocaesium from the accident was still predominant in radiocaesium concentrations in the algae here after five years. The general fallout pattern was still reflected, with highest concentrations in the southern Bothnian Sea. Between 1987 and 1991 the radiocaesium concentrations decreased, except for the area between Sweden and Denmark at the outflow from the Baltic Sea, where increased concentrations were found at a number of sites. This may indicate that radiocaesium is being transported out of the Baltic Sea. F. vesiculosus is a good and often used indicator of radioactive discharges into Nordic coastal waters, but in the northern Baltic Sea it is absent because of the low salinity (< 5 L).In this area we used instead epilithic diatom communities, which proved to be a good alternative to F. vesiculosus for monitoring radioactive discharges in areas with low salinities. INTRODUCTION Algae are widely used for monitoring radionuclides in aquatic environments. In Nordic marine waters the brown macroalga Fucus vesiculosus L. (bladder-wrack), is mainly used (e.g. Nilsson et al. 1981 Aarkrog 1985, Boelskifte 1985, Ilus el al. 1988, Carlson & Holm 1992). Epilithic diatoms (attached microalgae) were also found to be good monitoring organisms for radionuclides in Swedish coastal areas and can be used in areas where F. vesiculosus is absent, e.g. in the Gulf of Bothnia (Snoeijs & Notter 1993a). Earlier, collections of F. vesiculosus Results of measurements on Fucus vesiculosus have been supplied by: E. Ilus, Finnish Centre for Radiation and Nuclear Safety (Finland), H. Dahlgaard, Riser National Laboratory (Denmark), G . Christensen, Institute for Energy Technology (Norway), S. Magnusson, The National Institute of Radiation Protection (Iceland), E. Holm, Department for Radiation Physics, Lund (Sweden) and the Swedish Environmental Protection Agency (Sweden).

106

65

Iceland

69

Fig. 2.7.1. Sampling sites for Fucus vesiculosus (1-81) and diatoms (Dl-D4) in Nordic coastal waters during summer 1991.

107 were made along the Danish, Finnish, Norwegian and Swedish coasts during the eighties to study the radioactive contamination of the marine environment in the Nordic countries as a result of global fallout and of discharges by nuclear power plants and nuclear fuel-reprocessing plants. An evaluation of the data on radionuclides from monitoring programmes in the Baltic Sea is presently being carried out by the HELCOMIMORS group and will be published in 1994 (H. Nies: pers. comm.). At the beginning of the eighties, before the Chernobyl accident, the activity concentrations of I3'Cs in F. vesiculosus were fairly constant in the whole of the Baltic Sea (5-12 Bq kg-I dw). Along the Norwegian west coast there was a gradient from south to north (from ca 13 to ca 6 Bq kg-' dw), and around Iceland concentrations were less than 1 Bq kg-' dw (Ilus et al. 1981, 1983, Aarkrog 1985, Duniec et al. 1985). These L37Cs concentrations were mainly attributed to discharges from Sellafield in Great Britain (Aarkrog 1985). Measurements of radiocaesium concentrations in epilithic diatoms near the Forsmark nuclear power plant in 1984 (southern Bothnian Sea, Sweden) gave results that were always below the detection limit, which was 200 Bq kg-I in the presence of 'lomAgin the discharge of the nuclear power plant (Notter & Snoeijs 1986). After the Chernobyl accident the radionuclides were inhomogeneously distributed in the Baltic Sea. The highest concentrations in F. vesiculosus were found adjacent to the land areas that received the highest fallout deposition (see e.g. Persson et al. 1987, Carlson & Holm 1992). Since the accident radionuclide levels, mainly of radiocaesium, have been followed in F. vesiculosus from the Nordic coasts and in diatoms from the northern Swedish coast (e.g. Ilus et al. 1988, Carlson & Holm 1992, Snoeijs & Notter 1993a,b).

This paper presents results from a joint programme of the Nordic Liaison for Nuclear Safety Research carried out in 1991. Collections of F. vesiculosus were made in all Nordic countries during the summer, and diatoms were collected during the spring in Sweden. The radiocaesium concentrations in the algae sampled in 1991 are compared with previous measurements on the same algal taxa from the same sites. METHODS Fucus vesiculosus plants were collected and measured by the following institutes: the Finnish Centre for Radiation and Nuclear Safety (Finland), the Risra National Laboratory (Denmark), the Institute for Energy Technology (Norway), the National Institute of Radiation Protection (Iceland), Department for Radiation Physics, Lund (Sweden) and the Swedish Environmental Protection Agency (Sweden). In 1991 the samples were collected from the same sites as those visited for earlier surveys in the different countries (Fig. 2.7.1). The collections were made during June in Iceland, July in Finland and Sweden, August in Denmark, and August-

108

TABLE 2.7.1. Concentrations of 134Cs,137Csand 4% in Fucus vesiculosus from Nordic waters

in summer 1991 (Bq kg" dw). Measured uncertainty <20%. (For locations of the sampling sites, see Fig. 2.7.1.) Site 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43

134Cs I3'Cs 9 9 10 9 8 8 7 7 5 6 6 6 8 5 18 13 18 16 65 99 86 13 12 13 11 19 18 11 12 12 7 7 83 58 59 70 64 79 58 62 74 80 85

79 84 87 81 74 77 68 65 47 55 61 51 80 52 170 120 160 140 7 11 9 120 110 110 98 170 171 111 112 127 70 69 9 7 7 7 7 8 6 7 8 9 9

4oK

Site

990 900 900 850 820 890 800 750 670 770 530 760 880 700 840 830 690 720 730 670 760 760 640 670 550 610 999 605 681 801 691 668 605 639 643 507 515 708 540 741 729 796 1010

44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60 61 61a 61b 61c 61d 62 63

134Cs 137Cs 7 6 -

3 2 2 1 1

1

-

64 65 66' 67' 68' 69% 70 71 72 73 74 75 76 77 78 79 80 81

-

-

<1

-

6 5

9

71 52 33 25 23 17 14 10 8 6 5 5 5 10 3 3 3 2 2 2 4 11 2 1 1 1 0.4 0.4 0.5 0.4 5 11 7 6 8 13 19 26 37 57 5 76

4%

808 672 696 952 978 822 937 955 740 673 701 750 1240 755 990 750 960

795 725 790 lo00 1206 807 1089 933 1063 926 804 923 978 676 972 1361 1026 755 1017 812

~

*

Samples from Iceland: stated overall uncertainty for the 662 KeV 137Csline is 3%.

109 September in Norway. Epilithic diatoms were collected from the Gulf of Bothnia in April-May by P. Snoeijs and analysed by the Swedish Environmental Protection Agency. Complete plants of F. vesiculosus were collected by hand in the uppermost metre, but in Finland the samples were collected from greater depths (1-3 m) by SCUBA diving. Visible epiphytes were removed from the plants, which were then dried and homogenized. The diatoms were collected from stones from 0.2-0.7 m depth, and then dried and homogenized. Water samples for the analysis of radiocaesium concentrations were also collected at some of the sites. The gamma-emitting radionuclides were measured by gamma spectrometry using high purity detectors. Activity concentrations for '34Cs,L37Csand 40K in the algae are given as Bq kg-' dry weight (dw) for the algae and as Bq m-3for water.

RESULTS AND DISCUSSION

Radiocaesium in Fucus vesiculosus The Baltic Sea was significantly contaminated by the Chernobyl accident in April 1986. Radiocaesium from the accident was still predominant in the radiocaesium concentrations in

Fucus vesiculosus samples from the Baltic Sea in 1991. The concentrations in the summer of 1991, five years after the accident, were still higher than in 1983 (Fig. 2.7.2). The highest concentrations (> 100 Bq kg-' dw) were observed in the Bothnian Sea: at the northernmost sampling sites for F. vesiculosus along the Finnish and Swedish coasts and around h a n d (Table 2.7.1). In Finland the concentrations of radiocaesium in F. vesiculosus had decreased to a greater degree along the Gulf of Finland (south coast of Finland) than along the Gulf of Bothnia (west coast of Finland) from 1987 to 1991. This may indicate that the west coast of Finland was still receiving radiocaesiurn from the Kumo river, which has its drainage area in Tavastland, an area that in 1986 received a large part of the total fallout in Finland (E. Ilus: pers. c o r n . ) . The Baltic Sea currents run northwards along the Finnish west coast and the radiocaesium in the river discharge is thus transported to the north. The relatively high concentrations around the Finnish island of h a n d (Sites 15-18) probably originate from the Baltic currents that flow southwards along the Swedish coast. An additional possible explanation for the faster decrease of radiocaesium in the Gulf of Finland is the water turnover rate which is higher here than in the Gulf of Bothnia. This was also observed in the water samples that contained 137Csconcentrations of 80-110 Bq m-3 in the Gulf of Finland and 110-200 Bq m-3in the Gulf of Bothnia in 1991 (E. Ilus: pers. c o r n . ) . Along the Swedish east coast the highest concentrations of radiocaesium in F. vesiculo-

110

Fig. 2.7.2. Activity concentrations of 137Cs(Bq kg-l dw) in Fucus vesiculosus in Nordic coastal waters in 1983, 1986, 1987 and 1991.

111 sus in 1991 were measured at the northernmost sampling sites, e.g. at SimpMs (Site

27: 171 Bq kg-' dw). In 1987 concentrations had increased in the southern part of the Swedish

east coast since 1986 (Carlson & Holm 1992). In 1991 the radiocaesium concentrations had decreased since 1987 everywhere along the Swedish coasts, except at some sites on the south coast where concentrations were somewhat higher than in 1987, indicating the presence of radiocaesium in the outflow from the Baltic Sea. The largest deposition of Chernobyl fallout in Sweden was measured in the area of Gavle, north of the sampling sites for F. vesiculosus (Site D2, Fig. 2.7.1). The radiocaesium was most probably transported to the Bothnian Sea via the rivers with drainage areas in the surroundings of Gavle and then southwards with the Baltic currents that run southwards along the Swedish east coast. The radiocaesium concentrations in F. vesiculosus from Danish coastal waters were similar to the Swedish ones in adjacent areas. The highest concentrations were measured at Bornholm (Site 81) and west Zealand (Site 79). In 1986 and 1987 only the Zealand sites were visited, and in 1991 the concentrations of radiocaesium had increased since 1987 in east Zealand (Sites 77 and 78). This indicates the presence of radiocaesium in the outflow from the Baltic Sea. In Norway (Sites 57-65) the radiocaesium concentrations in F. vesiculosus were low compared with those in the Baltic Sea region in 1991, mostly below 5 Bq kg-l dw. A clear decrease in activity concentrations since 1987 was observed. The concentrations of 137Csin 1987 did not differ significantly from those measured in 1981, five years before the Chernobyl

accident, but 134Csconcentrations were relatively higher in 1987, indicating the Chernobyl origin of radiocaesium in F. vesiculosus along the Norwegian coast. The highest concentrations of radiocaesium were measured in F. vesiculosus from the Hardanger fjord (Sites 61a-d). A concentration of 11 Bq kg" dw was found in F. vesiculosus from the innermost site in this fjord, where the highest radiocaesium concentrations in Norwegian F. vesiculosus were measured also in 1987. The lowest radiocaesium concentrations in F. vesiculosus, 0.4-0.5 Bq '37Cskg-' dw, were observed in Iceland (June 1991). South of Iceland the 137Csis mainly old fallout from nuclear tests, whereas releases from the nuclear industry are noticeable north of Iceland (blafsdottir ef af. 1992). '"Cs

was not detected in any of the Icelandic samples. It may be concluded that

radiocaesium from the Chemobyl accident is negligible in the Icelandic marine environment. The '34Cs/'37Csratio in F. vesiculosus varied between 0.103 and 0.108 in the Baltic Sea in the summer of 1991, which is in agreement with the calculated decrease in this ratio through physical decay of the Chernobyl fallout for June-August 1991, based on the value 0.54 for the ratio in April 1986 (Aarkrog 1988). This supports the assumption that the high concentrations of radiocaesium in the Baltic Sea region in 1991 still originate from the Chemobyl accident.

112 1000 3 n

-u 4

2

100:

CT

m

W

C

0 .-tJ

c

lo

c, C (u

0 C

8

1 i

I

I

I

I

Fig. 2.7.3. Activity concentrations of 137Cs(Bq kg-' dw) in Fucus vesiculosus in Nordic coastal waters in 1991.

Radiocaesium in water The Nordic coastal waters vary much in salinity. The Baltic Sea has a relatively stable salinity gradient from < 1%0in the north to ca 10%0in the south. From the Oresund and the Danish Straits to the Skagerrak there is a more fluctuating gradient from ca 10 to ca 2 5 7 ~ and , further northwards around Norway and Iceland there are full marine conditions (ca 3 5 L ) . The uptake and release of radionuclides and their stable isotopes are affected by environmental factors such as salinity and temperature (e.g. Dahlgaard 1984, Carlson & Erlandsson 1991) and therefore concentration ratios too depend on these factors. In Sweden concentrations of 137Csin water varied from 180 Bq m-3 at the northernmost sampling site in the Baltic Sea to 30 Bq m-3 on the west coast in July 1991. In Finland concentrations in water varied from 80-110 Bq m-' in the Gulf of Finland to 110-200 Bq m-3 on the Finnish Bothnian Sea coast. After the Chernobyl accident 137Csconcentrations in Baltic Sea water were up to 3700 Bq m-' at Forsmark in the summer of 1986 (Grimis et al. 1986), ca 500 Bq m-3 in the h a n d Sea in October 1986, and concentrations gradually decreased southwards to about 35 Bq m-3 in the southern Baltic Sea (DHI 1987). At an offshore sampling station in the Gulf of Finland the highest concentration of L37Cs in water was 1100 Bq m-3 in August 1986 (Ilus el al. 1991), and in a coastal area of the Gulf of Finland the maximum value was 5200 Bq m-3 in May 1986

113 (Ilus el al. 1989). In 1987 a southward transport of radiocaesium in water since 1986 was observed, but in 1991 only slightly increased concentrations of 137Cswere found in the southern Baltic Sea as a result of water movements from the more contaminated areas in the northern region (Ikaeimonen et al. 1988, H . Nies: pers. c o r n . ) .

Concentration ratios for radiocaesium in Fucus vesiculosus Ratios between radiocaesium concentrations in algae and water are often calculated when using algae as monitoring organisms. Concentration ratios are however affected by several factors such as season and salinity and these must be taken into account. In general, in areas with lower salinities higher concentration factors are found than in areas with higher salinities (Aarkrog 1985). The ratios reported here for F. vesiculosus are based on measurements on algae and water from the same sites. The observed concentration ratios for I3'Cs between algae and water on this sampling occasion were ca 710 in the Baltic Sea proper (n=6), ca 300 in the southern Kattegat (n=3) and ca 190 in the Skagerrak (n=3). These concentration ratios (on a dry weight basis) are in agreement with those observed at the beginning of the eighties by Aarkrog (1985), who mentioned a concentration ratio of 200 for F. vesiculosus in temperate ocean water.

TABLE 2.7.2. Concentrations of '34Cs,137Csand 4% in epilithic diatom communities from the Gulf of Bothnia in 1991 in Bq kg-' dw (% measured uncertainty). Site (Sample)

Area

'34Cs

137cs

4%

Forsmark Forsmark Forsmark

230 (2) 210 (4) 80 (10)

2070 (1) 2120 (1) 700 (2)

690 (3) 970 (6) 400 (10)

Bight of Bight of Bight of Bight of Bight of Bight of

340 (3) 190 (3) 150 (2) 130 (6) 400 (2) 180 (3)

2930 (1) 1840 (0) 1440 (1) 1120 (1) 3770 (1) 1600 (1)

490 (8) 370 (4) 400 (4) 710 (8) 1040 (4) 390 (9)

60 (14) 50 (8) 30 (13)

640 (2) 480 (2) 330 (2)

480 (1 1) 580 (7) 440 (7j

20 (12j 20 (12) 10 (35)

240 (2) 90 (3) 210 (2) 80 (4) 90 (4)

470 (6) 690 (4) 480 ( 5 ) 790 (4) 770 (5)

Gavle Gavle Gavle Gavle Gavle Gavle

N Bothnian Sea N Bothnian Sea N Bothnian Sea Bothnian Bay Bothnian Bay Bothnian Bay Bothnian Bay Bothnian Bay

114

Radiocaesium in diatoms from the Gulf of Bothnia The concentrations of radiocaesium in epilithic diatom communities in the Gulf of Bothnia five years after the Chernobyl accident still clearly reflected the original fallout pattern (Table 2.7.2). The mean concentration of 137Cswas highest in the area adjacent to the highest fallout, the Bight of Gavle (southern Bothnian Sea), 2120 Bq kg-I dw in 1991 (Table 2.7.3). At Forsmark, ca 50 km south of Gavle, the mean concentration was also high: 1630 Bq kg-I dw. Immediately after the Chernobyl accident maximum concentrations of 137Csin diatoms were as high as 110 kBq kg-' dw in the Bight of Gavle and 30 kBq kg" dw at Forsmark (Grimis et al. 1986, Snoeijs & Notter 1993b). Further to the north the mean 13'Cs concentrations in diatom samples in 1991 were significantly lower, 480 and 140 Bq kg' dw in the northern Bothnian Sea and the Bothnian Bay, respectively. The 134Cs/137Cs ratio in the diatoms was 0.114 in April/May 1991, in agreement with those established for F. vesiculasus (see above). In epilithic diatom communities radiocaesium is being recycled to a large extent; it was calculated that at Forsmark on average ca 20% of the initial high mean concentration in April/May 1986 still occurred in epilithic diatoms in 1991 (Snoeijs & Notter 1993b). Since the summer of 1986 the caesium isotopes have mainly disappeared through physical decay after initial steep drops a few months after the accident in the early summer of 1986. The high recycling concentrations may be related to the fact that diatoms occur throughout the year and multiply by cell division, but they cannot be explained completely as there are also large loss rates of cells, e.g. by the scraping of ice, and after periods of high biomass (e.g. after the spring bloom).

TABLE 2.7.3. Mean concentrations of '%?s, I3'Cs and % in epilithic diatom communities from the Gulf of Bothnia in 1991 (Bq kgl dw). CV = Coefficient of variance. Area

n

134cs mean (CV)

137CS mean (CV)

mean (CV)

Forsmark (Dl) Bight of Gavle (D2) N Bothnian Sea (D3) Bothnian Bay (D4)

3 6 3

170 (39%) 230 (43%) 50 (27%) 20 (41 %)

1630 (40%) 2120 (44%) 480 (27%) 140 (48%)

690 (34%) 570 (43%) 500 (12%) 640 (22%)

5

4%

115

Different algae as monitoring organisms Different kinds of algae provide different kinds of information when following fallout of the Chernobyl type over a longer period of time. Immediately after the fallout, surface adsorption is the main form of uptake and concentrations in algae depend mainly on their surfacelvolume ratios. With time other factors become important too, such as physiology of the algae, storage from one year to another in macroalgal tissues and release to the environment (Carlson & Holm 1992), as well as recycling in epilithic diatom communities (Snoeijs & Notter 1993a). Snoeijs & Notter (1993b) compared radiocaesium concentrations in different macroalgal species with those in epilithic diatom communities in the Baltic Sea. They found that in 19861987 radiocaesium concentrations were mainly related to the surface/volume ratios of the algae. Diatoms have high surface/volume ratios, and they always had the highest radiocaesium concentrations. Pilayella liftoralis (thin uniseriate filamentous brown alga) had ca 80% of the concentration in diatoms from the same site, Cladophora glomerata (uniseriate filamentous green alga) ca SO%, Ceramium gobii (filamentous red alga) ca 40%, Fucus vesiculosus (brown alga with broad, thick thallus) ca 20%, and Enferornorpha intestinalis (filamentous green alga with relatively broad filaments) ca 15%. In 1990-1991 this pattern had changed, now the algal group was most important; red and brown macroalgae had the highest radiocaesium concentrations related to diatoms (Pilayella: 11%, Ceramium: 9%, Fucus: 8%), and green algae the lowest (Cladophora: 4%, Enteromorpha: 2%). Diatoms show little selectivity for kind of radionuclide and less dependence on physiological and seasonal cycles than do macroalgae. Immediately after Chernobyl, radiocaesium concentrations were extremely high in diatoms compared with other algae, and they were directly related to the fallout pattern and water turnover time of the sampling site. Six months after Chernobyl about 20% of the initial high concentrations were being recycled in the diatom communities, and they still were in 1991; the concentrations decreased only according to the physical decay of the radionuclides (Snoeijs & Notter 1993a,b). The advantage of using diatoms is thus their high concentration potential for radiocaesium; for the Forsmark area a concentration ratio of 8300 (on a dry weight basis) was calculated for 137C~ (Snoeijs & Notter 1993a). On the other hand it is relatively hard to gather epilithic diatoms in the field, contrary to F. vesiculosus, which can be recognised without difficulty and provides enough easily gathered material for analysis throughout the year. Diatoms may be used as monitoring organisms for radiocaesium especially in the northern Baltic Sea which has very

low salinity and where diatoms are abundant. In environments more distinctly marine in character it might be more difficult to gather enough diatom material, and F. vesiculosus is the best alternative here.

116

CONCLUSIONS (1) In 1991 radiocaesium from the Chernobyl accident still dominated the radiocaesium concentrations in algae in the Baltic Sea, but along the Norwegian coast the impact of Chernobyl still measurable in 1991 was very small, and in Iceland negligible. (2) The general pattern of fallout from Chernobyl in Nordic coastal waters was still reflected in the algal samples. The highest 137Csconcentrations in Fucus vesiculosus (110-170 Bq kg-' dw) were found in the southern part of the Bothnian Sea, where also the highest concentrations were found in epilithic diatom communities (1100-3800 Bq kg-' dw). In the Gulf of Finland and the Baltic Sea proper, 137Csconcentrations in F. vesiculosus were ca 50-100 Bq kg.' dw, and they decreased rapidly along with the outflow from the Baltic Sea through the Kattegat and Skagerrak down to 2-3 Bq kg-' dw along the Norwegian coast, and reached background levels of about 0.4 Bq kg-' dw around Iceland. Along the Swedish side of the Gulf of Bothnia, radiocaesium concentrations in epilithic diatom communities were significantly lower in the northern Bothnian Sea (330-640 Bq 137Cskg-' dw) and the Bothnian Bay (80-240 Bq 137Cskg-' dw) than in the southern Bothnian Sea. (3) Indications were found of the transportation of radiocaesium out of the Baltic Sea. Radiocaesium concentrations in F. vesiculosus have decreased since 1987 at most of the sampling sites, but increased concentrations occurred at the outflow from the Baltic Sea in 1991. (4) Epilithic diatom communities are good monitoring organisms in areas with low salinity, where F. vesiculosus is absent and epilithic diatoms are abundant. Results of both algal groups could be compared as I3'Cs concentrations in F. vesiculosus were on average 8 % of those in epilithic diatoms 4-5 years after the Chernobyl accident.

ACKNOWLEDGEMENTS This study was supported by the Swedish Institute of Radiation Protection (SSI) and the Swedish Evironmental Protection Agency (SNV).

REFERENCES Aarkrog, A. (1985). Bioindicator studies in Nordic waters. Nordic Liaison Committee for Atomic Energy, Ris0 Denmark, Report NKA REK 5B.74 pp.

117

Aarkrog, A. (1988). The radiological impact of the Chemobyl debris compared with that from nuclear weapons fallout. J . Environ. Radioactivity, 6 , 151-162. Boelskifte, S. (1985). The application of Fucus vesiculosus as a bioindicator of '%o concentrations in the Danish Straits. J . Environ. Radioactivity, 2, 215-227. Carlson, L. & Erlandsson, B. (1991). Effects of salinity on the uptake of radionuclides in Fucus vesiculosus L. J . Environ. Radioactivity, 13, 309-322. Carlson, L. & Holm, E. (1992). Radioactivity in the Baltic Sea following the Chemobylaccident. J. Environ. Radioactivity, 15, 231-248. Dahlgaard, H. (1984). Transuranics, rare earths and cobalt, zinc and caesium in Fucus vesiculosus (seaweed): Laboratory exercises and field realities. In: Proceedings of the International Symposium on the Behaviour of Long-lived Radionuclides in the Marine Environment, A. Cigna & C. Myttenaere (editors). Commission of the European Communities. pp. 347-354. DHI. (1987). Die Auswirkungen des Kernkraftwerkunfalles von Tschernobyl auf Nord- und Ostsee. Meereskundliche Beobachtungen und Ergebnisse, 62. 1-67. Duniec, S . , Carlson, L., Hallstadius, L. & Holm, E. (1985). Fucus vesiculosus (L.) as a bioindicator for 137Csin the Baltic Sea and Kattegat. Proc. Seminar on the behaviour of radionuclides in estuaries, Renesse, The Netherlands 1984. C.E.C. pp. 229-240. Grimis, U . , Neumann, G. & Notter, M. (1986). Tidiga erfarenheter av nedfallet frin Tjemobyl - Radioekologiska studier i svenska Kustvatten. Naturvdrdsverket Rapport, 3264. 65 pp. Ik&eimonen, T.K., Ilus, E. & SaxCn, R. (1988). Finnish studies on the radioactivity in the Baltic Sea in 1987. Finnish Centre for Radiation and Nuclear Safety. Helsinki, Report STUK-A, 82. 35 pp. Ilus, E., Ikaheimonen, T.K., SaxCn, R., Sjoblom, K.-L., Klemola, S. & Aaltonen, H. (1989). Review of the Finnish studies on radioactive substances in the Baltic Sea in 1986-1988. Baltic Sea Environment Proceedings, Helsinki Commision, Helsinki 31. pp. 91-93. Ilus, E., Ikaheimonen, T.K., SaxCn, R., Suomela, M., Gavrilov, V.M., Gedeonov, L.I., Gritchenka, Z.G., Ivanova, L.M., Tishkov, V.P. & Reshetov V.V. (1991). Study of radioactive substances in the Baltic Sea in 1986-1987. Finnish Centre for Radiation and Nuclear Safety, Helsinki, Report STUK-B-VAL0 69. Ilus, E., Klemola, S . , Sjoblom, K.-L. & Ikaheimonen, T. (1988). Radioactivity of Fucus vesiculosus along the Finnish coast in 1987. Finnish Centre for Radiation and Nuclear Safety, Helsinki. Report STUK-A83 Supplement 9 (to Annual Report 1987). 36 pp. Ilus, E., Ojala, J., Sjoblom, K.-L., & Tuomainen, K. (1981). Fucus vesiculosus as bioindicator of radioactivity in the Finnish coastal waters, 1. Gulf of Finland. Institute of Radiation Protection, Helsinki, Report SIZL-B-TUTO 14. Ilus, E., Ojala, J . , Sjoblom, K.-L. & Tuomainen, K. (1983). Fucus vesiculosus as bioindicator of radioactivity in the Finnish coastal waters, 2. Archipelago Sea and Gulf of Bothnia. Institute of Radiation Protection, Helsinki, Report STLL-B-TUTO 18. Nilsson, M., Dahlgaard, H., Edgren, M., Holm, E., Mattsson, S. & Notter, M. (1981). Radionuclides in Fucus from inter-Scandinavian waters. In: Impacts of radionuclide releases into the marine environment. IAEA-SM-2481107. pp. 501-513. Notter, M. & Snoeijs, P. (1986). Radionuklider i bentiska kiselalger - Ett irs studier i Biotestsjon, Forsmark. Natumirdsverket Rappor? 3213. 30 pp Olafsdottir, E.D., Magnusson, S.M. & PBlsson, S.E. (1992). Radiocaesium in seawater, fish and Fucus in Iceland. Proc. 6th Nordic Seminar on Radioecology, The Faroe Islands. 10 PP, Persson, C., Rodhe, H. & De Geer, L.-E. (1987). The Chernobyl Accident - A meteorological analysis of how radionuclides reached and were deposited in Sweden. Ambio, 16, 20-31. Snoeijs, P. & Notter, M. (1993a). Benthic diatoms as monitoring organisms for radionuclides in a brackish-water coastal environment. J . Environ. Radioactivity, 18, 23-52. Snoeijs, P. & Notter, M. (1993b). Radiocaesium from Chernobyl in benthic algae along the Swedish Baltic Sea coast. Swedish University of Agricultural Sciences, Department of Radioecology, Rapport SLU-REK 72. 21 pp.

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1 I9

2.8 POLONIUM-210 AND RADIOCAESIUM IN MUSCLE TISSUE OF FISH FROM DIFFERENT NORDIC MARINE AREAS.

ELIS HOLM* Department of Radiation Physics, Lund University, Sweden

S#MARY Po and radiocaesium were analyzed in t h e muscle tissue of different species of fish f r o m the Baltic Sea, the Norwegian Sea and lcelandic waters. On the basis of the results a dose assessment was made which demonstrates t h a t the 210 dose t o the population from Po originating from the consumption of fish from t h e Baltic Sea is similar t o that from radiocaesium, even a f t e r the Chernobyl accident. For the other areas the dose from radiocaesium is smaller but of t h e same importance as t h a t from 210Po. Differences in salinity are of minor importance compared t o the food chains with respect to 210Po. INTRODUCTION

Polonium-210 belongs t o t h e natural uranium series starting with 238U, but itsdistribution radionuclides

in the environment depends t o a large extent on t h e other in

210

atmosphere

Pb

the

decay series,

originates

mainly

continents. The low concentrations of

such

as

from 226

226

the

Ra

but

'"Rn

also

'l0Pb.

emanating

In the

from

the

Ra in sea water mean t h a t relatively

small amounts emanate from from the world oceans (less than 2% of t h e total 222

Rn). Polonium-210 is relatively short-lived

decay of

210

Pb(Tl,2=22 yl via 'loBi

138 dl and formed by the

(T1,2=5 dl. Taking radioactive decay into

account and considering t h a t the residence time of 'loPb between 4 and 29 days, the total inventory of

in the atmosphere is

210

Pb in the atmosphere is

estimated t o be lOI5Bq. : Data has been made available by: Gordon C. Christensen, Instltute for Energy Technology, KJeller, Norway (analysis); Lars Foyn, Instltutc of Marlne for Research. Bcrgen, Norway (sample collection); Erkkl Ilus, Flnnlsh Contra Radlatlon and Nuclear Safety, Helsinki, Flnland (sample collectlon); Ceorg Neumann, Swedish Unlverslty of Agricultural Sclences. Uppsala, Sweden (sample collectlon); Slgurdur Magnusson, National Institute of Radlatlon Protectlon,

ReykJavlk. Iceland (sample collection)' Ryszard Oceanology. Sopot, Poland (unpubllshed data for aloPo).

BoJanowskl,

Institute

of

120 There are other sources of 'loPb aeolian

resuspension

of

soil,

the

in the atmosphere such as

and 'loPo

burning

of

fossil

fuel,

manufacture

of

phosphate fertilizers, volcanic eruptions, forest fires, etc. The deposition of

210

Pb can be estimated at 220 Bq m-'per

year in areas

situated f a r f r o m the sea. The average deposition over the earth i s about 70 per year per year. The atmospheric flux i s controlled by t h e t r a n s p o r t

Bq m-'

of aerosols and wet and dry deposition. The atmospheric flux is controlled by t h e transport

of

aerosols and wet and dry deposition. Regional differences

occur and also differences between the hemispheres i.e. the flux is smaller in the southern hemisphere because of the smaller land mass. The global cycle of 210

Po has been described by El-Daoushy (1988). The residence time of '"Pb

in the atmosphere makes t h a t 210Po does not

reach radioactive equilibrium with

210

Pb. The atmospheric 210Po/210Pb r a t i o is

in t h e order of 0.1. In sea water the ratio is 0.5-1.0, of '"Po

the principal source

210

being the radioactive decay of

Pb (Cherry and Heyraud, 1988).

MATERIALS AND METHODS Fish samples of different species were collected during 1991, at t h e end of 1992 and at t h e beginning of 1993 in Iceland, Finland, Norway and Sweden. The sampling stations are shown in Fig. 1. Only the muscle tissue was retained and dried.

The species collected,

although somewhat

different

in

the different

countries, were Cod (Gadus rnorruha), Herring (Clupea harengus), Pike (Esox

lucius), Perch (Perca f l u v i a t i l u s ) , Plaice (Pleuronectus p l a t e s s a ) , Whitefish (Coregonus lavaretus), Capelin (Mallotus villosus) and Common mussel ( M y t i l u s edulis). After gamma-spectrometric analysis f o r radiocaesium, the samples (10-20 g dry weight) were wet ashed with nitric and perchloric acid. Polonium was plated [spontaneous deposition) onto nickel or silver discs.

209

Po was used as

radiochemical yield determinant '09P0 was used and the samples were measured f o r 2-10 days by solid s t a t e alpha spectrometry. The solutions remaining were kept f o r possible further analysis of '"Pb months

after

collection.

As

'loPb

too. Analyses were performed 1-6

concentrations a r e generally

factor of 2 0 lower ( in the order of 0.15 Bq kg-' 1992))

than

production

of

those

for

210Po

210Po as a

concentrations,

result

of

the

this

decay of

at

least

a

dry weight (Bojanowski, is

possible,

'loPb

is

in

and

the

this

case

negligible. A concentration f a c t o r of 20 000 f o r polonium and 100 f o r lead is given by the IAEA (1985). In the calculations of the activity concentration a t collection, a f a c t o r of 20 between the

concentrations of

210

Po and 'l0Pb was

121

Fig. 2.8.1 Sampling sites f o r fish (and Mytilus) in different Nordic areas.

122 used. This w a s supposed t o be the case too f o r the mussels which were analyzed under t h e same conditions and using the same procedures.

RESULTS AND DISCUSSION Detailed studies of 'loPb

in the marine environment show t h a t polonium plays

an important role f o r the absorbed dose t o man in sea food consumption. Very f e w d a t a and dose estimations have been found however f o r the Baltic Sea in spite of the f a c t t h a t this is an important source of fish consumption. The Baltic has an important fish catch. The Baltic Sea o f f e r s a possibility t o study

different

parameters

such

as salinity and

seasonality,

and

connects

with t h a t of a marine environment. Relative t o water

fresh-water radioecology

volume, t h e Baltic is a marginal sea with a comparatively large input of industrial waste, which might contain natural radionuclides such as

210

Pb and

210

Po. 210

Biological processes lead t o preferential accumulation of organisms and

starting from the

f i r s t trophic

level

the

Po in living

210

Po/210Pb

ratio

becomes g r e a t e r than unity. The ratio is of the order of 2 in particulate organic

material

and

feacal

pellets

of

planktonic

phytoplankton, about 25 in the whole body of

animals,

zooplankton

about

5

in

and 100 in the

hepatopancreas of pelagic crustaceans (Cherry and Heyraud, 1983). The distribution of zlOPo varies greatly within the fish depending on t h e type of tissue. Thus, the highest levels have been invariably recorded in t h e pyloric caecal masses, liver, viscera and stomach content and the lowest in muscle tissue. The respective activity concentrations may vary by three orders of magnitude (Shell et al., 1973). Polonum-210 concentrations a r e amplified through successive food chains, but at higher trophic levels they a r e largely deposited in internal organs. 210

Pb shows no increasing concentration through successive

as in the case of

food chains but,

210

Po, the highest concentrations occur in internal organs.

The 210Po/210Pb r a t i o covers a wide range of 10-100 depending on t h e organ and is highest in the muscle tissue. Four species of fish in the Baltic Sea account f o r over 80% of commercial catches: -Baltic herring (Clupea harengus) -Cod (Gadus rnorhua) -Flounder (Platichthys f l e s u s ) -Sprat (Clupea sprattus).

123 TABLE 2.8.1. Polonium-210 in fish muscle from different Nordic areas. Species

Flounder

Herring

Cod

Plaice Herring Cod

Pike Herring Perch Whitefish Herring

Cod

Place of

Site

collection

on map

Year of

Bay of Gdansk

1

Southern Baltic Bay of Gdansk Gulf of Finland Bay of Gdansk

2

Activity conc.

Reference

Bq kg-' dry w.

1988 1990 1990 1991 1970 1988 1990 1991 1988 1990 1991 1991

16f2 4.2f0.9 3.9f0.4 2.720.3 3.6 23f3 1.9f0.2 4.821.1 3+1 3.3f1.6 0.920.1 1.9f0.6

Oslo Fjord 21 MBre, Norway 5 Finnmark, Norway 6 Oslo Fjord 4 Finnmark, Norway Oslo Fjord 21 More, Norway Finnmark, Norway

1992

12f2 llf2 822

20 Sipoo, Finland Pietarsaari, Finland 19 Helsinki, Finland 18 Porkkala, Finland 17 Pietarsaari Finland Helsinki, Finland Pietarsaari, Finland

1993

3.8f0.7 1.7fO.3 4.821.1 0.2fO.l 0.2fO.1 2.9f0.2 0.8f0.1

Landsort, Sweden 7 Utlangan, Sweden 8 Fladen, Sweden 9 S . Landsort, Sweden 11 12 Utklippan. Sweden bingskarrsklubb, Sweden 13 Fladen. Sweden

1991 1992 1992 1992 1992 1992 1992

1.6f0.1 2.1fO.1 4.4f0.4 2.220.5 9.6f3.3 5.2f0.1 3.2f0.3 3.0f1.4 3.920.6

1993 1993 1993

0.920.3 (n=101 6.420.8 (n=15) 5.3f0.7

3

Southern Baltic Bay of Gdansk Southern Baltic Bay of Gdansk

SE Gotland, Sweden Cod Plaice Capelin

collection

Iceland, 64'N, Iceland, 66'N, Iceland, 64'N,

14

15 18'W 16 24OW

24'W

(n=3) (n=41 (n=ll (n=7) (n=l) (n=51 h=8) (n=9) (n=l) (n=3) (n=l) (n=4)

(n=25) (n=25) (n=251 0.5+0.1 (n=25) 5f1 (n=25) 4fl (n=25) 1.3f0.3 (n=25) 0.9f0.2 (n=25)

(n=15) (n=15) (n=15) (n=151 (n=15) (n=15) (n=15) (n=61 (n=6)

n= number of individuals per pooled sample. Generally only 1 analysis. References: 1 = Skwarzec, 1988. 2 = Bojanowski, 1992. 3 = Kauranen and Miettinen, 1970. Uncertanities indicate total estimated e r r o r s (this work).

124 In t h i s study results f o r the Baltic Sea were compared t o those f o r other Nordic a r e a s i.e t h e Norwegian Sea and areas around Iceland. Table 1 gives t h e 210

results f o r

Po in fish muscle tissue compared with other results available

from t h e Baltic. 210

Po are in general about 1 t o a f e w Bq per

Activity concentrations of

kg dry weight or less than 1 Bq per kg wet weight. Values higher byalmost a f a c t o r of 10 were reported by Skwarzec (1988). These samples came from t h e Bay

of Gdansk. Significantly higher concentrations were found in blue mussels from t h e Swedish west coast (sites nos 10 and 14 on the map); 103f11 and 10357 Bq kg-'

137

(n= 30). The corresponding activity concentrations of

134

Cs and

Cs

were 2.6f0.8 and 4.150.8 Bq kg-' respectively. The following is t h e sequence of polonium concentration in t h e types of samples analyzed. Perch < Cod < Pike

As

mentioned

spectrometry.

The

radiocaesium

earlier

Whitefish the

5

Herring < Flounder << Mytilus

samples

results

are

given

concentrations

are

lower

were in

in

also

Table the

2.

North

measured It

is

Atlantic

by

gamma

obvious

that

(Iceland

and

Norway) than in the Baltic. This i s expected since water concentrations in t h e Baltic are about 80-100 mBq I-'

compared t o a few mBq I-'

in t h e North

Atiantic. Furthermore the salinity of the Baltic is lower thus enhancing t h e uptake of caesium. The highest activity concentrations a r e found in pike and perch, representing a higher trophic level.

DOSIMETRY The annual limit of intake f o r

137

Cs is 4 lo6 Bq and f o r

210

Po

lo5

Bq i.e 40

times lower f o r polonium than f o r the caesium isotope. The f a c t o r s f o r t h e Committed Effective Dose from intake of radionuclides a r e f o r zlOPo 4.3 Sv Bq-'

and f o r 137Cs 1.2 lo-* S v Bq-' (ICRP-30, NRPB-R245). A representative

value f o r t h e polonium concentration in fish from the Baltic Sea is about 3.6 Bq kg-'

dry weight and f o r 137Cs around 200 Bq kg-'

dry weight. This gives

t h a t t h e dose from 137Cs is about 50% higher than the dose from

210

Po by

consumption of fish from the Baltic Sea. For other Nordic marine areas t h e dose f r o m

210

Po i s about 30 times higher than t h a t from 137Cs.

125 Table 2.8.2. Radiocaesium in fish from different Nordic marine environments. Species

Place of collection

Site on map

Plaice Plaice Plaice Herring Herring Cod Cod

Oslo Fjord

4 More, Norway 5 Finnmark, Norway 6 Oslo Fjord Finnmark, Norway Oslo Fjord 21 More, Norway

Pike Pike Herring Perch Perch Whitefish Whitefish

Sipoo, Finland 20 Pietarsaari, Finland Helsinki, Finland Porkkala, Finland Pietarsaari, Finland Helsinki, Finland Pietarsaari, Finland

Herring Herring Herring Herring Herring Herring Herring Herring Cod Cod

Landsort, Sweden 7 Utlangan, Sweden 8 Fladen, Sweden 9 Haruf jarden, Sweden S . Landsort, Sweden Utklippan, Sweden h g k a r r s k l u b b , Sweden Fladen, Sweden Fladen, Sweden SE Gotland, Sweden

Cod Plaice Capelin

Iceland, 64'N 24'W Iceland, 66'N, 18'W Iceland, 64'N 24'W

19 18 17

22 11 12 13 14 15 16

cs

Year of collection

Bq kg-'d.w.

1992 1992 1992 1992 1992 1992 1992

3.121.0 3.7f0.8 3.4t0.7 5.4f0.8 2.0f0.9 8.1k1.2 8.1t1.4

137

1993 1993 1993 1993 1993 1993 1993

153f6 599f24 63f3 248f10 781t31 107f4 148t6

1991 1991 1991 1992 1992 1992 1992 1992 1992 1992

112f4 5923 7.5t0.4 9525 65f3 61f3 152f6 4.8f0.6 1321.1 89f4

1993 1993 1993

1.2f0.2 0.3fO.l 0.620.2

134

cs

Bq kg-'d.w.

-

8.6f0.5 35f1.6 3.7f0.3 1420.7 45t2 5.920.5 8.3f0.6 9.2f0.4 5.1f0.5

-

7.2f1.5 4.lf0.7 4.220.7 1OkO.9

-

6.1t0.5

-

REFERENCES Bojanowski, R., 1992. Personal communication and unpublished data. in t h e Cherry, C.D.; & Heyraud. M., 1988, Lead-210 and polonium-210 world's oceans, 1n:Inventories of Selected Radionuclides in t h e Oceans, IAEA-TECDOC-481, pp 139-158. El-Daoushy, F., 1988, The Pb-210 global cycle: Dating and tracing applications environmental radioactivity. In: Low-level measurements and their application t o environmental radioactivity, M. Garcia LBon and G. Madurga (Edsf, World Scientific, Singapore, pp 224- 273. Heyraud, M.; & Cherry, R.D., 1983. Correlation of 'loPo and 'loPb enrichments in the sea-surface microlayer with newton biomass. Continental Shelf Research 1 pp. 283-293 IAEA, 1985. Sediment K s and Concentration Factors f o r Radionuclides in t h e Marine Environment, IAEA Tech. Rep. Ser. No 247.

126 1982, Annals of t h e ICRP. Limits f o r Intakes of Radionuclides by Workers, P a r t 1, Pergamon Press, Oxford, United Kingdom. Kauranen. P. & Miettinen J.K., 1970, Polonium and radiolead in some aqueous ecosystems in Finland. In: Radioactive Foodchains in t h e S u b a r c t i c Environment, Annual Report, 1969, Department of Radiochemistry, University of Helsinki, paper no. 27. NRPB, 1991, Committed Equivalent organ Doses and Committed Effective Doses f r o m Intake of Radionuclides. A. W. Phipps, G.M. Kendall, J.W. S t a t h e r and T.P. Fell (Edsl. National Radiological Protection Board. NRPB-R245, Chilton, Didcot, United Kingdom. Shell, W.R.; Jokela, T.; & Eagle, R., 1973, Natural 'loPb and zlOPo in a marine environment. In: Radioactive Contamination of t h e Marine Environment, IAEA-SM-158/47. Skwarzec, B., 1988. Accumulation of 'loPo in selected species of Baltic fish. J. Environm. Radioact. 8 pp 111-118. ICRP,

127

2.9. RADIOCAESIUM AS ECOLOGICAL TRACER IN AQUATIC SYSTEMS A REVIEW

-

MARKUS MEILI Uppsala University, Inst. of Earth Sciences, Norbyvagen 18 B, S-752 36 Uppsala. Sweden

ABSTRACT A review is given of recent as well as pioneering studies based on the use of radiocesium as in-situ tracer of abiotic and biotic processes in aquatic systems. Special attention has been paid to dynamic aspects of lacustrine sediments and food webs. The review is focused on studies of I3’Cs in Northern Europe after the Chernobyl nuclear accident in 1986.

INTRODUCTION Radionuclides introduced into ecosystems can be used by geochemists and ecologists as in-situ tracers of processes that are difficult to assess by other means, either because they necessitate frequent and simultaneous measurement of many variables, or because they occur over large scales of time and space (e.g. Santschi et al. 1988, Santschi 1989, Santschi & Honeyman 1989). This was recognized already several decades ago when nuclear weapon testing were being tested, resulting in a world-wide dispersal of radioactive fallout. Two of the most common applications of radionuclides in aquatic ecosystems are as particle tracers, and as tracers for studying the trophic dynamics in organisms and food webs. SEDIMENT AND PARTICLE DYNAMICS Many radionuclides are metals which in natural waters have a high affinity for particulate matter. Caesium is often found in the clay fraction of sediments (Ritchie et al. 1973, Francis & Brinkley

1976, Evans et al. 1983, cf. Broberg 1994). Accordingly, 137Cs was earlier used as a tracer of particle transport and redistribution in lake waters and sediments (Ritchie et al. 1970, Edgington & Robbins 1976, 1990. Schell & Barnes 1986, Wan et al. 1987). In sediment studies, I3’Cs has been most commonly used as a time marker in order to measure or corroborate rates of sediment accumulation (Pennington et al. 1973, Ritchie et al. 1973. Robbins & Edgington 1975, Jaakkola et al. 1983). This approach may, however, he biased by various simultaneous processes, such as

I28

.

m

m

m

.

m 8

-rE

5

0

2

10

15 W

TI

..

I..

I

i

+ O

A

'

'

'

..m W

8 8

m

-fE

-

m

2

n

u

5 -

-

: 10.

15

-

Figure 2.9.1: Contamination of sediments with *3'Cs (Bq m-2) as a function of water depth in two forest lakes in central Sweden six years after the Chernobyl fallout. The initial deposition in 1986 was around 20 Bq m-* in both lakes. The water mixing depth during the summer period of stratification is about 3 in in Loppesjon and 2 m in Blacksktjam. After Meili et al. (1993).

129 vertical mixing and redistribution of sediments by mechanical forces (Nilton et al. 1986) or zoobenthic bioturbation (Robbins et al. 1977), or by secondary nuclide input from other parts of the lake (Eadie & Robbins 1987, Meili et al. 1993) or from the catchment area (Walling & He 1992). Postdepositional redistribution is of particular interest in soft organic sediments (Davis et al. 1986. Anderson et al. 1987). After the fallout from the Chernobyl nuclear accident, In7Csand In4Cswere used in several studies to trace the transport and redistribution of particles in aquatic systems. As expected, radiocaesium entering lakes was rapidly and strongly bound to particles (Santschi et al. 1988, 1990, Kansanen et al. 1991, Hammar et al. 1991, Schuler et al. 1991, Wieland et al. 1991, 1993, Robbins et al. 1992) and transported from the water co!wnn to the sediment (Meili 1988, Hammar et al. 1991, Kansanen et al. 1991, 1992, Broberg & Andersson 1991, Andersson & Meili 1994). Accordingly, Chernobyl 137Cshas been used, preferably in combination with other radionuclides, to estimate particle settling fluxes and velocities for particles of different type and origin (Schuler et al. 1991, Wieland et al. 1991, 1993, Robbins et al. 1992). From 137Cs activities in sediment traps and water columns, in-situ distribution coefficients (Kd,concentration ratio of In7Cs in particulate and in dissolved phase, in units of (L H2O) (kg dw)-l, or "dimensionless" as (kg H20) o99% of the 1 3 7 C ~ inventory in Nordic lakes (e.g. Broberg & Andersson 1991, Kansanen et al. 1991). despite potential remobilisation especially in soft organic sediments (Davis et al. 1986, Anderson et al. 1987, Heit & Miller 1987, Evans et al. 1983, Comans et al. 1989, Lindner et al. 1993). which are very common in Nordic lowland lakes. By comparing sediment inventories of 137Cswith the fallout in spring 1986, the efficiency of particle retention in lakes duiing periods of high water flow may be compared (Fig. 2.9. I ) . In

-

130

lakes with a long hydraulic residence time, the sediment inventory may be similar to the initial deposition on the lake surface, while it is usually less in lakes with a rapid water turnover (Meili et al. 1989. 1993, Kansanen et al. 1991). The same relationship applies to the initial contamination of fish (Hllanson et al. 1992). Man-made differences in hydraulic residence time may also p a d y explain the high inventories of 13’Cs observed in reservoirs as compared to natural lakes in a Swedish mountain area (Hammar et al. 1991). Studies of pre-Chernobyl 137Cs,which was deposited during all seasons, showed that inventories in soft-water lakes with highly organic sediments generally are lower than the cumulated deposition on the lake surface, whereas inventories of large reservoirs with large drainage areas are enhanced (Pennington et al. 1973, Heit & Miller 1987). This may be a combined effect of differences in particle input. particle retention and postdepositional rernobilisation of I3’Cs (Davis et al. 1986, Anderson et al. 1987, Heit & Miller 1987, Evans et al. 1983, Comans et al. 1989, Edgington & Robbins 1990, Lindner et al. 1993).

Figure 2.9.2: Decay-corrected activities of 134Cs from the Chernobyl fallout in sediment profiles from two Finnish lakes, together with fitted model curves. From Kansanen & Seppiilii (1992).

131 Sedimentary 137Csprofiles provide a measure of the active sediment pool size which can be derived from the thickness of thc surficial sediment laycr that is affected by mcchanical or biological processes of mixing (Eadie & Robbins 1987). This part of the sediment may interact with the water column and eventually with pelagic food chains by acting as a secondary source not only of 137Cs,but also of other particle-associated contaminants as well as nutrients. Vertical distributions of sedimentary Chernobyl 13’Cs have shown that the magnitude of this layer typically is in the order of a few mm or cm. but can vary a hundredfold even within lakes (Meili unpubl. data). Another important variable which can be quantified from sediment profiles of 137Cs is the actual burial rate of sedimentary contaminants and nutrients. This is the rate at which the concentration at the sediment surface, i.e. in the potentially bioavailable layer, decreases with time. This rate is determined by the combined rates of supply and mixing of both particles and contaminants (cf. Anderson et al. 1987). Kansanen & Seppala (1992) modified and applied a sediment mixing model (cf. Robbins & Edgington 1975, Robbins e t al. 1977, Anderson et al. 1987) to observed profiles of Chemobyl 137Cstaken three years after the fallout. The model can be used to calculate the expected vertical d i s t r i b u t i o n and mixing of 137Cs in lake sediments given a temporal input function, sedimentation rate, and mixing rates along the profile (Fig. 2.9.2). Such models are particularly useful today when profiles of Chernobyl I3’Cs can easily be compared both between and within lakes to assess patterns of horizontal and vertical particle movements and of mixing in different environments. However, sediment models still need to be improved to account for resuspension, compaction, density gradients, fluctuations in sedimentation rates, seasonal aspects, etc. It has repeatedly been found that already few months or years after fallout, 137Cs inventories show substantial horizontal variations, not only between different parts of lakes (Meili et al. 1989, Meili et al. 1993, Hongve e t al. 1993, cf. Eadie & Robbins 1987, Edgington & Robbins 1990) but even locally within small areas of accumulation bottoms (Kansanen et al. 1991). This indicates that particle sedimentation in lakes is not evenly distributed, which can be the result of particle focusing, near-shore wave action, intermittent complete mixing of the water column and other processes (Hilton e t al. 1986). The variations also imply that a large number of samples is required to obtain a reasonable estimate of a whole-lake inventory of e.g. 137Cs. Part of this spatial variation is caused by periodical r e s u s p e n s i o n of wind-exposed sediments in shallow water (Broberg & Anderson 1991, Meili et al. 1993). Sediment resuspension, which has been recognized to be important not only in small shallow lakes (Hilton 1985, Bengtsson et al. 1990, Evans 1993) but also in large lakes (e.g. Eadie et al. 1990), is followed by an evenly distributed settling of the same material, e.g. with respect to t37Csactivities (Meili et al. 1993). This process results in a net transport (focusing) of particles towards the centre of lakes where the accumulation of sediment as well as 137Csand other contaminants is most pronounced (e.g. Robbins & Edgington 1975, Edgington & Robbins 1976, 1990, Cahill & Steele 1986, review by Eadie & Robbins 1987). The ratio between maximum and mean inventory within a lake (max. focusing factor) provides a quantitative measure of lateral particle movement naturally integrated over

132 "C

Eq g-ldw 30

r 20

25 20 15

10

5 0

25

'

C

20 .

-

15.

20%

_c_____________

10. 10%

5. 0Apr

May

Jun

Jul

Aug 1986

Sep

Oct

Nov

Figure 2.9.3: Observed (*) and interpolated (-) activities of *3'Cs (Bq kg-1 dw) in zooplankton (>250 pm), and the surface water temperature ("C) (----) during the growing season after the Chernobyl nuclear accident in I986 (B) and (C): Observed activities of I3'Cs (Bq kg-I dw) in the muscle tissue of roach (B) and perch (C) (*), and predicted activities at different feeding rates (----). expressed as fractions of the maximum feeding rate (100%). The actual mean feeding rate was estimated at 14% of the maximum rate for roach Vertical bars show 95% confidence intervals of observed mean and at 38% for perch (-). activities, together with the number of observations. Data from Lake Blackslstjlm, a small mesohumic forest lake in central Sweden. After Meili (1991).

133 months or years (Edgington & Robbins 1990). Focusing of radionuclides was confirmed after Chemobyl from mass balance calculations in large lakes (Wieland et al. 1991, 1993). A sediment survey in spring 1992 showed a significant redistribution of Chernobyl 137Cseven in small shallow lakes (Meili et al. 1993, Fig. 2.9.1). In studies of other systems, 137Csfrom the Chemobyl accident has also been used as a tracer of fluvial particle transport in flood plains (Walling et al. 1992). into lakes (Walling & He 1992). and into coastal areas (Brydsten & Jansson 1989). In the dissolved phase, 137Cshas been used to quantify slow rates of vertical mixing of deep water bodies (Santschi & Honeyman 1989). Large-scale studies of marine currents have also been based on Chernobyl radionuclides (Aarkrog et al. 1988, Dahlgaard et al. 199 1).

FOOD WEB STRUCTURE AND DYNAMICS Caesium is readily taken up by aquatic biota and transferred along food chains, probably due to its similarity to potassium (SaxCn 1994, Andersson & Meili 1994). In fish, most 137Csis taken up from the food (e.g. King 1964). Even in brackish water, where distribution coefficients are much lower and the dissolved fraction higher than in lakes, the proportion of direct uptake from the water is <20% in invertebrate-feeding fish (Evans 1988) and probably even lower in fish predators. As a result, I3’Cs can be used in various ways to study the physiological turnover and trophic transfer of biomass in natural and experimental food webs. Such studies are usually based on laboratory studies of the physiology and kinetics of 137Cs in organisms, including the uptake from water by invertebrates (King 1964, Forseth & Nreumann 1993) and fish (e.g. Williams & Pickering 1961, King 1964. Hewett & Jefferies 1976). assimilation efficiencies for ingested 137Csin invertebrates (e.g. King 1964) and fish (e.g. Hewett & Jefferies 1978, Forseth et al. 1992) and elimination rates of accumulated 137Csin invertebrates (e.g. King 1964) and fish (e.g. Hewett & Jefferies 1976, Evans 1989, Mailhot et al. 1989, Ugedal et al. 1992). The use of radionuclides as tracers of trophic dynamics, i.e. biomass turnover and transfer, in contaminated aquatic ecosystems was proposed already several decades ago (Davis & Foster 1958. Odum & Golley 1963). It was also recognized early on that feeding habits largely control the uptake of 137Csin fish (Hannerz 1968). The first application of 137Csin freshwater was to assess natural feeding rates of fish with a mixed diet in a waste-contaminated pond (Kevern 1966). Under more controlled conditions, radionuclides have been used to assess feed intake rates (Storebakken et al. 1981) and food evacuation time (Peters & Hoss 1974) in fish. After the Chernobyl accident, 137Cswas used in two studies as in-situ tracer for assessing food consumption by fish in natural systems. The first study was based on the initial nonequilibrium phase immediately after the fallout (Meili 1991). Two different species of fish (roach and perch) of similar size, living in the same environment and feeding on a similar diet showed a significant difference in 137Csaccumulation, which could only be explained by different ingestion rates (cf. Fig. 2.9.3). From comparisons with previous laboratory data, both relative and absolute feeding rates could he calculated, which also explained the slow hut different growth rates.

134

Figure 2.9.4: Mean weight-specific daily ration (+SD)for brown trout at different months estimated by (a) the radioisotope method and (b) the Eggers gut volume method. Estimates from several years are pooled. The broken line denotes the water temperature. From Forseth et al. (1992).

135 The second study was based on observations at a later stage, i.e. during a period approaching steady state (Forseth et al. 1992). The food consumption of brown trout was quantified from 137Cs concentrations in the fish and its prey, with results comparable to estimates from a conventional method based on gut fullness (Fig. 2.9.4). After a single contamination event. a pollutant is gradually transferred between and distributed among ecosystem components. A comparison of the magnitude and timing of the resulting concentration pulses can be used to study the trophic structure of food webs. As a result of the serial coupling of trophic levels in food chains, the transfer of contaminants from the abiotic phase to large animals is subject to a delay which cannot be explained by a lower metabolic rate alone. The time to reach maximum activity therefore reflects the trophic level of animals within food webs (cf. Fig. 2.9.3, Meili 1991, Andersson & Meili 1994, Fig. 2.5.1). In the case of a continuous contamination (steady state), concentrations of pollutants can either increase or decrease along food chains, depending on the type of pollutant. Mercury and many organic pollutants increase in concentrations. whereas most heavy metals decrease. As a result, pollutants can be used to assess trophic levels in natural food webs (Meili 1991). In the case of 137Cs, however, the biomagnification factor (Cs in predator I C s in prey) varies around 1 or 2 (Thomann 1981, Mailhot et al. 1988, Rowan & Rasmussen 1994). Consequently, predators in a given ecosystem can often not be clearly distinguished from their prey based on 137Csconcentrations alone. Moreover, biomagnification factors vary both between and within systems, e.g. due to differences in turbidity, or due to differential enrichment between, within or along pelagic and benthic food chains.

REFERENCES SEDIMENT A N D PARTICLE DYNAMICS Aarkrog, A., L. Carlsson, Q.J. Chen, H. Dahlgaard, E. Holm, L. Huynh-Hgoc. L.H. Jensen, S.P. Nielsen and H. Nies (1988). Origin of technetium-99 and its use as a marine tracer. Nature 335, 338-340. Anderson, R.F., S.L. Schiff and R.H. Hesslein (1987). Determining sediment accumulation and mixing rates using *loPb, 13’Cs, and other tracers: Problems due to postdepositional mobility or coring artifacts. Can. J. Fish. Aquat. Sci. 44, 231-250. Andersson, T. (1993). Mercury and radiocesium in Swedish lakes. Ph.D. Thesis, Gerum 18, University of Umel, Sweden. Andersson, T. and M. Meili (1994). The role of lake-specific abiotic and biotic factors for the transfer of radiocaesium to fish. Chapter 2.5 in: Nordic Radioecology - The Transfer of Radionuclides through Nordic Ecosystems to Man. Ed. H. Dahlgaard. Elsevier Science Publ., Amsterdam (this volume). Bengtsson, L., T. Hellstriim and L. Rakoczi (1990). Redistribution of sediments in three Swedish lakes. Hydrobiologia 192, 167-181. Brittain, J.E., H.E. Bjamstad, B. Salbu and D.H. Oughton (1992). Winter transport of Chernobyl radionuclides from a montane catchment to an ice-covered lake. Analyst 117.515-519. Broberg, A. (1994). The distiibution and characterization of 137Cs in lake sediments. Chapter 2.3 in: Nordic Radioecology -The Transfer of Radionuclides through Nordic Ecosystems to Man. Ed. H. Dahlgaard. Elsevier Science Publ., Amsterdam (this volume).

136 Broberg, A. and E. Andersson (1991). Distribution and circulation of (3-137 in lake ecosystems. In: The Chernobyl Fallout in Sweden. Ed. L. Moberg. Swedish Radiation Protection Institute, Stockholm, pp. 151-175. Brydsten, L. and M. Jansson (1989). Studies of estuarine sediment dynamics using 137Csas a tracer. Estuar. Coast. Shelf Sci. 28, 249-259. Cahill, R.A. and J.D. Stcele (1986). 137Cs as a tracer of recent sedimentary processes in Lake Michigan. Hydrobiologia 143, 29-35. Comans, R.N.J., J.J. Middelburg, J. Zonderhuis, J.R.W. Woittiez, G.J. De Lange, H.A. Das and C.H. Van Der Weijden (1989). Mobilization of radiocaesium in pore water of lake sediments. Nature 339, 367-369. Dahlgaard, H.. Q.J. Chen and S.P. Nielsen (1991). Radioactive tracers in the Greenland sea. In: Radionuclides in the study of marine processes. Eds. P.J. Kershaw and D.S. Woodhead. Elsevier Applied Science, London, pp. 12-22. Davis, R.B., Hess, C.T., Norton, S.A., Hanson, D.W., Hoagland, K.D. and Anderson, D.S. (1986). 137Cs and 21°Pb dating of sediments from soft-water lakes in New England (U.S.A.) and Scandinavia, a failure of 137Cs dating. Chem. Geol. 44, 151 - 185. Eadie, B.J. and J.A. Robbins (1987). The role of particulate matter in the movement of contaminants in the Great Lakes. In: Sources and Fates of Aquatic Pollutants. Eds. R.A. Hites and S.J. Eisenreich. American Chemical Society, Washington, D.C., Advances in Chemistry Series 216, 319-364. Eadie, B.J., H.A. Vanderploeg, J.A. Robbins and G.L. Bell (1990). Significance of sediment resuspension and particle settling. In: Large Lakes: Ecological Structure and Function. Eds. M.M. Tilzer and C. Serruya. Springer Verlag. New York, pp. 196-209. Edgington, D.N. and J.A. Robbins (1976). Patterns of deposition of natural and fallout radionuclides in the sediments of Lake Michigan and their relation to limnological processes. In: Environmental Biogeochemistry, Vol. 2, Metal Transfer and Ecological Mass Balances. Ed. J.O. Nriagu. Ann Arbor Science Publ., Michigan, pp. 705-709. Edgington, D.N. and J.A. Robbins (1990). Time scales of sediment focusing in large lakes as revealed by measurements of fallout Cs-137. In: Large Lakes: Ecological Structure and Function. Eds. M.M. Tilzer and C. Serruya. Springer Verlag, New York, pp. 210-223. Evans, D.W., J.J. Alberts and R.A. Clark (1983). Reversible ion exchange fixation of I3’Cs leading to mobilization from reservoir sediments. Geochirn. Cosrmchim Acta 47, 1041- 1049. Evans, R.D. (1993). Various phenomena observed in lakes which suggest sediment resuspension as an important mechanism. Verh. Internat. Verein. Lirnriol. 25, 3 18-319 (Abstract) and Hydrobiologia (in press). Francis, C.W. and G.S. Brinkley (1976). Preferential adsorption of 137Csto micaceous minerals in contaminated freshwater sediments. Nature 260.51 1-513. Hammar, J., M. Notter and G. Neumann (1991). Radioaktivt cesium i rodingsjiiar - effekter av Tjernobylkatastrofen. (Radiocesium in arctic cham lakes - effects of the Chernobyl accident). Institute of Freshwater Research of the Swedish National Board of Fisheries, Drottningholm, Sweden, Report no 3/91, 152 p. (in Swedish, English summary). Heit, M. and K.M. Miller (1987). Cesium-137 sediment depth profiles and inventories in Adirondack Lake sediments. Biogeochemistry 3,243-265. Hilton, J. (1985). A conceptual framework for predicting the occurrence of sediment focusing and sediment redistribution in small lakes. Lirnnol. Oceanogr. 30, 1131-1 143. Hilton, J., J.P. Lishman and P.V. Allen (1986). The dominant processes of sediment distribution and focusing in a small, eutrophic, monomictic lake. Limnol. Oceanogr. 31, 125-133. Hongve. D., LA. Blakar and J.E. Brittain (1994). Radiocesium in the sediments of 0 v r e Heimdalsvatn, a Norwegian subalpine lake. Submitted to J. Environ. Radioact. Honeyman. B.D. and P.H. Santschi (1988). Critical review: Metals in aquatic systems. Predicting their scavenging residence times from laboratory data remains a challenge. Envir. Sci. Techno/. 22, 862-87 1. Haanson, L. and T. Andersson (1992). Remedial measures against radioactive caesium in Swedish lake fish after Chernobyl. Aqiraric Scierices 54, 141- 164.

I37 IAEA (1985). Sediment Kd's and concentration factors for radionuclides in the marine environment. International Atomic Energy Agency, Vienna, Techn. Rep. Ser. 247,73 p. Jaakkola, T., K. Tolonen, P. Huttunen and S. Leskinen (1983). The use of fallout 137Cs and 2399240Pufor dating of lake sediments. Hydrobiologia 103, 15-19. Kansanen, P.H., T. Jaakkola, S. Kulmala and R. Suutarinen (1991). Sedimentation and distribution of gamma-emitting radionuclides in bottom sediments nf southem Lake Paijanne, Finland, after the Chemobyl accident. Hydrobiologia 222, 12 I - 140. Kansanen, P.H. and J. SeppalS (1992). Interpretation of mixed sediment profiles by means of a sediment-mixing model and radioactive fallout. Hydrobiologia 243/244,37 1-379. Kansanen, P.H., T. Jaakkola. J. Seppala, M. Hiikka, R. Suutarinen and S. Kulmala (1992). Chemobyl fallout nuclides used as tracers of sedimentation and sediment mixing in four Finnish lakes. Proc. Int. Syrnp. on the Use of Isotope Techniques in Water Resources Development. IAEA (International Atomic Energy Agency), Vienna, Austria, 11-15 March 1991. Lindner, G., S. Kaminski, I. Greiner, M. Wunderer, J. Behrschmidt, G. SchrBder and S. Kress (1993). Interaction of dissolved radionuclides with organic matter in prealpine freshwater lakes. Verh. Intemat. Vereirz. Lirnnol. 25. 238-241. Meili, M. (1988). Radioactive caesium in Swedish forest lake ecosystems after Chernobyl: Zooplankton 1986, Sediment 1988. Proc. 5th Nordic Seminar on Radioecology (Rattvik, Sweden, August 22-26, 1988). Meili, M., A. Rudebeck, A. Brewer and J. Howard (1989). Cs-137 in Swedish forest lake sediments, 2 and 3 years after Chemobyl. In: The Radioecology of Natural and Artificial Radionuclides. Ed. W. Feldt. Verlag m V Rheinland GmbH, Kiiln, Germany, Publ. Ser. Prog. Radiat. Prot. 22, 306-311. Meili, M., K. Konitzer, L. Braf, S.B. Baines and T. Andersson (1993). Seesedimente als langfristige Sekundarquelle von 137Csin schwedischen Fischen (Lake sediments as a long-tern source of 137Csin Swedish fish). In: Urnweltradioaktivitat, Radiookologie, Strahlenwirkungen. Eds. M. Winter and A. Wicke. Verlag TUV Rheinland GmbH, Koln, Publ. Ser. Prog. Radiat. Prot. 25,617-621 (in German, English summary). Nyffeler, U.P., P.H. Santschi and Y.H. Li (1986). The relevance of scavenging kinetics to modeling of sediment-water interactions in natural waters. L i m o / . Oceanogr. 3 1,277-292. Pennington, W., R.S. Cambray and E.M. Fisher (1973). Observations on lake sediments using fallout 137Csas a tracer. Nature 242,324-326. Ritchie, J.C., M.R. McHenry, A.C. Gill and P.H. Hawks (1970). The use of fallout cesium- 137 as a tracer of sediment movement and deposition. Proc. Miss. Water Resour. Conf. 1970, pp. 149-162. Ritchie, J.C., M.R. McHenry and A.C. Gill (1973). Dating recent reservoir sediments. Limnol. Oceanogr. 18,254-263. Robbins, J.A. and D.N. Edgington (1975). Determination of recent sedimentation rates in Lake Michigan using Pb-210 and Cs- 137. G e o c h b . Cosnwchim Acta 39.285-304. Robbins, J.A., J.R. Krezoski and S.C. Mozley (1977). Radioactivity in sediments of the Great Lakes: Post-depositional redistribution by deposit-feeding organisms. Earth Planet. Sci. Lett. 36, 325-333. Robbins, J.A., G. Lindner, W. Pfeiffer, J. Kleiner, H.H. Stabel and P. Frenzel (1992). Epilimnetic scavenging of Chernobyl radionuclides in Lake Constance. Geochim. Cosmochim. Acta 56, 2339-236 1. Santschi, P.H. (1989). Use of radionuclides in the study of contaminant cycling processes. Hydrobiologia 1761177.307-320. Santschi, P.H. and B.D. Honeyman ( 1989). Radionuclides in aquatic environments. Radiat. Phys. Chem. 34, 213-240. Santschi, P.H., S . Bollhalder, K. Farrenkothen, A. Lueck. S. Zingg and M. Sturm (1988). Chernobyl radionuclides in the environment: Tracers for the tight coupling of atmospheric, terrestrial, and aquatic geochemical processes. Envir. Sci. Technol. 22, 510-5 16. Santschi, P.H., S. Bollhalder, S. Zingg, A. Lueck and K. Farrenkothen (1990). The self-cleaning capacity of surface waters after radioactive fallout - evidence from European waters after

138 Chernobyl, 1986-1988. Emir. Sci. Techno/. 24, 5 19-527. Saxen, R. (1994). Behaviour of 137Cs in lake basins. Chapter 2.4 in: Nordic Radioecology - The Transfer of Radionuclides through Nordic Ecosystems to Man. Ed. H. Dahlgaard. Elsevier Science Publ., Amsterdam (this volume). Schell. W.R. and R.S. Barnes (1986). Environmental isotope and anthropogenic tracers of recent lake sedimentation. In: Handbook of Environmental Isotope Geochemistry. Eds. P. Fritz and J.Ch. Fontes. Elsevier, Amsterdam, Vol. 2, pp. 169-206. Schuler. C., E. Wieland, P.H. Santschi, M. Stumm, A. Luck, S. Bollhaldzr, J. Beer, G. Bonani, H.J. Hoffmann, M. Suter and W. Wolfli (1991). A multitracer study of radionuclides in Lake Zurich, Switzerland. 1: Comparison of atmospheric and sedimentary fluxes of 7Be, l0Be, 2I0Pb, *loPo and 137Cs.J. Geophys. Res. 96. 17051-17065. Sundblad, B., U. Bergstrom and S. Evans (1991). Long term transfer of fallout nuclides from the terrestrial to the aquatic environment. Evaluation of ecological models. In: The Chernobyl Fallout in Sweden. Ed. L. Moberg. Swedish Radiation Protection Institute, Stockholm, pp. 207-23 8. Walling, D.E. and Q. He (1992). Interpretation of caesium-137 profiles in lacustrine and other sediments: the role of catchment-derived inputs. Hydrobiologia 2351236,219-230. Walling, D.E., T.A. Quine and J.S. Rowan (1992). Fluvial transport of Chernobyl fallout radionuclides. Hydrobiologia 235/236, 23 1-246. Wan, G.J., P.H. Santschi, M. Sturm, K. Farrenkothen, A. Lueck, E. Werth and C. Schuler (1987). Natural (ZlOPb, 7Be) and fallout (137Cs,239.24oPu, 9OSr) radionuclides as geochemical tracers of sedimentation in Greifensee, Switzerland. Chew. Geol. 63, 181-196. Wieland, E., P.H. Santschi and 1. Beer (1991). A multitracer study of radionuclides in Lake Zurich, Switzerland. 2. Residence times, removal processes, and sediment focusing. J. Geophys. Res. 96, 17067-17080. Wieland, E., P.H. Santschi, P. Hohener and M. Sturm (1993). Scavenging of Chemobyl 137Csand natural 210Pb in Lake Sempach, Switzerland. Geochim. Cosmochim. Acta 57,2959-2979. FOOD WEB STRUCTURE AND DYNAMICS Davis, J.J. and R.F. Foster (1958). Bioaccumulation of radioisotopes through aquatic food chains. Ecology 39,530-535. Evans, S. (1988). Accumulation of Chernobyl-related 137Cs by fish populations in the biotest basin, northern Baltic Sea. Studsvik Nuclear, Nykoping, Sweden, Studsvik Report NP-88/113,70 p. Evans, S. (1989). Biological half-time of Cs-137 in fish exposed to the Chernobyl fallout. Clearance of Cs-137 in roach exposed to various potassium concentrations in the water. An experimental study. Studsvik Nuclear, Nykoping, Sweden, Studsvik Report NP-89/74, 17 p. Forseth, T., 0. Ugedal, B. Jonsson, A. Langeland and 0. Njistad (1991). Radiocaesium turnover in Arctic charr (Salvelinus alpinus) and brown trout (Salmo trtitfa) in a Norwegian lake. J . Appl. Ecol. 28, 1053-1067. Forseth, T., B. Jonsson, R. Nreurnann and 0. Ugedal (1992). Radioisotope method for estimating food consumption by brown trout (Salmm truttu). Can. J. Fish. Aquat. Sci. 49, 1328-1335. Forseth, T. and R. Nreumann (1993). Radiocesium turnover in some freshwater invertebrates. Report to NKS (in Norwegian). Hammar, J., M. Notter and G. Neumann (1991). Radioaktivt cesium i r6dingsjb.r - effekter av Tjernobylkatastrofen. (Radiocesium in arctic charr lakes - effects of the Chernobyl accident). Institute of Freshwater Research of the Swedish National Board of Fisheries, Drottningholm, Sweden, Report no 3/91, 152 p. (in Swedish, English summary). Hammar, J., M. Notter and G. Neumann (1991). Northern reservoirs as sinks for Chernobyl cesium: Sustained accumulation via introduced Mysis relicta in arctic char and brown trout. In: The Chemobyl Fallout in Sweden. Ed. L. Moberg. Swedish Radiation Protection Institute, Stockholm, pp. 183-205. Hannerz. L. (1968). The role of feeding habits in the accumulation of fall-out 137Csin fish. Rep. Inst. Freshw. Res. Drottningholm 48. 112-119. Hewett, C.J. and D.F. Jefferies (1976). The accumulation of radioactive caesium from water by the brown trout (Sobno trrrtta) and its comparison with plaice and rays. J. Fish. B i d . 9. 479-489.

139 Hewett, C.J. and D.F. Jefferies (1978). The accumulation of radioactive caesium from food by the plaice (Pleuronectesplatessa) and the brown trout (Salmo tnttta). J. Fish. Biol. 13, 143-153. Kevern, N.R. (1966). Feeding rate of carp estimated by a radioisotopic method. Trans. Am. Fish. SOC. 95, 363-371. King, S.F. (1964). Uptake and transfer of cesium- 137 by Chlainydonwnas, Daphnia, and bluegill fingerlings. Ecology 45, 852-859. Mailhot, H., R.H. Peters and R.J. Cornett (1988). Bioaccumulation of cesium by aquatic organisms. Verh. Internat. Verein. Limnol. 23, 1602-1609. Mailhot, H., R.H. Peters and R.J. Cornett (1989). The biological half-time of radioactive Cs in poikilothennic and homeothermic animals. Health Physics 56,473-484. Meili, M. (1991). In situ assessment of trophic levels and transfer rates in aquatic food webs, using chronic (Hg) and pulsed (Chernobyl 13?Cs) environmental contaminants. Verh. Internot. Verein. Limnnl. 24, 2970-2975. Odum, G.P. and F.P. Golley (1963). Radioactive tracers as an aid to the measurement of energy flow at the population level in nature. In: Radioecology. Eds. W. Schultz and A.W. Klement Jr. Reinhold Publ., New York, pp. 403-411. Peters, D.S. and D.E.Hoss (1974). A radioisotopic method of measuring food evacuation time in fish. Trans. A m Fish. Soc. 103. 626-629. Rowan, D.J. & J.B. Rasmussen (1994). The bioaccumulation of radiocesium by fish: the influence of physico-chemical factors and trophic structure. Can. J. Fish. Aquat. Sci. (in press). Storebakken, T., E. Austreng and K. Steenberg (1981). A method for determination of feed intake in salmonids using radioactive isotopes.Aquaculture 24, 133- 142. Thomann, R.V. (1981). Equilibrium model of fate of microcontaminants in diverse aquatic food chains. Can. J. Fish. Aquat. Sci. 38, 280-296. Ugedal, 0..B. Jonsson, 0. Njlstad and R. Nzumann (1992). Effects of temperature and body size on radiocaesium retention in brown kout (Salmno trutta). Freshwar. B i d . 28, 165-171. Williams, L.G. and Q. Pickering (1961). Direct and food-chain uptake of cesium137 and strontium*' in bluegill fingerlings. Ecology 42,205-206.

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Chapter 3 AGRICULTURAL ECOSYSTEMS

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I43

3.1. INTRODUCTION TO RADIOECOLOGY OF THE AGRICULTURAL ECOSYSTEM

PER STRAND Norwegian Radiation Protection Authority, P.O.Box 55, N - 1345 Osterfis, Norway.

When agricultural areas are contaminated by radioactive fallout, the doses to the population from the intake of food will depend on a large number of factors, such as: Climate, soil conditions, agricultural production, local consumption of the food produced, the season of the fallout, official guidelines and control, countermeasures, etc. Looking at the geographical location of the Nordic countries, from latitude 5 5 O N (crossing the middle of Labrador) in the south of Denmark to 72' N in the north of Norway (which is about 560 km north of the Bering Strait), one would expect very poor climatic conditions for agriculture.

Thanks to the Gulf Stream this is not so, although the conditions'are not as good as further south. Over the centuries these Nordic conditions, especially in the northernmost parts and in the mountain areas, have led to a type of agriculture where available cultivatable areas are utilized intensively for their most economic purpose, depending on whether the natural conditions favour cultivation of different sorts of crops, animal husbandry or forestry. The climate and the geographical conditions vary considerably within the Nordic area itself. The northernmost part (north of 62" N) consists mainly of grassland and mountains and is dominated by animal husbandry, while the southern part is used primarily for cultivation of grain crops, sugar beet, potatoes, oil seeds, legumes, etc., although animal husbandry is important in the southern part too, especially in the valleys. Owing to the slow growing and short vegetation, the amount of fungi and lichens, and the intensive utilization of favourable areas for animal husbandry, the transfer of radioactive substances from the soil via the vegetation to animals is higher than in countries further south. Unfortunately, the weather conditions on the very day of the Chernobyl accident made the Scandinavian countries (especially Sweden and Norway) the ones most exposed to fallout outside the Soviet area. For these reasons, many of the local communities in the these two countries have been especially exposed to the effects of radioactive fallout, depending on the local structure of agriculture, and the accidental local concentration of the fallout.

144

The soil, the farms and the working methods are more or less the same all over the Nordic countries. However, local differences do exist, and this complicates the situation for each of the countries with regard to the effects of radioactive fallout on farm management and on agricultural production in the long term. In the short term the situation may be still more complicated, since the development of the crops and the time of ripening varies, depending on their geographical position in the climate zone. The largest differences between the Nordic countries as far as the effects of radioactive fallout are concerned, are the main structures and quantities of agricultural foodstuff and meat production, and how much of the production is consumed locally, combined with the official levels of acceptance in each country. One example is the consumption of relatively highly contaminated reindeer meat. However, each of the Nordic countries has imposed strict control of contaminated foodstuffs, and has taken action to bring the radioactivity levels in animals below the acceptance levels before slaughtering. During the first two years after the Chernobyl accident the amounts of condemned meat were substantial in Norway and Sweden and the costs high. Today these costs have been brought down to a minimum. Since the accident, the costs of measures to reduce the amount of condemned meat have been reduced year by year. The dominant part of food production in the Nordic countries is intensive. However, since the transfer of radiocaesium may be much higher in the marginal system, this could become the main contributor to the dose from food to groups of the population in some areas. Experience since the Chernobyl accident indicates that the intake of radioactivity through food is an important source of the dose to the population in the Nordic countries. In the long term the problems will be connected to contamination of various sensitive ecosystems. For the authorities it is important to have up to date knowledge of migration, transfer and fixation of radionuclides in the different ecosystems in the Nordic countries, to be able to implement the most effective countermeasures. It is also important to be able to give both the Nordic population and the world as a whole a description of the contamination situation. The main objective of this programme is to obtain better qualitative knowledge about the transport and uptake of radioisotopes (specially radiocaesium) in different agricultural ecosystems since the Chernobyl fallout. An important factor is the effect of the different agricultural parameters (e.g. soil type, fertilization) on the uptake of radiocaesium by the roots. The long-term behaviour of radiocaesium in the different Nordic ecosystems, and the sensitivity of different ecosystems to direct fallout, are of importance. The structure of the productive and the unproductive parts of the Nordic countries total area, and the structure of the production of different crops and animals, are shown in Figure 1-4 below.

I45 Research carried out over many years during and after the fallout from nuclear weapons testing has provided important knowledge about the behaviour of the various radionuclides in the different ecosystems. However, some topics became far more important after the Chernobyl accident than after the nuclear weapons tests. Accordingly plans for the RAD-3 project were directed towards topics generated by the fallout from the Chernobyl accident, rather than duplicating the research already carried out. Topics of special interest were defined as: 1. Food production in semi-natural ecosystems 2. Long-term behaviour of radionuclides and their mobility 3. Seasonality 4. Countermeasures

The research after the nuclear weapons testing was connected mainly to intensive agriculture production and the lichen-reindeer food chain. Less attention was given to cattle and sheep grazing

on uncultivated land, or to studies of the long-term behaviour of radiocaesium in the soil and the transfer of activity from soil to plants and further to animals after a single fallout. Therefore, one of the major products within the RAD-3 programme was to study the use of uncultivated land for food production. The transfer of activity in the chain soil-plants-sheep was to be studied at one particular spot in each of the Nordic countries over several years in order to describe the possible differences in transfer factors and the long-term behaviour of the radioisotopes within Nordic territory. It was hoped that this study would contribute to the understanding of important parameters that controlled the transfer and long-term behaviour of radiocaesium in the food chain. This study is described in the chapters 3.6 and 3.7. A similar study on a minor scale was also to be carried out on transfer of radiocaesium to cattle. This is described in chapter 3.5. The fallout from the Chernobyl accident differs from the faltout from the nuclear weapons tests in several ways. The most obvious difference is that the fallout from the nuclear weapons testing lasted over several years, while the fallout from the Chernobyl lasted for hours or days, depending on geographical location. Consequently the time of the year when the Chernobyl accident occurred was of marked significance for the activity levels in food. The influence of the time of the fallout is described in chapter 3.2. Since the Chernobyl fallout was concentrated to a short period of time, it created an opportunity to study the short and long-term behaviour of the transfer of radioisotopes in annual crop systems. The project focused on following five specific agricultural products in each of the Nordic countries: barley, potatoes, cabbages, carrots and peas. This project is described in chapter 3.4. The transfer and mobility of radionuclides within an ecosystem depend on the physico-

146

148

chemical form of the deposited radionuclides. The source of the fallout from the Chernobyl reactor accident was different from the weapons testing. This fact, and also the distance from the Chernobyl reactor, could result in different behaviour of for example radiocaesium. In that respect the influence of the physico-chemical form on the transfer of radiocaesium in the ecosystems is addressed in chapter 3.3. Countermeasures were taken to a much larger extent after the Chernobyl accident than were needed during the fallout period of the nuclear weapons tests. This made it necessary to develop countermeasures suited to the Chernobyl fallout situation, directed specifically at the semi-natural ecosystem. Chapter 3.8 contains a review of available countermeasures, focused mainly on those developed in the Nordic countries in connection with natural pastures. The long-term behaviour of the radioisotopes was estimated in terms of "effective ecological half-life (Teff)". This term is defined as the time needed to reduce the activity level in an ecosystem to one half it's original value. It consists of the physical half-life (T,) and the ecological half-life (TecJ according the equation: (Teff)-'= (TJ1 +(Teco)-'. Reasonable estimates of effective ecological half-lives depend on the results of many years of study, and a comparable sampling strategy is needed for the yearly monitoring, with regard to spot and time of the year. Therefore the use of this term involves many limitations when studying historical data. However, when studying the end products of an ecosystem there is no better way as yet of describing the observed behaviour of the radionuclides. An objective of the studies of long-term behaviour is the parameterization of the ecological half-life expression. Another term used is the "aggregated transfer coefficient (T,,)", which is used to describe the transfer of radionuclides from soil to the end product in an ecosystem. Tag(m2/kg) = activity in product (Bq/kg) / deposit (Bq/m2). This aggregated transfer coefficient is the product of the

transfer coefficient at all the transfer levels leading to the end product. The compiling of information on aggregated transfer coefficient for different food products, combined with respective effective ecological half-lives for relevant radionuclides and knowledge of which parameters are the most important for describing the Tag is necessary in order to assess the consequences of fallout on the Nordic countries.

This information, combined with

identification of important pathways of radionuclides to man and how one may alter the transfer by introducing countermeasures, will provide the information required to develop a management strategy for contaminated areas. This will be the knowledge base for how to reduce the dose to man, and optimize the situation.

149

3.2 DIRECT CONTAMINATION - SEASONALITY

ASKER AARKROG Rk0 National Laboratory, DK-4000 Roskilde, Denmark.

SUMMARY Direct contamination is the primary pathway to terrestrial vegetation in the first period after an activity release to the atmosphere. All radionuclides are able to be transferred via this pathway. Deposition, interception and retention are the three processes involved in direct contamination of crops. Wet deposition is more important than dry deposition in temperate regions. Resuspension and rainsplash both belong to secondary direct deposition and became evident for e.g. radiocaesium after the Chernobyl accident. Seasonality is the varying response to radioactive contamination of crops according to the time of the year when the contamination occurs. Shorlived radionuclides (as I3’I) and those that mainly enter the foodchain by direct contamination (e.g. I3’Cs) are especially important in this connection. In particular, the contamination of cereal crops is influenced by seasonality. As a result of seasonality the impact of the Chernobyl accident on the radioactive contamination of human diet was for the same deposition density higher in southern than in northern Europe. INTRODUCTION Plants are the primary recipients of radioactive contamination to the food chain from the abiotic environment. Vegetation may be subject to

direct and

indirect contamination. In the case of

terrestrial vegetation, direct contamination implies an uptake of radioactive debris from the atmosphere by the above-ground parts of plants. Indirect contamination is a sorption of debris from the soil by the root system of plants. The direct contamination of plants may be of two types: a primary one involving transfer direct from the source via the atmosphere to the plants, and a secondary one by which activity already deposited on the ground may be resuspended, e.g. by the wind, and thus be transferred to the plants. The response to direct contamination of the plants depends on the development of the plants at the time of contamination. In this context the concept seasonality is defined as the varying response to radioactive contamination of vegetation according to the time of the year when the contamination occurs.

150 Direct contamination Transfer of radionuclides from the atmosphere to vegetation involves three processes: DeDosition, interceution and retention. In the case of primary direct deposition these processes are the only ones involved. For secondary direct deposition resuspension should also be taken into account. Deposition of radionuclides from the atmosphere may occur either as

& or as

wet

deposition. Dry deposition occurs continuously, while wet deposition involves the intervention of rain or some other form of precipitation. Dry deposition is usually described applying the deposition velocity, (Vg), (Chamberlain 1960)

vg = -

F C

where F is the fallout rate of the depositing radionuclide to a unit area of land, (Bq m-' s-'), and C is the concentration in ground level air over the area of land considered, (Bq m-3).The SI unit of Vg thus becomes m s-', but cm

s-l

is often used instead.

The deposition velocity varies with the aerodynamic diameter of the particles deposited. Particles with a diameter of between 0.1 and 1 pm show a Vg of about 0.02 cm s-'; those between 1 and 10 pm show a deposition velocity increasing from 0.02 to about 5 cm s-' (Harrison 1993).

It also varies with the type of surface and with the chemical and physical characteristics of the radioelements involved. This is illustrated in Table 3.2.1. Table 3.2.1.

Radionuclide

Deposition velocities (cm s-') measured at Rise, in April 1986 after the Chernobyl accident. (Aarkrog, 1988") Rain collector (smooth plastic surface)

90Sr '2r lo3Ru

0.10 0.08 0.05

'37Cs I4%a I4Ve 239Np

0.004 0.07 0.08 0.07

1311

Grass field 0.14 0.08 0.04 0.3 0.04 0.07 0.09 0.07

Apart from 137Cs,for which the deposition velocity on the smooth surface of the raincollector was an order of magnitude less than that of the grass, it appears that the two surfaces showed nearly the same deposition velocities.

151 Caesium and ruthenium showed in general lower deposition velocities than the other elements. This was also seen in Sweden for grass samples collected 28-30 May, 1986 (Devell, 1991), but the Swedish deposition velocities were nearly an order of magnitude higher than the Danish ones for April. The deposition velocity of global fallout 90Srin Denmark during 1962-1974 was determined to 0.02 cm s-' (Aarkrog, 1979). Wet deposition occurs during precipitation. The "wash out ratio" W (Engelmann, 1971) is defined as the ratio between the radionuclide concentration in precipitation in Bq I-' and the concentration in ground level air in Bq m3. For global fallout W is about 1000 (Aarkrog, 1979), but for Chernobyl 137Csthe ratio was about 1.4-2 times higher (Aarkrog et al., 1988). Experience from global fallout studies has shown that around 90% of the total deposition of 90Sr and 137Cshas occurred as wet fallout. In the case of an accident the major deposition usually takes place within a few days. The Chernobyl accident demonstrated that high rainfall during the cloud passage results in deposition rates an order of magnitude higher than those observed for dry conditions (Edvarson, 1991). Interception is the process by which the radionuclides deposited are caught by the vegetation. It depends both on the physical characteristics of the deposit and the growth form of the plants. The subsequent fate of the deposit, i.e. the retention, is influenced by these factors and by the rate at which the material is removed by precipitation and other processes, called weathering or field loss. The fraction (f) of material intercepted by the crop canopy was expressed by Chamberlain (1970) as f =

1 - e-fifl

y is an empirical parameter, depending on the physic0 chemical properties of the deposit,

the manner of deposition, the morphology of the crop and the meteorological conditions. R is the dry weight of crop per unit area (kg rne2). p

usually falls within the range of 0.2 to 4 m2 kg-'

(Chamberlain and Garland, 1991). For moderate values of pB f = pR or f/B = p Chamberlain (1970) has defined the so-called "normalized specific activity" NSA as (Bq kg-' dry weight of crop) . (Bq d-' m-2of ground).'. NSA is thus a rate factor with the unit m2kg-' d. Values between 20 and 40 m2kg-' d have been observed for 137Csand 90Srfor herbage in good growing

152

conditions by Eriksson (1977) and Aarkrog (1979). Chamberlain found that winter grass showed 2-3 times higher NSA values than summer grass. The weathering or field loss is expressed by M/Mo = e-'7 where Mo and M are the quantities retained on the crop initially and after time t, and r is an empirical constant. During the growing season r is in the order of 2 weeks, in the winter period r increases to about 8 weeks. In case of rain the field halflife may be short. This is illustrated in Fig. 3.2.1. showing a comparison between predicted and observed 137Cslevels in grass after the Chernobyl accident. According to the model, the field halflife during rain was only 2 days compared with 14 days for dry conditions. One of the important lessons learnt from the Chernobyl accident was the importance of resuspension, which resulted in secondarv direct contamination of the crops. The Chernobyl accident was a tropospheric event, i.e. the debris was injected into the lower part of the atmosphere. The mean residence time in the troposphere is 20 days (UNSCEAR, 1962). Hence it was to be expected that the Chernobyl fallout had decreased to insignificant levels within a few months after the accident. This was not the case, however.

150

z

3

100

9

n

W

0

> t Y W

0

I

I

50 O/O

/

/ O

0

0

0 0

50 100 PREDICTED VALUES

150

Figure 3.2.1. Comparison between model predictions and observations for Bq '37Csm-' grass at R i s ~in May 1986. The model assumes an initial uptake in grass of 28% of the deposition. Furthermore, it is assumed that the field loss during rain corresponds to a field halflife of 2 days. During dry weather the halflife is 14 days. Aarkrog et al., 1988)

153 The resuspention factor RF is defined as the atmospheric radionuclide concentration in Bq

m-3divided by the ground contamination in Bq m-'. RF thus has the unit m-'. Figure 3.2.2. shows that the resuspension factor measured at Risa and on Bornholm since the Chernobyl accident has decreased according to a Power function.

0.5

I 100

I

I

I

I

200

500

1000

2000

Days since 7 May 1986 Figure 3.2.2. Caesium-137 resuspension factors after Chemobyl, July 1986 - December 1991 (quarterly values). R F = 9 . 3 ~ 1 0D-' - ~ 17. (D=days). Aarkrog et al., 1992)

Figure 3.2.3. shows how the deposition of '"Cs has decreased in Denmark (10 stations), the Faroe Islands (2 stations) and Greenland (1 station) since the Chernobyl accident. From July to December 1986 about 10% of the original deposition from the Chernobyl accident was redeposited by resuspension. This redeposition had decreased to about 1 % by the end of 1991, i.e. 5.5 years after the accident. It appears that the availability of resuspended

is less than that of primarily deposited

activity (Aarkrog, 1989), i.e. the transfer factor for primary direct contamination is higher than that of secondary direct contamination. There may be two reasons for the lower availability for resuspended particles compared with directly deposited fallout. Firstly, we must expect a higher field loss than for global fallout. Secondly, I3'Cs adheres, e.g. to clay minerals, and the suspended material may make the radiocaesium less available for absorption by the crops and thus for translocation to the grain. A

1 54

special case of secondary direct contamination of crops is rain-splash which may occur during heavy showers, when the rain drops by recoil carry contaminated soil to the surface of the vegetation. As to resuspension this manner of direct contamination is expected to be less efficient with regard to Contamination of the plants than is the primary one.

3E-01

-

DENMARK

FAROE ISLANDS

GREENLAND

0

A

0

f

f A H

A

O

H

"

0

0

A

I

1E-04 100

I

I

1

200

300

500

A

A

I 1,000

I

2,000

3,000

Days since 26 April 1986 Figure 3.2.3. Wet deposition of Cs-137 relative to Chernobyl deposit. (1 is Kannikeghd on Bornholm, 5 the five stations in Jutland and 4 the four stations on the Islands). (Aarkrog et al., 1992)

Figure 3.2.4. Summarizes the various types of direct contamination.

Direct contamination Primary

Figure 3.2.4. Types of direct contamination

Secondary

155

Seasonality The outcome of direct contamination of crops depends heavily on the season of the year when the contamination occurs. Seasonality means the varying response to radioactive contamination of environmental samples according to the time of the year when the contamination occurs. The seasonalitv factor 6)is the relative standard deviation (RSD) of the monthly time integral concentrations of a given radionuclide in a given sample observed throughout 12 months of a year. A total deposition of I Bq m-’ in a month (i) of radionuclide (p) results in an infinite time

concentration integral in sample item (a) of I,, Bq kg-’ year. The variation of Iipa between the 12 months of the year is a measure of the degree of seasonality. This variation is quantified as the RSD, i.e., the RSD of the 12 Iips values, which have an annual mean of I,,. Hence,

If S > 0, seasonality is present. If S = 0, no seasonality is observed. A deposition event during winter, when the fields lie fallow and livestock are housed, will result in a significantly lower radiological impact than if similar contamination were to take place on a mature crop in summer. Seasonality is generally of greater importance in temperate regions like Scandinavia than in the sub-tropics where soil can be cultivated throughout the year. Seasonality is of particular importance in connection with the contamination of cereals. The effect of seasonality on the contamination of grains has been studied by Middleton (1959) and Aarkrog (1983). It appears that the two important factors influencing contamination of the grain are the initial retention and the translocation from the vegetative parts of the seeds. Initial retention is largely independent of which radionuclide is involved in a soluble form, whereas translocation depends strongly on the radioelement. The fraction of initial intake 3 months before harvest is about 5% of the activity deposited over a field with a barley crop, and 1 month before harvest it reaches a maximum of 36%. If the field is contaminated with, for example, 137Cs,3 months before harvest (1 kBq m-’), the mature

156

grain will contain 2 Bq kg-*;if the contamination occurs 1 month before harvest, 100 Bq kg-I is found in the mature grain (Figure 3.2.5 .). For 90Sr,the corresponding concentrations would have been 0 and 20 Bq kg-I, respectively, demonstrating that i37Csis translocated to a much greater extent than is 90Sr(Figure 3.2.6.).

5

10’ I

100

$ r/)

0

1

lo-’

Q 0

50 100 DAYS BEFORE HARVEST

Figure 3.2.5. Percentage (p Cs) of radiocaesium, applied at various dates to 1 m2 of a barley field, recovered per kilogram of mature grain at harvest assuming no radioactive decay (number of determinations and 1 standard error are shown). Curve calculated from p(t) = 9 . 8 ~ 1 0 e-00013(t-34’? -~

157

10' (

100

$ lo-' L

cn

1

lo-*

T\

lo-;

0

50 100 DAYS BEFORE HARVEST

Figure 3.2.6. Percentage ( p Sr) of strontium, applied at various dates to 1 m2 of a barley field, recovered per kilogram of mature grain at harvest assuming no radioactive decay (number of determinationsand 1 standard error are shown) unless the standard error is less than the radius of the circle. Curve calculated from p(t) = 4 . 5 ~ 1 0 .e-o.ooos5a-2)2. ~

158

The highest levels in grain are expected when the contamination takes place in the final month before harvest. The lowest levels will be seen if the fields are contaminated before sprouting. In general, the Chernobyl accident confirmed qualitatively the observations on seasonality made before the accident. With respect to milk (Fig. 3.2.7.), it was seen that the radioiodine concentrations were low in northern Europe where the grazing season was just beginning, whereas they were significantly higher in southern Europe where the cows had been grazing for months. This latitudinal effect, which is a result of seasonality, was also seen for 137Csin grain. (Fig. 3.2.8.) (UNSCEAR, 1988)

Figure 3.2.7. Integrated concentrations of iodine-131 in milk per unit iodine-131 deposition density. Based on UNSCEAR (1988).

159

In Denmark the relative 137Csconcentrations of the grain species (rye, barley, wheat and oats) were compared with those observed in the years before the Chernobyl accident. (Aarkrog, 1988b). It appeared that the rye levels were an order of magnitude higher than those seen in the other species after the Chernobyl accident. This observation reflected the precocity of rye compared with the other species, resulting in a higher retention of '37Csby the more developed rye crops when the fallout from Chernobyl was deposited. UNSCEAR (1988) calculated the transfer factor (Bq I3'Cs kg-' year per kl3q m-* '37Cs)in various diets after the Chernobyl accident. It appears that the factors were usually lower than those observed previously for global fallout of 137Cs,especially for northern Europe. One reason for this was seasonality, which in particular influences the levels for grain and for which the transfer factors post-Chernobyl were 1-2 orders of magnitude lower than those observed for global fallout. Another reason could be a lower availability of the Chernobyl 13'Cs than of that from global fallout with respect to contamination of crops. However, seasonality seems to be the most important factor. 110

10 n

cv

Grain

-5

€ 0-

5

b Q

2

- 2

0

1

11

\

rn x

5 )r

- 0.5

0- 0.5

rn

W

a 0

Jr

- 0.2

0.2

4 0.1

1 0.1

LL

0.05

-

0.02

- 0.02

0.05

LL

cn

z 4 a I-

0.01

20

I

I

I

I

30

40

50

60

'

0.01

70

LATITUDE (ON) Figure 3.2.8. Integrated concentrations of caesium-137 in grain in the first year after the accident per unit caesium-137 deposition density. Based on UNSCEAR (1988).

160 Prohl (1990) calculated 50 year time integral concentrations (hap values) for a number of agricultural products. Figures 3.2.9. and 3.2.10. show the results of those calculations for contamination of rye with "Sr and 137Csrespectively in the various months of the year. Prohl's calculations are based on a time-dependent simulation model SINK, which is related to the socalled ECOSYS model (Prohl et al., 1988). Table 3.2.2. shows examples of the seasonality factor calculated from Prohls (1990) 50 year time integral concentrations.

Table 3.2.2. Some S values calculated from Prohl (1990)

Milk Wheat Root vegetables Leaf vegetables

1.1

0.4 0.5

2.6

0 0.4

0.2

0.8

1.9 1.2 0.5

The degree of seasonality depends strongly on the ratio between indirect and direct contamination of crops. If this ratio is low there is a high seasonality and vice versa. It should be stressed that the seasonality factors shown in table 3.2.2. are based on German data and they may thus not be of general applicability. It is, however, believed that with reasonable approximation the figures are valid for Denmark and southern Sweden. It should also be noted that for certain products, e.g., cereal grain, there could even be a monthly seasonality variation, but for practical reasons the definition of the seasonality factor has been based on monthly data.

161

Figure 3.2.9. The 50 year timeintegrated 90Sr concentration in rye related to the month of deposition of 1 Bq 90Srm-2 in connection with l m m precipitation. The white part of the columns is the contribution from direct contamination and the dark part represents indirect contamination (root-uptake) (after Prohl, 1990).

162

0

Time integrated concentration ( Bq y/kg ) 0.05 0.1 0.15

0.2 Cs-137,Rye

Mar

1;: ? Jul

Figure 3.2.10.

The 50 year time integrated '37Csconcentration in rye related to the month of deposition of 1 Bq 13'Cs m-* in connection with l m m precipitation. The white part of the columns is the contribution from direct contamination and the dark part represents indirect contamination (root-uptake) (after Prohl, 1990).

REFERENCES Aarkrog A. Environmental Studies on Radioecological Sensitivity and Variability with Special Emphasis on the Fallout Nuclides "Sr and 137Cs.Riss-R-437.( 1979). Aarkrog A. Translocation of Radionuclides in cereal crops. Ecological Aspects of Radionuclide Release, edited by P.J. Coughtrey, J.N. Bell and T.M. Roberts (Blackwell Scientific Publications, Oxford), pp 81-90. (1983) Aarkrog A. Studies on Chernobyl Debris in Denmark, Environment International 14, 149-155. (1988a) Aarkrog A. The radiological impact of the Chernobyl debris compared with that from nuclear weapons fallout. Journal of Environmental Radioactivity, 6 , 151-162. (1988b) Aarkrog A., Bstter-Jensen L., Chen Qing Jiang, Dahlgaard H., Hansen H., Holm E., Lauridsen B., Nielsen S.P. & Ssgaard-Hansen J. Environmental Radioactivity in Denmark in 1986. RissR-549, pp 272. Riss National Laboratory, Roskilde, Denmark. (1988) Aarkrog A. Radioecological Lessons Learned from Chernobyl, Proceeding of the XVth Regional Congress of IPRA, Visby, Gotland, Sweden, 10-14 Sept., 1989, 129-134. (1989) Aarkrog A . , Bstter-Jensen L., Chen Qing Jiang, Dahlgaard H.,Hansen H., Holm E., Lauridsen B., Nielsen S.P., Strandberg M. & Ssgaard-Hansen J. Environmental Radioactivity in Denmark in 1990 and 1991. Rise-R-621, pp 173. Riss National Laboratory, Roskilde, Denmark. (1992)

163 Chamberlain A.C. Aspects of the deposition of radioactive and other gases and particles. Int. J. Air Pollut. 3, 63-88. (1960) Chamberlain A.C. Interception and retention of radioactive aerosols by vegetation. Atmos. Environ. 4, 57-78. (1970) Chamberlain A.C. and Garland J. A. Interception and retention of radioactive fallout by vegetation. IAEA Technology, Harwell, UK, Report AERE R 13826. (1991) Devell L. Composition and properties of the plume and fallout materials. In Moberg, L.(Ed) The Chernobyl Fallout in Sweden - Results from a Reseach Programme on Environmental Radiology: 29-46. Swedish Radiation Protection Institute. (1991) Edvarson K. Fallout over Sweden from the Chernobyl accident p 47-65 in The Chernobyl Fallout in Sweden edited by L. Moberg. The Swedish Radiation Protection Institute. Stockholm.( 1991) Engelmann R.J. Scavenging prediction using ratios of concentrations in air and precipitation. J. Appl. Meteorol. 1Q,493-497. (1971) Eriksson A. Fissionsprodukter i svensk milj0. (Inst. f. Radiobiologi Lantbrukshragskolan, Uppsala) 97 pp. (1977) Harrison R.M. (co-ordinator): Atmospheric pathways p. 56- 100 in Radioecology after Chernobyl. Edited by Sir Frederick Warner & Roy M. Harrison. John Wiley & Son, Chichester. (1993) Middleton L.J. Radioactive strontium and caesium in the edible parts of crop plants after foliar contamination. Int. J . Radiat. Biol. 4 387-402. (1959) Prohl G., Muller H., Jacob P. and Paretzke H.G. Paper presented at the 4th International Symposium on Radioecology, Cadarache, France, March 14-18, Centre d’Etudes Nuclhires de Cadarache, pp. B43-B50. (1990) Prohl G. Modellierung der Radionuklidausbreitungin Nahrungsketten nach Deposition von Sr-90, Cs-137 und 5-13 1 auf landwirtschaflich genutzte Flachen. GSF-Bericht 29/90. Neuherberg. pp 180. (1990) UNSCEAR Report of the United Nations Scientific Committee on the Effects of Atomic Radiation. United Nations, New York. (1962) UNSCEAR Sources, effects, and risks of ionizing radiation. United Nations, New York. (1988)

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165

3.3. I N F L W C E OF PHYSICO-CHEMICAL FORMS ON TRANSFER

DEBORAH H. OUGHTON AND BRIT SALBU Laboratory for Analytical Chemistry, Agricultural University of Norway, P.O.Box 5026, N-1432 As, Norway. SUMMARY The transfer and mobility of radionuclides within an ecosystem will be dependant on the physico-chemical forms of the deposited radionuclides. After a short summary of the influence of speciation and source term on the behaviour of anthropgenic radionuclides in aquatic and terrestrial systems, this paper focuses on appropriate sampling and analytical techniques that are applicable for speciation purposes. Particular emphasis is given to the significance of fuel particles in the releases from the Chernobyl reactor and, as an example of the application of the discussed techniques, results are presented from studies on 137Csand "Sr in soils collected from Nordic countries, and compared with results from soils collected in the SUS. The topics addressed include inhomogeneous deposition, the vertical distribution of radionuclides in soils, the use of sequential extraction to study the association of radionuclides with soil components, the use of NH,Ac extraction to distinguish between the mobile or labile fraction and the inert fraction of radionuclides in soils, aggregated transfer factors, and comparison of the distribution of radionuclides with naturally occumng stable isotopes. INTRODUCTION When artificially produced radionuclides enter the environment due to authorized or accidental releases from nuclear installations, information on transfer within the affected ecosystems is needed for assessing the short and long-term consequences. One of the major factors influencing transfer is the source term, namely the composition, activity levels, activity ratios and physico-chemical forms of the released or deposited radionuclides. Activity concentrations and speciation will depend on the release scenario, course of events, distance from source and on factors influencing the dispersion in different media (e.g. marine systems or atmosphere). Information on the source term is usually restricted to the total activities of radionuclides released (inventories) or deposited (Bq m-'), while information on speciation is scarce. However, the radionuclides may be present in different physico-chemical forms varying in size (nominal molecular weight), charge, redox or density properties. The mobility and biological uptake of radionuclides are influenced by their speciation, as low molecular weight (LMW)

166

simple species are more mobile and bioavailable than high molecular weight (HMW) forms, colloids and particles (Figure 3.3.1). However, in aquatic systems the retention of colloids and particles may be an important pathway for filtering organisms. After entering the environment, the original distribution of species can be altered with time due to interactions with naturally occurring components (e.g. through sorption or complexation) and as a result of climatic conditions (e.g. weathering processes). The transfer of radionuclides in aquatic or terrestrial ecosystems will, therefore, depend on the original distribution of species, transformation processes and kinetic parameters.

In aquatic systems, a number of fractionation techniques are available for separating trace elements into different physico-chemical forms (Salbu, 1985). Based on size-fractionation, a major fraction of activation products and transuranics in discharges from nuclear installations are present in HMW forms (Salbu et.al., 1987; Salbu et.al.,1993). When effluents enter coastal waters, interactions occur in the mixing zone, for example aggregated particles are removed by sedimentation, while transport to distant areas is attributed to the mobile fraction. However, the speciation and mobility of radionuclides in sediments can be influenced by changes in physical conditions (e.g. redox), chemical interactions (e.g. organic ligands) and bioturbation, and such processes can lead to a release of radionuclides from sediments.

ENVIRONMENTAL EFFECTS

t

BIOLOGICAL EFFECTS f accumulation f(t) BIOLOGICAL UPTAKE

t

Kl

BIOAVAILABLE SPECIES low molecular weight forms

K,

T

INERT SPECIES high molecular weight forms TRANSPORT SOIL/SEDIMENT INTERACTIONS

SOURCE FIGURE 3.3.1. Size distribution and bioavailable species (Salbu, 1988a).

167

In terrestrial ecosystems contaminated by atmospheric fallout, the characterization of readily labile (where labile is defined as being able to change chemically) or bioavailable species, or categories of species, of radionuclides is rather complicated. Information on categories of species can be obtained by distinguishing particles in fallout during the event, using cascade impactors or by fractionation of rainwaters (Salbu, 1988b), or after the deposition, by fractionation of soil waters or extraction of the reversible fractions in soil-waterplant systems (Figure 3.3.2). Changes, with time, either in observed transfer factors or in the fraction of radionuclides present in a readily labile form indicate that transformation processes affecting the mobility and bioavailability of radionuclides have occurred. Following the Chernobyl accident, radionuclides were released in different physicochemical forms, e.g. volatiles, condensed and fuel particles (IAEA, 1991). The activity levels and the relative distribution of species depended, among other factors, on the distance from the

source and climatic conditions (dry and wet deposition). Close to the source of release, the deposition of fuel particles was high, (Loshchilov et al., 1992), and the "relative biologic availability" of fallout 137Cswas low compared to ionic tracer 137Cs(Bondar et al., 1992). In soils containing fuel particles, the extractability of "Sr is low (Oughton et al., 1993), and the behaviour of 137Csand wSr is quite different from that of the corresponding natural stable isotopes (Oughton et al., 1992).

+

Biological uptake

-

4............................................................................................................

......................................

"

Run-off

I

I

.%.

REVERSIBLY BOUND lo" exchange

MOBILE FRACTION

k4

'-' IRREVERSIBLY BOUND Strong fixation

INERT FRACTION

FIGURE 3.3.2. Labile and Inert Fractions of Radionuclides in Soils.

168

With increasing distance from Chemobyl, a decrease is observed in: total activity levels; the relative activities of %r, refractory fission products and actinides; the size and number of fuel particles; and the fraction of radionuclides associated with fuel particles (Bovoroy et al., 1992; Oughton et al. 1992; Loshchilov et al., 1992). Condensed particles, largely composed of the more volatile nuclides (e.g. 137Csand '"Cs), formed a major source of contamination outside the 30 km exclusion zone. However, small fuel particles (diameter of a few pm) were carried over considerable distances and found in Sweden and Norway at distances of up to 1500 km from the reactor (Devell, 1986; Salbu, 1988b).

After deposition, the long-lived radionuclides will be subject to time-dependent transformation processes which may change the original distribution of species. Mobilization of radionuclides associated with fuel particles may occur due to weathering. Then mobilized ionic species may be available for biological uptake, may be transferred with run-off or percolating waters, or may associate with soil components and be retained in the surface soil layers (Figure 3.3.2). Within the 30 km zone, the rate of 137Csfixation in clay soils is greater than the rate of release from particles, while for %Sr the rate of fixation in soils is slower than the rate of release from particles and, hence, the mobility and the biological uptake of %Sr increases with time (Bobovnikova et al., 1991). However, the rate of weathering can be influenced by the physico-chemical properties of the soil, i.e. pH, Eh, organic content, microbial activity. Even in areas of considerable distance from Chemobyl (e.g. Scandinavia), the mobility and bioavailability of 137Csand 9 r in soil-water systems has changed since deposition in 1986. Feeding experiments and in vitro studies with rumen liquid have also revealed that the bioavailability of fallout 137Csin vegetation was low in 1986/1987, but increased with time after deposition (Hansen and Hove, 1991; Salbu et al., 1992). Furthermore, the transfer to animals from fallout 137Csassociated with soil components (surface contamination) is significantly lower than for I3'Cs incorporated in plant material (root uptake) (Singleton et al., 1992). Comparisons

of the in vivo uptake of radionuclides by animals to the extractability into various extraction agents, showed that the fraction of radionuclides that can be extracted into 1M NH,Ac gives a good approximation of the bioavailable fraction in both soils and vegetation (Salbu et al., 1992). The present work discusses the transfer of 137Csand %Srin natural soil-water-plant ecosystems with particular emphasis on sampling and analytical techniques which can be utilized in studies on the physico-chemical forms and transformation processes of deposited radionuclides. Much of the data presented is based on sequential extraction studies performed

on samples collected between 1989 and 1992 at selected sites in Norway (Riise et al., 1990; Oughton, et al., in prep). Similar techniques have been applied for samples collected in 1992

169 from natural pastures in Sweden, Finland, Denmark and the Faroe Islands, where activity levels were supposedly greater than 2000 Bq kg". The results obtained in Nordic countries are compared with those obtained in the Ukraine, Belarus and Russia (Oughton et al., 1992; Salbu et al., in press).

METHODS Soil and vegetation samples were collected in Norway, Sweden, Denmark, Finland and the Faroe Islands. At each site, a number of soil cores, taken to a depth of between 4 and 10 cm, were collected from within a relatively small surface area (ca. 1 m') and corresponding vegetation samples were taken. Activity levels were compared in order to obtain information on the homogeneity of deposition, both between sites and within sites. Extraction studies were carried out to provide information on the association of radionuclides with soil components. Soil-to-plant transfer coefficients were compared with the fraction of radionuclides in a labile form (mobility factors), and the influence of seasonal variations assessed. Sampling sites

Norway (1989-1992): Sampling was carried out on natural upland pastures. At Lierne, Middagshaugen, in July, August and September of each year, an area of 1 mz was split into 4 squares of 250 cm2, the vegetation was cut from each square and a soil core (depth 4-5 cm, diameter 10 cm) was taken at the centre of each square. The distance between the monthly sampling areas was 1 m. At Tjatta, Lienga, soil cores of 10 cm depth, 20 cm diameter were collected and sectioned into 1 cm layers. Sweden (1991-1992): Five different sampling sites were selected at Blomhojden. At each site, grass was cut from four squares of 0.5 x 0.5 m, and four soil cores of 10 cm depth were taken from within each square and split into 0-5 cm and 5-10 cm layers. Denmark (1992): A composite sample of soil and vegetation was collected from within 1 mz at Ribe. Finland (1992): Four soil cores (0-5 cm) and vegetation samples were collected from within 1 mz on natural forest pasture in Jokionen. Faroe Islands (1992): four soil cores and vegetation samples were collected at Skala, Sumba, and Hvalvik. In Finland and the Faroe Islands, sampling was Carried out as in Norway. Sample Preparation

Soil and vegetation samples were dried at room temperature and homogenised before radionuclide determination and extraction studies. Soil samples were sieved to remove large stones. Dry weight (105°C) was measured on soil aliquots, and all results are given in units of Bq kg" dry weight (DW) for vegetation and Bq m-* for soil.

170

Sequential Extraction Homogenised soil samples were weighed (2 g) and extracted sequentially according to standard procedures (Tessier et al., 1979; Riise, et al., 1990). All extractions were carried out by shaking the samples in polyallomer centrifuge tubes (50 ml, Nalgalene Centrifuge Ware, Nalge) with 20 ml of the following extractants: H20, 1M W A C , 1M NaAc, 0.04M NH20H.HC1in 25% HAc, 30%H,02 at pH 2 (HNOJ, 7M HNO,. The supernatant was separated from the solid by high speed centrifugation (Biofuge 17RS, 11OOO g). All the solid phases were washed

(10 ml distilled water) and centrifuged between each extraction step. Whole soil samples, supernatants, wash solutions and the residues were analyzed for 137Cs(gamma-spectrometry). After '"Cs analysis, the supernatants from a number of the sequential extraction studies were also analyzed for WSr,samples from the same month were combined in order to improve the detection limit. A number of soil, vegetation and extraction fractions were also analyzed for naturally-occurring stable isotopes of Cs, Sr and Ca using neutron activation analysis. For many soils, a simplified extraction scheme using only W A C and HN03 was employed, and increased soil masses (up to 1OOg) were used for low activity samples from Sweden, Finland and the Faroe Islands. Activities in soils from Denmark were too low to allow extraction.

Determination of the Readily Labile or Mobile Fraction 1M W A C extractions of soil and vegetation samples have been used to differentiate between the "inert" and "labile" or "mobile" fraction of radionuclides in the soil-water-plant system. The percentage of deposited radionuclides that is present in a labile or potentially bioavailabile form at a given time, i.e. the "mobility factor", has been calculated by summing the activities extracted from soil and vegetation. Mobility Factor =

Total labile '37Cs IBq m-2) Total deposited 137Cs(Bq m2)

Total Labile =

Labile,o,

+ Labile,,,

The total labile 137Cs(or ?Sr) in vegetation ( E q m-*)is defined as that transferred to vegetation via root uptake, i.e. it should not include activity arising from soil contamination. Thus, in order to make a rough correction for soil contamination, vegetation samples were extracted sequentially with two aliquots of 1M W A C (mass:volume ratio 1:20; extraction time 2hr at room temperature). The labile fraction in soil (Bq m-*) was calculated by summing the H20 and NH4Ac sequential extraction fractions.

171

Radionuclide Determination Soil, vegetation, liquid and solid fractions were analyzed with respect to '37Csand '"Cs using a Packard Miniaxi 3" NaI Gamma Counter (5000 Series). Although both isotopes were determined, only '37Cs activities, being the most precise measurements, have been quoted for the purpose of this study. wSr determinations were carried out by Cerenkov counting of 9oy using a Quantalus 1220 low-level liquid scintillation counter (LBK Wallac, Finland) (Bjmstad et al., 1992).

Neutron Activation Analysis Samples of soil, vegetation and soil extracts were analyzed for stable element concentrations using a routine neutron activation analysis technique (Oughton et al. , 1992; Oughton and Day, 1993). Dried soil and vegetation were weighed into polythene capsules (0.5 cm'), and aliquots of soil extract solutions were repeatedly evaporated into the capsules. The capsules were heat sealed and then irradiated at a thermal neutron flux of 3 x (100%)activates to l'Cs

n cm-' s-l for 8 hrs. Stable 13'Cs

and stable ?3r (0.56%) activates to "Sr. After a decay time of ca. 2

weeks the capsules were counted using gamma spectrometry. Although a large number of trace elements were determined by this technique, only the results of stable Cs, Ca and Sr are reported here.

RESULTS AND DISCUSSION

Deposition The activity concentrations of 137Csranged from 1.8 to 105 kBq m-' and showed large differences both between sites and within the sites (Table 3.3.1). Activity levels at the selected sites in Denmark and Sweden were rather lower than anticipated. From 1989 to 1992, the activity concentrations at Lierne varied by between a factor of 7 and 20 within a few m2 (Table 3.3.2). These results reflect the inhomogeneous deposition of radioactivity that originates, on a large scale, from the influence of climate (i.e. wind direction and rainfall) and, on a small scale, from the influence of microclimate and the presence of particles. Fractionation of rainwater, May 1986, indicated that 75% of Cs-isotopes and more than 90% of refractory elements were present as HMW forms (Salbu, 1988b).

172 TABLE 3.3.1. The deposition (137CskBq m-2),aggregated transfer factors (mz kg-'), mobility factors (labile Bq m-'/total deposited Bq m-*)and the percentage of labile 137Csin vegetation (vegetation Bq m-2/ labile Bq m-') for natural soil-plant systems in Norway, Sweden, Denmark, Finland and the Faroe Islands. All samples were collected in summer 1992. Veg .

T'pp

MF

DW

m2 kg-'

%

%MF in veg .

13.4 - 105 30 - 63

270-8400 1900-4200

0.04-0.17 0.03-0.07

2-10 2- 7

2-10 18-30

SWEDEN Site la Site lb Site l c Site 2 Site 3 Site 4 Site 5

3.4 - 8.8 2.4 - 18.0 5.3 - 11.8 3.8 - 12.3 12.8 - 20.1 4.5 - 7.2 8.4 - 12.0

2000 760 330 600 550 660 490

0.286 0.080 0.039 0.068 0.034 0.116 0.049

DENMARK Ribe (n=l)

1.8

< 100

C0.056

NA

FINLAND Jokionen

14.1 - 15.7

<70-130

<0.004-0.008

<0.5

FAROE ISL. Skala Hvalik Sumba

2.7 - 3.9 2.6 - 4.3 4.0 - 5.1

<80 120 <50

< 0.024 0.035 <0.011

<1 2.8 <0.9

Site

Soil 0-5cm kBq m-'

Bq kg-'

NORWAY Lieme Tjata (1991)

3

< 0.5

> 50

8.3 < 0.5 2.7 3.6

15 > 50 60 17

8

NA - Not Analyzed; n=1-4

TABLE 3.3.2. (3-137 activities in soil and vegetation from Lieme, Norway, 1989-1992. Aggregated Transfer Factors (m2 kg") and Mobility Factors (Labile Fraction Bq m-*/Total deposition Bq m2). Mean f SEM (range), n=12. Year

SOIL (Kbq m")

GRASS (Bq kg-'1

T.pp

(m kg-2)

MF %

1989

141 f 45 (31.4 - 648)

5340 f 3460 (1800 - 21800)

0.045 f 0.005 (0.024 - 0.080)

13 1 (11 - 19)

1990

52 f 11 (13.0 - 176)

3830 f 570 (1300 - 7800)

0.096 f 0.023 (0.04 - 0.33)

7.8 f 0.7 (6 - 10)

1991

85 f 20 (30.2 - 213)

3790 f 560 (2100 - 8200)

0.060 f 0.007 (0.015 - 0.107)

1992

43 f 8 (13.4 - 105)

( 270

3580 f 610 - 8400)

0.100 f 0.015 (0.017 - 0.169

7.2 f 1.2 (4 - 9) 4.2 f 1.1 (2 - 10)

173

In Lierne and Tjratta, the deposition of 90Srwas between 1-4% of the 137Cs,whereas the 90Srto 137Csratio in the total releases from Chernobyl has been estimated to be 0.22 and the ratio in fuel 0.70 (IAEA, 1986). This suggests that, in Norway, a relatively low fraction of deposited radionuclides was associated with fuel particles and that condensed particles formed the major fraction of deposited 137Csand 3 r . However, the presence of refractory nuclides (e.g. '"Ce and "Zr) in the fallout suggests that the total particle burden was relatively large. Due to the low activity levels in the samples from Sweden, Denmark, Finland and the Faroe Islands, 90Sranalysis was not carried out on these soils. When determining the activity level of deposited fallout nuclides at a site, the sampling strategy employed should be capable of reflecting the large variations resulting from inhomogeneous deposition (i.e. multiple sampling over the whole area). Although a composite sample will provide an estimate of the average deposition level, it will not give a true picture of the range or inhomogeneity of deposited fallout. Alternatively, in situ monitoring using a portable detector calibrated against soil samples is a rapid and inexpensive method for estimation of deposition level (Haugen et al., 1992) and has been shown to be well-calibrated over a wide range of activities (Salbu et al., in press). Finally, it should be stressed that laboratory analysis of soil samples using gamma spectrometry can produce erroneous results if a significant fraction of the activity is present as a few point sources (particles) within the bulk sample but one assumes that the activity is distributed homogeneously throughout the sample. Vertical Distribution in Soils The soil profile for Tjratta in 1990 indicated that 85% of the total 137Csactivity in a 10 cm deep core was retained within the upper 2 cm (Figure 3.3.3). Analysis of cores collected from Tjratta in 1991 and 1992, indicated that the 137Cshad been transported down the soil profile, with only

47-67% being found in the upper 2 cm. No significant difference was apparent between cores collected in 1991 and 1992 from Spellemannsakeren, nor between the Alstandhaug and Spellemanndkeren cores in 1991. For all cores, 89 f 9 % of the activity was retained in the upper 4 cm, with no significant difference between the years. Analysis of Swedish soil cores in 1991 and 1992 showed that, respectively, 86 f 6 and 88 f 5 % of 137Csin a 10 cm deep core was retained in the upper 5 cm, with no difference between the sites. Comparison of radionuclides with the stable elements can provide valuable information on the "equilibrium status" of the deposited radionuclides within the ecosystem and can be a considerable aid in prediction of future transfer (Oughton and Day, 1993). As 137Csis transported down the soil profile, and with binding to the inaccessible sites in the clay matrix, the plant-to-soil concentration ratio for '37Cswill be expected to approach that of stable Cs. Of

174

course, compared to InCs, the stable Cs concentrations are relatively constant down the soil profile. For Lieme soils, the concentration ratio (CR) for stable Cs (pg-' plant/pg-I 0-5 cm soil) is between 0.10 and 0.25, and this is the minimum CR which can be attained for " 'Cs in the 05 cm of Lieme soils. In 1990, the concentration ratio for 137Cs(Bq kg" plant/Bq kg-' 0-5 cm

soil) in Lieme soils was between 0.7 and 1.5, and could be expected to decrease with time. However, the rate of decrease is too slow, and the range in , T too large, to show significant differences from 1989 - 1992.

FIGURE 3.3.3. Soil profile showing the distribution of '"Cs as a function of depth for samples collected at Tjetta: a) Alstadhaug, 1990 ('"Cs: 45 kBq rn'2, n = l ) and 1991 (38 kBq m2,n = l ) ; b) Spellemannsakeren, 1991 (28 f 3 kBq m-*, n=6) and 1992 (61 f 4 kl3q m", n=6).

In natural pastures where activity has been retained in the upper soil layers, the observed

activity as Bq kg" will be strongly influenced by the depth to which the sample has been taken. In general, the soil activity as 137CsBq kg" at a particular site tended to decrease with sample weight. Hence, for natural pastures, results should be given as Bq m-' rather than Bq kg-'. Soil profiles collected in the Ukraine indicated that the transport of

down the soil

profile was greater than for '"Cs (Alexahkin et al., 1992; Salbu et al., in press). This may reflect the transport of %r associated with particles or the greater mobility of strontium compared to caesium.

I75

Sequential Extraction Studies Sequential extraction has been used to study the association of radionuclides with soil components and the influence of fuel particles on radionuclide mobility. Sequential extraction studies on Norwegian soils collected between 1989 and 1992 indicated that "Sr was more labile than W s . For both Tjnrtta and Lierne, between 47 and 90% of the %r was found in the

easily-extractable fractions, whereas for '37Csmore than 80% was found in the strongly-bound fractions (Figure 3.3.4). The sequential extraction technique gives very reproducible results, with less than 10% variation between soils collected at the same site at the same time.

FIGURE 3.3.4. Sequential extraction of 137Csand %r in soil samples from Lierne, 1989-1992. '"Cs values represent the mean of 12 samples for each year (4 x 3 months), "Sr values represent the mean of 3 samples for each year (1 composite sample per month).

The radionuclide distribution throughout the extraction fractions has been compared with the distribution of the naturally-occurring stable isotopes, in order to evaluate the degree of isotopic exchange Wise et al., 1990; Oughton et al., 1992). Although, in 1990, %r and stable Sr showed similar distributions between extraction fractions in Norwegian soils, soils collected from within the 30 km zone showed considerable differences in distribution (Figure 3.3.5). Similar results have been observed in Belarussian soils collected from areas up to 70 km from the reactor (Salbu et.al., in press), and reduced availability of Y3r as compared to stable Sr is thought to reflect the influence of fuel particles.

I76

FIGURE 3.3.5. Distribution of wSr and naturally-occumng stable Sr and Ca in sequential extraction fractions from soils collected in Lierne, Norway, and Bourykovka, Ukraine (15 km from the Chernobyl Reactor) 1990. (Redrawn from Oughton et.al., 1992)

Sequential extraction of Tjata soils indicated that the W A C extractable fraction decreased with depth (the mobile fraction was between 3 and 11 % in the 0-1 cm layer as compared to 1 to 4 Iin the 1-2 cm layer), this is probably related to the higher organic content in the upper soil layer. Hence the sampling depth, and the transport of '"Cs down the soil profile with time, would be expected to influence the observed lability of 13'Cs.

177

Transfer Factors and Mobility Factors In soils, the W A C extractable fraction is assumed to reflect the fraction of radionuclides (and stable elements) reversibly bound to ion-exchange sites in the soil, i.e. radionuclides which are in dynamic equilibrium with the soil solution (Figure 3.3.2), and has been used to calculate mobility factors. Various reagents, including NH4Ac, have been utelized to estimate the cationexchangeable fraction of elements in soils (Van Loon and Barefoot, 1993), the extractable fraction (i.e., displaced ions) is best classed as gperationally defined by the chosen extractant. Mobility factors have been measured on ca. 200 soil samples collected in Norway, and were observed to range from between <0.5 to 22 % for 137Cs,and from 43 to 95 % for %r. Corresponding aggregated transfer factors for 137Csranged from 0.008 to 0.33 and for ?3r from 0.02 to 0.35. The 137Csmobility factors and transfer factors for soil samples in the other Nordic countries were of the same order of magnitude as those seen in Norway (Table 3.3.1). The mobility factor represents the W fraction of the deposited radionuclide in the soil-waterplant system that is present in a labile form (i.e., including that already incorporated in plant material), The MF represents the fraction that is probably, at the time of analysis, readily available for uptake by plants; however, the & plant uptake will be dependant, among other factors, on vegetation type, season, microbial activity and available K and stable Cs levels. The 137Csin vegetation is usually less than 10% of the total labile 137Cs,and less than 1-2 % of the

total deposited *37Cs.Hence, when vegetation activities are not available, MF can be estimated from the W A C extractable fraction in soils (i.e. 137Csreversibly bound to ion-exchange sites). However, it should be stressed that when MF is less than 5 %, the 137Csin vegetation can often account for up to 40% of the total labile 137Cs(the value for one Swedish soil suggested up to

60%). The "Sr in vegetation is usually less than 1% of the total labile

and hence, for 90Sr,

the MF is usually equivalent to the a A c extractable fraction of "Sr in soils. Seasonal Variations In 1990, the 137CsBq m-' activity levels in vegetation from Lierne increased from July to August, although the Bq kg-I levels remained fairly constant (Table 3.3.3). In August, the increase in Bq m-' was largely due to the increase in vegetation biomass. The aggregated transfer factor also increased in August, but this was not significant. Mobility factors, however, remained constant from July to August indicating that the labile 137Csreservoir remained constant. Hence, in 1990, the bioavailability of the '"Cs in the Lierne soil did not appear to be affected by season, and differences in aggregated transfer factors do not seem to be attributed to differences in availability of 137Csin the soil. The fraction of the total labile 137Csthat is

found in vegetation increases in August and falls again in September, possibly reflecting

TABLE 3.3.3. Change in transfer of Cs-137 to vegetation over the growing season, Lierne, 1990 (n=4).

Month

July

soil

Veg .

kBq m-'

g m2

81 f 25 - 176)

20

36 f 8 (13 - 59)

66

(41

Aug

Sept

*

39 5 (33 - 56)

57

CS-137

Transfer Coeff.

Mobility Factor

kBq kg-'

kBq m-2

m2 kg-1

(Soil +Veg)

5.4 f 1.1 (3.5 - 7.8)

0.11 f 0.02 (0.06 - 0.16)

0.076 f 0.011 (0.04 - 0.11)

0.079 f 0.008

4.1 t- 0.4 (2.9 - 5.1)

0.26 f 0.03

(0.20 - 0.35)

0.158 & 0.051 (0.05 - 0.33)

0.078 f 0.007

2.0 f 0.2 (1.3 - 2.5)

0.11 f 0.13 (0.07 - 0.14)

0.053 f 0.008 (0.041 - 0.081)

0.077 f 0.007 (0.063-0.099)

VEGETATION

(0.056-0.092)

(0.056-0.090)

veg %MF

2.1 f 0.3 (1.5 - 3.1)

8.1 f 0.8

(6.2 - 9.9)

4.0 f 0.9 (3.1 - 6.8)

179 variable uptake of '"Cs from the available reservoir. This may reflect the influence of biomass, vegetation type, and/or mineral uptake and release by the plant roots over the growing season. It should be stressed, however, that other studies on seasonal variation in MF for Soviet soils found a significant seasonal variation in the mobile fraction for some soils (Salbu et al., in press). Hence, the MF can be influenced by seasonal variations in the soil-water-plant system (e.g. microbial activity, nitrate levels). Lone-term Van'ations 1989-1992 For the Lierne samples, aggregated transfer factors for '37Cs show wide variation over the three years, thus it is difficult to evaluate any trend (Table 3.3.2). The mean value of the mobility factor decreases from 1989 to 1992, possibly indicating a removal of 137Csto more inaccessible binding sites in the soil, however the decrease is only significant (p = 0.01) between 1989 and 1990. Although mobility factors tend to show a narrower range as compared with Tar, the MF range increases in 1992 (Table 3.3.2), and the 1992 results from all sites do not show any obvious relationship between T,

and MF (Table 3.3.1). The combination of the low activities

in many of the soil samples, together with the low extractability of 137Cs(decreasing with time after deposition), results in large uncertainties in MFs. Furthermore, with increasing time after deposition and as the '"Cs is transported down the soil profile, the influence of sampling depth on MF can increase. Other parameters such as microbial activity and mineral status may also be influencing the availability of '37Cs. For %r, both the activity concentration in vegetation (Bq kg-') and the mobility factor remain constant over the 4 year period (Table 3.3.4). The transfer coefficients are however less reproducible. This is largely due to differences in measured soil activities, which may reflect a sampling problem in 1990. Despite the higher MF for

as compared to 137Cs,the Twgand

CR for 90Sr in Lierne soils was of a similar order of magnitude as for 137Cs,which indicates a rather low root uptake of %Sr from the labile reservoir. It is possible that transfer of "Sr into run-off water will be the more significant transport mechanism for %r. Within the 30 km zone, the contamination of water by wSr has been shown to be increasing due to weathering of fuel particles, and was considered to be one of the more pressing problems in 1991 (Konoplyov et al., 1992). Vegetation samples from Lierne gave > 80% extraction with W A C , hence the wide range in vegetation activities is unlikely to be due to soil contamination. The Tj&ta samples showed evidence of soil contamination in September 1990 and 1991, the NH4Ac extractability being reduced to 40% in some samples. All vegetation samples gave a greater than 90% extraction of 90Srwith NH.,AC.

I80

TABLE 3.3.4. wSr activities in soil and vegetation from Lieme, Norway, 1989-1990.Transfer Coefficients (mZkg") and Mobility Factors (Mobile Fraction Bq m-*/Total deposition Bq m-*). Samples refer to a composite sample for each sampling month (i.e. n=3), for vegetation, n=12. Mean f SEM (Range) SOIL (Bq m")

VEGETATION (Bq kg-')

T

1989

970 f 100 (600 - 1800)

86 f 1 1 (50- 160)

0.087 f 0.010 (0.030- 0.160)

57 f 6 (46 - 63)

1990

1650 f 360 (1300- 2400)

88 f 15 (53 - 181)

0.058 f 0.008 (0.060 - 0.098)

54 f 9 (44 - 74)

1991

360 f 24 (310 - 410)

78 f 14 (43 - 127)

0.219 f 0.029 (0.096 - 0.296)

66 f 4 (59 - 75)

1992

1600 f 320 (1200- 2400)

63 f 11 (43 - 143)

0.045 f 0.010 (0.020- 0.058)

53 f 2 (54 - 57)

Year

(3kg")

MF %

Applications of Mobilitv Factors Transfer coefficients (i.e. T,

and CR)

mobility factors (MF) can be influenced by a

number of parameters, including: a) the physico-chemical form of the deposited radionuclides (e.g. ionic, particulate); b) the chemical properties of the radionuclide (e.g. solubility, binding strength to mineral or organic soil components); c) the soil characteristics (e.g. pH, organic content, microbial activity); and d) time (e.g. particle weathering, migration into clay mineral lattices, transport down the soil profile, seasonal variations). In addition, T,

and CR are &

affected by the vegetation type, growth rate, and bulk density; the mineral status of the soil (ie. level of exchangeable K); surface contamination of vegetation (e.g. by soil or fallout species); and the depth to which the soil sample has been taken. Calculation of the percentage of deposited radionuclides present in a labile form (i.e., mobility factor) should be seen as providing a Supplement and not a replacement for transfer factors. In certain cases, mobility factors may be less susceptible to "noise" (i.e. show less variation) or may help to isolate the possible reasons for large variations in transfer factors. Mobility factors can be useful when 5omparing radionuclide transfer and lability between different sites or soil types, or when 5omDaring different radionuclides at the same site. They are particularly useful for evaluating chanees in lability over time, both long-term and seasonal. Because a calculation of the labile fraction helps to identify the "source term" of the radionuclide, MFs are essential when modelling transport of deposited radionuclides in an ecosystem. Even on identical soils with identical vegetation types, there will be a marked

difference in the behaviour of "Sr incorporated with uranium oxide particles as compared to 90Srdeposited in an ionic or soluble form. Mobility factors have proved to be particularly usehl in studies on near-field soils. The effect of fuel particles on the lability of radionuclides, and the correlation between MF, Tqg and CR, can be clearly seen in analysis of 3 r in soils collected from the SUS (Table 3.3.5). The range of MF for I3'Cs in soils collected within the SUS between 1990 and 1992 is similar

to that observed in Norwegian soils, but the MF for "Sr in near-field soils contaminated by fuel particles is significantly lower than that observed in Norwegian soils. In those areas where the MF of "Sr is relatively low, the lability of 3 r and transfer to vegetation is likely to increase. In this case, evaluation of the labile fraction can provide a more concrete assessment of the future transfer of radionuclides to vegetation than calculation of transfer factors alone. TABLE 3.3.5. "Sr in vegetation from Norway, Russia, Ukraine and Byelorussia, 1990-1991. Site

"Sr

TW

Vegetation

m2 kg"

NORWAY Lierne Tjntta

0.02 0.01

88 10

0.058 0.028

0.80 0.24

66 75

UKRAINE Bourykovka' Poleskoe Rowna

0.19 0.07 0.01

NA 110 76

NA 0.016 0.033

NA 0.69 0.41

14 95 68

BELORUS Dublin Sudkova Slavgorod

0.15 0.12 0.03

965 360 NA

0.027 0.009 NA

0.19 0.14 NA

28 15 52

RUSSIA RIAF Komsomolets

0.01 0.02

138 466

0.029 0.035

0.40 0.57

60 50

Bq kg-' DW

CR

MF

9osr/137Cs Total Deposition

%

NA - Not analyzed. * - 30 km zone Uncertainties f 20% (n = 4-12)

CONCLUSIONS The activity levels of radionuclides (Bq m2) deposited in the Nordic countries showed considerable variation, and variations were seen between the countries, regions and within a single m2. Activities in vegetation and transfer factors also show variations between sites,

182

within sites, with time and between the different radionuclides. In 1989, studies on the mobility of radionuclides ('"CS and "Sr) in Norwegian soil-plant systems indicated that the fraction of radionuclides deposited as fuel particles was not having any significant effect on the transfer of 137Csor "Sr. The degree of isotopic exchange of '"Cs with the naturally-occurring stable element and removal to slow turnover binding sites was extensive in 1989, and studies from 1989 and 1992 suggest that '"Cs and "Sr lability is more dependant on the physical and chemical properties of the soil and the chemical properties of the element than the fallout speciation. Hence, the particle form of deposition from Chernobyl is not expected to be important for future transfer of radionuclides in the Nordic countries. In comparison, studies on soils collected from the 30 km zone between 1990 and 1992, suggest that the lability (or rather "non-lability") of "Sr is largely determined by the fraction of associated with fuel particles. Long-term monitoring of the changes in radionuclides' mobility in the near-field areas will provide valuable information on the rates of particle weathering and on the possible contamination of other parts of the ecosystem due to increased lability of radionuclides associated with fuel particles, including actinides. Prior to the Chernobyl accident, knowledge on the effect of particles was almost non-existent. In the event of another accident involving dispersal of irradiated fuel, knowledge on the environmental impact of radionuclides incorporated in fuel particles will be of great importance for reasons of preparedness. In the event of a nuclear accident, it seems that soon after deposition of radionuclides, an estimation of the labile fraction of radionuclides by W A C extraction will provide essential information on the speciation of radionuclides in fallout. Furthermore, the transformation processes can be followed by monitoring the change in the labile fraction over time. Studies on Norwegian soils suggested that both transfer factors and mobility factors are needed for a full understanding of the processes involved and for future predictions of radionuclides in the other parts of the ecosystem.

REFERENCES Alexakhin, R. M.; Sanzharova, N. 1.; Fesenko, S. V.; Sprin, E. V.; Firsakova, S. K.; Astasheva, N. P. , 1992. Radioactive contamination of agricultural ecosystems, In: Radeoecological Consequences of the Chernobyl Accident, 1.1. Kryshev (Ed), Nuclear Society International, Moscow, pp. 9-20. Bjmstad, H. E.; Lien, H. N.; Yu-Fu, Y.; Salbu, B., 1992. Determination of %r in environmental and biological materials with combined HDEHP solvent extraction, J. Radioanal. Nuc. Chem., 156, pp. 165-173.

183

Bobovnikova, T. I.; Virchenko, Y. P.; Konoplev, A. V.; Siverina, A.; Shkuratova, I. G., 1991. Soviet Soil Science, 23 pp. 5-52. Bondar, P. F.; Ivanov, Yu. A.; Ozornov, A.G. 1992. Estimation of relative biologic availability of '"Cs in fallouts and of its total biologic availability in soils at territory subjected to radioactive contamination. In: The Radiobiological Impact of Hot BetaParticles from the Chernobyl Fallout: Risk Assessment, Kiev, Ukraine, August 1991. IAEA, Vienna, Part I, pp. 58-67. Borovoy, A. A.; Demin, V. F.; Blinavoa, L. D. 1992. Radioactive releases originating from the Chernobyl accident. In: Radeoecological Consequences of the Chernobyl Accident, 1.1. Kryshev (Ed.) Nuclear Society International, Moscow, pp. 9-20. Devell, L.; Tovedal, M.; Eergstrom, U.; Appelgren, A.; Chussler, J.; Anderson, L., 1986. Initial observations of fallout from the reactor accident at Chernobyl, Nature, 321, pp. 817-819. Hansen, H. S; Hove; K., 1991. Radiocaesium bioavailability: transfer of Chernobyl and tracer radiocaesium to goat milk, Health Physics, 60,pp. 665. IAEA, 1986. Report of the USSR State Committee on the Utilisation of Atomic Energy to the IAEA Meeting on the Chernobyl Accident, 25-29 August, Vienna, 1986. IAEA, 1991. The Radiobiological Impact of Hot Beta-Particles from the Chernobyl Fallout: Risk Assessment, Kiev, Ukraine, August 1991. IAEA, Vienna. Konopylov; A. V.; Borzilov, V. A,; Bulgalov, A. A.; Nikitin, A. I.; Novitsky, M. A; Voicehovitch, O., 1992, Case study No 8 - Hydrological aspects of the radioactive contamination of water bodies following the Chernobyl accident. In: Hydrological Aspects of Accidental Pollution of Water Bodies, Operational Hydrology Report No. 37, W M O No. 754, WMO, Geneva, pp. 167-190. Loshchilov, N. A.; Kashparov, V. A.; Yudin, Ye. B.; Protsak, V. P.; Zhurba, M. A.; Parshakov, A. E. Physical-chemical forms of the radioactive fallout from the Chernobyl reactor accident. 1992. In: The Radiobiological Impact of Hot Beta-Particles from the Chernobyl Fallout: Risk Assessment, Kiev, Ukraine, August 1991. IAEA, Vienna, Part I, pp. 34-39. Oughton, D. H.; Salbu, B.; Riise, G.; Lien, H. N.; Ostby, G.; Nmen, A., 1992. Radionuclide mobility and bioavailability in Norwegian and Soviet Soils, The Analyst, 117, pp. 481486. Oughton, D. H.; Salbu, B.; Brand, T. L.; Day, J. P.; Aarkrog, A., 1993. Underdetermination of Strontium-90 in Soils Containing Particles of Irradiated Uranium Oxide Fuel, The Analyst, 118, 1101-1105. Oughton D. H.; Salbu, B. S. The influence of speciation on radionuclide transfer and mobility in Norwegian soils. In preparation. Oughton, D. H.; Day, J.P., 1993. Determination of Cs, Rb and Sc in biological and environmental materials by NAA, Journal of Radioanalytical and Nuclear Chemistry, 174, pp. 177-185. Riise, G.; Bjmstad, H. E.; Lien, H. N.; Oughton, D. H.; Salbu, B., 1990. A study on radionuclide association with soil components using a sequential extraction procedure. J. Radioanalytical and Nuclear Chemistry, 142, pp. 531-538. Salbu, B., 1988a. The Quality of Analytical Data for Modelling PurpOses, BIOMOVS Tech. Rep. 3, ISSN 100-0392, National Institute of Radiation Protection, Sweden, p. 79. Salbu, B., 1988b. Radionuclides associated with colloids and particles in rainwaters, Oslo, Norway, In: Hot Particles from the Chernobyl Fallout, H. von Philipsborn, F. Steinhausler, F. (Eds), proceedings of an International Workshop, Theuern: Bergbau- und Industriemuseums, Theuern, Vol. 16, pp. 83-84.

184

Salbu, B.; Bjmstad, H. E.; Lindstrem, N.; Lydersen, E.; Breivik, E. M.; Rambrek, J. P.; Paus, P. E., 1985. Size fractionation techniques for the determination of elements associated with particulate or colloidal material in natural fresh waters, Talanta, 32, pp. 907-913. Salbu, B.; Bjmstad, H. E.; Lydersen, E.; Pappas, A. C., 1987. Determination of radionuclides associated with colloids in natural waters, J. Radioanal. Nucl. Chem., 115, pp. 113-123. Salbu, B. S.; Ostby, G.; Garmo, T.; Hove, K. 1992. Availability of caesium isotopes in vegetation estimated from incubation and extraction measurements. The Analyst, 117, pp. 487-492. Salbu, B.; Bjmstad, H. E.; Svreren, I.; Prosser, S. L.; Bulman, R. A.; Harvey, B. R.; Lovett, M. B., 1993. Size distribution of radionuclides in nuclear fuel reprocessing liquids after mixing with seawater, Science of the Total Environment, 1301131, pp. 51-63. Salbu, B.; Krekling, T.; Oughton, D. H.; Ostby, G.; Kashparov, V. A.; Day, J. P., 1994. The significance of hot particles in accidental releases from nuclear installations. The Analyst, 119, pp. 125-130. Salbu, B., Oughton, D. H.; Ratnikov, A. V.; Zhigareva, T. L.; Kruglov, S. V.; Petrov, K. V.;Averin, B. V.; Firzakova, S. K.; Astasheva, N. P.; Loshchilov, N. A.; Hove, K.; Strand, P., in press. The mobility of I3’Cs and %r in agricultural systems Ukraine, Belarus and Russia, Health Physics. Singleton, D. L.; Livens, F. R.; Beresford, N. A,; Howard, B. J.; Barnett, C. L.; Mayes, R. W.;Segal, M. G., 1992. Development of a laboratory method to predict rapidly the availability of radiocaesium, The Analyst, 117, pp. 505-510. Tessier, A.; Campbell, P. C. G.; Bison, M., 1979. Sequential extraction procedure for the speciation of particulate trace metals, Anal. Chem., 51, pp. 844-852. Van Loon, J. C.; Barefoot, R.B. 1992. Overview of analytical methods for elemental speciation. The Analyst, 117, pp. 15-22.

185

3.4. CONTAMINATION OF ANNUAL CROPS

MORTEN STRANDBERG' Ris0 National Laboratory, Roskilde, Denmark. SUMMARY Results are presented from the Nordic countries dealing with the uptake of radiocaesium from soil in annual crops after the Chernobyl accident. Barley, potato, carrot, cabbage and pea were selected as suitable representatives of Nordic annual crops. The transfer of radiocaesium to man from these annual crops was generally low. Common experience was that levels after the first year decreased considerably in the agricultural ecosystems, because of the absence of fresh direct fallout and the rapid, strong fixing of caesium in most soil types. Thereafter the rate of decrease was very uncertain with a large variation between localities. Agricultural practices inhibit uptake and especially resuspension by deeper placement of the contaminated surface soil. Only in areas with highly organic soils, low in clay, potassium and pH, can considerable uptake through roots take place. Examples of such places with an enhanced uptake from soil are the Swedish peat study sites in the Gavle region, and the Faroe Islands. In such areas the addition of potassium can be recommended in cases of severe contamination. for content of radiocaesium in the treated Reliable effective ecological halflifes species cannot be calculated from the material available. A cautious estimate of TI,, of about 5-10 years in the period from 1987 and until today seems reasonable. Results indicate the longest for the Danish and Finnish mineral soils, and the shortest for the Swedish and Faroese organic soils. Aarkrog (1992) states that the ecological halflife for Chernobyl 137Csin the Danish total diet is 3 years. The content of radiocaesium is lower in barley grain than in the vegetable species. Carrots had a lower uptake to the edible parts than vegetable species where other parts than the root are used. These uptake patterns correspond well with what is generally assumed. INTRODUCTION Arable land in the Nordic countries received varying amounts of debris primarily containing '"Cs and 137Csafter the Chernobyl accident. This variation was the result of differences in rainfall, distance and direction relative to the winds from the place of the accident. The fallout distribution was very uneven in Sweden, Finland and Norway

. In Denmark, the Faroe Islands and Iceland

*: Data has been made available by: Aarkrog et al. (1988, 1989, 1991,1992), Bjerke (1987), Bjerke & Bakken (1988), Eriksson & RosCn (1989), Magnusson pers. comm., Mascanzoni (1986), Rantavaara pers. comm., Rantavaara (1987, 1991), Rantavaara & Haukka (1987), Rantavaara & Kostiainen (1993), R o l n & Feuerbach (1988), RosCn (1989), Roskn (1991), SIS of Norway (1990), SaxCn et al. (1987), STUK A-55 (1987).

I86

levels were low and the distribution was more homogeneous. The total variation was between approximately 0.1 and 200 kBq m", levels being lowest in Iceland and highest in parts of Sweden (Magnlisson et al. 1992, Moberg 1991).

On this background investigations were started on the uptake of radiocaesium from soil and the content of radiocaesium in annual crops in the Nordic countries. The work was carried out either in the framework of monitoring programmes, or in special investigations sometimes involving different kinds of experimental treatment. The following species were selected as suitable to represent Nordic annual crops in this kind of investigation: Barley (Hordeum vulgare), Potato (Solanum tuberosum), Cabbage (Brassica oleracea), Carrot (Daucus carob) and Pea (Pisum sativum). Measurements made on these crops from all the Nordic countries are included in the present review. RESULTS 1986 The results from 1986 are presented in Table 3.4.1.The differences between the Nordic countries are well documented in this year. Denmark received minor amounts compared to the other Scandinavian countries. Sweden had the highest fallout levels. However the south of Sweden (not included in the Table), where most annual crops are grown, received less than the area around Gavle and neighbouring regions, and levels and trends similar to the Danish can be expected in southern Sweden. Radiocaesium in Danish, Finnish and Swedish crops after Chernobyl. Danish values for radiocaesium in soil and crops (Table 3.4.2.) are means from 10 experimental state farms. Products from these farms have been included in the Danish monitoring programme on radionuclides in the diet since the early sixties. The Faroe Islands received fallout amounts TABLE 3.4.1 Values of 137Csin annual crops in the Nordic countries in 1986 (Bq kg' '1. All crop data are in Bq kg-' fresh weight, soil kBq m-* 1986

barley

potato

cabbage

carrot

pea

soil0-5cm

Denmark

1.52

0.197

0.21

0.103

0.19

1.2

Faroe Isl.

2

15.2

Finland

4.2

2.2

Norway

7.3

5.0

Sweden

70

20

2.5

1.5

2.2

16 9

13

7

I0

10

187

TABLE 3.4.2. Summary of results from Denmark for 1986 to 1991. Soil is kBq m-', crops Bq kg-' fresh weight. No peas were measured after 1989.

DENMARK

1986

1987

1988

1989

1990

1991

Soil 0-5cm

1.2

Barley

1.52

0.078

0.087

0.054

0.067

0.049

Cabbage

0.21

0.059

0.045

0.045

0.092

0.058

Carrot

0.10

0.060

0.043

0.058

0.087

0.034

Pea

0.194

0.068

0.196

0.168

Potato

0.197

0.134

0.094

0.114

0.088

0.078

similar to those in Denmark. The Faroese soil type makes the rate of uptake from soil much higher than in the rest of Scandinavia. The difference is demonstrated in Table 3.4.5.,where aggregated transfer factors (T,.) for potatoes from the Nordic countries are compared. The ratio between the concentration of radiocaesium in the crop and the deposition on the growth site is named the aggregated transfer factor (Tag. = concentration in plant(Bqkg-I) /deposition(Bqm")). T,, is a good measure for comparison between sites with different deposition, because the deposition is eliminated in the calculation. Hence the differences observed in tables

3.4.5. and 3.4.6. are only the result of differences in soil type between the sampling sites. Most post-Chernobyl measurements in Sweden were carried out in an area called the fallout area, which received the highest levels of contamination. It should be noted that Swedish measurements on barley presented in Table 3.4.4.)are obtained from a site with peat soil and a deposition of 200 kBq m-*. Vegetable results from Sweden for 1987-89 were obtained from a silt loam soil, with a deposition of 30 kBqm-' (RosCn 1991). Swedish results were recalculated to fit the mean deposition of 10 kBq m-2 in Sweden. Hence the Swedish results are calculated values rather than TABLE 3.4.3. Summary of results from Finland for 1986 to 1990. All crop results are Bq kR" fresh weight, soil is kBq m-*. No peas were measured after 1986. FINLAND

1986

1987

1988

1989

1990

Soil 0-5 cm

16

Barley

4.2

0.7

0.7

1.4

1.4

Cabbage

2.7

0.7

0.6

0.5

0.5

Carrot

1.6

0.5

0.4

0.4

0.4

Pea

2.4

Potato

3.0

1.o

1.o

0.8

0.7

TABLE 3.4.4.Summary of results from Sweden for 1986 to 1990.Barley is Bq kg-' dry weight, soil is kBq m.', the other crops is Bq kg" fresh weight. SWEDEN

1986

Soil 0-5cm

10

Barley

70

Cabbage

13

Carrot

7

Pea

10

Potato

20

1987

1988

1989

1990

50

18

8

6

7

11

7

8

8.5

26

experimentally obtained results. They can be supposed to be valid for barley and vegetables grown

on peat soil respectively silt loam soil, with a deposition of 10 kBq m-'. Potatoes When comparing potatoes, which are the only annual crop measured by all countries, the uptake rate is highest on the Faroe Islands, with Taggof 0.0076, Sweden follows with Twgof 0.005 in selected places, and then comes Norway. In Finland, Iceland and Denmark aggregated transfer factors are very low, between 0.0001 and 0.00015. Differences in soil properties between the countries, and thereby also between the measurement sites, are a likely explanation of the observed variation. For potatoes, the effective ecological halflives calculated include the year of the Chernobyl deposition, 1986. These halflives obviously consists of at least two halflives, one short immediately after the deposition and a longer halflife dominating the period from one year after the accident

Table 3.4.5. Aggregated transfer factors Tagg (m2kg.') and effective ecological halflifes T,,, (years) for potatoes in the Nordic countries.

1986

1987

1988

1989

1990

1991

1992

Denmark 0.00016 0.00011 0.00008 0.00009 0.00008 0.00007

T,,' 7

Faroes

0.0076 0.0018 0.0042 0.0035 0.0017 0.0020

3

Finland

0.0002 0.00005 0.00005 0.00004 0.00004

2

Iceland

0.0001

Norway

0.00054

Sweden

0.005

0.00054 0.002

0.0007

0.0008

1

189 and forward. The most striking result is the high rate of uptake on the Faroe Islands where the rate of uptake from the basaltic-type organic soil is higher than on the peaty sites studied by RosCn (1989 & 1991) and Eriksson & RosCn (1989). The difference is most distinct in 1988 and 1989, because

of a much faster decrease in uptake in Sweden. From Tables 3.4.1. and 3.4.5. it appears that the radiocaesium content in Faroese potatoes has generally been on a high level. The high Faroese uptake could be due to the basaltic soil type on the islands. Levels in the Faroese potatoes are almost as high as in the Swedish, despite much higher radiocaesium levels in Swedish soil, Aarkrog et al. (1988,1989,1992) and RosCn (1991)

Barley Aggregated transfer factors and effective ecological halflifes for barley are given in table 3.4.6. For barley, the trend is very similar to that observed for potatoes. In the calculation of TIReco for barley, 1986 is excluded because of the contribution from direct contamination in this year.

Differences between years The content of radiocaesium in Danish annual crops in the years following the Chernobyl accident is illustrated in Figure 3.4.1. (Note that pea is not included after 1989). The largest difference, a decrease of about 85 %, is between the harvest seasons of 1986 and 1987. The main part of this decrease can be assigned to barley, because winter barley received some direct contamination in April 1986.The differences between the Nordic countries are quite large as shown by Tables 3.4.1-3.4.6., hence it makes no sense to calculate common effective halflifes. In Denmark an effective ecological halflife of 7 years can be calculated for potatoes, the value is not statistically significant but at the 5 % level of confidence it is very close. If we look at the effective halflife for I3’Cs in barley the difference between countries is also demonstrated. This calculation gives an effective halflife in Danish barley of 17.3 years calculated on the basis Table 3.4.6. Aggregated transfer factors T,,, (m2kg-’) and effective ecological halflives T,,, (year) for 137Cs in barley in the Nordic countries Denmark

1986 1987 1988 1989 1990 1991 T112 0.0013 0.00007 0.00007 0.00005 0.00006 0.00004 6

Finland

0.0003 0.00004 0.00004 0.00009 0.00009

Norway

0.0010 0.00024 0.00015

Sweden

0.0070 0.0050

0.0018

0.0008

0.0006

1

190

of the years from 1987-1991 (the result is not significant). 2.5

2

1.5

rn Potato Pea

w Carrot rn

Cabbage Barley

1

0.5

0

Figure 3.4.1 Comparison of radiocaesium in annual crops in the years after Chernobyl in Denmark.

In Sweden the similar value calculated for the years 1987-1990 gives the result 3.4 years, which is significant on a 5 % level. In Norway the calculation gives an effective halflife of 2.8 years (not significant), here the years 1986 - 1988 are included in the calculation, and the value is too low because of the contribution from direct contamination in 1986. In Finland no decrease was observed in average barley measurements from 1987-1990, Table 3.4.3. In a short span of years after an accident like that in Chernobyl, the Swedish effective halflife of 3.4 years is the most reliable, but only valid for peat soils. An explanation of the shorter ecological halflife on peat soils could be a faster decrease in the availability of radiocesium from peat than from clay soils. On other soil types uptake is smaller but ecological halflives apparently longer. The variation between different localities can be considerable. Long effective halflives, close to the physical halflife, are not unlikely in the major part of Nordic arable land. In Denmark the ecological halflife of Chernobyl '"Cs in the total diet is 3 years according to Aarkrog (1992).

Species characteristics The uptake of radiocaesium in annual crops was generally lowest in cereals and highest in

191 vegetables where aboveground parts were used. Figure 3.4.2. shows that in 1986 carrots had the lowest caesium content in the three countries studied. Potatoes are not actual root vegetables, and in them the content is higher than in carrots. Perhaps diffusion takes place more easily place from soil to potato tubers than to the anatomically different root crops. The radiocesium content in cabbage is only slightly higher than in carrots. Table 3.4.2. shows that carrot and cabbage had the lowest uptake and pea and potato the highest, about a factor 2 higher, in the years after the Chernobyl accident. Malm et al. (1991) also found the highest uptake from soil in leafy vegetables and the lowest in the root vegetables.

I

Barley

+ Potato

+ m X 0

m Cabbage 0

Carrot X

Pea

I

%

0.1

n

Denmark

I

Finland

I

Sweden

Figure 3.4.2. Comparison of I3’Cs in annual crops in three Nordic countries in 1986. Barley

The annual decrease in content of radiocaesium in barley in Norway, Sweden, Finland and Denmark is shown in Figures 3.4.3.A. and 3.4.3.B. The decrease in Sweden and Norway is exponential, but not so in Denmark and Finland. The explanation could be that resuspension processes dominate over root uptake in Denmark and Finland. Probably this is not the case on the sites investigated in Norway and Sweden.

I92 +

+ + +

+

m

m

x

x

'

0.01 1985

1986

1987

Norway

+

1988

)1(

1989

Sweden

m

1990

Denmark

1991 0

1 32

Finland

-

100-

+ 0

0

I

0

+ n I

+

10: x

+

* %

m

1 1985 I

1986

Norway

1987

1988

+ Sweden

m

1989

*

1990

1991

1 92

Denmark u Finland

Figures 3.4.3.a. and 3.4.3.b. The content of '"Cs in barley in the Nordic countries. a: Absolute values, b: Normalised to 100 from the first year of measurement.

193

CONCLUSION After an event like the Chernobyl accident, a common trend is that levels decrease rapidly from the first to the second year. Thereafter the rate of decrease is more uncertain and it seems that long ecological halflives are possible in agricultural ecosystems. The uptake of radiocaesium from soil through roots to edible parts of annual crops is generally very low in Scandinavian agricultural ecosystems. Aggregated transfer factors~80iCpLnt) ranging between

and 10" mz kg-' seem to be the rule in the Nordic countries. Increased T,, values

are only observed on areas with very special soil types. These peaty organic or sandy soils are often used for purposes other than growing annual crops, e.g. animal or hay production. If contamination levels are high as they were in parts of Sweden there may be grounds for certain countermeasures to be taken, such as the addition of potassium or lime, see chapter 3.8. The most important pathway for the transfer of radiocaesium from annual crops to man is through direct fallout, because of the low uptake from soil. Therefore the season of the year is the most important factor determining the transfer to man after an event like the Chernobyl accident. The Chernobyl accident happened at a time when direct contamination was of minor importance as regards the contamination of annual crops (see chapter 3.2. for seasonality effects).

On the Faroe Islands the uptake is generally between one or two orders of magnitude higher than in the other Nordic countries. This could be due to the special soil properties on these islands, where the basic geological material is basalt. How the basaltic soil influences root uptake is not explained in the investigations. The high content of organic matter and sand may be part of the explanation. The results concerning radiocaesium in Nordic annual crops were not sufficient to calculate reliable effective ecological halflives for the Nordic countries. Radiocesium content is influenced by resuspension and local differences especially in soil properties. However some few examples of effective halflives have been given in the text. An effective halflife for radiocaesium content in

barley of between 5 and 10 years seems reasonable on common arable land soil types in the first years after an accident like that at Chernobyl. In potatoes a similar value of 6 years was calculated for Denmark. The determination and understanding of these halflives provide good arguments for further investigations within this field of research.

REFERENCES Aarkrog, A. 1983. Translocation of radionuclides in cereal crops. In "Ecological aspects of radionuclide release" pp. 8 1-90. Aarkrog, A.; Batter-jensen, L.; Chen, Q.J.; Dahlgaard, H.; Hansen, H.; Holm, E.; Lauridsen, Bente.; Nielsen, S.P.; & Seegaard-Hansen, J. 1988. Environmental radioactivity in Denmark in 1986. Risa-R-549

194 Aarkrog, A.; Nielsen, S.P.; Dahlgaard, H.; Lauridsen, B. & Sagaard-Hansen, J. (1988) Slutrapportering af Risas maeprogram i forbindelse med Tjemobylulykken Risnr-M-2692 Aarkrog, A.; Buch, E.: Chen, Q.J.; Christensen, G.C.; Dahlgaard, H.; Hansen, H.; Holm, E. & Nielsen, S.P. 1988. Environmental radioactivity in the North Atlantic region. Including the Faroe Islands and Greenland. 1986. RIS0-R-550 Aarkrog, A.; Buch, E.: Chen, Q.J.; Christensen, G.C.; Dahlgaard, H.; Hansen, H.; Holm, E. & Nielsen, S.P. 1989. Environmental radioactivity in the North Atlantic region. Including the Faroe Islands and Greenland. 1987. RIS0-R-564 Aarkrog, A.; Bater-Jensen, L.; Chen. Q.J.; Dahlgaard, H.; Hansen, H.; Holm, E.; Lauridsen, B.; Nielsen, S.P.; & Ssegaard-Hansen, J. 1991. Environmental radioactivity in Denmark in 1988 and 1989. Risa-R-570 Aarkrog, A.; Buch, E.: Chen, Q.J.; Christensen, G.C.; Dahlgaard, H.; Hansen, H.; Holm, E. & Nielsen, S.P. 1992. Environmental radioactivity in the North Atlantic region. Including the Faroe Islands and Greenland. 1988 and 1989. RISPI-R-571(EN) Aarkrog, A.; Bater-Jensen, L.; Chen. Q.J.; Dahlgaard, H.; Hansen, H.; Holm, E.; Lauridsen, B.; Nielsen, S.P.; Strandberg. M. & Saegaard-Hansen, J. 1992. Environmental radioactivity in Denmark in 1990 and 1991. Riss-R-621(EN) Bjerke, H. 1987. Radioaktivt cesium i korn 1986. SIS 1987:l Bjerke, H. & Bakken, E. 1988. Radioaktivt cesium i korn 1988. SIS ARBEIDSDOKUMENT 1988:7 Eriksson, A. & RosCn, K. 1989. Cesium transfer to agricultural crops for three years after Chernobyl. In "The radioecology of natural and artificial radionuclides" Proceedings of the XVth regional congress of IRPA Visby, Gotland, Sweden, 10-14 September, 1989. Magnusson, S.M. personal communication. Malm, J.; Rantavaara, A,; Uusi-Rauva, A. & Paakola, A. 1991. Uptake of Caesium-137 from peat and compostmould by vegetables in a greenhouse experiment. J. Environ. Radioactivity 14 (1991) 123-133. Mascanzoni, D. 1986. The aftermath of Chernobyl in Sweden: Levels of Cs-137 in foodstuffs Rapport SLU-REK-62 Moberg, L. (ed) 1991. The Chernobyl fallout in Sweden. Swedish Radiation Protection Institute, 1991 Puhakainen, M. & Ylaranta, T. 1990. Uptake of radionuclides by spring wheat and barley from cultivated soils added with sewage sludge. NJF-UtredninglRapport Nr. 59, 1990 ISSN 03331350 "Deposition and transfer of radionuclides in Nordic terrestrial environment" Rantavaara, A. Personal commmunication. Rantavaara, A. 1987. Radioactivity of vegetables and mushrooms in Finland after the Chernobyl accident in 1986. STUK A-59. Helsinki 1987 Rantavaara, A. 1987. Transport av radiocesium till livsmedel fran tjernobyl nedfallet -JAmforelse mellan de tv2 forsta 2ren. Data presented at "Det Femte Nordiska Radioekologiseminaret2225 Augusti 1988, Rattvik, Sverige Rantavaara, A. 1991. Radioactivity of foodstuffs in Finland in 1987-1988. STUK-A78. Helsinki June 1991. Rantavaara, A. & Haukka, S. 1987. Radioactivity of milk, meat, cereals and other agricultural products in Finland after the Chernobyl accident in 1986.STUK A-58. Helsinki 1987 Rantavaara, A. & Kostiainen, E. 1993. Radioactivity of foodstuffs in Finland in 1989-1990. STUK A-95, Helsinki (to be printed in 1993). Roca, V. ; Napolitano, M.; Speranza, P.R. & Gialanella, G. 1989. Analysis of radioactivity levels in soils and crops from the Campania region (South Italy) after the Chemobyl accident. J. Environ. Radioactivity 9 pp. 117-129. RosCn, K. & Feuerbach, P. 1988. Faltforsok visar: Kalimagnesia minskar vaxters upptag av cesium. Biodynamisk tidsskrift 2 pp. 22-23.

195

Rostn, K. 1989. Effects of potassium on the cesium transfer to the crops after Chernobyl. In "The radioecology of natural and artificial radionuclides" Proceedings of the XVth regional congress of IRPA Visby, Gotland, Sweden, 10-14 September 1989. (Ed. W. Feldt). RosCn, K. 1991. Effects of potassium fertilization on caesium transfer to grass, barley and vegetables after Chernobyl. Department of Radioecology, Swedish Univ. of Agnc. Sciences, Uppsala, 1991, pp.305-322. In "The Chernobyl fallout in Sweden" Ed. L. Moberg. Swedish Radiation Protection Institute. SIS 1990. Malinger fra sesongen 1990, unpublished. SaxCn, R.; Taipale, T.K. & Aaltonen, H. 1987. Radioactivity of wet and dry deposition and soil in Finland after the Chernobyl accident in 1986. STUK A-57. Helsinki 1987 SSI-RAPPORT 1988. Projekt Tjernobyl - Lagesrapport 3. SSI-rapport 88-13 STUK A-55 1987. Studies on environmental radioactivity in Finland in 1986. STUK A-55 Strand, T.; Strand, P. & Baarli, J. 1987. Radioactivity in foodstuffs and doses to the norwegian population from the Chernobyl fall-out. Radiation protection dosimetry 20 pp. 221-229

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197

3.5. TRANSFER OF 13'Cs TO COWS' MILK IN THE NORDIC COUNTRIES

HANNE SOLHEIM HANSEN' and INGER ANDERSSON'" Agricultural University of Norway, P.O.Box 5025, N-1432 As, Norway.

* Swedish University of Agricultural Sciences, P.O.Box 59, S-230 53 Alnarp, Sweden.

SUMMARY A comparison has been made of the transfer of Chernobylderived 137Csto cows' milk in the different Nordic countries. A compilation is given of data on 137Cslevels in both dairy milk and milk from individual farms. In 1986 and 1987 the levels of 137Cswere highest in Finland and Norway, intermediate in Sweden, the Faroe Islands and Iceland (137Csfrom global fallout only) and lowest in Denmark. The aggregated transfer coefficient (T,) to cows' milk was 2-10 times higher in the Faroe Islands, Iceland and Norway compared to that in Denmark, Finland and Sweden for all years after 1986. The effective ecological half-life (T,/l,,) for dairy cows' milk ranged from 1-2.3 y for all countries, except Iceland where the T, was 18.4 y (global fallout). It was therefore concluded that cows' milk production in the Faroe Islands and Norway was most sensitive to the Chernobyl "%s fallout. Though milk production systems and management systems change over time and could alter the sensitivity to 13'Cs fallout, it is concluded that the Faroe Islands, Iceland and Norway would be most susceptible to future 13'Cs fallout. INTRODUCTION Following the Chernobyl nuclear accident in 1986 several studies were made in the Nordic countries Denmark, the Faroe Islands, Finland, Iceland, Norway and Sweden on the transfer of 137Csfrom feed to cows' milk (Aarkrog, 1992 a, b, Aarkrog et al., 1992, Rantavaara & Haukka, 1987, Rantavaara, 1991, Rantavaara & Kostiainen, 1993, Prllson et al., 1993, Strand, 1994, Strand & Hove, 1992, Bjaresten, 1986, Samuelsson & Josefsson, 1986, HAkansson et al., 1987, Bertilsson

et al., 1988, Karlen et al., 1991, Alskog, 1992, Suomela & Melin, 1992). The purpose of these studies was mainly to measure the level of 137Cscontamination of cows' milk and to predict the dose to humans via cows' milk. Action guidelines for the content of 137Csin cows' milk for human

*

Data has been available by Asker Aarkrog, Research Center Risa, Denmark, Sigurdur Prllson, Geislavarnir Rikisins Reykjavik, Iceland, Aino Rantavaara, Finnish Center for Radiation and Nuclear Safety, Finland, Gunnel Karlh, Swedish University of Agricultural Sciences, Sweden, Inger Bjiresten, County Agricultural Board of Jhtland, Sweden, Christer Samuelsson, University of Lund, Sweden, Per Strand, Norwegian Radiation Protection Authority, Norway.

I98

consumption were in 1986 1000 Bq L“ in Finland and Denmark, 300 Bq L-I in Sweden and 370 Bq L-’ for 137Cs+ 134Csin Norway. In the studies in question the content of 137Csin milk was related in different ways to the deposition in the area: as the transfer coefficient (ratio of I3’Cs concentration in milk to activity ingested daily by the cow), as an aggregated transfer coefficient (ratio of 137Csconcentration in milk to I3’Cs deposited per m2 ground surface), and in some cases as the effective ecological half-life. The effective ecological half-life was estimated in production systems unchanged over the years as the time needed for a 50% reduction of the I3’Cs concentration in milk. The results given in the literature indicated some differences between the countries in the transfer of 137Csto cows’ milk. Therefore it was considered necessary to review these results, make additional estimations, and to collect additional data in 1992. The purpose of the present study was to estimate the sensitivity of cows’ milk production to ‘37Cs fallout in the Nordic countries by comparing estimated data on activity transfer and effective ecological half-lives, respectively.

MATERIALS AND METHODS The material was based on data from measurements of 137Csin 1) milk sampled from dairies or dry-milk factories and 2) in milk from individual farms. The samples from dairies or dry-milk factories were mainly collected for surveying the 137Cslevels in milk or milk products for human consumption. The 137Cslevels in milk from individual farms were measured as part of a survey

or research programmes to estimate the radiocaesium transfer to cows’ milk, or the effective ecological half-life in the present production systems.

Sampling schedule and measurements of I3’Cs activity concentrationin dairy milk or dry-milk Denmark Data from Denmark were given by Aarkrog (1992 a, b, Table 3.5.1). Results were based on pooled milk samples collected in 1987-1991 at dry-milk factories supposed to represent milk production in the whole country (Table 3.5.1). Data on milk yield and feed intake of the cows used in the calculations of the transfer coefficients were based on the mean values given by Danish Agricultural Statistics. The ground deposition of L37Csfrom the Chernobyl accident in 1986 was considered to be uniform over the country with a mean of 1.2 kBq m-’ (Table 3.5.1). In the calculations the ‘37Csfrom global fallout was separated from the Chernobyl 137Cs. The Faroe Islands Data on measurements of 137Csin milk from three dairies in the Faroe Islands for 1987 to 1991

Table 3.5.1. Mean ground deposition of 13’Cs in each country corrected for areal distribution of milk production. Sampling schedule for measurement of 13’Cs in cows’ milk from dairies or dry-milk factories. The sampling represented 50-100%of milk production in the different countries. Country

Mean deposition of 13’Cs in 1986, kBq m-*

Denmark Faroe Islands Finland

1.2 1.3 15.5-17.2

Period

No of dairies

Sampling frequency

Mean value based on

1987-1991 1987-1991

7 3

Monthly Monthly

1986-1991

13

Daily-monthly 34 w,12 mo

12mo 12 mo

Iceland Norway

2.0” 9.8

1986-1992 1986-1991

2 1-125

Monthly Monthly

Sweden

7.0

1986-1990

32-50

2 weeks-monthly 3 mo

a

Global fallout

x Estimations made in the original work z Estimations made in the present work

12 mo 1 mo

F,

x

Estimations T, T,,

X

x

X

x

X

z

Z

z

Z

z

Z

x

References

Aarkrog, 1992 a, b Aarkrog, 1992 a, Aarkrog et al., 1992 Rantavaara & Haukka, 1987, Rantavaara, 1991, Rantavaara Kostiainen, 1993 Pason et al., 1993 Strand, 1994 Backe et al., 1986 Suomela & Melin, 1992

&

200 were given by Aarkrog (1992 a, Table 3.5.1). The milk samples represented milk production in the whole country. The ground deposition of Chernobyl 137Cswas considered to be uniform over the areas with a mean of 1.3 kBq m-' (Aakrog et al., 1992). In the calculations the '37Cs from global fallout was separated from the Chernobyl 137Cs. Finland Data from Finland for 1986 to 1991 were based on results given by Rantavaara & Haukka (1987), Rantavaara (1991) and Rantavaara & Kostiainen (1993) (Table 3.5.1). The milk sampling of dairy milk represented about 50% of milk production in the country. The ground deposition of Chernobyl 137Csin 1986 in Finland was patchy and varied between 0 and 67 kBq m-'. Because of this patchiness the mean ground 137Csdeposition was corrected to 15.5-17.2 kBq m-', according to the areal distribution of milk production. The data for transfer parameters refer to L37Csfrom Chernobyl fallout only. Iceland Data from Iceland were given by PBlson et al. (1993) (Table 3.5.1). Dry-milk samples from two factories were collected from 1986 to 1992. The milk samples represented the total milk production on Iceland. The mean ground deposition of '37Cswas 2.0 kBq m-'. The 137Csground deposition from the Chernobyl accident was negligible, and the I3'Cs recorded in milk was thus considered to originate from global fallout only. Norway Pooled milk samples from dairies were measured for 137Csfrom 1986 to 1991 (Strand, 1994, Table 3.5.1). The single dairy was represented on one sampling occasion only. The ground deposition of 137Cswas patchy in Norway and varied from 1-103 kBq m-' in the different counties (Backe et al., 1986). The given values included 137Csfrom both global fallout and the Chernobyl accident. The contribution from global fallout was, however, negligible compared to that from Chernobyl accident. Because the fallout deposited patchily the mean ground '37Csdeposition was corrected according to the areal distribution of milk production, the mean being 9.8 kBq m-* (Backe et al., 1986, Agricultural Statistics, 1993, Table 3.5.1). Sweden Data on 137Csconcentration in milk samples collected by The Swedish Radiation Protection Institute from dairies in 1986-1990 were given by Suomela & Melin (1992) (Table 3.5.1). All these values include 137Csfrom global and from Chernobyl fallout. The contribution from global

20 1 fallout was, however, negligible compared to that from Chemobyl. The Chernobyl 137Cswas deposited unevenly in Sweden with the lowest levels in the south (0-2 kBq m-*)and the highest in the east and north ( > 80 kBq m-’) (Swedish Geological Company, 1986). Because the fallout was deposited patchily, a mean ground 137Csdeposition was calculated for the present study according to the areal distribution of milk production, the mean being 7.0 kBq m-’ (Table 3.5.1).

Sampling schedule and measurements of ‘37Csactivity concentration in cows’ milk from individual farms Norway Norwegian data were available from studies of four dairy farms (farms A-D, Table 3.5.2) during the grazing periods in 1988-1992 (Strand & Hove, 1992). These farms all used unimproved mountain pasture at an altitude of 900-1000 m. Soil and grass samples were taken from the pastures in 1989 and 1990. Sweden Swedish data were based on studies of 12 different farms (farms E-P, Table 3.5.2). One of these was an experimental farm (farm E, Table 3.5.2, Bertilsson et al., 1988), where the cows were fed individually with green-cut grass contaminated by Chemobyl fallout. Individual milk yield recording and measurements of I3’Cs were carried out. The other studies (farms F-P, Table 3.5.2) were made under pasture conditions. The herd size, milk yield and feed rations of the herds were recorded and used to estimate the quantity of pasture consumed daily per cow. The frequency of grass and bulk-milk sampling is given in Table 3.5.2.

Calculations and units used The transfer coefficients (Fm),aggregated transfer coefficients (Tag),and effective ecological halflife (Tlhmol) of 137Csto cows’ milk were estimated for the different materials according to Tables 3.5.1 and 3.5.2. The THmol-values were estimated from samples of milk from the same farms or areas where similar conditions prevailed each year.

RESULTS Caesium-137 activity concentration in milk The mean 137Csactivity concentration in milk from the dairies varied from 0.6 to 20 Bq L-’ in the different countries in 1986 and 1987 (Fig. 3.5.1). The content of *37Csdecreased in all countries from 1986 or 1987 to 1991 or 1992. In this period the content of 137Cswas highest in milk from Norway and Finland, lowest in milk from Denmark, and intermediate in milk from Sweden, the

h)

Table 3.5.2.Characteristics of individual farms and sampling schedule for 137Csmeasurements in feeds and cows’ milk. Country Farmlherd code. (Number of farms)

Ground Year deposition of 137Csin farms or areas, kBq In-2

Norway A,B,C,D 55-200 (4)

Sweden

E (1)

30-60“

Sweden

F (1)

1.1

Sweden

G,H,I (3)

0-2”

Sweden

J, K (2)

85, 10-30”

Sweden

L,M,N, 0 , P (5)

1988 1989 1990 1991 1992 1986

Comments

Fm

Uncultivated working farms. Grazing seasodmountain pasture, 5-6weeks. Weekly milk sampling. Soil and grass on one occasion in 1989 and 1990. In 1988 only farm A was studied.

Cultivated experimental farm. Two groups of 10 cows individually fed grass cut at stubble height of 5 and 15 cm, respectively. Sampling of feed daily and milk from each cow twice daily. 1986 Cultivated working farm. Grazing season, 4 months. 1992 One and three samples of grass, one of bulk milk.

1986 Cultivated working farms. Grazing season. 1992 Three samples of grass, three of bulk milk from each farm. In 1992 one herd was housed and fed greencut grass; two herds grazing.1-3 samples of grass and bulk milk from each farm. 1986 Cultivated working farms. Grazing season. 1-3 1992 samples of grass and bulk milk from each farm.

1987 Cultivated working farms. Grazing season. Daily 150 (farm P) 1988 sampling of grass and milk.

30-60,”

According to the Swedish Geological Company (1986). x Estimations made in the original work z Estimations made in the present work a

Estimations

Tag

X

X

8 Reference

T112scol

x

Strand&Hove, 1992

Bertilsson et al., 1988

Z

X

Z

z

H b s o n et al., 1987 Anderson & Molin,

X

Z

z

Samuelsson & Josefsson, 1986,Anderson & Molin, 1992

X

Z

z

Bjtiresten, 1986 Anderson &

1992

Molin,

1992 X

Z

Karlen et al., 1991 Alskog, 1992

203 Faroe Islands and Iceland (Fig. 3.5.1). 0 Finland

Norway Sweden Faroe Islands

0.05

1

I

I

I

I

I

I

I

1986 1987 1988 1989 1990 1991 1992

Fig. 3.5.1. Mean ‘37Csactivity concentration in cows’ milk from dairies or dry-milk factories in the Nordic countries from 1986 to 1992. Values refer to Chernobyl 137Cs, except in Iceland where global 137Cs only was included. Transfer coefficient The F,-values estimated for Danish conditions based on mean 137Csconcentrations in dry-milk and mean feed consumption ranged from 0.60*10-2to 1.20*102d L-’ from 1987 to 1991 (Table 3.5.3). The F,-values estimated for farm E in Sweden in 1986 were 0.19*102 and 0.67*10-2d L-* for grass harvested at stubble heights of 5 cm and 15 cm, respectively. The overall mean F,-value valid for the other farms ranged from 0.45*10-2to 2.85*10-’ d L-‘ from 1986 to 1992 (Table 3.5.3). Farm K had in 1992 a considerably higher F,-value of 6.82*102 d L”, which resulted in the high standard deviation of the mean for this year (Table 3.5.3). In 1992 the 137Cscontent in milk was below the limit of detection (2 Bq L-’) on farms G, H, and I, and therefore F,-values were not estimated. When comparing individual data for the farms from 1986 to 1992 a tendency, though not statistically significant, to increasing F,-values was observed (Table 3.5.3). The general mean F,-

204 value for the whole material, including data for dry-milk, dairy milk and milk from individual farms, independant of year of measurement and country, was 0.94*1.32*10* d L-’ (N=23). Table 3.5.3. Mean transfer coefficients (F,-values, d L-’) for 137Csfrom feed to cows’ milk based on mean values for 137Csin dry milk from factories in Denmark and on mean values for *37Csin milk from individual farms in Sweden. Country Farm

.

1986

1987

1988

Year 1989 1990

1991

1992

E F G H I J K L M N 0 P

0.0060 0.0120 0.0110 0.0100 0.0100 0.0019,0.0067 0.0043 0.0100 0.0039 0.0042 0.0037 0.0014 0.0073 0.0097 0.0682 0.0060 0.0070 0.0050 0.0050 0.0036 0.0076 0.0112

Overall mean SD N

0.0045 0.0066 0.0080 0.0110 0.0100 0.0100 0.0285 0.0027 0.0026 0.0036 0.0344 8 6 3 1 1 1 3

Denmark Sweden

-

= 137Csactivity concentration in milk was below the detection limit

Aggregated transfer coefficient The Tagestimated as mean for the whole country (Fig. 3.5.2) decreased with time after 1986 for all countries. The Tagwas lowest for Denmark, Finland and Sweden and about 2-10 times higher for the Faroe Islands, Iceland and Norway (Fig. 3.5.2). The Tag estimated for individual farms (Fig. 3.5.3) showed the same pattern as for whole countries with decreasing values from 1986 to 1992. For the Swedish farms, the mean values decreased from 2.9*10” to 0.02*10-3m2 L-I from 1986 to 1992, and for the Norwegian farms the mean values decreased from 4.7*103 in 1988 to 3.4*1B3m2 L-I in 1991. The T, was not estimated for Swedish farms G, H and I in 1992 because the 137Cslevels in milk were below the detection limit.

205

4.0

1

3.0

-

2.0

-

Denmark Faroe Islands fl Finland Iceland 69 Norway Sweden

I

-

d 1

=l 4

-

N

E

m I

0 d

*

F

E

1.0

0.0

1986

1987

1988

1989

1990

1991

1992

Fig, 3.5.2. Mean aggregated transfer coefficients (T,) of Chernobyl I3’Cs to cows’ milk in the Nordic countries (global 137Csin Iceland) from 1986 to 1992. Estimations based on mean ground deDosition. where needed corrected according to the areal distribution of milk production and mean 137kscontent in dairy milk.

8

E Sweden 0

.

1986

1987

1988

1989

a Norway

1990

1991

1992

Fig, 3.5.3. Mean aggregated transfer coefficients (Tag)and standard deviations of 137Csto COWS’ milk estimated at individual farms in Norway and Sweden from 1986 to 1992. See Table 3.5.2 for references.

Effective ecological half-life The TjheOl for the transfer of ‘37Csto cows’ milk estimated for the whole countries (Table 3.5.4) ranged from 1 to 2 years, except for Iceland where the T,heolwas estimated to 18.4 years. The of individual farms in Sweden was Icelandic data refer to global fallout only. The mean THecol similar to that of the whole country (1.0 year) from 1986-1992 (Table 3.5.4). For the farms studied in Norway the mean TIheco, was 4.8 years for three of the farms and 30 years for one farm (Table 3.5.4). Table 3.5.4. Effective ecological half-life (THecol) for Chernobyl 137Csin cows’ milk sampled from dairies and dry-milk factories or individual farms. Country

TIE, years

SD

Range

Years

Comments

1987-1991 1987-1991 1986-1991 1986-1992 1986- 1991 1987-1991 1986- 1990

Global fallout

1988-1992 1989-1992 1986-1992

Farm A Farm B-D Farm F-K

Milk from dairies or dry-milk factories Denmark Faroe Islands Finland Iceland Norway Sweden

1.6 1.6 1.4 18.4 2.0 2.3 1 .o

Milk from individual farms Norway

30

4.8 Sweden

1 .o

1.7 1.4

2.7-6.7 0.2-3.7

DISCUSSION Transfer coefficients The mean F,-values ranged from 0.45*1U2to 2.85*lO-’d L-I, with a general mean of 0.94k1.32

*lo-’ d L-’. These values were in agreement with values reported in the literature ranging from 0.25 to 0.41 (Ward et al., 1967) and with values reviewed by Anderson (1989) and Bertilsson et al. (1988) ranging from 0.05-1.42*lo-’ d L-’. Results from the present material showed that the transfer of 137Csfrom vegetation to cows’ milk did not increase significantly from 1986 to 1992, which was in contrast to other studies showing increased transfer coefficients to lamb meat and goat milk after the first harvest following the deposition of the fallout (Howard et al., 1989; Hansen & Hove, 1991). The present results with stable F,-values with an overall mean of

0.94*10-’ d L-I indicate that transfer from vegetation to cows’ milk did not depend on the year.

207 Aggregated transfer coefficient The Tagdecreased both for milk from individual farms and for milk from dairies (Fig. 3.5.2, Fig. 3.5.3) showing that less and less of the deposited 13’Cs was transferred to cows’ milk each year after the deposition. Assuming that F,-values were similar for the period of time in which measurements were made, the decreasing T, indicated net fixation or washout processes for 137Cs in the soil. Estimates of Tagfor 137Cswere highest in the Faroe Islands, Iceland and Norway and lowest in Denmark, Finland and Sweden, indicating that cows’ milk production in the Faroe Islands, Iceland and Norway was considerably more sensitive to 137Csfallout than it was in the other Nordic countries. Similar results were observed in a study of transfer of radiocaesium to lambs’ meat (Hove et al., 1994), where the Tagvalues for Norway, Iceland and Sweden were about 10 times higher than in Denmark, Finland and the Faroe Islands. The observed shift for Sweden from the low Tagvalue for cows’ milk to the high Tag value for lambs’ meat can be explained by the study site for the lambs’ meat experiment. The lambs grazed an uncultivated mountain pasture grown on either peat or gravelly and sandy moraine, where high T, values would be expected. The reason for the observed shift for the Faroe Islands from high T, value for cows’ milk to low Tagvalue for lambs’ meat remains unclear, but is possibly due to interplay between soil types and the different production systems for cows’ milk and lambs’ meat. The Tagfor milk from individual farms in Norway was 3-5 times higher than T, for dairy milk in this country. The dairy milk was produced mainly on farms with intensive use of high quality roughage and concentrates, and less than 5% of the milk was produced on uncultivated pastures (Agricultural Statistics, 1993). However, the individual farms all used uncultivated mountain pastures during the measuring periods. Assuming stable F,-values these results indicated considerably higher transfer from soil to vegetation on uncultivated pastures than on cultivated pastures. This is in agreement with reports from Underdal et al. (1967) where a high degree of correlation was observed between intensive farming and feeding practices and reduced levels of 137Csin milk. At the individual farms no preussian-blue was mixed in the concentrate to reduce the radiocaesium levels in milk. This countermeasure has only been used on about 0.5% of all Norwegian dairy cows since 1989 (Helle, 1994). This small amount of use would therefore not be expected to affect the mean T, values estimated for dairy milk. Fallout deposited as a result of power-plant accidents leading to patchy deposition of 137Cs would affect the values of Tagestimated as means for the whole countries. The transfer of 137Cs from soil to vegetation varies with soil type (Eriksson & RosCn, 1991) and as the present compilation indicates also with management practice. The values of T, estimated in the present study for the Chernobyl fallout are therefore not necessarily valid for a future fallout depositing

208 differently compared to the Chemobyl fallout. Effective ecological half-life for Chemobyl 137Csin cows’ milk measured in dairy milk ranged between 1 and 2.3 The Tthecol years in all countries except for Iceland (Table 3.5.4). The THecol in cows’ milk of less than 30 years indicated that net fixation processes or wash-out were taking place in soil. The results from Sweden showed both for milk from individual farms and for dairy milk that the TIAccol was 1 year. Corresponding data from Norway showed twice as long a TBA,, for milk from individaul farms using uncultivated pastures than for dairy milk mainly produced on cultivated pastures. This is in agreement with results from Sweden where a THW,of 3 years was observed in milk from some dairies receiving most of their supplies of milk from farms where uncultivated pastures were used (Suomela & Melin, 1992). Effective ecological half-lives were estimated in some studies for global fallout. On Iceland it was estimated to 18.4 years. Effective ecological half-lives for global fallout of 1 . 5 , 2.8, 7 . 1 and

4.4 years were reported for Sweden, the Faroe Islands, Denmark and Norway, respectively (Suomela & Melin, 1992, Aarkrog, 1992 a, Hove et al., 1989). The Swedish and Norwegian estimates refer to periods of time when global fallout was still being deposited. The values of T,h,, corrected for continuous 137Csdeposition were therefore shorter than those given above. A longer TIh,, for global 137Csfallout compared to Chernobyl 137Csfallout indicates that future estimates of the Tghecol for Chernobyl fallout may increase with time, and possibly reach values similar to those observed for global fallout.

CONCLUSIONS The present study showed that the transfer of ‘37Csto cows’ milk related to ground deposition was highest in the Faroe Islands, Iceland and Norway and lowest in Denmark, Finland and Sweden. The effective ecological half-life for Chemobyl ‘37Cs ranged from 1-2 years for all the Nordic countries and was 18.4 years for global 137Csfallout in Iceland. It was therefore concluded that in the Nordic countries cows’ milk produced in the Faroe Islands, Iceland and Norway was most sensitive to 137Csfallout.

ACKNOWLEDGEMENT We are indebted to the Norwegian Research Council and the Swedish University of Agricultural Sciences which partly financed this report. We highly appreciate the researchers Asker Aarkrog, Inger Bjaresten, Sigurdur Pilson, Aino Rantavaara, Gunnel KarlCn, Christer Samuelsson and Per Strand who have shared their results with us in order to prepare this report.

209 REFERENCES Aarkrog, A. (1992 a). 0kologiske halveringstider i Faererske og Danske landbrugsskosystemer. Det Sjette Nordiske Radioerkologi Seminar, Torshavn, Foroyar, 14-18 Juni 1992. (In Danish). Aarkrog, A. (1992 b). Personal communication. Riser National Laboratory, Roskilde, Denmark. Aarkrog, A., Buch, E., Chen, Q . J., Christensen, G. C., Dahlgaard, H., Hansen, H., Holm, E. & Nielsen, S. P. (1992). Environmental Radioactivity in the North Atlantic Region Including the Faroe Islands and Greenland. 1988 and 1989. Riser-R-571(EN). Riser National Laboratory, Roskilde, Denmark. 97 pp. Agricultural Statistics 1991. (1993). NOS C 71, Statistisk Sentralbyri, Oslo, Kongsvinger. Alskog, E. (1992). Lokala undersokningar i jordbruket efter Tjernobylolyckan 1986. I. Overforing av radiocesium frHn betesmark till mjolk i Gavleborgs Ian 1987. Rapport, SLU-REK-69 Department of Radioecology , Swedish University of Agricultural Sciences, S-75007 Uppsala, Sweden. (In Swedish). Andersson, I. (1989). Safety precautions in Swedish animal husbandry in the event of nuclear power plant accidents. Report 181, Dissertation, Department of Animal Nutrition and Management, Swedish University of Agricultural Sciences, S-75007 Uppsala, Sweden. Andersson, I. & Molin, M. (1992). Concentrations of 137Csin grass and milk and calculated transfer coefficients in six dairy herds in Sweden studied in the pasture periods of 1986 and 1992. Unpublished report given at the working group meeting within the Nordic Nuclear Safety Research RAD 3 programme, Oslo, 24-25 November 1992. Backe, S . , Bjerke, H., Rudjord, A. L. & Ugletveit, F. (1986). Nedfall av cesium i Norge etter Tsjernobylulykken. SIS rapport 1986:5. (In Norwegian). 49 pp. Bertilsson, J., Andersson, I. & Johanson, K. J. (1988). Feeding green-cut forage contaminated by radioactive fallout to dairy cows. Health Physics 55; 855-862. Bjaresten, I. (1986). Personal communication. County Agricultural Board in the county of Jiimtland, Ostersund, Sweden. Eriksson, A. & Rostn, K. (1991). Transfer of caesium to hay grass and grain crops after Chernobyl. In: The Chernobyl Fallout in Sweden. Results from a research programme on environmental radiology. Ed by L. Moberg, 291-304. The Swedish Radiation Protection Institute, Stockholm, Sweden. Hansen, H. S. & Hove, K. (1991). Radiocesium bioavailability: Transfer of Chernobyl and tracer radiocesium to goat milk. Health Physics 60;665-673. Helle, A. M. (1994). Personal communication. Governmental Foodinspector for Valdres, 2943 Rogne, Norway. Hove, K . , Strand, P. & IZlsteris, 0. (1989). Varighet av cesiumproblemet i norsk husdyrproduksjon. Rapport, Institutt for husdyrfag, Norges Landbrukshcrgskole. (In Norwegian). 27 PP. Hove, K . , Liinsjo, H., Andersson, I . , Sormunen-Cristian, R., Hansen, H. S . , Indridason, K., Joensen, H. P., Kossila, V., Liken, A., Magnusson, S . , Nielsen, S. P., Paasikallio, A., PBlsson, S. E., Rosen, K., Selnes, T., Strand, P. & Vestergaard, T. (1994). Radiocaesium transfer to grazing sheep in Nordic environments. In: Nordic Radioecology. Ed by H. Dahlgaard, Elsevier Science Publishers, Amsterdam. Howard, B. J . , Mayes, R. W., Beresford, N. A. & Lamb, C. S. (1989). Transfer of radiocesium from different environmental sources to ewes and suckling lambs. Health Physics. 57;579-586. HHkansson, E., Drugge, N., Vesanen, R., Alpsten, M. & Mattsson, S. (1987). Transfer of ‘34Cs, 137Csand lS1Ifrom grass to cow’s milk. A field study after the Chernobyl accident. Report GU-RADFYS 87:Ol .Department of Radiophysics, The Sahlgren Hospital, S-413 45 Goteborg, Sweden. 47 pp.

210 Karlkn, G., Johanson, K. J. & Bertilsson, J. (1991). Transfer of cesium-137 from pasture to milk after Chernobyl. Investigationsof dairy farms in Sweden. In: The Chernobyl Fallout in Sweden. Results from a research programme on environmental radiology. Ed by L. Moberg, 343-360. The Swedish Radiation Protection Institute, Stockholm, Sweden. Phlson, S. E., Magnusson, S. M. & Olafsdottir E. D. (1993). Measurements of '37Cslevels in dry milk produced in Iceland 1986 to 1992. Report, Geislavarnir Rikisins Reykjavik, Laugavegur 118D, IS-150 Reykjavik, Iceland. 5 pp. Rantavaara, A. (1991). Radioactivity of foodstuffs in Finland in 1987-1988. Report STUK-A74. Finnish Centre for Radiation and Nuclear Safety, Helsinki, Finland. Rantavaara, A. & Haukka, S. (1987). Radioactivity of milk, meat, cereals and other agricultural products in Finland after the Chernobyl accident in 1986. Report STUK-AS8. Finnish Centre for Radiation and Nuclear Safety, Helsinki, Finland. 106 pp. Rantavaara, A. & Kostiainen E. (1993). Personal communication. Finnish Centre for Radiation and Nuclear Safety, Helsinki, Finland. Samuelsson, C. & Josefsson, D. (1986). Personal communication. Department of Radiophysics, University of Lund, Lund, Sweden. Strand, P. (1994). Radioactive fallout in Norway from the Chernobyl accident. NRPA Report 1994:2, Dissertation, Norwegian Radiation Protection Authority, 1345 0steris, Norway. Strand. P. & Hove, K. (1992). Long-term behaviour of radiocaesium in Norwegian semi-natural ecosystems, unpublished. Suomela, J. & Melin, J. (1992). Forekomsten av cesium och strontium-90 i mejerimjolk for perioden 1955-1990. SSI-rapport 92-20. Statens strilskyddsinstitut (The Swedish Radiation Protection Institute). Stockholm. Sweden. (In Swedish). 15 pp. Swedish Geological Company. (1986). Map of 137Csground contamination in Sweden. Results from aerial surveys May to October 1986. Uppsala, Sweden. Underdal, B., 0degaard, 0. A. & 0verland N. 0. (1967). The influence of farming and feeding practices on Cs13?and SrWconcentration in fodder, milk and excreta. Acfu Vet. Scund. 8;89-97. Ward, G. M., Johnson, J. E. & Sasser, L. B. (1967). Transfer coefficients of fallout cesium-137 to milk of dairy cattle fed pasture, green-cut alfalfa or stored feed. J. Dairy Science 50;1092-1096.

21 1

3.6. RADIOCAESIUM TRANSFER TO GRAZING SHEEP IN NORDIC ENVIRONMENTS

*Knut Hove', *Hans LiinsjcF. Inger Anderssonl, Riitta Sormunen-Cnstian', Hanne Solheim Hansen', K h i Indndasond, Hans Pauli Joensenb, Vappu Kossila', Andrew Liken', Sigurdur M. Magnhsond, Sven P. Nielsen", A j a Paasikallio', Sigurdur E. Pblssond, Klas Rasing, Tone Selned, Per Strand, Jbhann Thorssod, W g v i Vestergaardb. Riss National Laboratory, DK-4000 Roskilde, Denmark. Department of Natural Sciences, University of the Faroe Islands, FR-100 Torshavn. Agricultural Research Centre of Finland, FIN-3 1600, Jokioinen, Finland. National Institute of Radiation Protection, IS-150 Reykjavik, Iceland. Department of Animal Science, Agricultural University of Norway, 1432 As, Norway. Norwegian Radiation Protection Authority, 1345 OsterBs, Norway. g Department of Radioecology , Swedish University of Agricultural Sciences, S-75007 Uppsala, Sweden. Southern Animal Experimental Station, S-230 53 Alnarp, Sweden. Agricultural Research Institute, IS-1 12, Reykjavik, Iceland.

a

J

* Authors to whom correspondence should be addressed. ABSTRACT Radiocaesium transfer in the soil-herbage-lamb food chain was assessed in a four-year trial conducted in sheep production locations of the Nordic countries. Radiocaesium contamination of the topsoil ranged from 3 to 30 kBq m-*and was predominantly of Chernobyl origin in Finland, Norway, and Sweden, whereas in Iceland 137Cswas primarily of nuclear weapons test origin, and in Denmark and the Faroe Island contamination was derived from both sources. Soil-to-herbage radiocaesium transfer factors were high on the organic and acidic soils of the Faroe Islands, Iceland, Norway, and Sweden, averaging 18-82 Bq 137Cs kg-' herbage on a soil deposition of 1 kBq 137Csm-', and much lower on the sandy soils of Denmark and clay soils in Finland (0.4-0.8). Herbage-to-lamb concentration factors were generally more homogeneous, with values ranging from 0.25-0.70, indicating that the absorption of radiocaesium from herbage was similar in each of the countries. A I3'Cs deposition of 1 kBq m-* soil gave rise to much lower meat radiocaesium concentrations at the sites in Denmark, the Faroe Islands, and Finland (0.5-3.0 Bq kg-I) than in Iceland, Norway, and Sweden (20-47 Bq kg-'). Major factors which will determine the time-integrated dose of radiocaesium transferred to man are levels of consumption of lamb meat, aggregated transfer factors from soil to meat, and effective ecological halflives of 137Csin the production system. It is concluded that among the Nordic countries the soil-herbage-lamb pathway is clearly of greatest importance in Iceland and Norway, intermediate in the Faroe Islands, and of comparatively lesser importance in Denmark and Sweden.

212

INTRODUCTION In order to study variation in the transfer of radiocaesium from soil to vegetation and to grazing lambs, an inter-Nordic investigation was initiated in 1990 by the NKS working group RAD-3. By participation of all of the Nordic countries during the period 1990 to 1993, possible regional differences in the transfer of radiocaesium to grazing lambs, as influenced by variation in climate, soils, and other environmental factors, as well as agricultural practices, could be demonstrated.

Sheep farming in Nordic countries Sheep farming in Nordic countries is highly influenced by the latitude. Plant growth does not take place in sufficient quantities to sustain grazing for the full year in any of the countries. Winter feeding with hay or silage is therefore commonly practised, but the length of the feeding season varies from a couple of months in the southernmost and coastal parts of the area, including Denmark and the Faroe Islands, to 7 months or even longer in the northern and mountainous parts of Scandinavia, Finland and Iceland. Lambs are born in April to May and slaughtered at the end of the grazing season in September to October. The majority of lamb production in the Nordic countries occurs on uncultivated pastures or in semi-natural or natural environments. Plant species utilized by grazing sheep differ from the coastal and insular ecosystems in Iceland and the Faroe Islands, through the lowland and forest systems used in Finland and Sweden, to the mountain pastures in the central parts of Norway, Sweden and Iceland. The extent and economical importance of sheep farming varies considerably between the Nordic countries. According to government statistics, the 1991 summer sheep stock in each country was (in thousands): Denmark 111, the Faroe Islands 70, Finland 107, Iceland 700, Norway 2211, and Sweden 406. In Iceland, Norway and the Faroe Islands, the sheep is the most abundant domestic animal and sheep farming is of relatively greater importance than in the other Nordic countries. The annual per capita consumption of lamb meat in Finland (0.3 kg), Sweden (0.7 kg), and Denmark (0.8 kg) is low compared to Norway (5.2 kg), the Faroe Islands (10 kg), and Iceland (24 kg).

Radiocaesium in the Nordic environment The entire Nordic area has been contaminated by global fallout from atmospheric nuclear weapons tests. The integrated deposition density of I3'Cs through 1980 amounted to 4.6 and 2.8 kBq m-', averaged over the 50-60" and 60-70" northern latitude belts respectively (UN, 1982). Fallout from the 1986 Chernobyl accident affected Nordic countries to varying extents. In Iceland, no Chernobyl fallout has been recorded, based on the absence of detectable 134Cs(see

213 below). In Denmark and the Faroe Islands the deposition of I3’Cs from Chernobyl has been calculated to be of about the same order of magnitude as I3’Cs from global fallout. In Finland, Norway, and Sweden vast areas were contaminated, such that ’37Csground deposition increased many times above the present levels from world-wide fallout (Arvela et al., 1989; Backe et al., 1987; Swedish Geological Co., 1986). Radiocaesium transfer to grazing sheep A common denominator for most of the ecosystems associated with sheep grazing is a soil characterized by a low pH topsoil layer and low content of most plant nutrients. In addition the organic matter content is often fairly high, and when mineral soils are present, organic matter is located in a distinct topsoil layer. The clay content in these soils is normally low (Scott Russell, 1967; Fredriksson et al., 1966; Livens and Loveland, 1988; Ha& et al., 1973; Melin and Wallberg, 1991).

As most soils utilized for sheep grazing cannot be ploughed or tilled, deposited radiocaesium remains bound to organic matter in the upper few cm of the topsoil layer, where root density is highest, and from which layer the downward migration of Cs appears to be very slow. Such distribution, as well as the long topsoil residence time, also contributes to a longlasting plant radiocaesium availability. Although deposition plays a major role in determining radiocaesium transfer on different pastures, the soil types and other environmental location parameters do add to the variations observed in the body burden of 13’Cs in sheep. Radiocaesium accumulation and metabolism in sheep Biomass production on sheep grazing land is low compared to managed lands, and the vegetation is commonly much more varied with respect to the number of plant species. The sheep usually graze freely over large areas, and may over the grazing season select a large variety of different plant species. Species selection may be particularly broad in the mountain pastures where birch forest systems and alpine herb, grass and heather communities may be utilized (Hove et al., 1990; Mayes et al., in press). Both plant selection and differences in radiocaesium contents of plants from the different areas are reflected in the radiocaesium burden accumulated by individual lambs during the grazing season. In some of the Nordic countries particular emphasis has been placed on the high abundance of fungi in certain years, since several species of fungi may have radiocaesium concentrations up to 100 times greater than green vegetation. In the British Isles, heather and other ericaceous species have been noted to be particularly important in the transfer of radiocaesium to lamb meat,

Organic soils maintain their radiocaesium levels for prolonged periods, probably due to

214 continuous recycling of the element within the soil biomass. In clay soils, radiocaesium is strongly bound by clay mineral lattices, and so becomes much less available for root uptake compared to organic soils. On the pastures used for sheep production in the Nordic countries soil ingestion probably contributes very little to the body burden in lambs. Exceptions would be found in situations of high stocking density and low grass production. There are reports of low radiocaesium bioavailability in ruminants for feeds harvested during 1986 and partly in 1987 (Ward et al., 1989; Howard et al., 1989) probably related to direct deposition of fallout. Radiocaesium transferred to plant tissues through root uptake generally had a bioavailability approaching that of ionic radiocaesium in studies carried out over a longer time interval following the Chernobyl accident (Hansen and Hove, 1991), and in the course of the Nordic study it can be expected that a major fraction of the '37Csingested would have a very high availability. The true absorption coefficient (i.e. corrected for faecal excretion of endogenous matter) of ionic radiocaesium has been estimated at about 0.9 in recent experiments (Mayes et al., in press; Beresford et al., 1993). Caesium is therefore digested in a manner similar to the other alkali metals. Radiocaesium is transferred rapidly from feeds to animal products such as meat and milk. Reported biological halflives for the excretion of L37Cs from lamb muscle range from 2 to 3 weeks (Vandecasteele et al., 1989; Hansen and Hove, 1993). Transfer coefficients from feed to meat fall in the range 0.2-1 for lambs and about 25% lower for ewes.

As a result of the fairly rapid rate of radiocaesium excretion, the current methods of choice for reducing radiocaesium contamination of sheep prior to slaughter have been a period of controlled feeding of L37Cs-low diets, and the use of caesium binders (Andersson, 1989; Andersson and Hansson, 1989; Hove et al., 1993).

NORDIC STUDY - MATERIALS AND METHODS

General layout of the study The inter-Nordic study was planned as case studies on individual farms or locations in each country, with sampling of soil, herbage and lamb carcasses. Site locations are indicated in Figure 3.6.1. In the Faroe Islands a number of localities were selected to represent the various islands. No counter-measures have been taken on farms or study sites located in Chernobyl fallout areas. From the activity concentration data obtained, the deposition density and 137Cstransfer factors between soil, herbage, and lamb muscle tissue have been calculated; from these data the expected dose contribution to man can be estimated. In order to study possible time changes in

215 plant radiocaesium availability, particularly with respect to the fraction attributable to Chernobyl fallout, the investigation was carried out over a four-year period from 1990 to 1993. As far as possible, the same sampling and analysis techniques have been used in each country. Details of the sizes and locations of sampling sites are provided in Table 3.6.1, and information on altitudes, soil types, and vegetation types for the various locations is given in Table 3.6.2. Tables 3.6.1 and 3.6.2 indicate significant variation in environmental locations of the grazing areas studied. The locations in Denmark, the Faroe Islands, Iceland, and Norway are located in coastal - or with respect to the climate, maritime - environments, although at varying altitudes, while the Swedish location is situated in the eastern part of the Scandinavian mountain chain, just below the limit of cultivation. The Finnish farm is situated inland, where the climate is more continental compared to that of the other locations.

Figure 3.6.1.

Locations used in the RAD-3 study. Details of each location are provided in Tables 3.6.1 and 3.6.2.

216 Table 3.6.1.

Size and localisation of the grazing areas selected in the various countries.

~~

Country and code

Grazing area size, kmz

Number of sampling sites

Localisation of the area: Province etc. Latitude

Denmark

(DEN)

0.09

1

Jutland

55.3"N

8.7"E

Faroe Islands

(FAI)

29

9

Six different islands

62.O"N

7.0"W

Finland

(FIN)

0.02

2

SW Finland

60.9"N

23.5"E

Iceland

(ICE)

0.12

1

Borgarfjordur

64.4"N

21.4"W

Norway

(NOR)

0.004

1

Nordland

66.6"N

12.4"E

Sweden

(SWE) 10

1

Jamtland

64.4"N

14.4"E

Longitude

Table 3.6.2. Altitudes, soil types, and vegetation types of the grazing areas. Country code

Environment

DEN

Coastal

2

Sandy

Permanent grassland

FAI

Coastal

50-240

Peaty

Permanent grassland

FIN

Inland

110

Clay

Natural pastureb)

ICE

Coastal

20

Peat, gravelly

Lowland mire

NOR

Coastal

10

Peaty

Permanent grassland

SWE

Mountain forest

580-770

Peat, gravelly, sandy moraine

Mountain moor, Betula forest, permanent grassland

a)metres above sea level.

b,

Altitudea) Soil type@)

Vegetation type@)

50% grass-growing old field, 50% forest meadow.

The dominant soil types in each of the grazing areas, shown in Table 3.6.2, are sandy soils (Denmark), peaty soils (Faroe Islands, Iceland and Norway), and coarse moraine soils (Sweden), all typical soil types used for sheep grazing. Whereas each of these soils was low in clay content, the Finnish location had an extreme clay soil (> 50% clay).

217

Climatic data is provided in Table 3.6.3, illustrating summer temperature variations between 9 and 15°C. The Swedish site had the highest summer rainfall. All locations had suitable precipitation and temperature characteristics to favour the continuous production of herbage.

Table 3.6.3.

Data on precipitation and temperature from weather stations adjacent to the respective grazing areas.

Country PreciDitation May-SeDtember, mm code 1990 1991 1992 1993 19611990

Mean temDerature. MaySeptember, "C 1990 1991 1992 1993 19611990

DEN

467 272 312 267

332")

13.9

13.5

15.2 13.8

13.9')

FA1

330 340 393 286

426b'

9.9

10.2

9.8

8.3

9.2')

FIN

279 325 245

313

310

12.4 12.2

13.7 11.8

12.6

ICE

325 301 319 336

291

9.0

7.9

8.1

8.1

NORd'

297

434

11.4 10.4

-

-

11.0

SWE

328 345

466

9.7

8.5

9.5

626

-

-

346 466

9.3

9.3

9.6 _

a)

1971-1990 data

b,

1961-1981 data

c,

_

~

_

_

1961-1988 data d ) N o 1992-1993 data available.

Soil characteristics Data from the chemical analysis of soil samples are provided in Table 3.6.4. Chemical analyses of the 1990 soil samples were performed according to conventional analytical techniques employed at the Agricultural University of Norway or the Swedish University of Agricultural Sciences. The organic matter content, as determined by loss on ignition methods, ranged between 8-16% for mineral-type soils in Denmark, Sweden and Finland, while the peaty soils in the Faroe Islands, Iceland, and Norway yielded 50-60% organic matter. In the mineral soils, the major portion of organic matter was located in the upper 0-5 cm of the soil profile. All soil samples were acidic to strongly acidic, with pH values ranging between 6.0 and 4.6, and contained moderate amounts of plant-available potassium (KAJ . The relatively high content of plant-available potassium in the

Faroe Islands soils may be due to deposition of potassium, as well as other elements, by storm splash from the Atlantic; these soils have also been shown to have high Na contents. This form of element deposition may also occur to varying degrees in the other coastal grazing areas. The extremely high content of calcium in the Norwegian soil samples was attributed to seashell fragments in marine deposits.

218 Table 3.6.4.

Chemical data for the top soil layer (0-10 cm) of the grazing areas (based on 1990 soil samples). Meanfstandard deviation for the sampling sites. PHH,O

me per lOOe drv soiln) CaAL KAL KHNO,

8.2k0.6

5.1i-0.7

39k6

21k5

C)

FA1

50.8f21.2

4.8fO.4

C)

52f22

c>

FINb)

16.6 13.5

5.0 5.8

87 225

22 37

155 230

ICE

52.0

6.0

50

7.6

25

NOR

60.1f25.5

5.lf0.8

737f259

34f21

53519

SWE

10.0f4.0

4.6f0.3

3659.7

12.1k4.5 32f10

Country code

Organic matter % wiw

DEN

"'Modified after Egner et al. (1960). b, Finnish data for both pasture types provided. Upper row data refer to the forest-type pasture, while lower row data refer to the field-type pasture. ') No data available.

Sampling of soil, vegetation and meat Vegetation types of the grazing areas are given in Table 3.6.2. Herbage samples in all locations were taken annually from 4-6 plots of approximate area 0.25 mz, and were dried and ground prior to radiometric and chemical analyses. Samples of individual species grazed by the sheep were collected in some locations; only a limited number of species were common to all study locations. Soil samples were taken annually from the same areas of the pastures during the grazing period. On some plots, exclusion cages or fencing were utilized to prevent grazing near the soilsampling areas. Soil cores were taken to 10 cm depth from field plots following harvesting of the grass. Soil cores were sectioned into 2 or more depth zones prior to oven- or air-drying and measurement of radiocaesium activities. Samples of meat were taken from study animals immediately after slaughter. Muscle from neck, leg, or abdominal wall was used for 137Cs measurement. In Norway live-monitoring data was used in place of meat data. For all soil, plant and meat samples, radiochemical analyses were performed by the individual participating countries, using NaI, Ge-Li or HPGe detector systems. Differences in management and sampling schedules thought to influence radiocaesium levels of the study lambs are detailed below:

219

Denmark Samples of soil and vegetation were collected from a coastal farm in west Jutland. Lambs were younger at slaughter (3 months old) than in the other studies.

Faroe Islands Samples of soil and grass were collected from 9 locations, covering various islands. Plant specimens were collected in July-August, and meat samples taken from the neck of slaughtered lambs in October. Live weights of slaughtered lambs were about 30 kg. In 1993 a feeding experiment involving twin lambs was conducted to measure grass-to-meat transfer factors. The lambs were removed from pasture in late September, and fed herbage cut from the same pasture for four weeks prior to slaughter.

Finland Lambs were grazed on about 2 ha of a pasture which consisted of approximately equal parts of an unmanaged field on a heavy clay soil and a forest meadow with juniper and birch. Lambs were put on pasture in early June, and were slaughtered at the end of September. In 1993 the pasture was physically divided such that one group of sheep grazed only on the field, while another group grazed only on forest meadow. Carcass weights were around 15-17 kg.

Iceland Lambs grazed at a stocking rate of 3.3 animals per hectare. Plant specimens were collected in June-October, and the lambs were slaughtered in October at an average carcass weight of 11 kg.

Norway Lambs were grazed in a fenced field of area 0.4 ha. Ewes and their twins were put on pasture during the last week of May, and slaughter occurred in August-September when pasture became insufficient. Plant specimens were collected monthly during the grazing season. Lamb livemonitoring data were collected weekly and corrected against meat values in 1990 to 1992.

Sweden Lambs were grazed on the fields and naturally forested slopes surrounding the farm. Birch and spruce were the dominating tree species up to the tree line at about 700 m. Mountain pastures, including Carex bogs, were accessible to an altitude of about 800 m. The sheep utilized an area of about 10 km2. The lambs were put on pasture in mid June and slaughtered in late September

at 5-8 months age. Carcass weights were 16-22 kg.

220

Calculations Transfer factors of 137Csfrom soil to grass, grass to lamb, and soil to lamb (aggregated transfer factor) were calculated on the basis of average I3%s activities of soil, grass and meat samples. Where possible, individual transfer factors were calculated before an average was reached; individual transfer factors to lamb from soil and grass were also calculated in this way for Faroe Islands data. Trends in the year-to-year changes of 137Csconcentrations in soil, herbage and meat were analyzed using standard General Linear Model (GLM) statistical procedures, with Duncan’s t-test at (r=0.05.All statistical differences between yearly values were classified as significant at the 0.05 level.

RESULTS AND DISCUSSION

Deposition of 13’Cs in soil The accumulated deposition levels of radiocaesium in soil samples ranged from approximately 3 to 37 kBq m-’. Variation between sampling sites in individual countries accounted for approximately half (range 34-77%) of the total variance, demonstrating a high degree of heterogeneity in soil radiocaesium deposition. The 1990 to 1993 soil 137Csvalues for the respective grazing areas are provided in Table 3.6.5, with average values given in Table 3.6.6. The 137Csdeposited in the Nordic countries by nuclear weapon test would, in 1993, have decayed from values in the range 4-15 kBq m-’in the 1960s, to approximately 2-7 kBq me*.In Iceland levels of 134Cswere undetectable, indicating that all soil I3%s was of nuclear weapon test origin. Also in Denmark and the Faroe Islands, a large proportion of the deposition was attributable to nuclear weapon testing. In Norway, Sweden and Finland, the majority of I3?Cswas of Chernobyl origin. Deposition densities of 137Csin both Norwegian and Swedish study locations approximated to between 1530% of the peak 137Cscontamination levels observed in these countries after the Chernobyl incident (Backe et al., 1987; Swedish Geological Co., 1986). Values for soil, herbage and lamb meat in the present study represent the combined values for remaining global radiocaesium and Chernobyl radiocaesium The relative distribution of *37Cs between the 0-5 cm and 5-10 cm depth layers is shown in Table 3.6.7. It can be seen that in Denmark and the Faroe Islands approximately 60-65% of accumulated 137Csin the 0-10 cm layer was located in the upper 5 cm. In Finland, Norway, and Sweden the corresponding fraction was in the range 8590%. In Iceland, however, only about 30% was located in the 0-5 cm layer, with the other 70% being present in the 5-10 cm layer. These differences reflect higher levels of weapon-test fallout in Denmark, the Faroe Islands, and

22 1 Table 3.6.5.

Deposition of 137Csand its transfer from soil to herbage to lamb in 1990-93. Activim data refer to the sampling date in the respective year.

Year, country code

kBq m-2a)

137Csin soil,

'37Csin herba e, Bq kg-*fiM

137Csin lamb meat Bq kg-I FdJ

M

3.4

5.4

5.0

FAI

5.551.5

155593

FIN")

23.2f4.4 17.151.3

41.4524.9 6.8f2.0

ICE

_--

NOR

Number T F

of

lambs

'-t)

CF g-;

s-m d)

1

1.6

0.90

1.50

26.1f22.1

30

32.6

0.18

5.50

14.652.1

10

1.8 0.4

108

56.1f3.3

8

----

0.52

----

31.2f5.7

1904f1430

1068f194

20

73.1

0.56

34.2

SWE

17.6f2.2

1175f604

1090f196

13

54.5

0.93

61.9

&

3.5f0.4

1.4

1.2f0.4

3

0.4

0.85

0.36

FA1

5.4fl.l

106f65

19.5f18.7

34

19.5

0.29 2.21

8.2 0.6

---- 1.41

17.6

---- 0.73

FIN"

25.0f5.1 19.5f0.7

204f78 10.856.4

31.4f10.4

10

ICE

3.7f0.4

65.1f5.6

53.1f5.7

7

NOR

36.6f5.7

2453f1392

1501*150

18

72.0

0.62 41.0

SWE

14.2f2.7

1100f338

668

1

74.1

0.61 47.0

IsEm

3.4f0.4

2.7

0.5 fO. 1

3

0.8

0.19

0.15

FA1

5.4f0.6

63.1k52.0

11.2f6.5

30

12.2

0.29

2.21

0.9 0.1

----

0.35

28.8

0.53

15.2

1992

0.82

14.4

FIN"

26.7f3.1 24.9f3.6

22.6k4.6 1.4f1.2

9.lf1.9

10

ICE

3.650.3

103.7f10.5

54.8f6.9

6

NOR

29.157.8

2230f1581

1145f132

10

96.5

0.50

38.0

SWE

15.954.8

920f597

515f65

6

62.3

0.56

32.4

ITEm

2.6f0.6

1.2

0.5f0.1

3

0.5

0.40

0.19

FAI

5.1f0.8

45.3547.2

10.1f8.3

38

8.7

0.32

1.97 11.0 0.28

1993

FINO

21.6f9.6 13.3f2.2

31.8f30.3 4.4k3.6

237f65 3.7f0.9

6 6

1.8 0.6

7.45 0.84

ICE

___

80.752.5

51.7f7.5

6

___

0.64

---

NOR

32.9f13.0

1982f1548

1405f175

10

89.0

0.71

42.7

SWE

13.1f3.7

840f481

597f117

7

58.6

0.71

45.6

,d) See corres onding footnotes to Table 3.6.6. All Finnish soil a n i grass data calculated individually for the forest pasture (upper row values) and the field pasture (lower row values). However, lambs moved freely between the two pastures, and so only one value is iven for lamb meat 37Cs. Tagvalues were calculated by first averaging field and forest I3'Cs levels. Cb values were not calculated in ths way due to the extreme differences between field- and forestcontent. herba e '37Cs 1993 Finnish data differ from those of other years in that lambs were not permitted to move between the two pastures, and so separate lamb meat Cs values are available.

a) ,b) ,')

222 Table 3.6.6. De osition of I3’Cs and transfer from soil to herba e to lamb, averaged over 1980 to 1993. Means and standard deviations calcukted from yearly averages. Country 137Csin code soil, kBq

137Csin herbage, Bq kg-I DM

I3’Cs in meat, Bq kg-’ FW

TF s-g b)

CF g-m c)

Tag s-m d)

DEN

3.2f0.4

2.7f1.9

1.8f2.2

0.82f0.54

0.58f0.35

0.55f0.64

FA1

5.350.2

92.4f48.9

16.7f7.5

18.2f10.6

0.27f0.06

2.97f1.69

FIN‘)

24.1f2.2 18.4f5.9

75.0f86.3 5.8f4.0

18.4O f11.6

3.18f3.38 0.42f0.24

---

0.83 f0.54

ICEg)

3.65f0.07

89.4f20.1

53.9f1.9

23.2f7.9

0.63f0.14

14.8f0.6

NOR

32.4f3.2

2142f249

1280f206

82.6f12.1

0.60f0.09

39.0f3.7

SWE

15.2f2.0

1009f155

718f256

62.4f8.4

0.70f0.16

46.7f12.1

~

a)

~~~

In the layer 0-10 cm.

Transfer factor (TF) = Bq kg-I DM (dry mass) grass per kBq m-*soil. Concentration factor (CF) = Bq kg” FW (fresh weight) meat per Bq kg-’ DM grass. d, Aggregated transfer factor (Tag)= Bq kg-’ FW meat per kBq m-*soil. e, Finnish data: upper row values are for forest pasture, lower row values are for field pasture. For details on T, and CF values, see footnote (e) to Table 3.6.5. O Finnish data : lamb meat Cs averaged from 1990, 1991 and 1992 data. g) Icelandic data :soil, TF, and T, data averaged over 1991-1992. Herbage, lamb meat, and CF data averaged over 1990-1993. b, ‘)

Iceland; this fallout has migrated vertically down through the soil profile to a greater extent than did the Chernobyl fallout and is most noticeable in Iceland. No significant increase in the fraction below 5 cm was recorded during the experimental period. 137Csin sampled herbage

Data pertaining to the I3%s activities of herbage from 1990 to 1993 are provided in Table 3.6.5, with average values given in Table 3.6.6. High within-year variations from Norwegian and Faroe Islands data were indicated by coefficients of variation (CV) between 57 and 105%; such variations can be attributed both to differences in botanical composition of sample plots and to heterogeneity in soil 137Csdeposition. Such heterogeneity in Chernobyl I3’Cs soil deposition has previously been demonstrated (Haugen et al., 1991; Rodn et al., in press). Variation between yearly averages of herbage 137Csactivity concentrations yielded CVs between 12 and 115%, with the lowest CVs in the countries with high levels of Chernobyl deposition.

223 Table 3.6.7.

Relative distribution of 137Csin the 0-5 and 5-10 cm depth layer of the topsoil profile, per cent.

1990

1991

1992

1993

-

58.9 41.1

68.0 32.0

64.1 35.9

63.75 36.3”

4.6 4.6

0-5 5-10

63.3 36.7

63.1 36.9

68.2 31.8

57.8 42.2

63.1 39.9

4.2 4.2

FIN

0-5 5-10

93.7 6.3

94.1 5.9

92.2 7.8

89.4 10.6

92.4 7.6

2.1 2.1

ICE

0-5 5-10

-

-

30.3 69.7

31.0 69.0

-

30.6b) 69.4b)

0.5 0.5

NOR

0-5 5-10

89.0 11.0

93.5 6.5

88.0 12.0

69.0 31.0

84.9 15.1

10.9 10.9

SWE

0-5 5-10

87.6 12.4

87.4 12.6

89.8 10.2

89.5 10.5

88.6 11.4

1.2 1.2

Country code

Depth layer, cm

DEN

0-5 5-10

FA1

Mean of 1991-1993 data.

b,

Mean

1990-93

SD

Mean of 1991-1992 data.

Soil-to-grass transfer factors can be divided into two groups: high values (18-83) were evident in the countries with high organic matter (peaty) soils, while low values (0.4-3.2) were observed in Denmark and Finland. The soil types in Denmark and Finland (sandy soil and heavy clay soil respectively), in combination with more favourable soil chemical properties tending to reduce radiocaesium uptake, explain the comparatively low transfer factors seen in these locations. The soil-to-grass transfer factor varied greatly depending on the particular species of plant assessed. Particularly high transfer to Rumex acetosu has been noted previously in other locations. Where the same species has been sampled in different countries, soil types are of great importance. Low transfers were typical for herbs and grasses in Finland, while high transfers were observed in samples from Iceland, Norway, and Sweden (Table 3.6.8). The difference between the transfer to the leaves of the tree-sized Salix spp. in Norway and the shrub-sized Sulk

spp. in Sweden most likely reflects a difference in plant root distribution.

Meat 13’Cs activity concentrations High meat radiocaesium concentrations in Norway and Sweden reflect the high levels of Chernobyl deposition, as shown in Table 3.6.5. Also evident is the large difference in lamb radiocaesium content between Denmark and Iceland, two countries with approximately the same

224

Table 3.6.8. Average 137Cscontents of individual species of pasture plants, Bq kg" DM Plant group

Countrylspecies

Festuca (grass)

FAI -F. rubra

46

Number of years b, 3

FIN -F. rubra -F. pratensis -F. ovina

5 <1 103

1 2 1

0.35 0.07 7.2

ICE -F. SPP.

3.7

3

1.o

NOR-F. SPP. FIN -C. SPP.

2285 140

+

4

70.5

202

1

14.2

ICE -C. bigelowii -C. lyngbyei -C. nigra

3.2 74.4 119

4 4 4

0.89 20.7 33.1

SWE-C. SPP. FA1 -D. flexuosa

1739 276 33

+

4 1

114.4 6.2

FIN -D. caespitosa -D. flexuosa

23 172

1

1

1.6 12.1

ICE- D. caespitosa

4.9

4

1.4

NOR-D. caespitosa

1281k465

4

39.5

SWE-D. caespitosa -D. flexuosa ICE -A. capillaris

856f 186 740,130

4 4

56.3 48.7

3.4

4

0.95

FIN -A. tenuis

6

1

0.42

NOR-A. SPP.

1882

2

58.1

SWE-A. capillaris NOR-R. acetosa -R. longifolia -R. SPP.

825 +642 6210 1193 1284

3 1 1 2

54.3 191.7 36.7 39.6

SWE-R. acetosa FIN -T. spp.

1765f685

4

116.1

30

2

2.1

NOR-T. SPP. NOR-S. spp.

1590+703 42

3 2

49.1 1.3

Carex (sedge)

Deschampsia (grass)

Agrostis (grass)

Rumex (herb)

Trifolium (herb) Salix (shrub)

Mean content

TF s-g" 8.6

SWE-S. SPP. 549f118 3 36.1 For Finland, both field- and forest-type pastures used. Only data from the foresttype pasture is used here. b, Number of years of available data used in calculating the mean, )' Soil-to-grass transfer factor calculated using the average 1990-1993 soil 137Cs concentration for each country.

a).

225 level of deposition. The concentration factor (CF) between grass and meat generally reflects the availability of feed radiocaesium for absorption in the animal. The CFs were similar in all study locations (0.58-0.70) apart from the Faroe Islands, which exhibited a much lower value (0.27). The availability of radiocaesium from grass harvested from one site in the Faroe Islands was assessed in a feeding experiment involving two lambs restricted to a pasture for two weeks, and subsequently fed grass harvested from the field for three weeks. In this experiment concentration ratios were calculated as 0.46 for hindleg muscle and 0.36 for foreleg muscle. The transfer factor to hindleg muscle was 0.94. From these data the low concentration factors observed in the Faroe Islands study seem difficult to explain. The aggregated transfer factors (T,) between soil and lamb meat, given in Tables 3.6.5 and 3.6.6, express the combined effects of soil factors and vegetation differences on the transfer of radiocaesium from soil through grass to lamb. The T, values in the Nordic countries fall clearly within two groups. Low values of 0.3-3 Bq kg-' meat on a deposition of 1 kBq m-2soil in Denmark, the Faroe Islands and Finland ( Finnish field pasture, 1993 data) contrasted with values between 14.8-46.7 Bq kg-' meat in Iceland, Norway and Sweden. The forest pasture in Finland showed a high T,, 11.0 Bq kg-' meat, when studied separately in 1993 (Table 3.6.5). Since the grass-to-meat concentration ratios were of similar magnitudes in most of the study locations, the comparatively high radiological sensitivity to caesium contamination observed in the Finnish forest pasture, Iceland, Norway, and Sweden must be related to differences between soil types and the uptake of caesium into herbage. In addition, higher caesium transfer factors from organic soils may also be accentuated by the more varied diets consumed by sheep grazing in forested or mountainous areas. Fungi, heather, and certain Curex species, which may all be found in abundance in natural and semi-natural environments, have been noted for their high capacity to transfer radiocaesium to sheep (Howard et al., 1991).

Year-to-year changes The majority of year-to-year changes among soil, herbage, and lamb meat radiocaesium levels were not significant when the within-year variation was taken into account. Significant differences were, however, observed for lamb 137Csin Denmark, with significant decreases from 1990 to 1991 to 1992, but with no significant difference between 1992 and 1993 levels, and for Faroe Islands herbage 137Cs,with a significant decrease from 1991 to 1992, but with no difference between 1992 and 1993. Although effective ecological halflives can theoretically be calculated from the data given in Table 3.6.5, such values are of limited significance with only four years of data available, and with the high sample variance observed. The series of observations from

226 the RAD-3 programme will, however, provide a well-defined starting point for follow-up studies relating to the long halflives previously observed in lamb production in some of the Nordic countries (Hove and Strand, 1990).

Doses to man from sheep in Nordic environments The lifetime dose to the population through a distinct nutritional pathway such as lamb meat depends on the average consumption of lamb meat, the ecological halflife of radiocaesium in the production system, and the transfer from soil to the herbage consumed by the animal. All of these factors appear to differ between each of the Nordic countries. High levels of lamb meat consumption occur in Iceland and Norway, both of which have high aggregated transfer factors from soil to lamb. In the Faroe Islands, where lamb meat consumption is also high, lower Tag values and shorter effective halflives combine to reduce the long-term dose to man. In Sweden, while T, values are high, lamb meat consumption is low, and thus the dose contribution of radiocaesium from lamb meat is of lesser importance than in the other Nordic countries. This is also the case in Denmark, where both lamb meat consumption and Tagvalues are low.

References Andersson, I. (1989). Transfer of i37Csfrom feed to lambs' meat and the influence of feeding bentonite. Swedish J. Agric. Res. 19:85-92. Andersson, I. and Hansson, I. (1989). Distribution of 137Csto different muscles and internal organs of lambs fed contaminated hay. Swedish J. Agric. Res. 19:93-98. Arvela, H., Markkanen, M., Lemmela, H., and Blomqvist, L. (1989). Environmental gamma radiation and fallout measurements in Finland, 1986-1987. Supplement 2 to Annual Report STUK-A74. 32p. Helsinki. Backe, S . , Bjerke, H., Rudjord, A.L., and Ugletveit, F. (1987). Fallout pattern in Norway after the Chernobyl accident estimated from soil samples. Radiat. Prot. Dosim. 18:105-107. Beresford, N.A., Barrett, C.L., and Crout, N.M.J. (1993). Radiocaesium variability within sheep flocks in the restricted area of Cumbria. Institute of Terrestrial Ecology Report, August 1993. Fredriksson, L., Gamer, R.J., and Russell, R.S. (1966). Caesium-137. In: Radioactivity and Human Diet. Russell, R.S. (ed.), Pergamon Press, Oxford. pp 319-348. Haak, E., Eriksson, A., and Karlstrom, F. (1973). Studies on plant accumulation of fission products under Swedish conditions. XIII. Entry of '%r and i37Csinto the herbage of contrasting types of pasture. Research Institute of National Defence FOA Report C 4525-A3. Hansen, H.S. and Hove, K. (1991). Radiocaesium bioavailability: transfer of Chernobyl and tracer radiocaesium to goat milk. Health Phys. 60(5): 665-673. Hansen, H.S. and Hove, K. (1993). The effect of exercise on I'Cs retention in lambs. J. Environ. Radioactivity 19:53-66. Haugen, L.E., Garmo, T.H., Bjemstad, H.E., and Pedersen, 0. (1992). Different approaches for estimating the deposition of radiocaesium on mountain pasture in

221

Southern Norway. Analyst 117529-532. Hove, K. and Strand, P. (1990). Predictions for the duration of the Chernobyl radiocaesium problem in non-cultivated areas based on the behaviour of fallout from the nuclear weapons tests. In: Flitton, S., Katz, E.W., eds. Environmental contamination following a major nuclear accident, Proceedings of an International Atomic Energy Agency Conference. Vienna: IAEA; IAEA-SM-306140. Vol. 1:215-223. Hove, K., Pedersen, O., Garmo, T.H., Hansen, H.S., and Staaland, H. (1990). Fungi: a major source of radiocaesium contamination of grazing ruminants in Norway. Health Phys. 59(2): 189-192. Hove, K., Strand, P., Voigt, G., Jones, B.E.V., Howard, B.H., Segal., M.G., Pollaris, K. ; and Pearce, J. (1993). Countermeasures for reducing radioactive contamination of farm animals and farm animal products. The Science of the Total Environment 137:26 1-271. Howard, B.J., Mayes, R.W., Beresford, N.H.. , and Lamb, C.S. (1989). Transfer of radiocaesium from different environmental sources to ewes and suckling lambs. Health Phys. 57579-586. Howard, B.J., Beresford, N.A., and Hove, K. (1991). Transfer of radiocaesium to ruminants in natural and semi-natural ecosystems and appropriate countermeasures. Health Phys. 61(6):715-725. Livens, F.R. and Loveland, P.J. (1988). The influence of soil properties on the environmental mobility of caesium in Cumbria. Soil Use and Management 4:69-75. Mayes, R.W., Beresford, N.A., Lamb, C.S., Barnett, C.L., Howard, B.J., Jones, B.E.V., Eriksson, O., Hove, K., Pedersen, O., and Staines, B.W. (1993). Novel approaches to the estimate of intake and bioavailability of radiocaesium in ruminants grazing forested areas. In Press. Melin, J. and Wallberg, L. (1991). Distribution and retention of cesium in Swedish boreal forest ecosystems. In: The Chernobyl fallout in Sweden. Moberg, L. (ed.), Sundt Artprint, Stockholm. pp 467-475. Rosin, K., Andersson, I . , and Liinsjo, H. (1993). Transfer of radiocaesium from soil to vegetation and to grazing lambs in a mountain area in northern Sweden. In Press. Scott Russell, R. (1967). Uptake and accumulation of radioactive substances in terrestrial plants - the radiobiological aspect. In: Aberg, B., Hungate, F.P., eds. Radioecological concentration processes. Oxford, Pergamon Press, pp. 367-382. Swedish Geological Co. (SGAB). (1986). Caesium-137 kBq m-2 ground surface. Results from aerial surveys, May to October 1986. Uppsala. UN. (1992). Ionizing radiation: sources and biological effects. United Nations Scientific Committee on the Effects of Atomic Radiation, 1982 Report to the General Assembly, New York. pp 220-222, 230. Vandecasteele, C.M., Van Hees, M., Culot, J.P., and Vankerkom, K. (1989). Radiocaesium metabolism in pregnant ewes and their progeny. The Science of the Total Environment 851213-223. Ward, G.M., Keszthely, Z., Kanyar, B., Kralovanszky, U.P., and Johnson, J.E. (1989). Transfer of 137Csto milk and meat in Hungary from Chernobyl fallout with comparisons of worldwide fallout in the 1960s. Health Phys. 57587-592.

This Page Intentionally Left Blank

229

3.7

DYNAMIC MODEL FOR THE TRANSFER OF CS-137 THROUGH THE SOIL-GRASS-LAMB FOODCHAIN

SVEN P. MELSEN’ Ris0 National Laboratory, DK-4000 Roskilde, Denmark

SUMMARY A dynamic radioecological model for the transfer of radiocaesium through the soil-grass-lamb foodchain was constructed on the basis of field data collected in 1990-1993 from the Nordic countries: Denmark, Faroe Islands, Finland, Iceland, Norway and Sweden. The model assumes an initial soil contamination of one kilobecquerel of I3’Cs per square metre and simulates the transfer to grass through root uptake in addition to direct contamination from resuspended activity. The model covers two different soil types: clay-loam and organic, with significantly different transfers of radiocaesium to grass. The implementation of the metabolism of the lamb includes an assumption of a biological halflife of three weeks for radiocaesium.

INTRODUCTION The present work focuses on modelling of the transfer of I3’Cs through the soil-grass-lamb foodchain based on the results of field studies carried out in the Nordic countries during the years 1990-93 under the NKS/RAD programme. It was decided to build a simple dynamic model that would incorporate the basic transfer processes and represent average conditions for the Nordic countries. The approach adopted is similar to that of Simmonds et al. (1979). Recently, developments of more detailed, dynamic sheep models have been reported by Assimakopoulos (1991) and Galer (1993). Such models could provide greater detail but, considering the variability of the field data, it was the objective of the present work to keep the model simple and still permit

Inger Andenson*, Hans UnsjO3, Klas RosCn’, Knut Hove4, Hanne Solheim Hansen4,Hans Pauli Joensen’, T. Vestergaard’, Sigurbur Emil P;ilsson6, A j a Paasikallio7 Southern Animal Experimental Station, S-230 53 Alnarp, Sweden Department of Radioecology, Swedish University of Agricultural Sciences, S-750 07, Uppsala, Sweden Department of Animal Science, Agricultural University of Norway, N-1432 As, Norway Department of Natural Sciences, University of the Faroe Islands National Institute of Radiation Protection, 150 Reykjavik, Iceland Agricultural Research Centre of Finland, SF-31600, Jokioinen, Finland

‘Contributors:

230 reliable estimates to be made of '"Cs concentrations in lamb meat. The endpoint for the modelling is calculation of radiation doses to humans from consumption of lamb contaminated with radiocaesium.

FIELD DATA The details of field sites, sampling, analysis and results are described by Hove and Liinsjo in chapter 3.6 and in the annual status reports from the participating laboratories. For the present purpose, only aggregated results are presented here. Table 3.7.1 gives the annual means of the data for the transfer of radiocaesium from soil to grass from each site in terms of observed ratios (in units of Bq kg-' dry weight grass per kBq m-' soil). Table 3.7.2 gives the corresponding data for the transfer from grass to lamb also in terms of observed ratios (Bq kg-' lamb per Bq kg-' dry weight grass). TABLE 3.7.1. Data from field studies in the Nordic countries on the transfer of radiocaesium from soil to grass. The data are given as average annual observed ratios for each site. All grass values refer to dry weight conditions. 1990

1991

1992

1993

Bq kg" grass per kBq m-* soil Denmark

1.6

0.41

0.81

0.47

Faroes

28

21

13

10

Finland

1.6

5.1

0.69

0.42

Iceland

59

30

Norway

61

55

77

60

Sweden

69

75

64

63

It is noted that the data on the soil-to-grasstransfer seem to fall in two different categories, one with a high transfer (a mean value of 48 Bq kg-' grass per kBq m-' soil f 50%, 1 sd) at the sites in the Faroes, Iceland, Norway and Sweden and one with a low transfer (a mean value of 1.4 E3q kg-' grass per kBq m-' soil f 60%, 1 sd) in Denmark and Finland. This difference is due to

the different soil characteristics: a soil with a high content of organic matter provides a high root uptake to the vegetation, whereas a mineral soil with a low content of organic matter provides a low root uptake to the vegetation. Furthermore, it is noted that the variability between years for each site is as great as the variability between sites within each of the two soil-type categories, and the transfer to the grass is generally reduced with time.

23 1 TABLE 3.7.2. Data from field studies in the Nordic countries on the transfer of radiocaesium from grass to lamb. The data are given as average observed annual ratios for each site. All grass values refer to dry weight conditions. 1990

1991

1992

1993

Bq kg-' lamb per &I kg-' grass Denmark

0.93

0.87

0.19

0.40

Faroes

0.19

0.19

0.18

0.24

Finland

0.63

0.29

0.75

0.84

Iceland

0.47

0.42

Norway

0.56

0.62

0.50

0.71

Sweden

0.93

0.61

0.56

0.71

For the grass-to-lamb transfer, the data display rather high variations in view of the (presumably) not very different metabolic processes in the lambs at the different sites. We find an average observed ratio of 0.5 Bq kg" lamb per Bq kg" grass f 35%, 1 sd.

MODEL STRUCTURE AND PARAMETEXS The dynamic model is based on first-order differential equations where the transfer of '"Cs between the model compartments is based on parameters called rate constants. The rate constants determine the fractional transfer of activity between the compartments for each time step. The model uses time steps of one day, and the rate constants thus have the dimension of reciprocal days. The differential equations have the following form: P

d% = n-1, nzm

kn, A,, -

A , - A, A,,,

+

A,,

where A, and & are the instantaneous amounts of material in compartments m and n at time t;

lgpn is the rate constant for the transfer from compartment m to n; is an effective rate constant for transfer out of the system from compartment m which is used to describe loss of material, e.g. by radioactive decay or transfer to a stable sink; Ao,, is the rate of input of activity into compartment m; and p is the total number of compartments. The schematic model structure is shown in Figure 3.7.1. The model incorporates an atmospheric compartment, a soil compartment, two grass compartments (one for the external plant and one for the internal plant), and two lamb compartments (one for the gastro-intestinal tract and one for the meat) representing a single animal. The transfer processes are indicated in Figure 1

232

I'

ATMOSPHERE

I

dl

MEAT

FAECES k60

'k27 I

Fig. 3.7.1. Schematic structure of the dynamic model for the transfer of radiocaesium through the soil-grass-lamb foodchain. Transfers between the model compartments are indicated by the arrows.

with arrows. The atmosphere, soil and grass compartments are associated with a surface area of one square metre. The yield of the grass biomass, Y, is assumed to be 0.15 kg dry weight per square metre. The soil compartment represents the rooting zone to a depth of 5 cm. Physical decay is taken into account for all of the model compartments. The model assumes an initial contamination of the surface soil of 1 kBq of 13'Cs m-*. From the soil compartment a fraction of the activity is resuspended (k2J into the atmospheric compartment which is limited by an atmospheric mixing height of loo0 m. The simulation of the resuspension process is based on a resuspension factor of

m-'.

From the atmospheric

compartment the activity is transferred to the ground through the application of a generic deposition velocity of 0.023 m s" representing both wet and dry processes. The deposition is split between a fraction of 30% intercepted by vegetation surfaces (d,) and the remainder which is returned to the soil compartment (4). From the soil compartment the bioavailable activity is reduced through the application of a removal process (kz7)corresponding to a fixation halflife of 3 y. Root uptake from soil to grass is governed by the rate constants k, and

b2which are based

on concentration factors of the soil-to-grass transfer. The activity on the external surfaces of the grass is transferred to the soil through the rate constant k3*which corresponds to a weathering halflife of two weeks. The transfer of the activity in the lamb is simulated by two compartments only. The daily intake of radiocaesium by the lamb is calculated from the current concentration in the grass and the amount of grass eaten per day. This intake is supplied each day to the compartment of the gastro-intestinal tract from where there is a fast transfer through faecal and urine excretion (ks0)

233

and to the meat compartment (ka).From the meat compartment there is a transfer through other corresponding to a biological halflife of radiocaesium in the lamb of three metabolic processes (b) weeks. Faecal excretion is included in the model at a rate of 70% of the total excretion; the remaining 30% is accounted for by urine excretion. The modelling of faecal excretion is included for the purpose of permitting additional field data to be compared with model predictions. The daily intake of grass by the lamb is assumed to increase linearly with time from 0.5 kg d-' at an age of 3 months to 1.4 kg

a' at an age of 12 months.

This may be expressed in the

following way, where I is the daily intake of grass in kg dry weight and M is the age of the lambs in months: I = 0.2

+ 0.1

M. Correspondingly, the carcass weights of the lambs are assumed

to increase linearly with time. From the field data the following relation was obtained, where m, is the carcass weight in kg and M the age of the lamb in months at slaughter: mkt) = 2.6

+ 2.4

.M. The rate constant kZ1for the resuspension process is obtained by multiplication of the resuspension factor with the mixing height, since the mixing height gives the ratio between the inventory of radiocaesium in the air over a unit area and the concentration in the air

.

For root uptake the concentration ratio gives, for the equilibrium situation, the ratio between the concentrations of radiocaesium in the grass and in the soil. This means that the two rate constants for this transfer kX and

b2are related

to the concentration ratio CR, the grass

biomass yield Y, the soil density p and the depth rd of the rooting zone in the following way:

The implementation of the model includes two soil types: organic and clay-loam. For the organic soil, use is made of a concentration ratio between grass and soil of 2 and a soil density of

500 kg mJ, whereas for the clay-loam soil there is a concentration ratio of 0.2 and a soil density of 1500 kg m-3. The data on concentration ratios and soil densities are taken from Ohlenschlaeger

(1991). For the rooting-zone depth parameter rd, a value of 0.05 m is used. These parameter values are consistent with the observed transfer of 137Csto the grass (Table 3.7.1). The rate constants k50,k5dand

b determine the dynamics of the transfer of radiocaesium

in the lamb. They are interrelated in the following way:

Here Ff is the meat transfer coefficient (d kg-') and m, the carcass weight. Ff is calculated by dividing the observed ratio between the I3'Cs concentrations in lamb and grass with the daily

234

intake. With this calculational procedure, the transfer coefficient becomes dependent on the age of the animal with values that agree well with observations (Galer et al. 1991). The rate constant determines the biological halflife for radiocaesium in the lamb. The rate constant k,, is chosen arbitrarily at a value of 1 d-I ensuring a rapid turnover in the gastro-intestinal tract.

RESULTS The model is implemented as a FORTRAN code in the TIME-ZERO modelling system (Kirchner 1989). This system provides a convenient environment for building and running compartmental models on personal computers and inspecting the model calculations in numerical and graphical form.

As input to the model information is needed on soil type, start of growing season, age of lambs when sent on pasture and age of lambs at slaughter. The soil type determines the root uptake to the grass as explained in the previous section. The start of the growing season sets the start of the root uptake, whereas resuspension starts from the first day of the simulation. The

lambs intake of contaminated grass starts from the day the animals are sent on pasture and continues until slaughter.

I

,~

-1 .. ...

I

... .

,

__

....-I-.----.-------.-----

~

. =""..

.,".

.. .I

TIME (Days after birth of lambs) I

Fig. 3.7.2. Model calculations of the transfer of 137Csfrom organic soil (1 kBq m") to grass (dry weight) and lamb meat.

Figure 3.7.2 shows the time course of the radiocaesium concentrations in grass and lamb meat from a simulation for organic soil where the growing season starts 90 days after the birth of the lambs; 10 days later the lambs are sent on pasture and then slaughtered at 250 days. The resulting concentrations at time of slaughter are about 45 Bq kg-' dry weight for the grass and

235 about 20 Bq kg-I for the meat. The corresponding results for clay-loam soil are about 1.7 Bq kg" dry weight for the grass and about 0.8 Bq kg-' for the meat as shown in Figure 3.7.3.

0

60

wi

100

200

TIME (Days after blrth of lambs)

Fig. 3.7.3. Model calculations of the transfer of 137Csfrom clay-loam soil (1 kBq mZ)to grass (dry weight) and lamb meat.

A comparison of the results of the model calculations with those of the observations is shown in Table 3.7.3 which gives the observed ratios for the soil-to-grass transfer (Bs kg" grass per kBq m-' soil) and for the grass-to-lamb transfer (Bq kg-' fw meat per Bq kg" dw grass). The results show that the model gives nearly the same observed ratios as the mean of the observations. This is not surprising because the selection of the model parameters in the dynamic model was designed for this purpose. TABLE 3.7.3. Comparison of the transfer of 13'Cs from soil to grass and to lamb meat found from the observations and estimated by means of the dynamic model. Soil-to-grass transfer

Grass-to-lamb transfer

Organic soil

Clay-loam soil

(Bq kg-I grass per

(Bq kg-I grass per kBq m-* soil)

(Bq kg-' meat per Bq kg-' grass)

kBq

mZsoil)

Observations

48

1.4

0.50

Dynamic model

44

1.7

0.45

For the purpose of testing the dynamic performance of the model, the Icelandic data from 1990 were used for comparison with the model calculations. The Icelandic data were selected

236

because they comprise multiple sampling of grass, faeces and meat over the grazing season. The lambs were on pasture from birth on 12 May 1990 until slaughter 132 days later. The results are shown in Figs. 3.1.4 and 3.7.5. Figure 3.1.4 shows the predicted and observed levels of 137Csin grass, and Fig. 3.7.5 shows the predicted and observed levels in faeces and lamb. The model is seen to provide a reasonable reproduction of the observed levels. For the grass data we find an average value of the predicted-to-observed (P/O) values of 0.9i-0.1 (1 sd), for the faeces data there is an average P/Ovalue of 1.0k0.2 (1 sd), and for the meat values

an average P/Ovalue of 0.8f0.2 (1 sd).

70 0

0

M)

I? D

P t

D90

.-C

h

c! 20

8

10

0

1

I

20

40

I

I

I

im TIME (Days after 12 May 1990) M

80

1

120

Fig. 3.7.4. Comparison of the model predictions of '37Cslevels in grass (Bq kg dw) with observations from Iceland in 1990.

LLI FEE8

im

s I

*

*

* 2

*

237

CONCLUSIONS A dynamic radioecological model for the transfer of radiocaesium through the soil-grass-lamb foodchain was constructed on the basis of field data collected in 1990-1993 from the Nordic countries. The model assumes an initial soil contamination of one kilobecquerel of 137Csper square

metre and simulates the transfer to grass through root uptake in addition to direct contamination from resuspended activity. The model covers two different soil types: clay-loam and organic, between which there are significantly different transfers of radiocaesium to grass. The implementation of the metabolism of the lamb includes an assumption of a biological halflife of three weeks for radiocaesium in the meat. The parameter values of the model were selected to permit a realistic representation of the transfer of radiocaesium in the Nordic countries through the soil-grass-lamb foodchain. When tested with a data set collected in Iceland in 1990, the model predictions agree within 20% with the mean values of the observations.

ACKNOWLEDGEMENT This work was supported by the Nuclear Inspectorate of the Danish Emergency Management Agency.

REFERENCES Andersson, I., H. Lonsjo and K. RosBn. Transfer of Radiocaesium from Mountain Pasture Land to Vegetation and to Grazing Lambs in Northern Sweden. 1990, 1991 and 1992 Status Reports. Swedish University of Agricultural Sciences, Sweden. Assimakopoulos, P.A., Ionnides, K.G. and Pakou, A.A. (1991). A general multiple-compartment model for the transport of trace elements through animals. Health Physics 61, 245-253. Galer, A.M., Crout, N.M.J., Beresford, N. A., Howard, B.J., Mayes, R.W.,Barnett, C.L., Eayres, H.F. and Lamb, C.S. (1993). Dynamic Radiocaesium Distribution in Sheep: Measurement and Modelling. Journ. Env. Rud. 20, 35-48. Hove, K., H. S. Hansen, P. Strand, K. Barvik and H. H. Hansen. Report from the NKS-RAD-3 Experiment in Norway, 1990-1992. Agricultural University of Norway, Norway; 1992. Joensen, H. P. and T. Vestergaard. Transfer of Radiocaesium in Uncultivated Pastures in the Faroe Islands. 1990, 1991 and 1992 Status Reports. University of the Faroe Islands; 1992. Kirchner, T. B. (1989). TIME-ZERO, The Integrated Modeling Environment. Ecological Modelling 47, 33-52. Magnusson, S. M., S. E. Palsson, K. Indridason, 0. Gudmundsson, B. Magnusson and J. Torsson. Transfer of Radiocaesium from Soil and Plants to Lamb in Iceland. 1990, 1991 and 1992 Status Reports. National Institute of Radiation Protection, Iceland; 1992. Nielsen, S. P. Radiocaesium in Soil, Grass and Lamb Collected at Ribe. 1990, 1991 and 1992 Status Reports. Rise National Laboratory, Denmark. Paasikallio, A., R. Sormunen-Cristian and V. Kossila. Transfer of Radiocesium from Soil to Pasture Plants and to Lamb’s Meat. 1990, 1991 and 1992 Status Reports. Agricultural Research Centre, Finland. Simmonds, J., Linsley, G. S. and Jones, J.A. (1979). A General Model for the Transfer of Radioactive Materials in Terrestrial Food Chains. NRPB-R89, National Radiological Protection Board, UK. Ohlenschlreger, M. (1991). The Transfer of Radionuclides in the Terrestrial Environment. Rise-M-2934.

This Page Intentionally Left Blank

239

3.8. STUDIES ON COUNTERMEASURES AFTER RADIOACTIVE DEPOSITIONS IN NORDIC AGRICULTURE

IUAS ROSEN Department of Radioecology, Swedish University of Agricultural Sciences, Uppsala, Sweden.

SUMMARY A description of the short- as well as the long-term agricultural situation after a deposition of radioactive nuclides provides the background for a review of experimental work to develop efficient countermeasures, carried out since the sixties. Experiments have mainly concerned two strategies: 1. Ploughing in a contaminated surface soil layer to place it at a considerable depth in the soil profile. External radiation is reduced, as too root accessibility and plant availability of the nuclides. 2. Fertilization and liming to dilute plant-available caesium and strontium in the soil, thereby reducing root uptake of the nuclides.

It was found that efficient placement below root depth can be achieved by means of two-layer ploughs and by deep ploughing equipment. However, soil type and moisture conditions in the soil during ploughing will influence the quality of the work. In practice loose, sandy soils and heavy clays will be more difficult to handle than other soil types. On soils with low clay content such as sandy soils and peat soils, fertilization with up to 200 kg potassium per ha can efficiently reduce caesium uptake by both grass and arable crops. These soils have low potassium reserves and need new potassium dressings during crop rotation. Heavy clays generally need no extra potassium dressings to reduce crop uptake of caesium.

INTRODUCTION Transfer of radiocaesium in the agricultural system Fallout of radionuclides from the atmosphere may be deposited on the ground and lead to increased radiation exposure of the population living in the area. There is a risk of both external and internal doses. The latter result fiom a more or less rapid transfer of fallout nuclides to man via the soil > vegetation > animals > food chain. One consequence of the mobility of radionuclides in the agricultural environment is that they contaminate food and thus endanger the production and economy of farmers.

In uncontrolled releases from nuclear facilities, nuclides with long as well as short half-lives reach the environment. If fallout occurs some time before the growing period (as in the Chemobyl case) there will be fewer problems for agricultural production than if fallout is deposited on growing

240 crops. These problems become increasingly severe the closer it is to harvest time (Eriksson et al. 1990; Eriksson 1991). The short-term situation

In the short-term, during the first year, the caesium transfer to grass produced on grassland is comparatively high. In general it is considerably reduced during subsequent years. This is followed, in the long-term, by a lag period with transfer on a lower level determined by root uptake from contaminated superficial soil layers. However, on temporary grassland this period may be shortened by the regular ploughing for other crops. After 3 to 5 years the plant availability of caesium is findamentally reduced by normal procedures. On ploughed land, plant availability has an ecological half-time in the range of 10 - 20 years.

In the short-term, the transfer of fallout nuclides to crop products depends on timing and on the magnitude and composition of the fallout. As mentioned above, the sensitivity of growing crops is high and even moderate deposition may lead to the total loss of a crop. Grass crops and grassland constitute the critical points of entry into the food chain. Arable crops cultivated for grain are somewhat less sensitive, as grain develops late in the season. Grasslands, natural as well as temporary, are products of the environment. When used in farming, the quality and development of the grass cover depend on the management of the land, the grass species used, pasturing, fertilization and the local climate. The grass cover formed, its density and growth, largely determine the capacity for interception and retention of fallout during the season. Retention of fallout by the crop at the time of harvest is influenced by the stage of development at the time of fallout. The growth up until harvest dilutes intercepted fallout nuclides and there is time for falloff. Moreover, the decay of the radioactivity also reduces the nuclide content per unit weight. Contamination of grasslands creates an immediate risk of contamination of milk and dairy products too. This risk can be counteracted by grazing restrictions for some time to allow the decay of nuclides with short half-lives and to allow reduction of the contaminant content of grass by falloff and dilution by growth. The long-term situation In the long-term, agricultural production is less influenced by the timing of the fallout. Greater influence is exerted, for instance, by the binding of caesium to the soil material and by the types of cropping and animal husbandry.

The type of soil, especially the content of clay and organic matter, is of great importance firstly for the content of plant-available potassium and secondly for the plant availability of caesium deposited in the soil. The mineral soils thus tend to bind caesium more efficiently and to have more potassium available for the crops than the organic soils. The extremes are heavy clays and peat soils. One experience resulting from the Chernobyl fallout was that caesium transfer to grass on clay-soil grassland was comparatively low and was rapidly reduced, while on peat soils the transfer was higher from the beginning and remained comparatively high for several years.

24 1

Research on countermeasures

During the three last decades research in the field of radioecology has increased our understanding of the effects of radioactive fallout on agriculture and on our possibilities to take remedial action. Interest has been concentrated on possibilities to reduce external radiation from the soil surface and to reduce the uptake of nuclides by crop plants. The work has developed along several lines, described in the following sections.

EFFECTS OF PLOUGHING AND PLACEMENT OPERATIONS ON EXTERNAL RADIATION AND PLANT UPTAKE

Ploughing is a process normal to soil management in agriculture and as such it has influence upon deposits on the soil surface. The deposit may either be mixed into the plough layer soil with time or be placed below it. The equipment available will decide the placement of a contaminated layer of surface soil. With ordinary ploughs, the contaminant will be contained within the whole plough layer and thereby reduce plant uptake and external radiation compared to superficial placement. With special deep-ploughing equipment the contaminated soil layer will be placed at a greater depth. It will be even less available for crop root uptake and it will contribute even less to external radiation on the ground than in the case of ordinary ploughing. Theoretical work and practical tests on the efliciency of deep placement Hedeman Jensen (1979) worked with various models to calculate how the external exposure rate from deposited 13’Cs was influenced by ploughing and found different placement effects. The following cases were studied: Model 1. Activity distributed evenly in the 25 cm plough layer. 2. Activity distributed evenly in the 10 - 35 cm soil layer. 3. Activity distributed evenly in the 25 - 50 cm soil layer. 4. Activity evenly diminishing down to 0 in layer 0 till 25 cm. 5. Activity reduced exponentially to a half every 5 cm downwards.

Result

Reduction factor:

6

Reduction factor: 18 Reduction factor: 75 Reduction factor:

4

Reduction factor:

3

Laboratory studies on the effects of nuclide placement on plant uptake

Andersen (1967) and Haugen & Uhlen (1992) carried out such studies. Andersen (1967) used pot experiments to study the effect on plant uptake of placing 90Sr in layers at various depths in different sandy soils. The layers were: A, 0-5 cm, B, 0-20 cm, C, 20-25 cm, D, 40-45 cm and E, 8085 cm. The placements were intended to simulate different soil management operations, superficial

242 treatment of the sowing layer, normal ploughing and two deeper ploughings, respectively. The uptake by barley, rye grass, red clover and root crops was studied. The experiment showed that placement at a depth of 40-45 cm, compared to that at 0-5 cm, reduced plant uptake by a factor varying from 2 to 5 depending on soil type. Placement deeper than 40-45 cm had little additional effect. Figure 3.8.1 shows the total 9% uptake from one sandy clay soil during the five-year experimental period.

1

0 Q)

L

v)

Figure 3.8.1. Accumulated 90Sr uptake from a sandy clay soil during five years of cropping. 90Sr placement: A = 0-5 cm, B = 0-20 cm, C = 20-25 cm, D = 40-45 cm and E = 80-85 cm. Deposition: 0.5 mCi per m2 carrier-free 90Sr was applied per plot. Crop rotation: Barley, grass, barley, root crops, barley. (Andersen 1967) Haugen & Uhlen (1992) also used pot experiments to study plant (Italian Ryegrass) uptake of caesium from contaminated soil placed at different depths in uncontaminated soil. The experimental soil was taken from the upper layers of two peat soils and one mineral soil (20 % clay) at sites in the mountains that were contaminated by Chernobyl fallout. The experiment comprised three placement treatments: In one the contaminated soil was placed on the soil surface, in the second it was placed at a depth of 15 cm, and in the third it was mixed into the whole soil volume. The plant uptake of caesium was higher from the peat soils than from the mineral soil, and more so in the third placement treatment than in the others. The difference between placement at 15 cm depth and homogeneous mixing in the profile was not particularly large, although the uptake was less from the mineral soil in the latter case.

243

Microplot field studies on the effects of nuclide placement on plant uptake A series of field Microplot experiments on nuclide placement was carried out by Fredriksson et al. (1969), Eriksson & Fredriksson (1972) and LLinsjd & Haak (1986). The microplots, 0.25 m2 in size, were delimited in the plough layer horizon by iron frames.

In the early study, the aim was to determine how placement of 9oSr at different depths and liming, both treatments achieved by simulation of ordinary soil management operations, could influence nuclide transfer to a crop in the long-term situation. The studies continued for a seven-year period at four different sites in different districts and with different soils (Table 3.8.1). Expt. 1 was carried out on a sandy loam, Expt. 2 on a silty loam, Expt. 3 on a soil with high content of organic matter (10.2 %), and Expt. 4 again on a sandy loam. The nuclide was deposited in four layers at depths of 0-5 cm, 0-10 cm, 0-20 cm and 20-25 cm. The local crop rotations differed somewhat. The crops studied were wheat, oats, barley, peas and vegetables (lettuce, radishes, spinach). Some experiments included treatments with liming, 10 ton CaO per ha. The experimental results indicated that the deeper the placement of 90Sr, the lower was the plant uptake (Table 3.8.1). Placement in the 20 - 25 cm layer thus reduced the uptake to 25 - 50 per cent of that placed in the 0 - 5 cm layer. The placement effects, i.e. reductions in nuclide uptake, varied somewhat in magnitude from site to site and between years. Variations were thought to represent influences of soil characteristics and differences between districts with regard to climate and weather conditions during the growing periods. Soil and water conditions at the crop site have influence on root penetration and feeding depths. In addition liming had a considerable effect on strontium uptake on one soil with a low pH. Table 3.8.1. Percent of applied 90Sr absorbed by crops during the experimental period 1959-1965 (Fredriksson et al., 1969) Placement depth of 9OSr, cm

0- 5 0 - 10 0 -20 20 - 2 5

Expt. 1

Expt. 2

Sandy loam

Silty loam

control liming

control liming

2.0 1.8 1.4 0.6

1.1 1.1 0.9 0.3

1.4 1.4 1.0 0.6

0.9 0.9 0.7 0.3

Expt. 3

Expt. 4

Sand

Sandy loam

3.5 3.5 2.6 1.3

5.1 4.1 3.0

5.3

Eriksson & Fredriksson (1972) carried out a model experiment under field conditions to study the root uptake of 1 3 7 0 by different crops when the nuclide was placed in thin soil layers at different depths: 7, 15, 35,60, 85 and 110 cm below the soil surface. The uptake of caesium was studied for five years in five crops: peas, barley, white mustard, oats and rye grass. The crops differ with regard to normal root-feeding depths and, as was found, with regard to caesium uptake.

244

Caesium uptake was strongly diminished with increasing placement depth. Uptake from the 7 cm level decreased in the sequence: peas > white mustard >>oats > rye grass > barley. Relative numbers for different crops and placement depths are given in Table 3.8.2. It can be noted that placement of the nuclide at a depth of 35 cm reduced crop uptake by 55 - 95 per cent compared to placement in the sowing layer at 7 cm. Considering uptake from the 35 cm level, crops were ranked as follows: oats > barley > peas >rye grass, white mustard. Table 3.8.2. Relative plant content of 137Cs, obtained by root uptake from different depths (Eriksson & Fredriksson 1972) Placement Barlev depth (cm) seed straw ~

~

&& seed straw ~

~

100 100

7 15 35

100 100

20 12

44 10

46 9

60 85

9

1

1

7 1

1

0

0

0

110

24 23 15 8 1

White mustard seed straw

seed straw

Rve e r a s 1st- 2nd-cut

100 100

100 100

oats

~

100 100

18 5 1 7 5

24 7 3 2 1

64 47 10 10 0

65 44 10

5 2

66 5 2 5 1

42 9 8 4 3

The investigation carried out by Lonsjo & Haak (1986) comprised Microplot experiments over several years under field conditions. The aim was to study the effects of two strategies on the caesium transfer to ordinary field crops. The strategies were deep placement of a contaminated soil surface layer and potassium fertilization. The nuclides, 134Cs, and in the first year also *9Sr, were used at two different sites with different soil types. The nuclides were deposited in two layers: A; homogeneously mixed into the top soil layer, 025 cm, and B; deposited at a depth of 27-29 cm and covered with uncontaminated top soil. Potassium fertilization, 250 kg K per ha, and controls were included four treatments. These treatments were intended to test different possibilities for remedial action to help reclaim contaminated agricultural land, and they represented four elements of possible field strategies:

AI (0-25 cm) potassium control, ploughing simulated. AII (0-25 cm) potassium fertilization (250 kg m a ) before ploughing. BI (27-29 cm) potassium control, burial of surface layer. BII (27-29 cm) potassium fertilization (250 kg m a ) before burial of surface layer

The treatments tested were very efficient. For the crops used, records of nuclide uptake gave the order: AI >MI > BI > BII. Thus, the combination of fertilization and deep placement of the contaminated surface layer (BII) gave the largest reduction of nuclide uptake. For the whole experimental period the efficiency of the treatments on the uptake of 134Csis shown by the relative uptake in the treatments: AI = 100, AII = 61, B1= 43 and BII = 16.

245

Field studies on the placement efficiency of available ploughs As shown by Meisel et al. (1991), ordinary ploughlng results in an inhomogeneous distribution of a

surface soil layer when buried. Experiments carried out in Austria for three years after the Chernobyl accident showed that the majority of the caesium in 1986 was deposited in the 0 - 5 cm layer. After ploughing in 1987, the highest concentration was found in the 15 - 18 cm layer. After a second ploughing in 1988 the highest concentration of caesium was found deeper, and the distribution was more even in the plough layer. Field studies on the quality of placement achieved with different equipment have been carried out by Roed (1982; 1991) and Nilsson (1983). Roed (1982) studied the effects of soil management on the gamma radiation level measured 1 m above the soil surface. Experimentally, the treatments were ploughing, harrowing and disc tilling on plots, 100 m* in size, that had been contaminated with 86Rb prior to the soil management operations. Experiments were made on different Danish sandy soils. For the ploughing operations, a normal 14" plough and a 16" deep plough were used. After ploughing, the plots were harrowed or disced as normal. It was found that the dose rate above ground was reduced by a factor of 5.6 - 8.3 after ploughing and after harrowing and discing by factors of 4.8 and 5.0, respectively. A considerable reduction in radiation level was thus obtained under practical conditions. Soil profiles were taken from the plots and each 2 cm layer was analyzed for the test nuclide. This part of the study showed that the normal plough distributed most of the nuclide to a depth of about 20 cm, while the deep plough distributed the nuclide evenly in the 0 - 30 cm layer. The aim of the experiments carried out by Nilsson (1 983) was to investigate the efficiency of the common types of plough in Sweden when used for turning down a contaminated soil surface layer. Different normal ploughs, deep ploughs and two-layer ploughs were tested in the study, which also included the influence of the speed of the ploughing equipment. Kieserite (MgSOq H20) was used as a tracer, simulating fallout on the ground. Kieserite is a water-soluble mineral used for magnesium fertilization. The experiments were carried out on three different sites in central Sweden. The sites were Kungsangen 1, a cultivated pasture on heavy clay, Kungshgen 2, arable land on heavy clay and Krusenberg 3, arable land on a sandy soil. Altogether 16 ploughs and different types of equipment were compared (cf Table 3.8.3). Eight soil profiles were taken from each ploughing treatment to study the distribution of the tracer in the plough layer. Table 3.8.3 presents the efficiency obtained in the different treatments as fraction of the tracer found below the 200 mm depth in the soil profile. As appears, the resulting fraction differed greatly between treatments, from 0.07 to 0.97. Under the dry and severe soil conditions prevailing in the experimental period, most of the ploughs commonly used in Sweden were not efficient enough to turn down a contaminated surface soil layer. However, those with a larger working depth and width (Treatment Nos. 14 and 15) gave the best results. The two-layer plough gave results almost similar to those of conventional ploughs. Operating speed did not influence the result.

246 Table 3.8.3. Layout of ploughing experiments and placement efficiency of different treatments. Sites: Kungsiingen I, Kungsbgen I1 and Krusenberg in the county of Uppsala (Nilsson, 1983) Exp. Furrow Detailsof No. ridge equipment width,

Plough depth mm

mm

15

350 400 400 400 400 400 400 400 400 400 400 400 500 600 600

16

550

1 2 3 4 5

6 7 8 9 10 11 12 13 14

skim coulter (1) 200 skim coulter 200 skim coulter (1) 200 skim coulter (3) 150 skim coulter (4) 250 skim coulter 200 non coulter 200 manure skimmer 200 trashboard 200 front plough 200 double-depth plough 200 double-depth plough(2)250 200 non coulter skim coulter 200 skim coulter 450 parallel plough 200

Fraction of tracer below 200 mm depth KungsKungsKrusen- Efficiency ftngen I hngen I1 berg order clay clay sand 0.69 0.51 0.40 0.30 0.42 0.71 0.37 0.27 0.60 0.15

0.36 0.59 0.44 0.64 0.97 0.14

0.07 0.08 0.10 0.25 0.05 0.41 0.20 0.12 0.15 0.08 0.42 0.35 0.62 0.38 0.70

0.27 0.17 0.3 1 0.21 0.09 0.36 0.49 0.52 0.43 0.35 0.08 0.62 0.62 0.55 0.83 0.10

8

13 11 12 15

5 7 9 6 14 10 4 2 3 1 16

(1) Trashboard on Kungsiingen I (2) no coulter on this plough (3) speed 3 km/h (4) speed 7 km/h (Normal speed 5 km/h). Roed (1991) describes experiments on sandy soils with a type of plough able to remove the upper 0 - 5 cm layer and place it below the 50 cm depth. The experiments were carried out on pasture land and on a field of new temporary grassland. After analyzing the distribution of the tracer, it could be concluded that most of the top soil layer (0-5 cm) had been placed at the bottom depth, 50 cm. In conditions more difficult than those on sandy soils, the placement was less efficient. REDUCTION OF PLANT UPTAKE BY POTASSIUM FERTILIZATION The potassium balance in agricultural soils In the long-term situation fanners should aim at retaining or improving the potassium status of their soils. The sum total of potassium released by weathering, returned with plant residues and manure and originating from fertilizers thus has to be higher than the amount removed from the agricultural system. In this case, the caesium available in the soil will be more greatly diluted with potassium for plant root uptake and plant uptake will successivelybe reduced. In the opposite case, there is a risk

241

of increasing the transfer of caesium to the crop products. Peat and sandy soils are soils of this kind with small natural potassium reserves, and they belong to the potassium-deficient soil groups from which high caesium transfers are expected. Clay soils have larger reserves of potassium-bearing minerals where this development is unlikely and from which the caesium transfer to crops tends to be much lower. The variation observed in these soil properties showed the need to counteract high caesium transfers by improving the potassium status of the deficient soils. Experimental work in this field started already during the fifties and sixties. Laboratory studies on effects of potassium fertilization

Andersen (1967) used 44 different plants and 22 different Danish soils in pot experiments for studies of the plant uptake of 137Cs and 8%. He found that the accumulation of caesium varied between crops and that the soil content of clay and organic matter influenced the transfer to the crops. After the 5th grass cut (after 5 months of cultivation), some treatments included potassium supplements and as a consequence there was a strong reduction in caesium uptake. However, there was then a slow increase up to the 10th cut and a sharper increase thereafter. This trend in caesium transfer may have depended on the depletion of plant-available potassium in the soil as a result of intense cropping. Haak & Eriksson (1973) studied the influence of potassium and nitrogen fertilization on the uptake of 137Cs by wheat and timothy in pot experiments. The crops were cultivated in six different soils at three potassium fertilization levels. It was found that at the lowest fertilization level, the crop uptake of caesium was determined by soil characteristics, the content of clay and organic matter. Potassium applications reduced the caesium uptake considerably. However, with increasing clay content this fertilization effect was reduced. Haugen et al. (1990) studied the availability to rye grass of Chernobyl caesium when returned to different soils, sand and peat, by goat dung. Also in this case, the influence of soil type was important. When placed in the sand with the dung the uptake of caesium was only one-fifth of that when placed in the peat soil. This experience shows that the dung should preferably be applied to mineral soils. In pot experiments Malm et al. (1991) studied the uptake of 137Cs by vegetables from a peat soil and from an artificial soil of compost as influenced by potassium fertilization. The soils used were contaminated by caesium from the Chernobyl fallout. The vegetables included were cucumber, tomato, parsley, radish and lettuce. All pots were given a basic application of NPK (1 1-11-18) and lime. The availability of caesium to plants was very different in the soils. The uptake from the peat was one order of magnitude greater than from the compost. The effect of the fertilization is shown in Table 3.8.4 No significant difference was created by the fertilizer treatment. The result may depend on the fact that the basic fertilization (A) covered the needs of the vegetables.

248

Table 3.8.6 shows the influence of time on the average transfer of caesium and strontium to the harvested grass. There is a marked reduction in the transfer of caesium during the first few years after the deposition. However, after 7 - 8 years a lag period with low transfer was reached for both nuclides at both sites. The difference between the sites depending on soil conditions is large and it increases up to the lag period, especially with regard to caesium transfer. At the natural site Lovsta it is thus about 7 times higher for caesium and about 3 times higher for strontium than levels on the clay soil. Table 3.8.6. Change with time in the transfer of nuclides to grass (m2(kg d.w.) * lom3)on a pasture on heavy clay (Risslinge) and on a natural pasture sandy soil (Lovsta). Haak et al., 1973; Haak, 1986 _____~

____~

_____~

~

Year

~

90Sr

1 3 7 ~ ~

Lovsta 1961*

Rissslinge

22.20 6.85 3.48 31.10 15.30 7.16 1.38 0.72

1962 1963 1964* 1965 1966

Mean 1967-1971 Mean 1972-1981

9.25 0.71 0.41 7.93 1.73 0.80 0.20 0.10

Lovsta

Risslinge

73.0 50.4 33.0 24.8 21.9 18.7 16.1 11.8

23.2 15.8 11.8 9.9 9.7 8.7 5.8 4.0

* deposition year Table 3.8.7. Effects of fertilization and liming on the nuclide transfer to grass at Lovsta, relative numbers. (Haak et al, 1973; Haak, 1986) Treatment

%

137cs 1967-71

1972-81

1967-71

1972-81

A B C D E

78 91 106 181

43 47 32 117

53

20

a) b) c)

125 114 68 95 100

59 57 44 47 52

171 45 66 98 124 148 84 99 70

127 19 50 91 79 95 64 70 65

100

73

4 Mean

249

Table 3.8.5. Relative effects of K-fertilization on transfer of 13’Cs from soil to various crops in microplots (Fredriksson et al. 1966; Lonsjd et al. 1990) Soi

K-fertilization Kg K per ha and year 1962 - 1981

1961

A

0

156 625

B

0

C

156 625 0 156 625

0 31 156 0 31 156 0 31 156

Relative. 137Cs transfer to grain and seed Oats

Peas

100 54 6 100 60 7 100 46 8

100

47 8 100 52 1s 100 50 14

White mustard 100 62 3 100 54

3 100 49 5

Pre-Chernobyl field studies on the effects of potassium

At two sites in the county of Uppsala, in Central Sweden, long-term experiments were started in 1950 as part of a programme for improvement of pasture land. One of the sites was a cultivated pasture on heavy clay (Risslinge) and the other was a natural pasture on a sandy soil (Lovsta). The treatments were the same up to 1960 and the experimental conditions then represented stable ecological environments with regard to soil and plant cover. These sites were taken over in 1961 for radioecological purposes, for studies of the behaviour in the soil profile of caesium and strontium deposited in early spring, and of the transfer to the pasture grass during a long-term period. Risslinge: A. Control B. 13 kg P (superphosphate)

C. 13 kg P (superphosphate) + 38 kg K. D. 13 kg P (superphosphate) + 38 kg K

Lovsta: A. Control B. 26 kg P (superphosphate) + 75 kg K + 3 1 kg N (ammonium sulphate) C. 26 kg P (superphosphate) + 75 kg K + 3 1 kg N (calcium nitrate) D.31 kg N (ammonium sulphate) E. 26 kg P (superphosphate) + 75 kg K a) unlimed. b) 4 ton CaC03. c) unlimed.

d) 4 ton Ca(OH)2. The fertilizer treatments were largely the same as those during the earlier period. The deposition of caesium was repeated in spring 1964. Samples for measurement of the nuclide transfers were taken from 2 - 4 grass cuts per year (Haak et al., 1973; Haak, 1986). Fertilizer treatments per ha at the pasture sites:

250 Table 3.8.4. Mean transfer factors (137Cs in plant d. w. / 137Cs in soil d.w.) for different growth media and vegetables (Malm et al. 1991) Vegetables

Treatment per m3 soil Peat A

Cucumber Tomato Radish Parsley Lettuce

1.5 kg 1.5 kg 1.2 kg 1.2 kg 1.2 kg

Transfer factors

Peat B NPK NPK NPK NPK NPK

+331 +331 +264 +264 +264

g g g g g

K2SO4 K2SO4 K2SO4 K2SO4 K2SO4

Peat A

Peat B

Compost mould

1.2 0.82

1.3 0.66

0.06

0.76

0.87 1.3 1.7

0.036 0.13 0.19

1.8

1.7

-

Microplot field studies on potassium effects Fredriksson et al. ( 1 966), and Lonsjo et al. (1990) worked with microplot experiments in the field to study the transfer of 137Cs to different crops when influenced by different levels of potassium on different soils. Three soils were used, A=loam, B=clay loam and C=silty clay. During the first year, caesium and the basic potassium applications were mixed into the plough layer. During the following years the fertilizers were mixed into the sowing layer. Treatment I

Control, no potassium added.

Treatment I1

Basic potassium: 156 kg K per ha (as KHCO3). During the following years 3 1 kg K per ha and year.

Treatment 111

Basic potassium: 625 kg K per ha (as KHCO3). During the following years 156 kg K per ha and year.

A single crop rotation comprised 3 years with oats, white mustard and peas. The experiment was continued for 7 crop rotations. The results can be generalized so far that plant uptake of caesium was lower on soils rich in clay and on soils given ample applications of potassium to cover the loss by cropping. During this time white mustard accumulated most caesium of all the crops, next came oats and last peas, representing oil seed crops, grain crops and leguminous seed crops, respectively. Table 3.8.5shows the reduction in caesium uptake with increased potassium supplementation. For an annual treatment using 3 1 kg potassium per ha, the average crop uptake was 46 - 62 per cent of that in the control. For the higher potassium treatment, 156 kg per ha and year, the average uptake was still lower, 3 - 15 per cent of that of the control.

25 1

At the natural pasture site, the fertilizations influenced the transfer of strontium. Treatments B and C were especially efficient in the long-term. The reason may be that in addition to providing calcium with the fertilizers, these treatments changed the composition of the plant community considerably compared to the control (Table 3.8.7). Field experiments after Chernobyl Methods to reduce the caesium transfer to crops by means of potassium applications were tested in many fallout districts after Chernobyl. It was felt that local data were needed to cover local conditions. In Norway, experiments with potassium fertilization were carried out at four sites during the period 1987-1991 and at five sites in 1988-1990. The aim was to study the influence of fertilization and environmental factors on the transfer of Chernobyl fallout caesium to grass. Two levels of potassium were used, 75 and 150 kg K per ha, and on two of the fields the number of treatments was increased to eight in a dose response study within a range of 0 - 525 kg K per ha, Figure 3.8.2. Also the residual effects of the fertilizer applications were studied at some of the sites (Haugen et al., 1992). A linear reduction (50 - 70 %) in caesium transfer was observed for the treatments 75 150 kg K per ha. However, the dose response study showed that the fertilization effect did not increase substantially at the higher potassium supplements, 225 - 525 kg K per ha (Figure 3.8.2).

100

Percentage (%)

80 60

40 20 0

0

75

150

225

300

Kg K

ha

375

450

525

Figure 3.8.2.The effect of different potassium fertiliziations on plant uptake of radiocaesium in two field experiments. The experiments were carried out on cultivated grassland on mineral soil, one loam, Vestre Slidre, and one sandy loam, Vinstern, fertilised in spring with 0-525 kg K per ha. The results for the first cut (1991) are given as percentages ofthe controls (Haugen et al., 1991). Uptake of radionuclides by spring wheat and barley from cultivated soils supplemented with contaminated sewage sludge, are described in a Finnish field experiments Puhakainen, M. and T. Yliiranta (1992). The experiments were conducted on clay, clay loam and sandy loam soils. The sludge was spread in spring on ploughed fields before the preparation of seedbed (expt.1) or in autumn prior to ploughing (expt. 2). The amount of 13’Cs distributed with the sludge was in the

252 range 10000-12000Bq/m* d.w. It was found that the application of sludge increased the content of 137Cs in grain by a factor 2-12.However, the absolute uptake was low, a few Bq per kg d.w. as was the fraction of caesium recovered in plant with a range of 10-150*lo4. The transfer of 137Csto the crops was lower on clay soil than on sandy loam and clay loam. Experiments with different levels of potassium fertilization have been carried out on contaminated land in the Chernobyl fallout regions in Sweden from 1987 onwards. During that period the fallout caesium has been reaching equilibrium in the agricultural environment. The experimental field sites are distributed in over four counties with relatively high fallout levels. Table 3.8.8 gives sites, soil types, crops, deposition and age of the grassland in 1986,duration of the experiment, and treatment (Rosen 1991). Table 3.8.8. County (Cty) and site of experiment, soil characteristics, deposition, crop, age of grassland in 1986,duration of experiment in years in 1991,and treatment (Rosen, 1991) Cty Expt Site No.

c c

1 2

c x

3 4

x 5 X x Y Y

Y Y Y

z Z

z ~~

6 7 8 9 10 11 12 13 14 15

Jarlisa

Soil

pH

Loss

type

aq

onignition% kBq/m2

Peat

6.2 6.0 ,I 5.5 Hille 6.6 I1 11 6.6 Tr6dje Sand 6.5 I, I1 6.5 Torsboda Sandy loam 6.0 Vilinger Loamy sand 5.6 Berge Loamy sand 6.0 Holmsta Loamy sand 6.4 Sundmo Silt loam 5.4 Hammarstr Silt loam 6.2 41 5.6 Sand Bllsjon Gravelly 6.1 sandy loam

,,

1,

4,

0,

~~

37 88 85 50 50 4 4

6 8 7 6 7

11

27 10

Dep

43 51 50

200 200

170 170 50

70 110 35 30 37 32 32

Crop

Ageof Dura- Treatgrassland

Hay Pasture It

Hay Barley Hay Barley Hay Pasture Hay

,,

3 15 10 3

3

2 6 4 1

Vegetables 3 Pasture 25 5 tt

I1

35

tion ment of expt. 5 3 5 4 4 5 3 1 3

5 3

4 4

6

K+Z K K+A K K K K K + CaO K + CaO K+Z+A+CaO K K + CaO K+Z K K+ Z

~

K = Potassium (50 % of K), Z = zeolite (aluminosilicate minerals, similar to clay minerals), A Adularia (potash feldspar with 8.3 % K, 3.0 % Mg and 4.3 % Ca), CaO = liming

=

The experimental crops were pasture grass, hay grass, barley and vegetables. The last included potatoes, carrots, leek, and lettuce. All experiments were given a basic annual fertilization with

253

nitrogen, 60 kglha, and phosphorus, 20 kglha. The experimental treatments were annual applications of potassium fertilizer (50 % of K) in a range of 0-200 kg K per ha. Table 3.8.9 gives the uptake of 13'Cs by the crops as a percentage of that deposited per unit area at the experimental sites in 1986-1989. Pasture grass on peat soils showed the highest uptake of 137Cs, 3-10 % . However, when increasing amounts of potassium were added to the same soil, the total caesium uptake could be reduced considerably. Thus, the transfer to the hay crops decreased from 2.5 to 0.03 per cent of the deposition per unit area. There were generally small differences in uptake between treatments with 50 and 100 kg m a , but major differences between that with 200 kg K h a and the other K-fertilization levels. Table 3.8.9. Ranges in uptake of 137Cs in grass and barley grain, in % of 137Cs deposited per unit area, at the different experimental sites during 1987-1989. 0-200 kg K per ha and year added (no effects of lime, zeolite, or adularia have influenced the data presented in the table) (Rosen, 1991) Cty Expt No

Site

C

1

Jarllsa

c c

2 3

X x X

4

5

x Y

6 7 8

Y

9

Y Y 2

10 11 13

2

14 15

Z

$9

I,

Hille (9

Trodje I,

Torsboda VBlinger

Crop

Hay Pasture

Recoverv. per cent of Cs deposited 1987

1988

1989

2.49-0.68

1.62-0.68 3.38-0.44

0.01-0.03 0.29-0.08 0.002-0.006

0.31-0.03 3.06-0.26 10.45-1.34 0.10-0.63 0.01-0.05 0.13-0.01 0,004-0.001

1.41-0.64

0.72-0.29

0.29-0.09 0.02-0.01 0.37-0.13 0.31-0.21 0.65-1.79

0.11-0.02 0.009-0.004 0.10-0.02

I,

* *

Hay Barley

0.48-1.05 0.03-0.05

Hay Barley Hay Pasture

0.10-0.42 0.002-0.003 0.18-0.04

Berge Hay Holmsta " Hammarstr. Pasture

,,

II

Bllsjon

"

1.55-0.55

*

0.007-0.005 0.42-0.64 0.76-0.29 1.16-4.60

*

0.32-0.96

*

*

0.34-0.02 1.66-0.35

* No experimental data this year Generally, the total uptake was higher in the control treatment than in those with potassium fertilization. The high potassium applications reduced the transfer of 13'Cs, except in 1987 on the peaty soils where high applications of N-fertilizers resulted in a high total caesium transfer.

Table 3.8.10. Transfer of 137Csfrom a peat soil (m2 * (kg d.w.)-l * at Jiirliisa (C 1) in the County of Uppland, to two cuts of hay grass in 1987, 1988 and 1989 at different levels of annual Kapplication. The deposition level was 43 kBqh2. Zeolite application after the 1st cut in 1988 was 4000 kg zeolite soil per ha (Rosen, 1991) Potassium applications. ke per ha

Potassium applications + Zeolite

Control 50

100

200

Control 50

100

200

1987 1st 1988 1st 1989 1st 1990 1st

33.3 41.1 15.6 9.30

20.7 13.7 6.05 4.42

22.3 8.14 4.19 3.02

9.07 1.86 1.86 1.40

10.7 7.60

3.02 3.84

2.56 2.33

0.47 0.74

average 1st.

24.9

11.2

9.42

3.55

9.15

3.43

2.44

0.61

1987 2nd 1988 2nd 1989 2nd 1990 2nd

28.1 10.2 5.35 4.65

19.3 3.26 2.56 2.79

11.6 1.40 1.65 1.40

6.74 0.70 0.23 0.23

16.0 5.35 5.81

6.51 2.79 2.79

3.72 1.86 2.19

0.93 1.40 0.42

average2nd. 12.1

6.98

4.01

1.98

9.07

4.03

2.59

0.92

Year Cut

No.

The caesium transfer from a peat soil, experiment C 1, (Table 3.8.10) may serve as an example of caesium transfer to hay grass. A tendency to an annual reduction in caesium transfer was observed during 1987 - 1990. The original level of plant-available potassium in this soil was low and, as a consequence, the TF-values were high in the control. The potassium treatments were efficient and had reduced the caesium transfer already by the first cut, one month after the dressing was applied. Whereas the effects of the potassium fertilizers were large, the addition of zeolite in 1988 had very little effect on caesium uptake from I988 and up to 1990. Experiment X 6 (Table 3.8.11) which took place at Trodje, in the county of Gavleborg, may serve as an example of caesium transfer from a sandy soil, low in plant available potassium. When compared with the data from the peat soils, it was found that the transfer level was 5-10 times lower on the sandy soil. Moreover, for this soil, the effect of potassium fertilization was considerable at the first harvest after application. The total caesium transfer was reduced from 0.03 to 0.01 per cent of the deposition (Table 3.8.5). Using an annual addition of 100 kg Wha extra after the first cut (100+100 kg Wha), the caesium transfer to the second cut was reduced compared with the single dressing of 200 kg Kiha in spring.

255 Table 3.8.11. Transfer of 137Cs fiom a sandy soil, (m2 * (kg d.w.)-l * lO-3), to hay grass and barley grain and straw on a sandy soil at Trodje (X 6), (X 7), County of Gavleborg, 1987-1991. The deposition level was 170 kBq/m. (Rosen, 1991) Year Grass cut

Potassium appl.. kg per ha

1987 1st 1988 1st 1989 1st 1990 1st 1991 1st

4.53 3.06 2.47 2.53

average 1st

3.15

1.47

0.73

0.76

1987 2nd 1988 2nd 1989 2nd 1990 2nd 1991 2nd

8.59 2.18 2.47 3.06

2.41 0.59 0.53 0.47

1.18 0.18 0.29 0.18

1.65 0.35 0.24 0.24

average 2nd.

4.07

Control 100 2.59 2.12 0.88 0.31

Barley

100+100 200 1.65 0.82 0.29 0.18

1.71 0.88 0.29 0.18

Control 100

200

0.41 0.82 0.29 no data 0.28

0.29 0.18 0.06

0.24 0.18 0.06

0.04

0.04

grain

0.53

0.16

0.09

straw

1.24 1.18 0.49 no data 0.07

1.21 0.31 0.24

0.81 0.39 0.18

0.11

0.06

0.74

0.47

0.36

grain I1 11

91

I1 ,I

9,

1.00

0.46

0.62

Potassium aDpl.. kg per ha

straw

The rate of transfer to barley after soil management is much lower than is that to hay grass. A reduction to one-tenth is common (Eriksson and Rosen, 1991), but the transfer of 13’Cs fiom peat soils to grain crops may still be noticeable. There are too considerable effects of time and of the annual potassium dressings. During the first year, in 1987, there was a significant difference between 100 and 200 kg Wha, but in 1988 to 1990 there were no significant differences between potassium levels. This was probably the result of an increasing level in the soil of plant-available potassium in both treatments. As straw has a higher mineral content than grain, caesium transfer to the former was twice as high as to the latter.

DISCUSSION AND CONCLUSIONS Placement effects

After heavy fallout, land that is in use may require a reduction of the external radiation level and of the transfer of fallout nuclides to feed and food. The burial of the upper contaminated soil layer by ploughing is one means of a chieving this. Ploughing is involved in regular soil management systems in agriculture and will, sooner or later, contribute to a deeper placement of a contaminated surface layer in the soil profile by mixing. Annual crops on arable land require annual ploughing, whereas

256

temporary grassland may be used for several years before being buried by the plough according to a the rotation scheme. Natural pasture land is not ploughed. An ordinary ploughing regime may thus be inadequate in such a situation. Of methods available for placement of a surface layer, ordinary ploughing and deep ploughing are important and the aim of both is the same: reduction of the radiation level by an efficient placement of the fallout nuclides in the soil profile and out of reach of plant root uptake. Studies have shown that at a moderate placement depth (25-30 cm), at the bottom of the plough layer, caesium uptake can be reduced by a factor of 2 - 4, and when combined with potassium fertilization by a factor of 8 - 10. Experimental data support placement of the potassium added on a caesium-contaminated layer before ploughing for annual crops because of the diluting effect. However, for perennial grass crops, this may concentrate the root systems in the contaminated layer and during following years, after depletion of the potassium applied in that layer, the caesium transfer to grass increased markedly. Enhancing root development in contaminated layers by the placement of fertilizers should therefore be avoided. A more efficient and deeper placement of a superficial soil layer can be achieved either by deep ploughs or two-layer ploughs. In practice, such work may be difficult. It may succeed on sandy soils under good moisture conditions, but will certainly be less successhl on heavy clays even if good equipment is used.

The effect of potassium fertilizers Much effort has been devoted to studies of the influence of potassium applications on the transfer of caesium from different soil types to various crops. Generally, it was found that soil properties, like the content of clay and organic matter, were important factors determining the accumulation of caesium and other minerals in plants, but also that the content in different crops and different crop parts could vary greatly. It was also established that, beside soil properties, the placement of the nuclide, the root feeding depth, the moisture, and other conditions in the soil profile are important factors which influence root uptake. Considering soil properties, it was found that heavy clays and peat soils are the extremes with regard to the supply of caesium to crops, being high and low recipients, respectively. The former soils tend to bind caesium efficiently to the soil material, to be rich in plant-available potassium, and to have a stable potassium status. The latter bind caesium loosely, have only small reserves, and an unstable and low status of plant-available potassium. The loams and sandy soils are grouped between these soil types. It is obvious that the need for potassium fertilization of cultivated soils increases in the sequence: Clays < Loams < Sandy soils < Peat soils, and that the stability of the potassium status decreases in the same order. As a result the crop uptake of caesium should increase markedly in the same order. It was found that the effect of potassium increases, but also that the depletion rate of the potassium that is plant-available in the soil increases in the same order. Consequently, to retain the balance and avoid high caesium transfers, annual applications of potassium are necessary on the poorer soils.

251

The range of potassium levels tested in experiments with contaminated soils is wide, from 30 to about 500 kg potassium per ha. Potassium levels up to 200 kg K per ha and year have efficiently reduced caesium transfer on both grass and arable land, higher levels have not shown krther reductions. Other countermeasures possible in Nordic agriculture

Agricultural countermeasures after a release of radioactivity to the atmosphere aim to reduce the radiation risks for farmers in the areas affected and for consumers of the food produced in these areas. Work for these purposes may include increasing the preparedness for such situations before any accident occurs as well as direct action taken afterwards. The countermeasures will deal with the problems arising in animal and plant husbandry and with the disposal of the contaminated surface soil layer. Deposition on this layer gives rise to external radiation doses, to resuspension and movement of radioactivity, and later to contamination of the next crop. To be efficient, the countermeasures have to be combined and put into action at the right season. In addition to burial of the contaminated upper soil layer by ploughing operations and the potassium fertilization discussed above, a list of countermeasures may include others equally important in both the short- and the long-term. 1. When crops are growing the crops the intercepted fallout nuclides are spread over a large amount of plant material. In addition, the radioactivity is reduced by decay and by falloff. Postponing the harvest, if a feasible option, therefore considerably reduces the radioactivity in the harvested crop. A factor of 2 - 3 may thus be achievable in grass cropping if early silage cuts are replaced by a hay harvest later in the season. 2. If at harvest the value of the contaminated crop is questionable, it must be removed and discarded. The efficiency of the decontamination of the land by early removal depends on the size of the fraction of fallout intercepted and retained by the crop. A dense crop can thus intercept 25 to 50 % of wet fallout and sometimes more of dry deposition. The method creates fairly concentrated wastes which have to be disposed of. 3. Change the method of harvest after early fallout. Lifting the cutting level when harvesting grass after Chemobyl in 1986 could thus reduce the caesium content in the grass by a factor of 10 (Hadders & Nilsson, 1987). 4. Changes in crops and varieties cultivated. It has been observed that caesium uptakes in barley

varieties can differ by 30 - 40 %. Also the uptake in different varieties of rye grass could differ by a factor of 2 ((ahlenschlaeger & Gissel-Nielsen, 1991).

258 5 . Scraping of the soil and removal of the soil surface layer can be quite an efficient method to reduce fallout on the land. Thus, on grassland, a flail-type forage chopper with equipment for removal of the crop and sod has been used experimentally with 90 % efficiency for artificial contamination, because part of the upper soil layer, altogether 4 to 10 tons per hectare, was removed as well (Menzel& James, 1971). To remove a 5 cm soil layer on each hectare of land, means transportation of a soil mass of at least 500 tons.

6. Changes in plant husbandq in the fallout district. The principle being that large areas of the grassland section in crop rotation are replaced by arable crops producing coarse grain. seeds and other utilities. In this way the direct, high-level nuclide transfer to the food chain is reduced. 7. Cease agricultural use of the land.

ACKNOWLEDGEMENTS This investigation was initiated and financial support by the Nordic Nuclear Safety Research (NKS) programme, working group RAD-3. Financial support was also given by the Swedish Radiation Protection Institute and the Swedish University of Agricultural Sciences.

REFERENCES Andersen, A.J., (1967). Investigation on the plant uptake of fission products from contaminated soil. I. Influence of plant species and soil types on uptake of radioactive strontium and caesium. Ris0 Report No. 170. Riss National Laboratory, Denmark. 32pp Andersen, A.J., (1967). Investigation on the plant uptake of fission products from contaminated soil. II. The uptake of radioactive strontium placed at different depths in the soil. Rise report no. 174. Risgi National Laboratory, Denmark. 19pp Eriksson, A (1991). Historical background to the studies of the transfer of Chernobyl fallout in Swedish agriculture. In: The Chemobyl fallout in Sweden. Ed. by L. Moberg, Swedish Radiation Protection Institute, Stockholm. 285-290. Eriksson, A. and L. Fredriksson (1972). Plant uptake of 137Cs from different depths in a homogeneous subsoil layer. Farsvarets forskningsanstalt avdelning 4 Report C 4491-A 4, Stockholm. 13pp. &son, A., K-J. Johansson, and H. Unsj6 (1990). Livsmedelsproduktion efter k#m-vapenkrig. Rapport Over fallstudie utfOrd pll uppdrag av Statens Jordbruksnhnd. Rapport SLU-REK-65. Swedish University of Agricultural Sciences. Department of Radioecology, Uppsala. 61pp. Eriksson, A. and K. Rosdn (1991). Transfer of caesium to hay grass and grain crops after Chernobyl. In: The Chemobyl fallout in Sweden. Ed. by L. Moberg, Swedish Radiation Protection Institute, Stockholm. 291-304 Fredriksson, L., A. Ekiksson, and H. Unsj6 (1966). Studies on plant accumulation of fission products under Swedish conditions, W. Uptake of 137Cs in agricultural crops as influenced by soil characteristics, and rate of potassium fertilization in a three year micro plot expriment. F6rsvarets forskningsanstaltavdelning 4 Report A 4486-4623, Stockholm.2Opp. Fredriksson. L., E. Haak and A. Ekiksson (1969). Studies on plant accumulation of fission products under Swedish conditions, Xl. Uptake of 90Sr by different crops as influenced by liming and soil tillage operations. Fbrsvarets forskningsanstaltavdelning 4 Report C 4395-28, Stockholm.57pp

259 Haak. E. (1986). Transfer of l37Cs to pasture grass in a long term situation. In: 1st international contact seminar in radioecology. Uppsala 8-11 July 1985. In: Ed. by A. Eiksson. Report SLU REK-61. Swedish University of Agricultural Sciences, Department of Radioecology, Uppsala. 123-130. Haak, E., A. Eriksson, and F. Karlstr6m (1973). Studies on plant accumulation of fission products under Swedish conditions. XIII. Entry of 90Sr and l37Cs into herbage of contrasting types of pasture. F6rsvarets forskningsanstaltavdelning 4 Report C 4525-A 3, Stockholm. 44pp. Haak, E. and A. Eriksson (1973). Studies on plant accumulation of fission products under Swedish conditions. XIV. Uptake of 137Cs by wheat and timothy from six different soils as influenced by rate of K-fertillization and by type and rate of N-fertilization in pot experiments. F6rsvarets forskningsanstaltavdelning 4 Report C 4557-A 3, Stockholm.37pp. Hadders, G. and E. Nilsson (1987). Sk6rd av foder som drabbats av radioaktivt nedfall. JTI-rapport 87. Swedish University of Agricultural Sciences, Jordbrukstekniska institutet Uppsala. 23pp. Haugen, L., H. Oskarsen., A. Karlsen, and L.Bruflot (1990). Plantetilgjenglighetav radiocesium pll ulike jordtyper. Ed. by T.GunnerBd and T. Garmo Seminarium 6-7 november 1990, NLWs forskningsprogram om radioaktivt nedfall. Norges LandbruksvitenskapeligeForskningsrlld. k, Norway. Nr. 28,37-57. Haugen, L. and G. Uhlen (1992). Radioaktivt nedfall fra Tsjernobyl-Ulykken. Felger for norsk landbruk, natunnilje og matforsyning. Ed. by T. Garmo and T. Gunnerd Norges Slutrapport fra NLVFs forskningsprogram om radioaktivt nedfall 1988-1991 Landbruksvitenskapelige ForskningsrAd. As,Norway. 65-80. Jensen, H. (1979). Dhpningsfaktorer f6r g a m m a s t r f i g fra deponeret radioaktivtet opnAet ved plejning af jord og asfaltering af veje. Rise Rapport 237-1212. Rise National Laboratory, Denmark. 41pp. Unsjb, H. and E. Haak (1986). Effects of deep placement and potassium fertilization on the crop uptake of caesium and shontium. Rapport SLU-REK-60. Swedish University of Agricultural Sciences, Department of Radioecology, Uppsala. 43pp. Unsj6, H., E. Haak and K. RosCn (1990). Effects of remedial measures on long term transfer of radiocesium from soil to agricultural products as calculated from Swedish field experimental data. In: Environmental contamination following a major nuclear accident, International symposium Vienna, Austria, 16-20 October 1989. V01.2: MA-SM-306132, Vienna. 151-162. Malm, J., A. Rantavaara, A. Uusi-Rauva and 0. Paakkola (1991). Uptake of 137Cs from peat and compost mould by vegetables in a greenhouse experiment.J. Environ. Radioactivity 14, 123-133. Meisel, S.,M. Gerzabek and H. Muller (1991). Influence of plowing on depth distribution of various radionuclides in the soil. Tiefenverteilung von Radionukliden, Weinheim. Z . PjZunzenerdhr. Bodenk., 154,211-215. Menzel, R.G. and P.E. James (1971). Treatments for farmland contaminated with radioactive material. Agriculture Handbook No. 395. Agricultural Research Service United States, Department of Agriculture. Washington D.C. Nilsson, J. (1983). Nedphjning av simulerad radioaktiv belllggning pll jordbruksmark. Rapport SLU-REK-56. Swedish University of Agricultural Sciences, Department of Radioecology, Uppsala. 27pp. Puhakainen, M. and T. YlPanta (1992). Uptaken of radionuclides by spring wheat and barley from cultivated soils supplemented by contaminated sewage sludge. Agric. Sci. Finl. 1,27-36. Roed, J. (1982) Reduktion af dosis ved nedplajning af gamma-aktive isotoper. Rise-M-2275. Rise National Laboratory,Denmark. 27pp. Roed, J., (1991). Radiation protection programme. Final report. Design and development of skim and burial plough for reclamation of contaminated land. Rise National Laboratory, Denmark. 9pp. RosCn, K. (1991). Effects of potassium fertilization on cesium transfer to grass, barley and vegetables after Chernobyl. In: The Chernobyl fallout in Sweden. Ed. by L. Moberg, Swedish Radiation Protection Institute, Stockholm. 305-322. 0hlenschleager. M. and G. Gissel-Nielsen (1991). Differences in ability for barley and rye grass varieties to adsorb caesium through roots. Acta Agric. Scand. 41,321-328.

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263

4.1. INTRODUCTION TO TERRESTRIAL SEMINATURAL ECOSYSTEMS

AINO RANTAVAARA Finnish Centre for Radiation and Nuclear Safety, P. 0. Box 14, FIN-00881 HELSINKI

Fairly constant radiocaesium contents were discovered in foodstuffs of terrestrial seminatural origin in successive years after 1986. Regardless of the ageing of the Chernobyl fallout, the long residence times of radiocaesium in these products indicated, that radioactive decay was often the only process reducing the amounts of bioavailable radiocaesium in forest and alpine ecosystems. Comparisons of postChernoby1 data with findings from the period after the fallout from nuclear weapons revealed a similar temporal pattern of radiocaesium contents in foodstuffs of wild origin. The importance of some forest and alpine ecosystems, for example those linked with reindeer herding, to the amount of dietary radiocaesium received by man has been known for decades. The novelty of post-Chernobyl findings to many scientists, and especially to the public, was that all food products of terrestrial seminatural origin contained significantly more radiocaesium than agricultural produce. The goal of the research programme was to obtain and review comparable quantitative information of relevance to human exposure to radiation in the Nordic countries. Both biological and radioanalytical expertise was represented in most research teams. The seasonal changes of radiocaesium contents in food products received from seminatural ecosystems and the reasons for the obviously constant bioavailability of radiocaesium in the soil were among the topics of interest. The need to find the most cost-effective measures for intervention lies behind the approaches chosen. The main emphasis is on the ingestion pathways, but the information on the distribution and redistribution of radiocaesium deposition also pertains to the estimation of external exposure. A simplified approach in the assessment of external doses does not consider gamma shine from radionuclides in canopies of trees. A marked difference in the dose rates occurring in a forest environment results when new direct deposition on trees is compared to a case of

aged fallout where radionuclide transport is associated with the nutrient cycle of the ecosystem.

264

The main components of seminatural terrestrial systems - soil, field vegetation, trees and game, including the semi-domestic reindeer, are discussed in sub-chapters. The typical features of the northern seminatural terrestrial ecosystems include an abundance of boreal coniferous forests in Scandinavia and in Finland, and varied topography connected with arctic conditions in the northernmost parts of each country. The importance of large mammals, such as elk, reindeer and roedeer, to the intake of radiocaesium of hunters and reindeer herders in Scandinavia and Finland led scientists to investigate the diets of these animals. The size of population receiving exceptionally high radiation doses through foodstuffs from forest ecosystems varies from country to country, and the efforts made to reduce the human exposure vary accordingly. Soils are dominated by a podzol soil profile, where the thickness of the organic litter and humus layer varies with climate. Field vegetation types reveal differences in nutrient status and

in other growing conditions. The major sources of variation in radiocaesium transfer to foodstuffs of wild origin are uneven surface contamination and differences in growing conditions. The remarkably uneven horizontal distribution of radiocaesium required careful analysis of experimental fields to avoid misleading estimates for different transfer parameters. The melting of snow following ‘deposition in spring 1986 certainly increased the variation in surface contamination by fallout. Interesting findings on different half-lives of radiocaesium in reindeer, probably caused by differences in the use of lichen fields for grazing of reindeers, were discussed. The ability of reindeer to adapt to feed other than lichen in circumstances where lichen is not available, has become evident in Iceland. New information was also obtained on the ecosystems that have been studied intensely since the early 1960s.

265

4.2. THE TRANSFER OF RADIOCAESIUM FROM SOIL TO PLANTS AND FUNGI

IN SEMINATURAL ECOSYSTEMS ROLF A. OMEN Department of Biotechnological Sciences Agricultural University of Norway SUMMARY

On natural pastures and forest areas in the Nordic countries about 90% or more of the deposited radiocaesium after the Chernobyl accident is found in the upper 4 cm of the soil four years after the fallout episode. A vertical distribution in the upper soil layer took place witin a short time, whereas an equelibrium between radiocaesium and stable caesium at least was established after three years with about 15%of the rddiocaesium content being easily extractable and most probably bioavailable. The uptake of radiocaesium in plants varied in different species and in different ecosystems. The transfer factor (T, ) of pooled plant samples from natural pastures varied in 1990 from 0.005 to 0.05 mz/kg d.w., while the variation in different acid forest soil ecosystems was between 0.08 and 0.3 mz/kg. The radiocaesuim content in fungal fruit bodies of different species varied considerably, but was for many species, all of them mycorrhizal basidiomycetes, more than 50 times higher than in plants sampled on the same site. The T, -factor for these fungi varied from 3 to 11 m2/kg d.w. and was approximately constant during the period from 1988 to 1992. A comparison between the T, -factor of r a d i d u m from the Chernobyl and the weapons tests fallout in fungi from acid forest soil, indicates that the biological and physical half-life in these ecosystems is more or less equal. INTRODUCTION

In the Nordic countries the radiocaesium fallout after the Chernobyl accident was deposited in

different terrestrial ecosystems. In Finland, Sweden and Denmark the major deposition took place on agricultural soil and in coniferous forest areas, whereas in Norway the most heavily contaminated areas, with few exceptions, were in upland and mountain regions in central parts of southern Norway. Thus economic interests and the differing usage of the contaminated areas in each Nordic country led to somewhat different attitudes to the implications for various terrestrial ecosystems in each country. In Norway, the upland and mountain regions are of considerable economic interest for farmers, and much effort was made to elucidate the effect the accident would have on the use of natural and upland pasture for grazing animals. It soon became clear that in addition to the transfer of radiocaesium from soil to plants and further to animals, the transfer through the

266

consumption by animals of fungi was of great importance in the years when a large mushroom crop occurred.

In Sweden studies on the transfer of radiocaesium in seminatural ecosystems were focused on the transfer from soil to game animals through consumption of grass, herbs and different heather plants in pine-forest areas, while in Finland much work was done on measuring the contents of different radionuclides in meat, berries and mushrooms used for human consumption.

In addition, studies of the transfer of radiocaesium to reindeer were a main object of interest in Finland, Norway and Sweden. In Denmark most activity was focused on the agricultural ecosystem and only few investigations were made of the transfer of Chernobyl fallout in seminatural ecosystems. Thus most of the results concerning the transfer from soil to plants and fungal fruit bodies in upland and mountain regions emerged from Norwegian studies.

DEPOSITION AND WDISTRIBUTZON OF RADIOCAESIUM FROM CHERNOBYL FALLOUT The amount of radiocaesium deposited in the different regions in the Nordic countries depended

on the meteorological conditions in the first weeks after the Chernobyl accident. Therefore there were variations in deposition in the different provinces in Finland, Norway and Sweden. A deposition map for the Nordic area shows that the main deposition took place in southern Finland and central Sweden and Norway, while minor amounts of “Chernobylcaesium” were deposited

on Danish soil. In Norway, a national survey of the amount of fallout was made in June 1986 by the National Institute of Radiation Hygiene. Soil samples from practically all municipalities (4 samples from each) were pooled, analyzed and the main deposition per mz calculated. In spite of the rough sampling technique used in this survey, the final map for the radiocaesium fallout pattern in different areas showed that the largest deposition had occurred in central Norway, in Oppland, Hedemark and Trrandelag provinces. It also showed that in the two former provinces the highest deposition was found in upland and mountain areas, while large deposition also occurred on agricultural soil in Trrandelag. The national deposition map showed that upland and mountains areas received on average more than 80 kBq/m2. It soon became clear that a large microscale variation in deposition had taken place, and that the variation was largest in natural pasture, and in high altitude areas covered in mountain birchldwarf birch. Figure 4.2.1 shows the frequency distribution of radiocaesium in 200 soil samples taken inside an area measuring

261

about 0.5 hectare, covered by mountain and dwarf birch. The radiocaesium content varied from about 10 to about 700 kBq/mZ, with mean and median values of about 80 and 60 kBq/m2 respectively. In lower-lying pine forest areas less variability was observed. Out of 30 Soil

FIGURE 4.2.1. FREQUENCY HISTOGRAM OF RADIOCAESIUM DISTRIBUTION IN SOIL FROM AN AREA COVERED BY MOUNTAIN BIRCH.

samples from a 0.5 hectare area, the highest deposition was estimated to about 100 kBq/m2. However most of the soil samples had a content corresponding to less than 50 kBq/m2 (Figure 4.2.2). A comparable distribution Pattern was observed in a Swedish pine forest soil (Bergman et al. 1991),

FIGURE 4.2.2. FREQUENCY HISTOGRAM OF RADIOCAESIUM DISTRIBUTION IN SOIL FROM A PINE FOREST AREA.

whereas in Denmark only small amounts of radio-caesium from the Chernohyl fallout, average amount about 3 kBq/mZ,were recorded in Danish forest ecosystems (Roos 1990, Stranberg 1992).

REDISTRIBUTION OF RADIOCAESIUM DEPOSITION Radioactive fallout can occur both as dry and wet deposits on the soil surface; radiocaesiurn from the Chernobyl accident took the form of a wet deposit on bare soil in the lowland and parts of the upland, but the mountain areas were still covered in snow in different stages of melting. The deposition took place as an episode, and the amount of deposition correlated with the

268

amount of precipitation in each locality. The deposition on soil of fallout from nuclear weapons tests is also well correlated to the amount of precipitation in each area, but this fallout material was continuously deposited over a long time period, and resulted in an approximately uniform distribution of radiocaesium per m2. The short, episodic nature of the Chernobyl fallout, together with the fact that deposition took place on soil with different coverages of snow, led to both large regional and local variations in the amount of deposited radiocaesium.

LATERAL TRANSPORT Wet deposition on a bare, unfrozen soil surface will normally lead to a direct infiltration of material dissolved in the rain, as is the case for radiocaesium from fallout following nuclear weapons tests. Wet deposition of contaminated precipitation on snow-covered or frozen soil on sloping ground will lead to deposited material heeing carried downwards with the meltwater to more level ground. The running meltwater makes pools in grooves in the soil surface or when its passage is blocked by obstacles such as stones or the base of tree-trunks, with the consequence that there are large site variations in the amount of radiocaesium deposited inside small areas. A “concentrating process” as mentioned above seems to have taken place in the mountain areas still covered with snow during the Chernobyl fallout period. Figures 4.2.3A and 4.2.3B show the large variations in the radiocaesium soil content in a 10x10 mz area. The number of “hot sites“ is largest on level ground, but they also Occure on sloping terrain covered by mountain birch. A thorough

FIGURE 4.2.3A. A 3-D PLOT OF THE RADIOCAESIUM DISTRIBUTION IN A MOUNTAIN-BIRCH-COVERED SLOPE.

269

investigation of the variation of radiocaesium fallout inside a subcatchment in a mountain valley (Heimdalen) in the Jotunheimen area was carried out by Haugen and Bjprrnstad (1990). This valley received large amounts of Chernobyl fallout and has a highly varied deposition pattern. In the subcatchment the amount ofradiocaesium was found to be higher in

FIGURE 4.2.3B. A 3-D PLOT OF THE RADIOCAESIUM DISTRIBUTION ON LEVEL GROUND.

the depression than on the ridges on both sides. They also observed a large content inside a 5x5 m2 area in the depression. A similar variability in radionuclide pattern in a forested area as a result of topography has been published by McGee et al. (1992) and on alpine tundra by Osburn (1963). VERTICAL TRANSPORT

The vertical distribution of radiocaesium is a result of downward transport in soil water, by diffusion and by adsorption of the caesium ion to mineral and to charged sites on organic particles, and most probably also by accumulation in fungal mycelia on sites having a large amount Of

fungal biomass.

FIGURE! 4.2.4. PERCENT OF RADIOCAESIUM DISTRIBUTION IN SOIL PROFILES DURING THE PERIOD 1986 TO 1990.

270

Depth distribution measurements of soil cores taken from different mountain upland and natural pasture localities showed (Figure 4.2.4) that practically all the radiocaesium content (-99%) was contained in the upper 5 cm of the soil profiles (-94% in the upper 3 cm), and that

no changes in the depth distribution had occurred between 1986 and 1990 (Haugen pers.com.). The bioavailability of the soil content of caesium, and thus radiocaesium, is strongly influenced by the chemical and physical properties of the soil. The caesium binding to negative sites on clay particles makes such bound caesium almost irreversibly fixed, whereas the exchangeability of caesium bound to negatively-charged sites on organic material in the soil is probably higher. At some sites in the organic layer of podzol up to 5oooO m of fungal hyphae have been recorded per g d.w., which is equivalent to about 1 kg fungal biomass per m2 (Oisen unpublished). With the high accumulating ability found in most fungi, it is possible that a large part of the soil radiocaesium content is accumulated in fungal structures in the litterlhumus layer. Thus caesium ions can be present as:

1) free ions in the soil solution, 2) bound, but in an easily exchangeable state,

3) irreversibly fixed to clay minerals, or 4) recycled within the living fungal biomass;

and the bioavailable caesiumiradiocaesium will be the amount present as either free- or easily exchangeable ions. By using a sequential extraction procedure, Riise et al. (1990) measured the relative distribution of stable caesium and radiocaesium in different extractions from the upper 0-2 cm of a soil sample from natural pasture in the Jotunheimen area (Figure 4.2.5). It was concluded that about 15% of the total radiocaesium content could be classified as easily

FIGURE 4.2.5. PERCENT DlSTRIBUTION OF '"Cs AND STABLE Cs (l'JCs) IN DIFFERENT EXTRACTIONS.

27 1

extractable and thus probably bioavailable. The approximately constant ratio between stable caesium (1J3Cs)and radiocaesium (lJ7Cs)in the different extracts indicated that equilibrium between stable caesium and radiocaesium had occurred 3 years after the fallout episode.

ACCUMULATION OF RADIOCAESIUM IN PLANTS ON NATURAL PASTURE

Radioactive fallout material can contaminate plants by two different mechanisms. -By direct contamination of dead and living plant parts. -By root uptake from the soil. The consequence of a direct contamination of plant surfaces will depend on the time of the fallout, and will have the most severe effect in the year following the deposition of fallout. Thereafter the radionuclide content of plants will depend on the root uptake from soil. Figure 4.2.6 shows the radiocaesium content in grass collected from a natural pasture in Heimdalen (Haugen unpublished). The high content in the autumn of A: Autumn

1986 must originate from

s:summer

direct contamination of last year’s dead leaf and of the plant anlage. The content in grass showed a considerable decrease in 1987, while

I

1986A

grass from this natural

l987A

19888A

3989A

I9WA

YtX

from 1988 onwards the

radiocaesium content in

I9873

FIGURE 4.2.6. THE RADIOCAESIUM CONTENT IN GRASS FROM A NATURAL PASTURE IN HEIMDALEN (Haugen unpublished).

pasture has shown a very slow decrease, which could be expected as a result of physical decay and stabilization of the amount of bioavailable radiocaesium in the soil. The difference in plant content that originates from soil factors is shown in Figure 4.2.7. Pooled plant material sampled on dry mountain upland pastures contained more radiocaesium in the first years after the Chernobyl fallout episode than did pooled plant sampled from peatlands; the latter having a very low content in 1986 and 1987 compared to plants from dry soil areas. After 1988 practically no changes in plants from peatlands have been observed. The

212

radiocaesium content in different plant species in

Moorland Peatland

+

natural mountain pastures, harvested in the autumn of 1987, varied considerably as shown in Figure 4.2.8 (Staaland et al. 1990) . FIGURE 4.2.7. RADIOCAESIUM CONTENT IN POOLED SAMPLES OF DIFFERENT PLANT SPECIES FROM NATURAL PASTURES IN 1987 (Staaland et al . 1990).

FIGURE 4.2.8. THE RADIOCAESIUM CONTENT IN POOLED SAMPLES OF DIFFERENT PLANT SPECIES FROM NATURAL PASTURES IN 1987 (Staaland et al. 1990).

These authors found that most plants, including herbs, grasses, rushes and sedges, contained less than 1.5 kBqlkg d.w. material, with common sorrel being the only exception containing 5.5 kBq/kg d.w.. The investigation was repeted three years later (1990) when 10 out of 19

species contained less than 50% of their 1987 contents; 7 species contained between 50% and 140%, but cotton grass and buckbean had an increased radiocaesium content corresponding to 478% and 236%, respectively, of their 1987 values. It is interesting to note that both these plants are growing on peatland.

273

TRANSFER FACTOR FROM SOIL TO PLANTS ON NATURAL PASTURE

The transfer factor (Tf) for pooled plant samples from natural pastures from "high fallout" provinces in Norway, calculated from measurements performed in 1986 and 198911990, is shown in Figure 4.2.9

-

(Haugen unpublished). In 1986 about 25% of the plant

0.8T

samples had T, greater than

1986 1989iW

0.18 m2/kg d.w. and about 2.5% had a T, of about 0.7 m2/kg. In 1989/1990 the transfer factors were drastically reduced with less

a 0

0.1 m2/kg. The high T, for

40

60

80

100

% of sampies

than 5% of the samples having a T, above

20

FIGURE 4.2.9. T, FOR POOLED PLANT MATERIAL FROM UPLAND AREAS IN 1986 AND 1989/1990.

many of the samples in 1986 is most probably explained by direct contamination of the plant material. The variations in T, in 1988/1990 are probably a consequence of soil type and root distribution. In nutrient poor-pcdzol soil much of the fine root biomass is concentrated in the upper organic layer, whereas roots penetrate to greater depth on richer soils. Differences in soil types on the growth site of the plants will also affect the ecological half-live as shown by Naeumann et al. (1990) for dwarf birch. They found twice as long an ecological half-life on podzol as on a richer soil.

THE TIME PERSPECTIVE OF THE TRANSFER OF RADICAESIUM FROM SOIL TO PLANTS IN NATURAL PASTURE AREAS

After the large decrease in the transfer factor in the first two years following the fallout episode, the T, has only decreased very slowly, and on peatland it has remained practically constant. As we have at present only about five years of experimence in measuring T, is it difficult to predict the future transfer rate and ecological half-life of radiocaesium in different plants from natural pastures. The first predictions based on measurements of the radiocaesium transfer in 1986 and 1987 concluded with an ecological half-life of 15 months. A further calculation made in 1990,

214

omitting the 1986 data, gave an estimate of 33 months, and if the calculation had been based on changes in the transfer of radiocaesium to plants on peatland from 1988 to 1990, a much longer ecological half-life would have been found. Thus we shall have to wait several years for more reliable values for the radiocaesium half-life in plants on different soil types.

ACCUMULATION OF RADIOCAESIUM IN FUNGI AND PLANTS IN UPLAND AND MOUNTAIN AREAS

In 1988 an investigation was started to study the accumulation of radiocaesium in the fruit bodies of different fungal genera and species. It soon became clear that fruit bodies of most of the common fungal genera had a very high radiocaesium content, in some cases up to 100 times more than that in the plant material. It was also observed that the content in fruit bodies of different genera and also of different species within a genus varied considerably. In addition fruit bodies within a species showed an extreme site variability, and most of the variation was caused by the differences in radiocaesium deposition at each site. For a better understanding of macrofungi as radiocaesium accumulators, a brief summary of their ecology might be helpful. With few exception the larger fruit bodies in areas covered by trees or shrubs belong to the group of basidiomycetes which forms ectomycorrhiza. The total fungal biomass related to each fruit body consists of the above-ground part, the fruit body, mycelial mats connecting the fruit body to one or more mycorrhiza, and varying amounts of mycelial "strands" which spread out into the surrounding soil volume. The amount of mycelium is specific for each fungal genus or species, and is very variable. Most of the total underground biomass is situated in the upper organic layer in stratified soils as podzol. In soils were the organic material is well mixed with the mineral material, the mycelium is more evenly spread deeper down in the soil profile. When we compare the total amount of fungal biomass above and below the soil surface for each fruit body, as much as 99% can be present in the soil. Because of the large variation in the radiocaesium content in fruit bodies of the same species, and between different genera, it was necessary to take soil samples at each sampling site in order to calculate the radicaesium transfer factor. It was also shown that the content in grass and herbs, taken at the same sites as the fungi, was much less affected by variations in site deposition than was that of the fungi, as shown in Figure 4.2.10. This sampling strategy made possible an integrated data analysis regarding size and variation of the transfer factor compared to the amount of deposition at each site. The different fungal species have great, but differing

275

abilities to solubilize chemically-bound ions. In our modelling studies the plant contents at each sample site were used as a measure of the bioavailability of radiocaesiurn. TABLE 4.2.1. RADIOCAESIUM CONTENT IN FRUIT BODIES OF THE MOST COMMON FUNGAL GENERA AND FUNGUS/PLANT RATIOS. Site

Ecosystem Genus

Leirungsdalen

Mountain birch

,I

Griningsdalen

Mountain birch

Griningsdalen

Pine

Sikkilsdalen

Pine

Amanita Cortinarius Lactarius Leccinum Rozites Russula Tricholoma' Xerocomus

Radiocaesium content kBqlkg d.w. Mean Min - Max

125 282 175 29 287 43 181 8

Fungudplant ratio Mean

Min - Max

1 - 657 14 - 1384 16 - 1077 3 - 76 122 - 687 5 - 182 7 - 759 6 - 10

16.1 61.8 63.0 9.5 59.0 12.1 84.5 1.8

0.2 - 78.0 9.5 - 222.5 3.5 - 262.6 0.2 - 48.2 29.3 - 109.5 1.1 - 42.0 2.1 - 218.4 0.3 - 4.0

Cortinarius Lactarius Leccinum Rozites Tricholoma'

64 56 0.3 61 6.3

1 - 143 0.2 - 87 0.2 - 0.6 26- 128 2.5 - 10.7

42.7 21.4 0.1 19.4 21.4

12.5 - 214.0 0.3 - 36.6 0.0 - 0.2 0.5 - 48.9 15.8 - 48.3

Cortinarius Lactarius Leccinum Russula Suillus

100 41 0.8 8 32

2 - 470 1 - 98 0.2 - 1.6 4 - 13 20 - 40

52.6 13.8

0.0 - 108.3 0.0 - 26.7

0.4 157.0 49.5 41.6 27.5 306 21.4 5.0 1.5

0.1 - 0.8 23.5 - 420.4 15.7 - 117.1 8.9 - 78.3 1.2 - 169.2 10.4 - 55.5 7.0 - 31.7 2.0 - 12.4 0.5 - 3.8

Amanita Cortinarius Lactarius , Suillus Tricholoma' Pine Cortinarius SkAbu Lactarius Suillus Tricholoma ' Large variation among species.

3 439 102 63 136 142 72 25 4

2-5 39- 1047 15 - 236 23 - 161 2 - 783 55 - 290 17 - 103 11 - 4 6 1 -~ - 9

The radiocaesium contents in different fungal species, collected in both mountain birch and pine forest ecosystems in the Jotunheimen region, have been published by Bakken and Olsen

(1990). In general all species of the genus Cortinarius together with Rozites cuperutu had a very

276

high radiocaesium content, the same was true for some species of the Lactarius and the Tricholoma genera and one species of Amanita.The other Amanita species, including the most common toadstool, Amanira muscuria, together with all investigated Russula and Leccinum species, had a low content irrespective of the soil content at the growing site.This result is in accordance with results obtained in other countries as shown by Seeger and Schweinshaut (1981), Byrne (1988), Haselwandter et al. (1988) and Dighton and Horrill (1988). A summary of the mean value and the variation of the radiocaesium content in different fungal genera, from birch and pine forest ecosystems, together with the fungal/plant ratios is shown in Table 4.2.1. The average radiocaesium content in the different genera varied considerably, from less than 10 kBq/m* in Xerocomus to more than 280 kBq/m* in Cortinarius and Rozites. Table 4.2.1 also shows the large variability within a genus. In some genera this is mainly caused by the large variability in deposited amounts of radiocaesium, but in other genera the variable contents are caused by the variability to accumulate radiocaesium of different species within the genus. The variability is

Kozites caperata

c

Ihricholoma album 0

most pronounced in Tricholoma and Amanita,

two genera in which most of the species had a very low radiocaesium content, but both having one species with a very high content. In comparison, grass and herbs at the same sampling site had only a low variation in the radiocaesium content compared to most fungi (see Table 4.2.2). In general all investigated fungal species showed a certain correlation

k13qlm2

FIGURE 4.2.10. COMPARISON OF THE RADIOCAESIUM CONTENTENT IN FRUIT BODIES OF "HIGH ACCUMULATOR" FUNGI AND PLANTS, AS A FUNCTION OF THE SOIL RADIOCAESIUM CONTENT.

271

between the soil content

Xerocomus subtomentosus

(kBq/m2) and the accumulated amount of radiocaesium, with the

40

Leccinum scaber

i

exception of fruit bodies of

Xerocomus subtomentosus. This is a typical fungal

0

species forming mycorrhiza in acid and nutrient-poor soil. All sampled fruit bodies of this fungus contained between 6 and 10 kBqlkg d.w., while the

0 00 0 0

100

I

200

l

l

300

I

l 100

l

1

200

kBq/m'

FIGURE 4.2.11. COMPARISON OF THE RADIOCAESIUM corresponding soil content at CONTENTENT IN FRUIT BODIES OF "LOW ACCUMULATOR" FUNGI AND PLANTS, AS A FUNCTION the different sampling sites OF THE SOIL RADIOCAESIUM CONTENT. varied from 37 to 265 kBq/mZ.Figure 4.2.10 and 4.2.1 1 show the variation in the fruit body and plant content as a function of the soil content for two fungi with a high ability, and two fungi with a low ability to accumulate radiocaesium. Fruit bodies of the different species of the genus Cortinarius showed comparable variation in their contents with lowest and highest amounts being 14 and

1385 kBq/m2 respectively. Most of the investigated fungal genera belong to the mycorrhizal group of the basidiomycetes. The other large group of macrofungi, the litter decomposers, have in most cases only very small to small fruit bodies, and because of that play practically no role in the transfer of radiocaesium from soil to grazing animals or to man. Fruit bodies of these litter-decomposing fungi have seldom been observed in the mountain and upland areas used in this investigation. Some very few samples have however been analyzed. Most contained low amounts of radiocaesium. More material must be analyzed though before a safe conclusion regarding the significance of this fungal group in radioecology can be drawn. Table 4.2.2 shows the average amount of radiocaesium per m2 in different ecosystems, together with the average content in grasslherbs at each sampling site.

278 TABLE 4.2.2. RADIOCAESIUM CONTENT IN SOIL AND FRUIT BODIES OF THE MOST COMMON FUNGAL GENERA. Mean Site

Ecosystem

Leirungsdalen Griningsdalen Griningsdalen Sikkilsdalen Skihu

Mountain birch

Pine

9,

PH (H20)

Mean PH (KC1)

Radiocaesium content kBq/m2

Radiocaesium content kBq/kg d.w.

Mean

Min- Max

Mean

4.9

4.0

85

6-688

4.6

0.4 - 2 3 . 0

5.0

4.3

25

11 -50

1.2

0.1 - 4.5

4.5

3.6

29

2 - 96

3.1

0.1 - 5 . 3

4.4 3.9

3.5 3.1

54 30

9 - 234 9 - 82

3.0 4.0

0.5 - 7 . 4 1.7 - 8.8

Min-Max

The highest mean deposition (85 kBq/m2) was observed in Leirungsdalen, an area covered by mountain birch and shrubs. The terrain is oriented north-west and the birch-covered part has an angle of inclination of 10-15O. The deposition varied considerably with min. and max. values

of 6 and 688 kBq/m2 respectively. Normally at least part of the area is covered by snow up to the beginning of June, and snow always remains much longer here than in the other areas investigated, of which those covered by pine lie about 200-300 m lower than the birch-covered areas. One of the areas in Griningsdalen is also covered by birch, but as it faces south the snow melts at least one month earlier than in Leirungsdalen. In this area, about 10 km from Leirungsdalen, the mean deposition value was 25 kBq/m2 and there was much less variability (min. 1 1 and max. 50 kBq/m2). In the investigated pine-covered areas, the mean amounts of deposition varied between 29 and 54 kBq/m2, with the highest variation found in Sikkilsdalen. Grass and herbs from these areas contained more radiocaesium compared to the soil content than could be expected from the values obtained in Leirungsdalen. The mean pH in soil samples was however significantly lower, being 0.5 to 1 .O unit lower than in Leirungsdalen. A correlation analysis of data from all the investigated areas showed a significant correlation between the In-transformed values of the radiocaesium content in fungi and plants and soil contents and pH, and the highest correlation coefficient, was found between the plant content and the soil data. A multiple regression analysis for all data from all areas showed that the soil content and pH accounted for 32% of the variations (R2=0.32). When the same analysis was performed on data from each of the different areas, higher regression coefficients were obtained. By using the data from the mountain-birch-coveredarea in Leirungsdalen, a

279

model based on the soil content per m2 and pH accounted for 46% of the observed variations (R2 = 0.46). Based on the 1988 data from Leirungsdalen, a preliminary prediction model for the radiocaesium content in plant material as shown below, was constructed (Bakken and Olsen unpublished) with the general form shown below.

In(P1ant content,,,,)

=

a

+ b * In(soi1 cont/rn2) - c * pH

The corresponding values for the fungal content were obtained from

In(funga1 cootentp,ear,,) = In(P1ant contentRedlct) * Fungal/plant-ratio (The values are corrected for physical decay.) A comparison between observed and predicted values for five of the most common genera

in 1989 and 1990, using the general model above, is shown in Table 4.2.3.

TABLE 4.2.3. COMPARISON OF PREDICTED AND MEASURED RADIOCAESIUM CONTENT (kBqlkg d.w.) IN FIVE COMMON FUNGAL GENERA (1988 -1990). Genus Cortinarius Lactarius Leccinum Rozites Russula

1988 Measured 267 181 22 287 42

1990

1989 Predicted 107 34 7

227 36

Measured 160 73 26 309 40

Predicted 89 114 17 178 48

Measured 195 135 5 220 68

The table shows that, except for the genus Leccinum in 1990, all the other predicted values were lower than those measured. The largest discrepancy between measured and predicted content was observed for Cortinarius in which the mean measured content is about twice the predicted. The predicted mean content in the different Leccinum species in 1990 is much less than predicted. However the number of samples taken of this genus in this year was low, a fact that makes the predicted value for Leccinum less reliable than for the other genera. In order to make a more realistic model for the future content of radiocaesium in different common fungi, more information seems needed about the behaviour of caesium in different soils and on uptake kinetics in different fungi under varying "soil environment conditions" as also better knowledge of their ability to make the chemically-bound radiocaesium hioavailable. With our present knowledge, it seems impossible to make a general model covering all fungal genera and different seminatural ecosystems. A better method seems to be the use of 2 or 3 genera of those fungi generally present in great numbers every year and which have shown the least variability

280

in their radiocaesium content. To describe the transfer of a radionuclide from soil to other parts of an ecosystem, the transfer coefficient (T, m'/kg d.w.) has been used. Although a useful parameter for describing the transfer of an evenly distributed compound in soil to plants, it is for obvious reasons less suitable for organisms such as fungal fruit bodies. Because of the lack of a better parameter for the comparing "transfer" from soil to fungi, and for comparing the transfer of radiocaesium in a long-term perspective, the transfer coefficient has been used for comparing the transfer from soil to h i t bodies of different genera and species in different ecosystems and in different years. In natural mountain pastures the T, -value for pooled grass samples dropped rapidly from

1986 to 1988, but thereafter the transfer coefficient has been relatively constant having a value of less than 0.1 m2/kg d.w. in 1989/1990 (Haugen pers.com.). The corresponding T,-values for grass/herbs in mountain birch and pine forested upland areas are shown in Table 4.2.4. The four first-mentioned areas in Table 4.2.4 had in 1988 a mean T, for the soil/plant transfer of

0.11 m2/kg d.w., which is practically identical to the transfer coefficient found for grass in a natural pasture area. Plant samples from one area (SkAbu), had a T, -value that was 3 times as high. The reason for this is unknown, but this area did have the lowest average soil pH values.

TABLE 4.2.4. COMPARISON OF THE T, -FACTOR FOR PLANT MATERIALS FROM DIFFERENT ECOSYSTEMS. Site

Ecosystem

Leirungsdalen Griningsdalen Griningsdalen Sikkilsdalen Skh

Mountain Birch Pine

T, soil/plant (m2/kg d.w.)

0.08 0.12 0.14 0.09 0.30

A mean T, for all fungal species with respect to radiocdesium would be meaningless

because of the large variations between genera, and sometimes hetween species resulting from the variation in their response to environmental factors; one such factor being the content of stable caesium and the specific uptake of this ion in different fungi. As shown in Figure 4.2.12, the transfer factor of radiocaesium is correlated with the fungal ability to accumulate stable caesium. Two species, Rozites caperufa and Lactanus utilis, deviate from this general pattern. They both have a higher transfer factor from soil to fruit bodies than expected from their content of stable caesium. This result is most probably explained by the fact that these two

28 1 fungi have most of their mycorrhiza and nutrient-

5 1

.

absorbing mycelia in the upper part of the soil organic layer, where the radiocaesium/stable caesium ratio is higher than deeper down in the 0

soil profile. A

1

2

3

4

5

6

7

8

9

10

11

ng slable &m per kg dw.

comparison of the Tf -

FIGURE 4.2.12. THE CONTENT OF STABLE CAESIUM COMPARED WITH THE T, -FACTOR OF RADIOCAESIUM FROM SOIL TO FRUIT BODIES OF DIFFERENT FUNGAL GENERA,

value for two genera having the least variation between species

is however possible. Table 4.2.5 presents T, -values for one "high-accumulating" and one "lowaccumulating" genus from different ecosystems.

TABLE 4.2.5. COMPARISON OF THE Tf -FACTOR OF A "HIGH AND A LOW RADIOCAESIUM ACCUMULATOR" GENUS FROM DIFFERENT ECOSYSTEMS. ~-

Site

Ecosystem

T, soil/fungi (m2/kg d.w.) Cortinarius

Leirungsdalen Griningsdalen Griningsdalen Sikkilsdalen SMbu

Mountain birch

, Pine

4.8

Russula 0.8

2.6 4.3 3.0 4.7

0.5 0.7

0.6 0.8

Except for the SMbu area, the soil/fungi transfer coefficients are 25 to 50 times higher than the corresponding soillplant T, -values.

TRANSFER OF RADIOCAESIUM IN A LONG-TERM PERSPECTIVE For a long-term perspective analysis of the ecological half-life of radiocaesium in fungal fruit

bodies, a comparison was made of the transfer coefficients for Chernobyl fallout and the nuclear weapons testing fallout. For this purpose soil, plants and different fungi were sampled in two areas with a low Chernohyl fallout deposition. One of the areas is in the south-eastern

282

region, the other on the west coast of southern Norway. In both areas the amount of Chernobyl fallout was about 2 k13q/mZand they had a comparable soil content of nuclear weapons testing fallout. In order to distinguish between the two different types of fallout in the sampled materials, a

ratio of "'CS to lS4Csof 2.0 was used (July 1986), published by Backe et al. (1987). The ratio was corrected for physical decay at the time of measurement (September 1989). The calculated T, of nuclear weapon radiocaesium from soil to different fungal fruit bodies was compared with the T, of the same genera sampled in the upland and mountain areas which had high Chernobyl fallout deposition. This comparison is shown in Table 4.2.6. TABLE 4.2.6. THE T, -FACTOR OF FRUIT BODIES OF DIFFERENT FUNGAL GENERA FROM SITES WITH A HIGH AND A LOW DEPOSITION OF CHERNOBYL FALLOUT RADIOCAESIUM.

site Leirungsdalen Griningsdalen Griningsdalen Sikkilsdalen Skabu Samnanger

EcoDepoAmanita 'Orti- Lactarius Rozites Suillus Leccinum Russula system sition narius Mount. 0.1 4.8 5.6 4.5 0.8 0.8 birch 85.0

Pine

25.0

-

2.6

3.0

3.5

29.0

-

4.3

2.0

-

54.0 30.0

0.1

3.0 4.1

3.7 2.5

3.1 3.1

0.2 0.3

4.0

3.7 3.4

AS

-

5.4

0.5 11.6

0.5

0.7

0.8

0.6

2.0

0.9

8.0 4.5 5.5

3.5

The total radiocaesium content in soil (kBq/mZ)and in fruit bodies from the two areas (As and Samnanger) that had a low content of Chernobyl fallout, divided into "Chernobyl radiocaesium" and "nuclear weapons testing radiocaesium", and their percent distribution in soil and fungi are shown in Table 4.2.7. The percent distributions of radiocaesium orginating from Chernobyl and nuclear weapons testings fallout in soil and fungal fruit bodies in both these areas are very similar with a distribution of close to 70%:30% for both types of fallout material. The same pattern concerning the transfer coefficient from soil to fungi is shown in Table 4.2.6 for different fungal genera in "high-Chernobyl"and "low-Chernobyl" areas.

283

TABLE 4.2.7. PERCENT DISTRIBUTION OF THE TOTAL RADIOCAESIUM CONTENT IN SOIL AND FUNGAL FRUIT BODIES ORIGINATING FROM CHERNOBYL AND NUCLEAR WEAPONS TESTING FALLOUT. Samnanger

AS

Soil Percent "Chernobyl Cs" Percent "nuclear weapons testing Cs"

Fungi

Soil

Fungi

%

%

%

%

73 27

74 26

65

58 42

35

The results indicate that the bioavailability of "Chernobyl" and weapons testing radiocaesium fallout is very similar. Thus the bioavailability of radiocaesium in upland and forest areas, having acid nutrient-poor soil, seems to be very little affected by time. Physical decay and a vertical transport of variable velocity are the main mechanisms causing a reduction in the fungal fruit body content. A similar conclusion has been reached by Selnaes and Strand (1992) for the transfer of radiocaesium from soil to plant and to grazing animals and dairy products. Hove and Strand (1 990) calculated the transfer coefficient of nuclear weapons testing radiocaesium from soil to grass and lamb meat, and concluded that the ecological half-life for this system was

about 20 years. At present it is impossible to perform a similar calculation for the soil to fungi system, because most genera in 1988-1991 contained more radiocaesium than could be expected from the soil content at the sampling sites (see Table 4.2.3). Although our prediction model needs refinements, it seems safe to conclude that the future fruit body content will be high with an ecological half-life close to the physical half-life of about 30 years. High radiocaesium contents in different fruit bodies have been observed in all countries which received Chernobyl fallout material, (Haselwandter et al. 1988, Battiston et al. 1989, Dighton and Horrill 1988). Several authors have also reported about the high ability of different fungi to accumulate radiocaesium from the fallout from nuclear weapons testings (Seeger and Schweinshdut 1981, Bum1 et al. 1989, Haselwandter et al. 1988). Other Nordic countries have had less interest in the transfer of radiocaesium from soil to fungi than has Norway. A summary of the radiocaesium content found in common fungal genera in Denmark, Finland and Sweden, together with the average amount of radiocaesium per m2, is given in Table 4.2.8.

284 TABLE 4.2.8. THE RADIOCAESIUM CONTENT (kBq/kg d.w.) IN DIFFERENT COMMON FUNGAL GENERA IN DENMARK, SWEDEN AND FINLAND. Denmark 1991 kBq1kg Genus 1.8 Amanita Cortinarius 6.6 Lactarius 1.4 Leccinum 0.4 Rozi tes 13.3 Russula 0.6 Suillus 1.2 Calculated per kg fresh weight

Sweden 1992 T,

1.3 3.3 1.3 0.3 9.3 0.6 0.8

kBq/kg 0.1 10.6 1.3

Finland' 1986 T* 0.1 2.1

kBqlkg

0.4

1.2 1.1

1.2 0.5

0.2 0.2

0.8 2.7

The radiocaesium content in the different fungal genera sampled and measured in Finland in the autumn of 1986 is difficult to compare with the results from the other Nordic countries obtained

2 to 4 years later. The low content in genera that showed a much higher content in other countries can perhaps be explained by the fact that only a few months after the fallout episode, the radiocaesium had not been washed down into the profile that harbours the fungal absorbing structures. The amount of radicaesium deposited in Denmark was very low and comparable with that in the "low Chernobyl fallout area" that was investigated in Norway. The content of radiocaesium in comparable genera is also of the same order of magnitude. The data from Sweden relate to a pine forest ecosystem in the Uppsala region, with an average amount of radiocaesium of about 11 kBq/m*. As found in other countries, members of the Cortinarius genus contained the highest amount of accumulated radiocaesium. In recent years the importance of the mycorrhizal fungi as radiocaesium accumulators has

been well documented and it is obvious that they play a significant role in the recycling of radiocaesium in different seminatural ecosystems. The biological half-life of this radionuclide is high in fungal h i t bodies, and probably near the physical half-life. A reliable prediction of fungal radiocaesium contents in the future is thus desirable because high contents are most probable,

CONCLUSIONS To sum up what we have learned about the transfer of radiocaesium From soil to plants and fungi and finally to grazing animals, the following factors have been found to be of great importance.

285 Time of deposition. The radiocaesium fallout from the Chernobyl accident was deposited in both areas covered in snow and on ground vegetation in different stage of growth development. In those areas covered by snow a redistribution took place by lateral transport, which gave rise to sites with a considerable higher radiocaesium content than sites nearby. In snowfree areas the radiocaesium deposition is more even and and is correlated to the amount of precipitation during the short period after the accident. Vertical transport. Earlier studies of the vertical transport of radiocaesium from the weapons tests fallout in different soils, mainly agricultural soils, showed a downward transport with time and an absorption to clay minerals, which reduced the amount of radiocaesium available to plants. In the mostly acid seminatural and forest soils in the Nordic countries practically no vertical transport occured, and more than 90 % is still bound in the top 3-4 cm organic layer with a very low content of clay minerals. Radiocaesium content in fungi and plants. The radiocaesium content in many fungal species is high and the patchy distribution of the radionuclides in some seminatural ecosystems have resulted in large variation in the radiocaesium content in fruit bodies of the same species situated in sites close to each other. Plants growing at the same sites contained less than 2 % of the fungal radiocaesium content. A minor reduction of the plant content have occured in the period after 1988, while in fungi no significant reduction has taken place and for some genera an increaced content was observed. Measurements of the amount of radiocaesium in fungi from weapons tests fallout indicate that the future content of radiocaesium from the Chernobyl fallout will be high. An implication of this prediction is that the transfer of radiocaesium from soil to products from animals grazing in seminatural and forest ecosystems, in at least some of the Nordic countries, will depend of the yearly production of fungal fruit bodies.

ACKNOWLEDGMENT The author wishes to thank Signe Dahl who was responsible for the data handling, graphical presentation and layout work for this manuscript. REFERENCES Backe S, Bjerke H, Rudjord A L and Ugletveit F. 1987. Fall-out pattern in Norway after the Chernobyl accident estimated from soil samples. Rudiut. h o t . Dosim. 18, 105-107. Bakken L R and Olsen R A. 1990. Accumulation of Radiocaesium in fungi. Can. J. Microbiol. 36,704-710. Battison G A, Degetto S, Gerbasi R and Sbrignadello. 1989. Radioactivity in mushrooms in the north of Italy following the Chernobyl accident. J. Environ. Rudioucf. 9 , 53-60.

286 Bergman R, NylCn T, Palo T and Lidstr~mK. 1991. The behaviour of radioactive caesium in a boreal forest ecosystem. In The Chernobyl fallout in Sweden. Results from a research programme on environmental radiology (ed. Moberg L.), pp 425-456. Bum1 K, Schimmack W, Kreutzer K and Schieri R. 1989. The migration of fallout lS4Cs, IJ7Csand lo6Ru from Chernobyl and IJ7Csfrom weapons testing in a forest soil. Z. Pflnzenernaehr. Bodenkd. 152, 39-44. Byrne A R. 1988. Radioactivity in fungi in Slovenia, Yugoslavia following the Chernobyl accident. J. Environ. Radioact. 6 , 177-183. Dighton J and Homll A D. 1988. Radiocaesium accumulation in the mycorrhizal fungi Lactarius rufus and Inocybe longicystis, in upland Britain following the Chernobyl accident. Trans. Br. Mycol. SOC.91, 335-357. Haselwandter K, Berreck M and Brunner P. 1988. Fungi as bioindicators of radiocaesium contamination: pre- and post-Chernobyl activities. Trans Br. Mycol. SOC. 90, 171-174. Haugen L E and Bjmstad H E. 1990. Transport of radiocaesium during snowmelting on a mountain pasture in Norway, spring 1989. In IUR Working group on Soil Plant transfer. Workshop on the Contamination of Crops because of Soil Adhesion. Uppsala Sweden, September 27-28, 1990. pp 139-142. Hove K. and Strand P. 1990. Prediction for the duration of the Chernobyl radiocaesium problems in non-cultivated areas based on reassessment of the hehaviour of fallout from the nuclear weapon tests. In Environmental Contamination Following a Major Nuclear Accident, Proceedings of an International Atomic Energy Agency Conference. (eds. Flitton S and Watz E.W.) Vienna 1990. 1, 215-223. McGee E J, Colgan P A, Dawson D E, Rafferty B and O'Keeffe C. 1992. Effects of Topography on Caesium-137 in Montane Peat Soils and Vegetation. Analyst 1 17, 461-464. Nieumann R, Steinnes E and Varskog P. 1990. Mobilitet og plantetilgjengelighet av radioaktivt cesium i naturlig jord. Informasjon fra Statens fagtJeneste for landbruket. 1990 1, 61-65 (In Norwegian). Osburn W S. 1963. The dynamics of fallout distribution in a Colorado alpine tundra snow accumulation ecosystem. In Radioecology. Proceedings of the First National Symposium on Radioecology. (eds. Schultz V and Klemet A W) Fort Collins 1961. pp 50-71. Riise G, Bjmstad H E, Lien H N, Oughton D H and Salbu B. 1990. A study on radionuclide association with soil components using a sequential extraction procedure. Journal of Radioanalytical and Nuclear Chemistry. 142 (2), 531-538. Roos N, Jensen M N and Jensen N L. 1990. Radioaktivt cesium i svampe. Rapport Roskilde Universitets Center RUC 90. (In Danish). Seeger R and Schweinshaut P. 1981. Vorkommen von Caesium in hBheren Pilze. Sci. Toral Environ. 19, 253-276. Selnas T D and Strand P. 1992. Comparison of the uptake of radiocaesium from soil to grass after nuclear weapons test and the Chernobyl accident. Analyst. 117, 493-496. Staaland H, Garmo T H, Pedersen 0 and Hove K. 1990. Endring i innhaldet av radiocesium i plantemateriale og beitedyr pi fjellbeite 1986-90 - fmebels resultat. Informasjon fra Statens fagtjeneste for landbruket 1990. 28, 58-63. (In Norwegian) Strandberg M. 1992. Investigation of radiocaesium in fungi in a Danish Scotch pine forest ecosystem. Det sjette Nordiske Radionrkologi Seminar 1992. Thorshavn Faeroyar.

287

4.3. RADIOCAESWM IN GAME ANIMALS IN TEE NORDIC COUNTRIES

KARL J.JOEtANSON Department of Radioecology, Swedish University of Agricultural Sciences P.O.Box 703 1, S-750 07 Uppsala, Sweden SUMMARY Forest ecosystems in the Nordic countries have been shown to be very important for the overall transfer of Chernobyl radiocaesium to man. One of the major pathways of radiocaesium to man is through game animals, mainly moose and roe deer. As most of the game meat is consumed by the hunters - about 300 000 in Sweden - and their families, these constitute the critical group. Aggregated transfer factors of between 0.01 and 0.03 m2 kg-1 were found for moose and 0.03 to 0.14 for roe deer. The considerally variation for roe deer is a result of the large seasonal variation in 137Cs activity concentrations in these animals with a peak value in August and September when the fruitbodies of fungi having high activity concentrations of 137Cs, are abundant. In Sweden where annually between 100 000 and 135 000 moose are harvested the calculated annual transfer of 137Cs by moose to man alter the Chernobyl accident has been between 2.4 and 2.7 GBq, corresponding to an annual collective dose of 30 and 33 manSv. The corresponding transfer by roe deer lay been between 560 and 1 400 MBq, corresponding to an annual collective dose of 7 to 17 manSv. During the post-Chernobyl years there has been very little or no decrease in 137Cs activity concentrations in moose and roe deer. The effective ecological half-life of radiocaesium in the forest ecosystems in the Nordic countries seems therefore to be very long and we suggest that the physical half-life is the best estimate. INTRODUCTION In the forest ecosystems of the Nordic countries an extremely efficient transfer of radiocaesium from soil to plants has been found. One of the most important soil parameters for this considerable transfer seems to be the high organic content of the upper 5-10 cm soil layer with high humic content, which is typical of many Nordic forests (Fawaris and Johanson 1994). Additionally in this layer there is often a low content of exchangeable potassium as well as a low pH. The forest ecosystems of the Nordic countries have been shown to be very important for the committed transfer of radiocaesium to man (Johanson and Bergstrom 1994, Bergman and Johansson 1989). The main food products originating in the forest ecosystems are game animals, berries and fungi. A. possible critical group is families of hunters, who often also gather berries and fungi

SOILS IN FORESTS IN NORDIC COUNTRIES In most studies of the vertical distribution of radiocaesium in the forests in Nordic countries 90 - 95 % of the total deposited 137Cs has been found in the upper 5 cm of the soil (Fawaris and Johanson 1994). Very small amounts of 137Cs activity have been found below 10 cm depth. In this upper 10

288

cm layer, 80 - 90 % of the soil often consists of organic matter. The normal pH in the upper part of the soil in typical Nordic coniferous forests is around 4 with some variations from 3.5 to 5.0. Compared to arable land in the same area, the pH in forests is low particularly in coniferous forests. The amount of exchangeable potassium is often low, which may be due to a fast transfer by the litter decomposing fungi to the mycorrhizial fungi. There are reports suggesting that a major fraction of the 137Cs activity found in the soil is in fact contained within the mycelium of the fungi (Olsen et al 1990). The major amounts of nutrients in forest soil are found in the upper organic layer as well as in the roots and the mycelium of fungi. As already mentioned, radiocaesium is also located in this zone. Bogs or mires are common in Nordic forests and they are often very poor in nutrients particularly the raised bogs that are common in the south and central areas of the Nordic countries. Mires that are more common in the north of the region may be nutrient-poor or nutrient-rich. For the transfer of radiocaesium to man by game animals from a certain area, nutrient-poor bogs within the area may be more important than differences in ground deposition due to the increased transfer from soil to plants on the bogs. Even areas within forests having rocky outcrops may increase the transfer of radiocaesium from soil to plants and therefore also to man by game animals. The predominant trees in Nordic forests are Scots pine (Pinus sylvestris) and Norwegian spruce (Picea abies) with some intermixture of deciduous trees. Of the last the most common are birch (Betula sp), mountain ash (Sorbus aucuparia), aspen (Populus tremula) and various species of willow (Salix spp). In the field layer various dwarf shrubs are found such as bilbeny (Vaccinium myrtillus), lingonberry (Vaccinium vitis-idaea) and heather (Calluna vulgaris). The ground is usually covered by a layer of mosses or lichens. Some herbs and grasses can be found depending on the nutrient status of the forest soil. On nutrient-rich areas herbs and grasses may be dominating in the field layer. Wavy hair grass (Deschampsia flexuosa) may be very common and another important grass is sheep's fescue (Festuca ovina). In clearings wavy hair grass and fireweed (Epilobium angustifolium) can be very common as well as raspberry (Rubus idaeus). In most Nordic forests there are abundant fruit bodies of fungi during autumn. A large fraction of the biomass in the soil consists of the mycelium of various species of fungi. They often live in symbiosis with various vascular plants - the most common are the mycorrhiza between fungi and various trees, for example Scots pine and Norwegian spruce. The biomass of fruiting bodies of fungi vanes very much from year to year. In recent years, 1988, 1990 and 1992 were good mushroom years at least in central Sweden, whereas 1987, 1989 and 1991 were rather poor years for mushrooms. GAME Mammals The most important game animal producing the largest quantity of game meat in the Nordic countries is the moose (Alces alces). The mean carcass weight of moose harvested in Sweden is 130 kg (Malmgren et al. 1976) and as about 120000 moose are harvested annually in Sweden (Bergstrom et al. 1992), the total amount of meat is about 13 million kg, assuming that 80 % of the

289

carcass is pure meat. During the last 30 years harvest numbers have increased from 32 000 in 1960 to 130 000 in 1990. In Norway 28 800 moose were harvested in 1990, showing an increase similar to that in Sweden from about 7 500 in 1960 (Hunting statistics 1989). In Finland 58 000 moose were harvested (Game and Fisheries Research Institute Finland 1991). Denmark and Iceland have no wild moose population. The normal hunting season for moose lasts from September to December, with some variations from region to region and also between countries. Sweden has about 300 000 moose hunters and assuming that each represents a family of 3 people, the critical group - that consuming the major part of the moose meat - would number 900 000 people. The combination of the large size of the moose and its comparatively high 137Cs activity concentration, means that the moose pathway is important for the transfer of radiocaesium from the forest ecosystem to man. Table 4.3.1 The annual harvests of some important game animals in the Nordic countries. (Bergstrdm et al. 1992, Central bureau of statistics of Norway 1989, Danmarks miljoundersogelser, Afdeling for Flora og Faunaokologi 1992, Game and Fisheries Research Institute, Finland 1991, Svensk jakt 1993). Species

Denmark

Finland

Moose Roe deer Red deer Fallow deer White tail deer Arctic hare Brown hare Wild rabbit Capercaillie Black grouse Hazel grouse Willow grouse Pheasant Canada goose Woodcock Wood pigeon Mallard

58 000 78 000 2 100 3 800

26 000 10 000

142 000 16 000

772 OOO* 16 000 24 000 350 000 673 OOO*

6 700 213 000 153 000 17 000 38 000 146 000 68 000 60 000 9 000 5 000 3 000 75 000 244 000

Norway

Sweden

128 000 54 000 9 700

125 000 28 000 26 000 63 000 8 000 697 000

82 000 70 000

312 000 700 3 500 160 000 101 000 72 000 36 000 48 000 25 000 40 000 19 000 17 000 35 000 87 000 81 000

*650 000 pheasants and 500 000 mallards are harvested by "put and take" hunting

Roe deer (Capreolus capreolus) can be found in most of the Nordic countries. In Sweden it is one of the most common game animals and some 300 000 were harvested last year. During the last 5 years there has been a sharp increase in the roe deer population size in Sweden because of the very mild winters. In 1986 150 000 roe deer were harvested so over 5 years the population has

290 doubled. In Norway the annual harvest has also increased up to 54000 in 1989. In Denmark harvests have vaned from 54 000 to 73 000 from 1986 to 1990. In Finland roe deer are only found on h a n d and in a small area near the Swedish border, east of the river Tome. The total harvest is only about 10 000 roe deer. There are no roe deer on Iceland. The carcass weight of a roe deer vanes from about 6 to 20 kg. The 137Cs activity concentration in roe deer is usually higher than in moose living in the same area particularly in the autumn. Red deer (Cervus elaphus) are quite common in western Norway where the annual harvests have numbered about 10 000 animals. They are very rare in Sweden and Finland and only small amounts of radiocaesium can be transferred to man by red deer. In Denmark the annual harvest has been less than 2 000 red deer. No red deer are found in Finland and Iceland. The carcass weight of a red deer is about 100 kg and therefore there may be very small populations of hunters having a high intake of radiocaesium as a result of transfer by this animal. Red deer are also kept as domesticated animals in fenced areas. Small populations of fallow deer (Dama dama) can be found in Sweden, Denmark and Norway. In Denmark the annual harvests numbered 2 300 to 3 700 animals from 1986 to 1989. Harvests in Sweden were on about the same level. These animals seem to be of minor importance for the transfer of radiocaesium to man. Norway and Finland have a small population of wild reindeer and the annual harvests in Norway have numbered around 8 000 animals. The radioecology of the reindeer is discussed in another chapter. Finland has quite a large population of white-tailed deer. This North American deer was introduced into some European countries and seems to have adapted particularly well in Finland. The annual harvests have numbered around 20 000 deer. In size the white-tailed deer lies between the roe deer and the red deer with a mean carcass weight of about 40 kg. Two species of hare are found in the Nordic countries - mountain hare (Lepus timidus) and brown hare (Lepus europeus). The mountain hare is generally associated with forested areas and the brown hare with open fields, but they can be found together in forest areas near fields and they can also mate with one another. The number of hares harvested in Sweden totalled about 160 000 arctic hare and 100 000 brown hare. In Denmark the annual harvests (brown hare) totalled 148 000 to 190 000 animals (1986-1990). The corresponding numbers in Norway for arctic hare have been 116 000 to 125 000. The normal weight of a hare is about 3 kg. The radioecology of the two species is quite different. The brown hare usually has rather low 137Cs activity concentrations. In contrast the mountain hare has rather high 137Cs activity concentrations, particularly during winters without snow. Wild rabbit can be found in some parts of Sweden and Denmark. They usually live on arable land and have quite low 137Cs activity concentrations in meat.

29 1 Birds One of the largest game birds in the Nordic countries is the capercaillie (Tetra0 tetrao) with a mean weight of about 3 kg. Annual harvests in Norway numbered from 20 000 to 26 000 birds. The black grouse (Lyrusus tetrix) is more common but smaller than the capercaillie and between 45 000 and 63 000 birds are harvested annually in Norway. The mean weight of a black grouse is slightly more than 1 kg. Willow grouse (Lagopus lagopus) is quite common in the mountain regions of the Nordic countries and particularly in Norway where the annual harvests number between 500 000 and 750 000 birds (probably both willow grouse and ptarmigan). They are rather small with a mean weight of 0.5kg. Ptarmigan (Ptarmigan mutus) too is quite common in the mountain regions and annual harvests in Sweden number between 900 and 5 100 birds. In size they are similar to the willow grouse. Wood pigeon (Columbia palumbus) is found in regions with arable land and forests. Annual harvests of wood pigeon are quite large - in Sweden, Denmark and Norway, some 75 000 to 90 000 birds, but they are rather small, 0.5 kg in weight. Of water fowl the most important is the mallard ( h a s platyrhynchos). The annual harvests of mallard in Norway number between 52 000 and 70 000 birds. The mallard weights about 1 kg. Other ducks give rather little transfer of radiocaesium to man. The Canada goose (Branta canadensis) is much larger than the ducks, weighing up to 5 kg. The annual harvests in Sweden number about 17 000 birds, but they mainly come from outside the most radiocaesium-contaminated regions. The woodcock (Scolopax rusticola) is quite a small bird (about 0.5 kg in weight) providing annual harvests of about 35 000 in Sweden. In 1986 rather high radiocaesium levels were observed in woodcock but in the following years levels were much lower. RADIOECOLOGY OF THE MOOSE Diet of the moose The diet of the moose has been studied by botanical analysis of rumen content, which shows that diet changes with the season (Cederlund et al. 1980). In winter, at least when the ground is covered by snow, its main food is twigs from Scots pine and deciduous trees, mainly birch. When the snow melts the diet changes to include more dwarf shrubs, mainly bilberry and heather. During late spring and summer, herbs, grasses and the shoots and leaves of deciduous trees, mainly birch, become most important. In summer, birch and fireweed may constitute more than 70 % of all feed intake (Johanson et al. 1994). In autumn, moose eat more dwarf shrubs, first b i l b e q in September and then heather in October (Cederlund et al. 1980). Often the botanical composition of a moose rumen is dominated by 2 or 3 species (Johanson et al. 1994). The 137Cs activity concentrations in moose muscle during the hunting period depend mainly on food intake from July to October. In a study of the diet of moose in Vasternorrland, Johanson et al. (1994) found that fireweed was the most important fodder plant during July and August making

292 up 55 % and 34 %, respectively, of the rumen content. In August and September, birch was also very important, comprising 31 % and 48 % of the rumen content. In October bilbeny 38 YO, mountain ash 22 % and birch 21 % were the most important fodder plants found in the rumen. During the period in question these three species were the dominating plants. In a study from central Sweden (Cederlund et al. 1980) leaves and shoots of trees and shrubs together with dwarf shrubs constitute about 70 to 80 % of the rumen content during the period leading up to the hunting season - August to October. Table 4.3.2 The most common plant species found in the rumen from July to October in moose from the county of Vasternorrland (Johanson et al.1994). Species

Betula sp. Vaccinium myrtillus Epilobium angustifolium

July

Percentage of rumen content August September October

17 1 55

31 12 34

48 14 8

21 38 0

Table 4.3.3 *37Cs activity concentrations in the most common fodder plants of the moose. The lant samples were collected in 1990 to 1992 from the Harbo area where ground deposition of p37Cs was around 35 000 to 40 000 Bq m- . (Fawaris and Johanson 1994, Kleiven et al. 1993, von Bothmer et al. 1990). Species

Plant parts

Scots pine Birch Heather Bilberry Lingonberry Fireweed Wavy hair grass Mixed fungi

current annual shoot current annual shoot green parts green parts green parts Upper 30 cm Upper 30 cm Fruit bodies

Bq kg-1 d.w. 2 300 2 500 10 300 3 000 4 000 400 2 400 50 000

TF (Bq kg-l Bq m-2) 0.07 0.08 0.32 0.09 0.12 0.01

0.07

1.4

Table 4.3.3 shows the transfer parameters for the most common plant species found in the rumen of moose and roe deer. Of the most important fodder plants, fireweed has rather low 137Cs levels with a transfer factor of about 0.01 m2 kg-1 (all plant and hngi weights are expressed as dry weight). Even birch and Scots pine, with transfer factors of 0.08 and 0.07, have comparable low levels of 137Cs. Bilberry has a higher transfer factor of 0.09 m2 kg-1 and heather has even higher transfer factors - 0.32 m2 kg-1. In the Vasternorrland study heather was found only in small quantities in moose rumen but in the study from central Sweden heather was quite important. Fruit bodies of fungi usually have much higher 137Cs levels than vascular plants growing on the same

293

site. Fungi, however, have been found only in small quantities in the moose rumen. In our Vasternorrland study 1 to 2 % was found in moose rumens (Johanson et al. 1994). Fungi pose a problem of method: how long is it possible to observe them in the rumen? In the reindeer rumen only small quantities of mushroom have been found but reindeer herders seem to agree that reindeer sometimes eat quite large quantities of fungi. The mean transfer factor for fungi is much higher than for any vascular plant - for all fungal species found in a study area north-west of Uppsala the transfer factor was 1.4m2 kg-l and compared to other forage plants for moose, hngi have a 5 times higher mean 137Csactivity concentration than heather, and nearly 20 times that ofbilbeny. Moose integrate the 137Cs content in fodder plants on the home range. In central Sweden a normal home range for moose in autumn is an area having a radius of 2.5 km. Inside this area there are many different stands and the moose, of course, choose the best parts which may differ from year to year and also from season to season (Palo 1990). In our study area in Harbo 40 km northwest of Uppsala there has been no decrease of the mean 137Cs activity concentrations in moose during the period from 1986 to 1993.In an area where deposition is around 35 000 Bq m2 the mean 137Cs activity concentration 1986 to 1991 was 750 for the 200 to 275 moose harvested annually. The daily feed intake is around 7.5kg dw during the growing season - June to August decreasing to 6 kg in September and in April and May. From October to March a daily intake of 5 kg has been reported (von Bothmer et al. 1990). The transfer parameters for the most common species of plant in the diet of the moose can be seen in Table 4.3.3. The summer diet consists of species showing rather low 137Cslevels, such as fireweed and birch. When the moose shift to more bilberry, there is an increase in the daily intake of l37Cs, which is further pronounced when the moose increases the intake of heather. During September, however, moose start to decrease their total intake of food, which partly compensates for the higher levels of 137Cs in some of the fodder plants. In late autumn (November and December) the decrease in food intake is still more pronounced and therefore the intake of 137Cs usually decreases. Many hunters have observed this decrease in 137Cs levels from October to December and some hunting teams in Gavle commune are now hunt moose very late in the year. In 1988 there was a drastic increase in the radiocaesium levels in moose harvested in September. This indicates that some species with very high radiocaesium levels may have occured in the diet in early autumn that year. It is very likely that the moose were eating fungi. There would be no need to eat large quantities - even 1 to 2 YOof fungi in the rumen content will increase the intake of radiocaesium by 50 YO. The winter diet - twigs of pine and birch - will not affect the 137Cs levels in moose meat during the hunting period. Thisfeed has medium levels of 1370 or a transfer factor of 0.06. Based on present knowledge of the diet of the moose and the levels of 137Cs in various species of its fodder plants, it is possible to estimate the daily intake of 137Cs for moose. Such calculations have been made for moose living in the county of Vastmanland in central Sweden and in the county of Vasternorrland hrther north. In summer,the daily intake of 137Csby moose in the Harbo area was estimated to be slightly more than 10 000 Bq. In autumn this increases to about

294 20 000 Bq with a maximum in October followed by a decrease during November and December. In December the levels was about 12 000 Bq (von Bothmer et al. 1990). The corresponding transfer coefficient (Ff ,Bq kg-1 /Bq d-1) was estimated to be 0.03 d kg-1. A similar estimation for moose from Vasternorrland gave a Ff of 0.07. It has to be realised that estimating the daily intake of the moose is rather complicated, and these calculations of transfer parameters should not be considered more than a rough estimate. There is also a large seasonal variation in the transfer parameters and the above values are calculated for the hunting period in October. Another possible way of calculating a transfer parameter involves use of the aggregated transfer factor (Tag, Bq kg-1 in moose meat/ Bq m-2 in the home range of the moose). In the Harbo area the mean value for this parameter from 1986 to 1992 is 0.023 m2 Bq-l (Johanson and Bergstrdm 1994). Similar calculations for Vasternorrland gave a Tag of 0.01 to 0.03 with a mean value of 0.02 (Johanson et al. 1994). The aggregated transfer factor can be used when it is difficult to estimate other transfer parameters due to lack of knowledge. It is, for example, possible to calculate the 137Cs activity concentration in a theoretically mean moose in Sweden where the mean 137Cs ground deposition is 10 000 Bq m-2, As the aggregated transfer factor is 0.02, the mean 137Cs levels would be 200 Bq kg-1 (Johanson and Bergstrom 1994). Similar calculations were made by Rantavaara et al. (1987). For 1986 and 1987 they found an aggregated transfer factor of 0.0012 to 0.030 m2 kg-1 (95 % confidence interval) with mean values of 0.0097 for adults and 0.015 for calves. The mean 137C.s activity concentration in moose meat in Finland was calculated to be 170 Bq kg-1. The 1 3 7 0 activity concentrations in moose calves have usually been higher than in adults harvested in the same area. In the north of Sweden where the hunting season starts in September, the calves have levels about 40 % higher (Bergman et al. 1990). In central Sweden where the moose hunting season starts in October levels are only about 10 % higher in calves (Johanson and Bergstrom 1989). Thus it seems to be the intake during August and early September which gives the differences. The most probable reason for the differences between calves and adults is some metabolic mechanism connected to the growth of the calves,which reaches a plateau at the beginning of October and then levels off until the following spring. Transfer to man The annual transfer of 137Cs from moose to man can be calculated on the basis of the previously determined aggregated transfer factor. In Sweden the mean 137Cs ground deposition was 10 000 Bq rnm2and the Tag was determined to be 0.02. This means that the mean moose in Sweden contained 200 Bq kg-1 in 1986. The moose harvested numbered between 113 000 and 135 000 during the period 1986 to 1992 (Table 4.3.1) and the total meat harvested annually was between 12.2 and 14.8 million kg. The total 137Cs activity transferred to man by moose has been between 2.4 and 2.7 GBq (Johanson and Bergstrom 1994). This corresponds to a annual collective dose of between 30 and 33 manSv, assuming a dose conversion factor according to ICW (1991). The transfer to the whole Finnish population through game meat was calculated to be 110 and 180 Bq per person during 1986 and 1987 respectively. Cervides contribute about 77 % of the

295

dietary intake via game meat. For the critical group - the hunter’s households (500 000 people) - the corresponding values were 1 000 and 1 800 Bq for 1986 and 1987. Ecological half-life The committed dose to man resulting from the transfer of Chernobyl 137Cs to man by moose depends on the transfer parameters, but also on the effective ecological half-lives of radiocaesium in the forest ecosystem. After the Chernobyl accident much interests has been shown in studies of the forest ecosystem and when knowledge increases new problems appear. One is the considerable transfer from soil to plants and another is the very long half-lives of radiocaesium in the system.

Table 4.3.4 The annual mean of 137Cs activity concentrations (Bq kg-1 fresh weight) in moose from various parts of Sweden over the years 1986 to 1992. The values from Heby and Nordanstig are the annual mean values for more than 100 moose, the values from Gavle are the mean values from 5 hunting teams (10 to 20 moose). (Johanson and Bergstrom 1993, Information From Nordanstig and Gavle communes). Area Heby Nordanstig Gavle Hytton Flaten Hillesddra Korsnas Hagsta

1986

1987

1988

1989

1990

1991

780 490

658 630

850 520

739 525

708 525

760 520

2070 2031 1656 1812

1735 1946 I759 1 123

1657 1661 2 115 1499

1398 1783 1250 1 112

1541 1036 1276 1291

1353 1501 1 127 1 093

1556

1915

2421

1536

1743

1389

In a study area in Heby commune in central Sweden there has been no decrease of the annual mean 137Cs activity concentrations in the harvested moose (Table 4.3.4). Each year about 200 to 250 samples of harvested moose have been analysed. As can be seen in Table 4.3.4 Heby and Nordanstig show rather small variations from year to year but no decreasing trend. The results from the individual hunting teams from Gavle indicate a decrease in levels. As the levels are high, changes in hunting management or the influence of other countermeasures cannot be excluded. We therefore suggest a half-life of 137Cs in the forest ecosystem similar to the physical half live of 137Cs. Bergman et al. (1990) suggested an ecological half-life of around 25 years when assessed by means of the 137Cs activity concentrations in moose. When applying the Tag value of 0.02 for moose and the cumulative ground deposition corrected for decay for the nuclear weapon 137Cs, we found that the predicted 137Cs activity concentration in 1985 would be 27 Bq per kg. The level found in central Sweden was 23 Bq per kg, and in northern Sweden 33 Gq per kg with a mean value of 28 (Johanson and Bergstrom 1994, Bergman et a1 1988). Rantavaara 1981) reported a value of 34 from 1977 whereas the predicted level was 30. As can be seen the predicted values agree very well with

296

those found - at least there will be no underestimation of the levels. Until there is more information on the development of 137Cs activity concentrations in moose, we suggest that the physical half-live of 137Cs should be used as effective ecological half-lives for 137Cs in moose in the Nordic countries. Possible countermeasures In the most heavily contaminated area in Sweden - Giivle commune - 35 % of the moose harvested in 1992 had 137Cs activity concentrations above 1 500 Bq per kg, the existing intervention level in Sweden. At such high levels it was obvious that countermeasures, even in the forest ecosystem, had to be discussed. We tried some possible methods. Theoretically the most promising method seems to be the use of salt lick containing Giese salt. This was applied in at least 3 areas in Sweden - in "our" Heby area, and in Gavle and Kramfors commune. For some hunting-team areas, the results are promising, in others less so. It may be possible to reduce the activity concentrations in moose by 20 - 25 YO- for some hunting areas it will be more, whereas for others there will be no reduction. There seems a lack of knowledge on the placing of the salt licks in the foresrt so that the moose use them optimally. In Gavle commune some hunting teams changed their hunting period from October to late December. Usually this gave a clear reduction of the 137Cs activity concentrations but it cannot be excluded that some years they will get higher values. More information is needed before a general recommendation can be given. A problem with moose hunting in December is of course the climate - in some parts of Sweden moose hunting is less attractive late in the year when it is cold, dark and there could be snow. RADIOECOLOGY OF ROE DEER Roe deer are adapted to nearly all ecosystems, at least in Sweden, and have during the last century spread practically over the whole country. Some live mostly on open fields, some spend all their lives in the forest. This of course influences the 137Cs activity concentration in these animals - the forest deer obtain much higher levels.

Feeding habit There are similarities and differences in the feeding habits of moose and roe deer. The roe deer seem to feed more actively than the moose and also have a more variable diet. However, even in the roe deer rumen rather a few species are predominant and many of the species ingested by the moose can also be found in the rumen of the roe deer (Bergstrom et al. 1992, Cederlund et al. 1980, Petersen and Strandgaard 1992). During summer herbs are predominant. In autumn the roe deer increase their intake of dwarfshrubs, mainly bilberry, lingonberry and heather, at least those living in forests. Even in winter these species are important during the snow-free periods that have been normal in recent years. In cold winters with much snow roe deer have difficulties in finding enough feed, and are often given supplementary fodder by hunters. At least in central and northern Sweden the roe deer population will be reduced during hard winters. The intake of conifertwigs is quite common in

291

January and February and in spring the intake of dwarf shrubs increases. In Denmark and probably in southern Sweden too roe deer mainly live in or near open fields, and they eat much grass and other fodder plants grown on arable land, In a Danish investigation large amounts of white wind flower (Anemone nemorosa) were found in the rumen of these animals and this seems to constitute up to half the total intake during winter and spring. When fruit bodies of fungi are available mainly during August, September and October, roe deer have a large intake of fungi. In the season the mean percentage of fungi in the rumen may be about 20 % and in individual roe deer as high as 80 % (Cederlund et al. 1980, Bergstrom et al. 1992). As most species of fungi have a very high 13'Cs activity concentration, intake gives a drastic increase in the 137Cs levels in roe deer when fruit bodies of hngi are abundant. Seasonal variations The 137Cs activity concentrations in roe deer show a pronounced seasonal variation. In the Harbo area of central Sweden a peak in autumn occurred annually after the Chernobyl accident. In 1988, 1990 and 1992 the peak appeared early in August, in 1989 and 1991 it appeared later in September. The explanation for the peak is the appearance of the fruit bodies of the fungi. The mean 137Cs activity concentrations in roe deer in August and September in our Heby research area (ground deposition about 35 000 Bq m2) were about 5 000 Bq per kg during the peak period. The highest mean level was found in August and September 1988 when 8 000 Bq kg-1 was found. When the fruit bodies of the hngi decrease, a corresponding decrease occurs in the roe deer. In December and January the mean levels have been about 1 500 Bq kg-1. During Spring the levels are still lower, or about I 000 Bq kg-1. Transfer parameters By using the data of Cederlund et al. (1980) on the rumen content of roe deer and the 137Cs activity concentrations in the major fodder plants and in muscle samples from roe deer, it was possible to estimate the daily intake of 137Cs by roe deer in the Harbo study area (Karlen et al. 1991). The daily intake of 137Cs, except that from fungi, varies over the year from about 1 000 to 3 000 Bq. During August, September and October an additional 4 000 to 5 000 Bq originaties from fruit bodies of fungi. The total intake during the peak period is between 6 000 and 7 000 Bq and gives about 5 000 Bq per kg in muscle. A rough estimation of Ffwill give a value close to 1 d kg-1. Probably it is more relevant to use Tag, which of course varies depending on the season. In autumn there are Tag values of around 0.14, in December around 0.04 and in spring 0.03 m2 kg-1 with an annual mean of 0.05. In certain areas there were pronounced spatial variations. In parts of the forest rich in bogs, the 137Cs levels in roe deer may be 5 times higher than in "normal" forest, and roe deer grazing arable land usually have still lower levels.

298

Transfer to man The annual roe deer harvest in Sweden has varied from 1 16 000 to 3 12 000 animals from 1986 to 1990. The corresponding amount of meat totals between 1.1 and 3.1 million kg. The Tag for roe deer, at least in the Harbo area, is 0.05 so a mean roe deer in Sweden would have a 137Cs activity concentration of 500 Bq kg-l during 1986. The annual transfer of 137Cs to man by roe deer has been from 560 MBq to nearly 1 400 MBq, corresponding to an annual collective dose of 7 to 17 manSv (Johanson and Bergstrom 1994). One of the greatest uncertainty in estimating the dose commitment for man resulting from roe deer is the prognosis of the number of roe-deer that will be harvested in the hture. Possible countermeasure There are more ways of trying to reduce the 1 3 7 0 activity concentrations in roe deer than there are in the moose. In Sweden successful use has been made of one very simple countermeasure and that is to change the time of the hunting season. Why hunt roe buck when they have the highest levels of 137Cs? This was the question we raised and the hunters in the most heavily contaminated regions get an "extra" hunting period in May and the first half of June. This simple countermeasure has given 137Cs activity concentrations in the spring-harvested roe bucks that are 5 times lower than those of the bucks harvested in August-September. In normal winters roe deer are usually given supplementary fodder, often by hunters. This will reduce the intake of 137Cs, and it is also possible to add some caesium binders to the fodder thereby obtaining an even better reduction. The problem for the last 6 winters is that there has been no snow in central Sweden. If there is no snow or if the temperature is above -5 degrees, roe deer seem not to use the fodder and instead eat quite large amounts of heather, lingonberry and bilberry that contain rather high activity concentrations. Like moose, roe deer use the salt licks and theoretically at least it is possible to reduce the levels by providing salt licks containing Giese salt. One problem when trying to assess the efficiency of the various countermeasures for reduction of radiocaesium levels in roe deer is the large "natural" variation - both temporal and spatial. Therefore it may be difficult to prove statistically the effects of a certain countermeasure. OTHER MAMMALS There are very few data on the 137Cs activity concentrations in other deer from the Nordic countries. Data from Scotland for red deer give approximate Tag values of 0.02-0.04 m2 kg-1 (Howard priv comm. 1992). As the habits of red deer are rather similar to those of roe deer and the size lies between that of the moose and the roe deer values of 0.02 to 0.04 seem appropriate. Tag values for white-tailed deer were estimated by Rantavaara (priv. communication) and found to be 0.017 in 1986 and 0.03 in 1990. Arctic hare has much higher 137Cs levels than the brown hare has because of differences in habitat. Probably the Arctic hare also has a pronounced seasonal variation resulting from its intake of dwarf shrubs in autumn and winter. Rantavaara (1990 and priv. communication) reported a mean

299 Tag value of 0.038 with a range from 0.006 to 0.104, Johanson et al. (1990) reported a mean Tag value of 0.03 with a range from 0.025 to 0.13, Last year the normal winter levels for Arctic hare in Gavle lay between 5 000 and 20 000 Bq kg-1. This resulted in a very pronounced decrease in hare hunting in this commune. Hare is the most important of the non-cervids but its contribution to the intake of 137Cs by man is less than 10 % of the contribution from the whole game bag. If we assume that the Tag for the Arctic hare is 0.04 and 2 kg meat is obtained from each hare, the total Arctic hare meat consumed in Sweden is 320 000 kg. The mean 137Cs leves will be 400 Bq kg-1 giving 128 MBq per year corresponding to 1.5 manSv. GAME BIRDS There are few data on the 137Cs activity concentrations in birds that can be used for obtaining information on the transfer from ground to muscle. Grouse in Norway give the highest amount of meat, nearly 350 000 kg per year. Pedersen and Nybti (1990) investigated the radioecology of the grouse. A seasonal variation in levels was observed with the highest level during August and September. The trend over the years seems to differ between adult and young grouse. Adults showed a decrease in the 137Cs levels from 1987 to 1990 but the young grouse had the highest levels in 1989. In August and September the adults had lower levels than the young grouse. The mean level in the autumn was about 250 Bq kg-1 and ground deposition in the area was around 45 000 Bq m-2. An aggregated transfer factor of 0.006 m2 kg-1 can be estimated. If we make an unrealistic assumption that all grouse in Norway contained 250 Bq kg-1 (upper limit) the total 137Cs activity transferred by grouse to man was 90 MBq per year corresponding to about 1 manSv. In Finland ca 250 000 kg of meat is obtained from mallard and a Tag value of 0.01 m2 kg-1 has been suggested (Rantavaara pers.comm.). A mean value of 100 Bq kg-1 for 137Cs in mallard in Finland can be calculated, which means that 2.5 MBq may be transferred to man by mallard. Other game birds only contribute to a minor transfer of radiocaesium to man - at least if we define the critical population as the hunters and their families. In individual birds levels of 137Cs above the intervention levels can probably be found. CONCLUSION 1. The forest ecosystems in the Nordic countries have been shown to be very important for the total transfer of radiocaesium to man. One of the most important pathways is via game animals, mainly moose and roe deer. The critical human group is the hunters - totalling about 300 000 individuals in Sweden - and their families. 2. The predominant fodder plants ingested by moose for 2-3 months before the hunting season starts are fireweed, birch and bilbeny. Moose seem to consume only small amounts of the fruitbodies of fungi -1-2 % of rumen content - but even such small amounts may significantly increase the daily intake of 137Cs.

300 3. Aggregated transfer factors for 137Cs lie between 0.01 and 0.03 m2 kg-1 for moose. As the mean ground deposition in Sweden was 10 000 Bq m-2, a mean moose in Sweden should contain about 200 Bq kg-1. 4. The annual transfer of 137Cs to man by moose lies between 2.4 and 2.7 GBq corresponding to an annual collective dose of 30 to 33 manSv. 5 . Roe deer consume much larger quantities of fungi than moose. In autumn their choice of fodder plants is rather similar to that of moose. 6. Because of their large intake of fungi, roe deer show a pronounced seasonal variation in 137Cs activity concentration with a peak value in August to October. 7. The aggregated transfer factor for 137Cs lies between 0.03 and 0.14 for roe deer. The large variation for roe deer depends on a very pronounced seasonal variation with a peak value in August or September. 8. The annual transfer of 137Cs to man via roe deer in Sweden lies been between 560 and 1 400 MBq, corresponding to an annual collective dose of 7 to 17 manSv. 9. There has been no significant decrease in the 137Cs activity concentrations in moose or roe deer during the post-Chernobyl years. The effective ecological half-life in the forest ecosystems in Nordic countries thus seems to be very long, and we suggest that the physical half-life is now the best estimate.

REFERENCES R.Bergman and L.Johansson. (1 989) Radioactive caesium in a boreal forest ecosystem and internally absorbed dose to man. Proc. XVth Congress of IRPq Progress in radiation protection. Ed W.Feldt. R.Bergman, T.Nylen, T.Palo and K.Lidstrom. (1990) The behaviour of radioactive caesium in a boreal forest ecosystem. In The Chernobyl fallout in Sweden. Ed L.Moberg. The Swedish radiation protection institute. R.Bergstrom, H.Huldt and U.Nilsson editors. (1992) Swedish game - biology and management. Almqvist and Wiksell Tryckeri, Uppsala. S.von Bothmer, K.J.Johanson and R.Bergstrom. (1990) 137Cs in moose diet; considerations on intake and accumulation. Sci.Total Environ. 91, 87-96 G.Cederlund, H.Ljungqvist, G.Markgren and F.St&lfelt. (1980) Foods of moose and roe-deer at Grimso in central Sweden, results of mmen content analysis. Swed.Wildl.Res. 11.4 Number of mammals killed in Denmark. Danmarks miljoundersogelser for Flora and Faunaekologi, 1992. B.Fawaris and K.J.Johanson. (1994) Radiocaesium in soil and plants in forest in central Sweden. J.Tota1 Environment in press. 0.Hjeljord. (1981) Elg-vinterbeiter og rettet avskytning. Nor.Skogsbruk 10, 11-13. Hunting statistics 1989. Central bureau of statistics of Norway. K.J.Johanson and R.Bergstrom. (1994) Radiocaesium transfer to man from moose and roe deer in Sweden. J.Total Environ. In press. K.J.Johanson, R.Bergstrom, 0.Eriksson and A.Erixon. (1994) Activity concentrations of 137 in moose and their forage plants in Mid-Sweden. J.Environ.Radioactivity.22, 25 1-267. K.J.Johanson and R.Bergstrom. (1 989) Radiocaesium from Chernobyl in Swedish moose. Environ. Pollution 61, 249-260 K.J.Johanson, R.Bergstrom, S.von Bothmer and G.Karlen. (1990) Radiocaesium in wildlife of a forest ecosystem in central Sweden. In Transfer of radionuclides in natural and semi-natural

30 1

environments. Ed. by G.Desmet, P.Nassimbeni and M.Belli. Elsevier Applied Sciencce, London. 183-193 G.Karlen K.J.Johanson and R.Bergstr8m. (199 1) Seasonal variations in the activity concentration of 137 in Swedish roe-deer and in their daily intake. J.Environ. Radioactivity. 14, 91-103 S.Kleiven, K.J.Johanson and A.Backhans. (1993) Variability on the caesium-137 activity concentrationsof some dwarf-shrubs in a forest area in central part of Sweden. Manuscript. T.Palo. (1990) Radioactive caesium in a boreal forest ecosystem. Ecological concepts in radioecology. In The Chernobyl fallout in Sweden. Ed. L.Moberg. The Swedish radiation protection institute. T.Palo, P.Nylen, T.Nylen and G,Wickman. (1990) Radiocaesium levels in Swedish moose in relation to deposition, diet and age. J.Environ.Quality20. H.C.Pedersen and S.Nybo. Radio cesium in lirype og fjellrype forirsaket av reaktor-ulykken i Tschernobyl. In Tchernobyl, Slutrapport fia NINA'Sradiookologi-program 1986-1990. M.R.Petersen and H.Strandgaard. Roe deer's food selection in two different Danish roe deer biotopes. CIC-symposium"Capreo1us"in Salzburg, April 1992 R.A.Olsen, E.Joner and L.R.Backen. (1990) Soil hngi and the fate of radiocaesium in the soil ecosystem - a discussion of possible mechanisms involved in the radiocaesium accumulation in fungi, and the role of hngi as a Cs-sink in the soil. Transfer of radionuclides in natural and semi-natural environments. Ed. G.Desmet, P.Nassimbeni and M.Belli. Elsevier Applied Science London, pp 657-663. A.Rantavaara, T.Nygren, K.Nygren and T.Hyvbnen. (1987) Radioactivity of game meat in Finland after the Chernobyl accident in 1986. Supplement 7 to Annual Report STUK-ASS. C.Schwarts, W.Regelin, A.Franzmann and H.Hubbert. (1987) Nutritional energetics of moose. Swed.Wildl.Res. Suppl.1 part I, 265-279

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303

4.4 PATHWAYS OF FALLOUT RADIOCAESIUM VIA REINDEER TO MAN

ELDAR GAARE~)AND HANS STAALAND~) 1) Norwegian Institute for Nature Research (NINA), Tungasletta 2, N-7005 Trondheim, Norway

2) Department of Biology and Nature Conservation, Agricultural University of Norway, P.O.Box 5014, N-1432 As, Norway

SUMMARY The Chemobyl accident triggered a world wide monitoring and research activity. In the Nordic countries a renewed interest was taken in the lichen reindeer food chain. A wealth of data has been put on record, data that confirm experience from nuclear weapons tests. Like other animals reindeer are exposed to radioactive fallout through their food, and because they alone utilise lichens as forage, a high level of radiocaesium is found in their muscle. The lichen intake of reindeer varies from 70-80% of the diet in winter to 10-20% in summer. This factor, coupled with the short biological half-life of caesium in the reindeer body (10-20 days), leads to a strong seasonal variation: a late winter high that in the first years is about five times the late summer low of radio caesium in the meat. The effective ecological halflives of reindeer food lichens are 5-7 years for species growing on ridgetops. Species from more sheltered habitats show longer halflives, 6-11 years. This is partly due to local resuspension. Predictions based upon experience so far show that it will take about 20 years before radio caesium burdens are the same as those prior to the accident. Estimates indicate that the effective half-life of radio caesium in the meat of grazing reindeer is 3-4 years for the post-Chemobyl period. This has justified early slaughter in the domestic reindeer industry and when hunting wild reindeer. Countermeasures based on fodder additives were tested successfully for domestic animals, both bentonite and ammonium iron(I1I)hexacyanoferrate may be given in several different ways prior to slaughter, in order to reduce the radio caesium burden to below the critical level in the meat. We suggest that more research is needed on ecosystem recycling of 137Cs as well as other resuspension in wilderness areas. The developing of countermeasures should also be continued, and a monitoring of fixed plots should be undertaken in major reindeer ranges. 4.4.1INTRODUCTION Throughout history in the north the reindeer or caribou (Rungifer turandus L.) has been the basis of man's survival and it is still an important part of the economy and culture of people living in this area. Unfortunately, since the end of the second world war, technological developments in the industrialized world have imposed a serious threat to the people in the north through the polluting of the lichen-reindeer-man food-chain by radioactive fallout and other airborne pollutants. This is

304 a matter of much concern because lichens obtain nutrients from airborne material, and they are the major food of most reindeer herds. These problems were recognized and extensively studied at the time of nuclear weapons testing and afterwards, but no countermeasures were developed. The basic mechanism of radiocaesium transfer through the lichen-reindeer-man food-chain was described, and studies made after the Chernobyl accident have increased our knowledge, but more detailed investigations are still needed to understand fully the behaviour of radiocaesium in arctic and subarctic ecosystems. The present review emphasizes a combination of basic reindeer nutritional ecology and radiocaesium transfer parameters.

4.4.2 SEASONAL MIGRATION PATTERNS AND HABITAT CHANGES Fennoscandian reindeer have different seasonal cycles of habitat changes. In northern Norway, Finnmark and Troms, there is a typical spring migration from the interior to the coast and coastal islands. In Sweden the typical spring migration pattern is from the forested area towards alpine mountain ranges closer to the Norwegian border. Some reindeer herds on Helgeland, Norway, spend the winter in coastal areas, but migrate to the interior mountains in the spring. In Finland reindeer have no long-distance migration, but alternate between winter and summer pasture areas within the same general area. Similar systems apply to reindeer-herding of the Saami as well as to the non-Saami herding in southern Norway. The wild reindeer of southern Norway usually have no long-distance migration, but there are shifts between winter and summer areas. Migration patterns and reindeer distribution are depicted on Figure 4.4.1. These migration patterns result in three general types of winter pasture: forest areas with ground and tree lichens (Sweden, Finland and a limited area in Norway), mountain plateaus, alpine environments with ground lichens (Northern Finland and Norway), and lowland areas at the coast in Nordland and Troms in Norway, where grazing is predominantly dwarf shrubs, graminoids, and small amounts of lichens. Summer pastures are located either in alpine environments or in coastal areas (Northern Norway, Finnmark). The months of April to June comprise the transitional period between winter and summer pasture, June through August is the summer grazing period, and September to October the transition to winter pasture. The precise timing of the move from winter to summer pasture and from a summer to a winter environment is difficult to predict, depending as is does on weather conditions, snow depth and geographical area. Predicting the time for the onset of the spring and autumn transitional periods is important because this marks the start of the decontamination or accumulation periods for the radiocaesium burdens of the reindeer.

4.4.3 FOOD SELECTION Domestic and wild reindeer are free-ranging animals obtaining all their food from natural pastures. Only in some parts of Finland are domestic reindeer regularly given supplementary food, mostly hay, in winter.

305

Figure 4.4.1. Distribution and migratory pattern of Fennoscandian wild and domestic reindeer. Redrawn from Staaland and Eichelmann (1991).

306 There have been many studies on seasonal variations in the food intake of reindeer, but usually they are not sufficiently detailed to give exact information on dietary composition. Only general knowledge of the seasonal variations in the diet of reindeer has been applied when attempts have been made to describe the pathways of radiocaesium to the lichen-reindeer-man food-chain in Fennoscandia. The general pattern of reindeer foraging ecology is well known. Information is usually based on observation of bite selection in the field, fractional composition of samples from oesophageal samples, rumen analyses and microhistological studies of faecal droppings. Rumen and faecal analyses give incorrect results with respect to diet composition because different plants are digested at different rates and to varying degrees. In addition the correct identification of plant fragments is difficult. For example, fungi are rarely observed in rumen samples, probably due to a short residence time, whereas mosses can be overrepresented because of their low digestibility. Moreover, these dietary studies give no direct information on the weight percentage of different plants or plant groups in the diet. This is important when estimating radiocaesium intake. Based on several studies (Gaare, 1968; Eriksson et al., 1981; Gaare and Skogland, 1975; Helle, 1981; 1984; Skjenneberg et al., 1975) we have attempted to produce a dietary table for each month of the year, Table 4.4.1. Data based on rumen content analyses are adjusted for differences in turnover times (Gaare et al., 1977). Reindeer are, however, opportunistic in their food choice, and diet varies with geographical area, season and year. In winter difficult snow conditions makes large areas inaccessible for grazing. The reindeer seek ranges with little or no snow both on a large as well as a small scale. Continental areas with little and dry snow are preferred to oceanic ones rich in snow where icecrusts are frequent. Locally, bare-blown ridges are preferred to leeward sides and snow beds. In Fennoscandia it is possible to differentiate three major winter-grazing habitats: coniferous forest, alpine areas above the forest, and the Norwegian coast. In Finland and Sweden reindeer graze in forested areas to a large extent, whereas in Norway most grazing is in subalpine-alpine regions or in coastal environments. The species composition of forage varies greatly between these three types of habitat, but in all areas lichens, when available, are of great importance. Overall lichens are most abundant in habitats with least snow. Ground lichen are eaten both in forested and alpine areas, but in the forest arboreal lichens may be important at times. In coastal areas rock lichens can be abundant in addition to ground lichens on coastal heaths. Lichens are eaten throughout the year, but in largest quantities during winter. In spring, after calving in May, the reindeer seek areas where the snow has melted to feed on the early growth of vascular plants low in fibre and high in nitrogen and potassium. Thus, during spring the intake of lichens decreases, and throughout the summer reindeer feed on graminoids, herbs and leaves of woody plants. These patterns are repeated through the snow-free season and a wide variety of species are grazed. Particularly in alpine regions, new-grown plant parts are available on the leeward slopes and snow-beds where spring comes late. When fungi become available from late July to the first frosts in late August/early September, reindeer consume substantial amounts in addition to vascular plants. As the vascular plants wither, lichens gradually increase in importance during September. On ranges where

307 stocking rates are within the carrying capacity and lichens therefore abundant, these become the dominant food in the late autumn and winter period. Graminoids and woody species, in about equal parts, amount to less than 10-20% of the intake in this period. Canadian experience confirms these data from the taiga region (Gauthier et al., 1989). In high arctic areas like Svalbard, mosses and grasses, herbs and arctic willow leaves are the major dietary components in all seasons, almost no lichens are available. If we consider the three major Fennoscandian grazing habitats mentioned above, there can be considerable differences with respect to species composition in the vegetation and in the selected diet. However, ground lichens eaten in the alpine area can be replaced by rock lichens on the coast or tree lichens, if the ground lichen carpet is depleted in the forest. From one range to another there can be a shift in the species of each plant group: i.e. in many wild reindeer alpine ranges the lichen Cetraria nivalis is the most important, whereas in a dry pine forest in northern Sweden Cladinu srelluris dominates the diet. The same applies to vascular plants too. If there is heavy grazing pressure for several winters, the lichen pastures can be depleted. A fully-grown lichen mat contains 1200-18OOg DM (dry mass)/m2, dependent on the composition of the species. When population densities are high the mat can be grazed down to about 25 g dry mass/m2. The lichen intake in winter-grazing reindeer then drops to 35% or less. Concomitantly the intake of graminoids, woody plants and litter increases. Today this most certainly applies to larger areas of the land used for grazing domestic reindeer in northern Fennoscandia. Table 4.4.1.Generalized Fennoscandian reindeer diet: percentages of intake.

Dietcomponent Jan Feb Mar Apr May Jun Jul Lichens 85 85 80 75 60 10 10 Graminoids 6 6 9 12 10 40 40 Woodyplants 6 6 9 12 30 20 10 Herbs 0 0 0 0 0 3040 Mosses 3 3 2 1 1 0 0 Fungi Litter

xx

xx

xx

xx

Aug Sep Oct Nov Dec 15 30 60 70 80 58 55 24 17 10 12 14 14 1 1 7 15 0 0 0 0 0 1 2 3 3 x

x xx

xx

xx

xx

x - may be present, but hard to detect in rumen samples xx - up to 10% present on depleted lichen grazing, negligible on rich ones

4.4.4 QUANTITATIVE FOOD INTAKE Exact information on the forage consumption of both domestic and wild grazing reindeer, or other animals, is hard to come by. Estimates of forage intake (relative to body mass, BM or metabolic body mass BM”’’) in grazing sheep and cattle usually fall within a range of 40-90 g dry m a s ~ / B M ~ ~ (Cordova ’~/d et al., 1978). Feeding experiments with reindeer under controlled conditions give some information about basic needs for maintenance and growth, but do not take into account the energy requirements for walking, reproduction, feeding, cratering etc. From May to September male yearlings (BM 55-85 kg) had a weight gain of 222 g/d, with a food intake that increased from 1.5 kg/d in March-April to 3 kgld in mid August. With a food intake of 1.9 kg/d, the animals in the period October-

308 September maintained a stable body mass of approximately 80 kg (Ryg and Jacobsen, 1982). Fed pure lichen ad lib, male reindeer calves (BM=40 kg) consumed 0.7-0.8 kg dry mass/d and lost about 47 g BM/d (Jacobsen and Skjenneberg, 1975). A pure lichen diet is usually too low in nitrogen and minerals to maintain the reindeer in nutrient balance and weight loss is inevitable (McEwan and Whitehead, 1970; Staaland et al., 1980). Usually reindeer loose body mass during winter, but regain it in spring and summer. Based on the accumulation of radiocaesium fallout from nuclear weapons testing, Hanson et al. (1975) estimated the winter lichen intake of free-roaming caribou in northern Alaska to be 4.98 f 0.40 kdd. Estimates based on data collected after the Chernobyl accident indicate an intake of about 3 k d d of dry lichens (Pedersen et al., unpublished). Some of these estimates appear high compared to the food intake measured in laboratory experiments. Estimates of the food intake of caribou during the summer at Prudhoe Bay, Alaska, are about 106 g Dm/BM0.”/d (White et al., 1975). Extrapolated to a reindeer with body mass of 70 kg, this would give a food intake of about 2.6 kg Dm/d. In Svalbard the summer food intake of adult reindeer males was estimated to 4.0-4.9 kg Dm/d, and for adult females about 3 kg Dm/d (Staaland, 1984). Forage intake will also depend on the available biomass and on type of vegetation too (Trudell and White, 1981). Trudell and White estimated the maximum food intake in reindeer with a body mass of about 70 kg to be about 4.5 kg dry mass/d when eating willow, and about 2.5 kg when feeding on different lichen species. A model made of an ‘average’animal in the Hardangervidda wild reindeer herd, Figure 4.4.2 (Rusten, 1975) depicts the seasonal fluctuation in food consumption, but gives very low values during the first summer months due to the effect of a large number of newborn calves. The variation in quantities of food intake is also influenced by the quality of the fodder and the possibility of selective feeding (White, 1983). Based on White’s modelling, winter food intake may vary between about 1 and 2 kg Dm/d for caribou/reindeer weighing 70 kg and summer food intake between about 2.5 and 3 kg Dm/d. In conclusion summer forage intake may be about twice that in winter. For a model animal of 70 kg body mass, the summer forage intake may vary between about 2.5 and 4 kg Dm/d, and the winter forage intake in a grazing reindeer between 1 and 2 kg. For further calculations in the present paper, we use median values and choose a daily food intake of 1.5 kg in the winter, 3.3 kg in the summer and 2.5 kg in spring and autumn, Table 4.4.2. It should be emphasized that different pasture composition, availability, etc., can greatly influence the daily food intake. Table 4.4.2. Estimated daily food intake of a reindeer with a body mass of 70 kg (kg Drn/d) at different seasons, grouped in the two food categories, lichens or vascular plants, according to Table 4.4.1. See text for further explanation.

Food intake Jan Feb Mar Apr May Jun Jul Aug Totalint. 1.5 1.5 1.5 1.5 2.5 2.5 3.3 3.3 Lichens 1.3 1.3 1.2 1.1 1.5 0.3 0.3 0.5 Vascul.pl. 0.2 0.2 0.3 0.4 1.0 2.3 3.0 2.8

Sep 2.5 0.8 1.8

Oct 44-Nov S N o v Dec 2.5 2.5 1.5 1.5

1.5 1.0

1.8 0.8

1.1

0.5

1.2 0.3

309

12 000

6 000 0

4

9

13 17 21 26 30 34 38 43 47 51

Weeks

Figure 4.4.2. Model of wild reindeer food consumption on Hardangervidda southern Norway from Rusten (1975).

4.4.5 RADIOCAESIUM ACTIVITY IN REINDEER FORAGE PLANTS Since 1986 a large number of measurements have been made of the radiocaesium content of forage plants taken from various reindeer ranges. Much remains unpublished or appears in reports with a limited distribution. It is common knowledge (e.g. Rissanen and Rahola, 1990; Erikson et al., 1991; Gaare, 1991; Haugen, 1992) that Chernobyl fallout is unevenly distributed on a large scale as well as a small. In the Dovrefjell-Rondane ranges measurements showed 11000-30000 Bq/m2 in regions exposed to rain on an eastem wind, but only 700-900 Bq/m2 in regions sheltered from the rain, e.g. h o t s d a l e n just 5 km away. Two Cludinu sfelluris samples collected 1 m apart on Dovrefjell showed 6000 and 66000 Bq/kg. Interspecific variation as well as seasonal differences add to the problem of estimating radiocaesium activity in reindeer forage. Without going into any detail, we conclude that there are several different reasons for the uneven distribution of the wet-deposited fallout. The donor, the contaminated air, contained uneven amounts of fallout, and the weather and topography determined where the rain showers fell. The receptor, the land covered in vegetation, did not have a uniform ability to receive and retain the fallout due to differences in soil and vegetation, individual plant characteristics (i.e.hairiness), possible prior wetness and patchy snow cover. Measurements of lichen species commonly eaten by reindeer, collected in Fennoscandian fallout areas in 1986 vary from 1000-2000 to 25000-45000 Bq/kg dry mass. Even higher values were common in areas such as eastern Jotunheimen, Bergefjell and Frostviken. A set of values from Griningsdalen and Derilen is given in Table 4.4.3. To facilitate regional comparison it was found necessary to present data for 13’Cs only accompanied by the time of measurement. When needed, I3’Cs content is calculated from given sums of radiocaesium (‘34Cs+ 13’Cs) based on the

310 relation 134Cs/I3’Cs = 0.54 on 1st May 1986 for the Dovremondane region, and W!s/137Cs = 0.63 in the east Jotunheimen region, and their physical halflives 30.17 and 2.06 yrs, Naeumann and Gaare (1991), Staaland et al. 1994 (in prep). Table 4.4.3. Radiocaesium (13’Cs) Bqkg dry mass in plant samples. Calculated from measurements of the sum of 134Csand I3’Cs. a. Griningsdalen, Jotunheimen in summer: June-AudSept.

Year 1986* 1987 1988 1989 1990

Graminoids 1454 f 1071 (5) 841 f 855 (33) 623 f 680 (36) 621 f 585 (36) 432 f 4 2 8 (36)

Herbs 1153 f 726 (5) 785 f 792 (36) 809 f 892 (36) 632 f 544 (36) 680 k 659 (36)

Foliage Lichens 281 f 154 (3) 29898 (1)** 328 f 185 (23) 38618 (1)** 270 f 243 (24) 25500 f 8200 (24)*** 260 f 223 (24) 23500 f 4100 (16)*** 257 f 218 (24) 15700 f 5300 (20)***

b. Dortllen, Rondane in summer: July/August.

Year 1986

Graminoids 7185 f 7410 (7)

Herbs 1694 f 990 (3)

Foliage 338 k 171 (3)

1990

1122f 1185 (11) 842+751(6)

1119+412(5)

219rt 154(3)

Lichens 34690f45063 (20)*** 11877+4732(17)***

593+80(3)

122+134(4)

6492f4646(14)***

1993

*

From August. Cludina stellaris Ground lichens. Graminoids: Alpine Hairgrass (Deschumpsiu ceaspitosu slat.), Wavy Hairgrass (0.flexuusu), Bottle Sedge (Curex rostrata), Common Sedge (Curex nigru), Common Cotton-grass (Eriophorum ungustifolium),Three-leaved Rush (Juncus trifidus). Herbs: Cow-wheat (Melampyrum spp.),Bogbean (Menyunthes trifoliatu), Common Sorrel (Rumex ucerosu), Arctic Milk-vetch (Astragulusfrigidus), Alpine Saussurea (Saussureu ulpinu), Goldenrod (Solidugo virgaureu), Wood Cranes-bill (Geranium silvuticum). Extremely high values in Common Sorrel are excluded. Foliage: Downy Birch (Betula pubescens), Dwarf Birch (B. nunu), Willows (Salix spp.), Aspen (Populus tremulu). Data from: a. Garmo et al. (1990), Staaland et al. (1990), and Pedersen et al. 1993, b. Gaare unpublished.

** ***

At the time of the fallout, the fodder plants of the reindeer ranges were for the most part covered with snow. Wind-exposed ridges in the alpine regions were bare, leeward sides and snowbeds were covered, but bushes and a few birches protruded through the snow. Samples of plants collected during the first years after the accident may show large values, probably originating from externally adsorbed radiocaesium. However, vascular plants take up minerals from the soil moisture through the roots only, and after some years different plant species may have a very different radiocaesium content as a result of differences in the situation of their active roots in the soil. In this connection the mycorrhiza may be considered as part of the root system. In alpine ecosystems trees and bushes tend to have a deep root system, and from 1988 almost without

311 exception these plants show low values. The position of the roots of species of dwarf bushes, graminoids and herbs may differ considerably and this often explains the vast variation in radiocaesium content that they show even today. Nevertheless, not all the interspecific variation can be explained in this way. A species such as Common Sorrel (Rumex acetosa) regularly shows 3-7 times the radiocaesium content of other herbs. This and similar variations are as yet unexplained. Lichens growing both on the ground and on other substrates are thalloid plants and lack roots, They obtain the water that they need, together with its mineral content, mostly by absorption of rain water over their whole surface. After rainy periods such organisms quickly dry out, retaining the minerals and possible contaminants accompanying the rain. The next shower will only wash out a minor fraction of the accumulated mineral supply stored on the inside cell surfaces in the lichen thallus. Species differences may be caused by small differences in habitat ecology, Figure 4.4.3. Coelocaulon divergens, Alectoria ochroleuca and Cetraria nivalis grows on snowfree exposed ridges, Cladina rnitis, C. stellaris and Stereocaulon paschale in more sheltered sides of a ridge (Gaare, 1987a, b; 1989; 1990 and 1991). This explains the large differences, often of an order of magnitude, between the 137Csload of vasculars and lichens. To predict the future content of 137Csof components in the ecosystem, the rate of reduction is more important than the exact value at any time. The effective halflives (Gaare, 1991) of some important lichens are shown in Table 4.4.4. Species from the top of the mountain ridge, 1-3, show the same relatively short halflife, 5.1-7.0 y r s (r2 > 0.3 in all cases). Except for Alectoria ochroleuca, all are important food species. Cetraria nivalis is a dominant in plant communities on the ridges. It is widespread on alpine and subarctic ranges in continental areas of Fennoscandia. Cladina stellaris is most important in forest ranges. Table 4.4.4. Effective halflives of 137Csin some ground lichens from the reindeer winter ranges on Dovrefjell in the low alpine region, 1000 m asl, updated from an earlier presentation (Gaare 1991). Most species are divided into a living (L) and a dead (D) part and are measured separately, one (no.2) is measured in toto (T). The estimates are based on a linear regression with time as an independent variable. A 95% confidence interval is given. A negative halflive means that the radioceasium content actually increased during the observation period, 1987-93.

Species 1 Alectoria ochroleuca 1 Alectoria ochroleuca 2 Coelocaulon divergens 3 Cetraria nivalis

3 Cetraria nivalis 4 Cladina rnitis 4 Cladina rnitis 5 Cladina stellaris 5 Cladina stellaris 6 Stereocaulonpaschale 6 Stereocaulonpaschale

n L D T L D L D L D L D

32 28 24 28 28 26 29 42 37 21 19

Halfl. yr 5.1 6.1 7.0 5.5 -29 8.2 -37 11.5 708 6.5 -317

95% c.i

>4.2 >4.5 >5.3 >5.4 <-4.9 <-110 <-2.6 <-29 <-5.2 >4.3 <-3.0

4.6 <11.8 41.7 4.7 >13.8 >4.8 >7.9 >5.7 >8.0 >6.0 >7.6

312

Radiocaesium in ground lichens: A upper living part

12

24

36

4%

60 72 04 96 Months after May 1986

24

36

48

60 72 84 96 Months after May 1986

B dead basis 15

0' 12

-

Dovre mountains, 1987 93

-

I

01*19*19111*11

Coelocaulon divergens Alectoria ochroleuca Cetraria nlvalls

-..--.-

-- -

-..-1.

I

Cladina mitis Cladina stellaris Stereocaulon pashale

Figure 4.4.3.Development of lichen radiocaesium activity in the Dovrefjell region. Species 1. Coelocaulon divergens, 2. Alectoria ochroleuca, and 3. Cetraria nivalis, grow on more or less snowfree ridgetops. Species 4. Cladina mitis, 5 . C. stellaris, and 6. Stereocaulon paschale need snow protection and grow in more sheltered habitats.

313 Species 4-6 grow on the leeward sides of the ridge. These have a longer halflife or even a negative one, i.e. the radiocaesium load increases. The basal part of these species has a negative halflife, and this is best explained by assuming a bioclimatologically driven resuspension. Wind (and surface water) carries lichen fragments and other plant litter from the windward top of the ridge to the leeward-growing species. The litter varies from 1-2 cm to microscopic size, and gets caught in the rather dense mat formed by these fruticose species. The Cladina spp. lack a bark and have a felt-like surface, Stereocaulon paschale does have bark, but is covered with more or less shell-like phyllocladia. Both can be characterized as good "dust catchers" even in the upper part of the lichen thallus. A sample series of Cladina stellaris from the V&g%mountains, 1986-1990, showed a lower halflife, 2.8 yrs. (Pedersen et al., 1993). Data from the forest region near the Rondane mountains gave 4.3-6.2 yrs. Material from the forest region of northern Sweden, for the period 1972-84 gave 7-10 yrs (Erikson et al., 1991), and 7-8 yrs was found near lake Rogen in an open mountain area in Sweden (Roos et al., 1991). Where resuspension is unlikely, it seems that the halflife is the same in all lichen species with similar growth form, but Chernobyl fallout disappears more rapidly than does fallout from tests of nuclear weapons, probably because the external dust contamination of the first year vanished more rapidly. The Cladinas and Stereocaulon are very common in the continental parts of Fennoscandia as dominants in the dry, open birch and coniferous (mostly pine) forest. In the alpine region they may dominate the leeward ridges. They constitute a widespread and very important winter food for reindeer, particularly the domestic ones. The transport of plant fragments from ridges to leeward areas explains an increase observed in radiocaesium loads per unit area snow-bed (Bretten, 1991). This is in agreement with the increase in many vascular plants growing in leeward plant communities and snow-beds observed in 1990-91. From being stored in the lichens of the winter pasture, radiocaesium is transported to the summer pastures of reindeer and other herbivores. The extent and future importance of this resuspension is unknown. Table 4.4.5 gives halflives for I3'Cs in some vascular plants or plant groups. Values are mostly in good agreement with those found elsewhere, e.g. Nylen and Nelin (1993). With some exceptions, most plants give an effective halflife of 3-5 yrs. The increase in recent years is also

Table 4.4.5. Effective halflives of 13'Cs in some vascular plants or plant groups from the Rondane mountains in Norway (Gaare unpubl and Staaland et al. 1990).

Species Graminoids Herbs Foliage Dwarf shrubs Dwarf BircWBetuula nana, last yr shoot BilberryJVaccinium myrtillus last yr shoot

Halflives 4.0-4.4 3.3-7.8 -3.6- 14.4

Observ. yrs 1986-90 1986-90 1986-90

4.2 9.7

1986-91 1987-91

314 observed in the forest region, but here the physiological decomposition processes are a more likely explanation than resuspension (Bergmann et al., 1991). In a limited number of samples of plants growing in poor fens in the Rondane mountains, Lindmo (1993), found that for cryptogams I3’Cs decreased from 1986-91, whereas vascular plants showed no significant change in the same period. The relative amount of I T S in the intake by reindeer of the food plant groups in Table 4.4.1, lichens, graminoids, herbs, dwarf bushes and mosses, may be multiplied by their respective radiocaesium contents. For each month one then obtains the load in one weight unit of the food. If no ecosystem process has been overlooked, and the halflives for I T S found for each different plant group also hold good in the future, we can predict the radiocaesium content of the food ingested by reindeer. Figure 4.4.4 shows this for some past and future years. Use was made of the following halflives for L37Cs: Lichens 4.5 yrs Herbs 12 yrs Graminoids 4.4 yrs Mosses 3 yrs Woody plants 7 yrs The prediction shows that the radiocaesium content of the winter and summer diet of the reindeer levels out. So far this is confumed by observation. It further shows that the 1985 level, that of the year before the accident, will be regained in the year 2005, about 20 years after the accident occurred.

Radiocaesium in natural reindeer food A 35

’I

B

Yoar:1087

=1 =O 25

J F Y A M J J A S O N D Month

C

Year:lW3

15

1 J F M A M J J A S O N D Month

D 35

E

Yaar:ls89

1

Yaar:2OOS

EE 10 5

0 J F M A M J J A S O N D

0 J

F M A M J J A

S O N D

Figure 4.4.4. Predicted seasonal variations in I3’Cs concentrations in reindeer food. Further explanation in text.

315 4.4.6 DISTRIBUTION OF RADIOCAESIUM IN REINDEER Muscle tissue contains the overall highest concentrations of radiocaesium and there is less than a 10% variation between different muscles (Rissanen et al., 1990; Ekman and Greitz, 1967). In equilibrium, muscle tissue concentrations are approximately 6.5 times red blood cell values (Garmo et al., 1990; Staaland et al., 1990a). Internal organs including bones, but excepting kidneys, the salivary glands and the pancreas, have considerably lower values (Rissanen et al., 1990, Figure 4.4.5). The exact distribution of the radiocaesium body burden is, however, difficult to establish accurately because there are no accurate weights of many reindeer organs given in the literature, and also the radiocaesium activity of the skin, head, etc., is missing. From estimates based on data in the literature, own unpublished data and assumptions, Table 4.4.6, it seems reasonable to assume that about 50% of the total body burden in reindeer is localized in muscle tissue. This is lower than the estimates 78.9% of the body burden in skeletal muscle and 9.4% in the smooth muscle - made by Holleman et al. (197 I). This leaves only about 12% of the total body burden to the remaining parts of the body.

A 18000

Muscle Tissue

i 8 w o 1 Internal Organs

16 000 14 000

E

.o, 12000 m 3 f

10000

2 0,8000 x c

d

6000

4000

2000 0

C

D

Figure 4.4.5. Distribution of 13’Cs in different reindeer organs (n=20). The indicator line show!; minimum and maximum concentration. Redrawn from Rissanen et al. (1990).

316 Table 4.4.6. Relationship between muscle and other organs radiocesium activity in reindeer (Bqkg organ weight) / (Bqkg muscle) and relative size of different organs (Weight of organ/BM*100). Relative organ weights is based on disection of 6 one year old male reindeer (Staaland et al. unpublished).

Organ Body weight (kg) Dressed weight Muscle tissue Bones Liver Kidney Heart Spleen Pancreas Lungs Rumen wall tissue Gastroint. tissue. Red blood cells Plasma Blood*** Rumen content Gastroint. tract Total measured Different unmeasured organs: Trachea Oesophagus Diaphragm Hide Head Antler Unmeasured Total unmeasured

Relative l37Cs activity

1.00 0.22 f 0.04 0.52 f 0.06 1.33f 0.08 0.72 f 0.04 0.67 f 0.04 1.22 k 0.16 0.43 rt 0.11 0.35 f 0.03 0.55 f 0.1 1 0.19 f 0.02 0.03 f 0.00 =0.3**** 4.7****

=0.5?

Relative organ weight 89.9 f 5.9 51.27 f 2.19 36.4 f 2.6* (3) 10.6 f 1.8* (3) 1.35 f 0.17 0.26 fO.O1 0.82 f0.09 0.18 f 0.06 0.27 f0.31 1.16 rtO.10 2.81 f0.21 2.29 f0.47 2.6** 5.4** 3.87 f0.37 13.52 f 3.11 3.95 f 0 . 3 3 81.61 0.43 f 0.03 0.16 rt 0.01 0.48 f 0.12 6.80 f 0.38 5.19 f 0.49 3.43 f 5.76 1.90 19.39

* One year old reindeer (Ringberg et al. 1981), ** Estimates based on an assumed blood volume of 8 percent of BM and hematocrite value of 32 (refering to sheep) (Frandson 1976). *** Blood collected during desections, **** Estimates based on data from Rissan et al. 1990. 4.4.7 SEASONAL VARIATIONS IN REINDEER RADIOCAESIUM BURDENS Typical of all the reindeer populations examined are the pronounced seasonal variations in radiocaesium burdens (Pedersen et al., 1993; Rissanen et al., 1990; Skogland, 1987; Hanson et al., 1975; Ahman et al., 1990; Allaye-Chan et al., 1990). The same seasonal trends, with highest radiocaesium activity in winter, are also observed for the consumers of reindeer meat; wolves (Holleman et al., 1990) and man (Linden and Gustavsson, 1967; Hanson, 1982; Tillander et al., 1992; Figure 4.4.6). Spring is the decontamination period and accumulation occurs in the fall. Most data are insufficient to describe the seasonal radiocaesium cycles in detail and usually show a minimum in summer and a maximum 'peak' in early winter. Data from the V&tI reindeer herd in Southern Norway (Pedersen et al., 1993) show an accumulation period that lasts from late

317 Augustlearly September to early November, i.e. =75 days, followed by a high plateau with slow decrease until late April early May, i.e.==165 days. Decontamination takes place from early May to the end of June, i.e. = 60 days, and then reaches a summer minimum lasting about 60 days. However, in years with large fungi crops (1988) accumulation starts in late July, Figure 4.4.6. Thus four periods can be recognized; 1: Autumn accumulation period, 2: Winter maximum period, 3: Spring decontamination, and 4: Summer minimum period.

3

4m0-

,g

3m2Oo01Oo01

d

o

.

.

.

.

.

.

.

.

.

.

.

.

.

.

a

B ‘O so

fii

E

Vilhelmina, Sweden

1

I

40

30

B

20 10 0

C

t

401

-----.

Rondane Knutshs

E 0I

0

1

I 0

2

I 0

I

3

0

I

4

0

I

5

0

Figure 4.4.6. A: Seasonal variation in red blood cell (RBC) I3’Cs activity in domestic reindeer from VBgB, Norway (redrawn from Pedersen et al. 1993). In meat from reindeer B: slaugthered in Vilhelmina, Sweden (redrawn from ahman and Ahman 1994) and C: wild reindeer, Norway (redrawn from Skogland et al., 1991).

318

In winter, adult females usually have a higher body burden per kg body mass than do calves, whereas the opposite is true in summer. The explanations for these differences are partly unknown. The winter situation could be explained by a higher metabolic rate in the smaller calves (Holleman et al., 1971) and/or by differences in the quality of food intake between adult females and calves (Pedersen et al., 1993). Similar observations were made on elk (Alces alces) where calves have significantly higher radiocaesium activity than do adults (Palo et al., 1991). No conclusive explanation of the phenomena has been presented. The period of transition (- 75 days) from summer grazing of green vegetation to lichendominated pasture starting at the end of August or early September initiates the accumulation of radiocaesium, which reaches its plateau in November. The length of this period corresponds to findings from experimental studies, Figure 4.4.7. When given constant doses either by feeding naturally radiocaesium-contaminated food or radiocaesium in ionic form, stable levels are reached after about 100 days. However, in years rich in fungi this accumulation period can start at the end of July, as exemplified by 1988. Through the winter there is a slow decline in radiocaesium levels which could be explained by a reduced food intake. Other possibilities could be the increased fibre content of the diet (Johnson et al., 1968). The period July-April is well documented, but spring and early summer less so. This is a period when sampling is difficult, partly because of problems of access to the reindeer and partly because these animals should not be disturbed in the calving period. It can be assumed that radiocaesium levels remain high, approximately the same as in winter up until early May, and that the summer minimum is reached by the end of June. Presumably this period lasts for = 60 days. When reindeer contaminated with radiocaesium are fed radiocaesiumfree fodder the halflife is usually approximately 20 days (Hove and Staaland, 1987; Holleman and Luick, 1975a). If reindeer in a 60 day-period (3 halflives) in spring ate radiocaesium-free food only, their radiocaesium burdens would be reduced by 88.5%, i.e. a reindeer having a winter burden of 20000 Bq/kg would reach about 1250 Bqkg body mass by the end of June. This is in good agreement with observations. Through the spring period reindeer do not, however, forage on completely radiocaesium-free fodder. They continue to eat lichens, and also the first growths of vasculars contain high levels of radiocaesium, usually higher than later in summer. It has been shown, however, that a high potassium (or sodium) intake increases excretion and reduces the retention time of radiocaesium, and thereby shortens the biological halflife (Holleman and Luick, 1975b). In experimental studies extra supplies of sodium or potassium fed to reindeer who had foraged on lichens, reduced the biological halflife of radiocaesium to 10-12 days (unpublished results). Thus high levels of potassium in the early growth of green vegetation may enhance radiocaesium excretion and thereby explain the rapid spring decrease in activity. Throughout the summer a combination of low intake of radiocaesium in fodder dominated by green vegetation and a potassium effect, maintain radiocaesium in reindeer on a low level. In late summer, however, large crops of fungi may considerably increase the radiocaesium level in reindeer. This happened in 1988, Figure 4.4.6, when the accumulation of radiocaesium in reindeer meat started in July/August. Seasonal variations in radiocaesium level have also been observed in elk and roe deer (Karlbn and Johansson, 1991; Bothmer et al., 1990; Palo et al., 1991).

319

A

75000 70 OOO

6

WOoo

5

5oOOO

2 -ga

.

40000

3 o m

2oow 10 Ooo 0

10

20

30

40

50

60

70

80

SO

100 110 120

80

SO

1W 110 120

Time (d)

16W1400

-

Dose: 28,2 KBq/d

10

20

30

40

50

60

70

Time (d) Figure 4.7. A: Accumulation of radiocaesium in red blood cells of reindeer fed daily equk. _loses of ionic radiocaesium (420 kBq/d) spread on lichens. Multiplying RBC values at equlibrium with 6.5 gives meat values (Bqkg meat) and approximate daily dose (redrawn from Roed, 1992). B: When the reindeer were fed 'naturally contaminated lichen' with equal daily doses of 19.7 kBq/ or 28.2 kBq/d multiplying RBC values at equilibrium with 6.5 also gives approximate meat values, but only about 30% of daily dose (redrawn from Hove and Hansen 1992).

4.4.8 INTAKE RESPONSE AND TRANSFER OF RADIOCAESIUM FROM FORAGE TO REINDEER Feeding experiments show that equilibrium is reached after about 100 days when continuously feeding daily equal-sized doses of ionic radiocaesium, Figure 4.4.7 (Hanson et al., 1975). The accumulation curves follow an exponential equation. The transfer coefficient (intake Bq/d) / (Bqkg meat at equilibrium) is approximately 1 when animals are fed ionic radiocaesium. When the reindeer are fed naturally contaminated lichens lower meat values are found, Figure 4.4.7, and the transfer coefficient is usually about 0.26. Fibrous food might also give lower absorption and lower values (Johnson et al., 1968). In feeding experiments the prediction of radiocaesiurn levels in reindeer can be fairly accurate provided that dose intake and food type are known. Under field conditions a major

320 obstacle to predicting radiocaesium burdens in reindeer is quantifying the food intake and its composition. Estimates of lichen intake based on radiocaesium intake generally give higher values than expected from feeding experiments. Hanson et al. (1975) estimated caribou lichen intake to 5.0 (3.7-6.9) kg dry mass/d. These figures appear high, because caribou presumably also feed on other types of vegetation. As pointed out by Hanson et al. (1975) it is important for these calculations that the animals are in equilibrium with respect to radiocaesium burden. A comparison of radiocaesium levels (Bq/kg dry mass) in lichen collected throughout the reindeer-herding districts of Finland and meat values (Bq/kg fresh weight) indicates slightly higher values in meat than in lichens (Figure 4.4.8, Rissanen and Rahola, 1990). Contrasting values for caribou meat in Alaska ranged from 2.9 to 5.5 times the lichen values (Hansson et al., 1975). Another possibility is to compare the ground deposition with the activity concentration in the animal. Ahman and Ahman (1994) estimated these aggregated transfer coefficients, Tag=(Bq/kg meat)/(Bq/m2), to 0.12f0.02 in September; 0.44f0.02 in November-December and 0.76f0.03 in January to April for the reindeer herds of Sweden. The differences in transfer coeffcients are caused by differences in reindeer diets throughout time, but can also be erroneous when the animals are in a radiocaesium-accumulating phase in the autumn or during a decontamination phase in spring. The Tag value (Bq/kg meat)/(Bq/m*) obviously increases during the accumulation phase when meat values increase and ground deposition remains constant. During winter, Tag, meat values and ground deposition of radiocaesium remain constant. During spring, decontamination transfer factors related to pasture contamination and meat radiocaesium activity also remain high because of the time lag between meat activity and forage activity. Transfer coefficients should therefore only be calculated when the reindeer is in radiocaesium equlibrium. E 2500

1

.m

Lichens

Relndeer meat

I

1960

62 64 66 68 70 72 74 76 78 80 82 8486 881990 Year

Figure 4.4.8. Radiocaesium activity 1960-1990 in reindeer lichens and reindeer meat in Finland (data from Rissanen and Rahola 1990 and Tillander et al. 1992).

321

4.4.9 ECOLOGICAL HALFLIVE OF 137Cs IN REINDEER Ecological halflives of I3’Cs in reindeer depend on the development of radiocaesium activity in soil and pasture. Major seasonal differences are found as well as differences between different geographical areas (Table 4.4.7). These differences may to a large extent depend on differences in pasture composition (Tables 4.4.3, 4.4.4 and 4.4.5). The ecological halflives is usually about 3-4 years in the post Chernobyl period, but about 6 years before the accident. Table 4.4.7. Estimated ecological halflife of lS7Csin reindeer meat before and after the Chernobyl accident.

t/2 (years) Comments Females August 1.6 Calves August 2.6 Females Nov. 3.5 Calves Nov. 3.1 Females April 3.3 Calves April 5.4 3.2 Sweden Sept-Nov 4.2 1986/87-92 Jan-Apr Finland 1966-76 -6 Finnmark, Norway 1966-83 6.4 Sweden 1961-65 4.3-6.6

Aredperiod VAgil, Norway 1987-1992

Authors Pedersen et al. 1993

, Ahman & Ahman 1994

,

Rahola & Miettinen 1977 Berthelsen 1986* LindCn & Gustafsson 1967

* Calculated from activity data given by Berthelsen 1986. 4.4.10 COUNTERMEASURE 4.4.10.1 Countermeasure for reindeer kept in corrals Radiocaesium is not firmly bound to any organ in the body of the reindeer, and it is therefore easily recycled and excreted in faeces and urine. When reindeer are fed non-contaminated food at the maintenance level, radiocaesium activity rapidly declines with a maximum biological halflife of about 20 days (Hove, 1986; Hove and Staaland, 1987; Holleman and Luick, 1975a and b, Figure 4.4.9). In practical feeding experiments using the commercially available reindeer feed RF71, a biological halflife of 18 days was found for blood (Sletten, 1987a). RF71 is a commercial, pelleted, reindeer feed based on grass meal, ground barley, ground oats and wheat bran to which are added 3% hydrogenated marine fats and 4 % sodium bicarbonate (Sletten and Hove, 1990). By adding bentonite, Zeolite or Ammonium iron (II1)-hexcyanoferrate (AFCF) to the food, fecal excretion of radiocaesium can be increased substantially (Ahmann et al., 1990b; Hove et al., 1991a). These minerals bind radiocaesium in the alimentary tract and prevent absorption by the body. Ionic radiocaesium is apparently continuously absorbed, secreted and reabsorbed in large quantities from the alimentary tract (Staaland et al., 1990). Thus caesium binders will presumably not only be bound to some of the radiocaesium ingested with food, but also to the radiocaesium recycled from the body. Experiments in Sweden (Ahmann et al., 1990b) indicate that 2% bentonite in the total food apparently reduces radiocaesium absorption by 85%. Zeolite ( 2 or 4% in the food) reduced absorption by 50% relative to the amounts ingested. By increasing potassium intake from 0.84 s/d to 16 g/d the biological halflife was reduced from 22 days to 11 days (Holleman and Luick, 1975b). Recent experimental studies at the

322 Agricultural University of Norway have confirmed these findings, and sodium was also found to have a similar effect (unpublished results). The addition of potassium to the forage would therefore appear useful in increasing radiocaesium excretion. Moreover salt blocks containing potassium could be used in theory, but reindeer have no innate craving for potassium as they do for sodium. Experiments have also indicated that a supplement of stable caesium has a positive effect on the excretion of radiocaesium (Oughton et al., 1991).

10 500

9000 A

P B

1

7500

C

.-'g w

6000

'C

$

5

-8

4500

0

5

K

3000

1500

0 0

20

40

60

Days of feeding Figure 4.4.9. Decontamination of reindeer meat by feeding reindeer 1 kg/d radiocaesium free pellet diet (RF 71). Redrawn from Hove and Staaland 1987.

4.4.10.2 Countermeasures in freely-grazing domestic reindeer In an area contaminated by radiocaesium there will always be some intake of radiocaesium in grazing reindeer. To reduce alimentary absorption and increase excretion of endogenous radiocaesium, AFCF can be used. A major benefit is that AFCF is active in small doses (50mg/d). Salt blocks containing 2.5% AFCF have been offered, and with an intake of about 20dd of the block containing NaCl a reindeer simultaneously ingests 500 mg AFCF. A positive effect on radiocaesium levels in reindeer has been observed. The major drawback of the method is that not

323 all animals will use the salt blocks, and reindeer in coastal environments do not have the urge for sodium that makes animals use the blocks. Salt blocks impregnated with AFCF can therefore reduce the average radiocaesium level in a herd, but individual animals may still remain on a high level (Sietten, 1987b). A rumen bolus was made by compressing a mixture of 15% AFCF, 10% beeswax and 75% barite into a 50 g tablet (50-60 mm long, diameter of 18-20mm) (Hove and Hansen, 1993). A special plier was developed to place this sustained-release bolus safely in the rumen of reindeer (Hove and Hansen, 1993b). A bolus remains in the reticulum and slowly desintegrates over 1-2 months releasing a daily dose of about 200 mg. This leads to a 60% reduction in radiocaesium activity in the reindeer, Figure 4.4.10. Usually two boluses are placed in the rumen. In sheep and goats the bolus can be placed on the back of the tongue for the animals to swallow. In reindeer the bolus must be placed directly in the rumen. The advantage of this method is a comparatively uniform reduction in radiocaesium activity in all the animals to be slaughtered.

I

EI

Controls

0 Treated in January (2 Boli)

0 Treated in January and March (2+2 Boli)

I

I I

I

I

3

Nov

Des

Jan

Feb Time

Mar

Apr

1

Figure 4.4.10. Decontamination of freely grazing reindeer by means of sustain release boli impregnated with Ammonium (111) ironhexacyanoferrate. Redrawn from Hove et al. 1990. The methods described above can all be used in the reindeer industry to combat high radiocaesium levels in the meat to be marketed. It will always be important to take into account the seasonal variations in radiocaesium activity as shown in Figure 4.4.6,and also to note the

324 effect of abundant fungi crops on radiocaesium activity in reindeer. If possible therefore reindeer should be slaughtered in late summer before the animals start feeding on radiocaesiumcontaminated lichen pasture. If activities in meat are above accepted levels, then this is obviously the best time to start feeding radiocaesium-free diets to the reindeer in order to obtain accepted activity levels in the meat. From the viewpoint of animal welfare as well as economy, it is obviously best to use the sustained-release AFCF bolus and then to release the animal that is to be slaughtered back to the herd again, corralling them again only when radiocaesium activity has reached acceptable levels. It should, however, be noted that the method to be selected, feeding radiocaesium-free diets or the use of sustained-release bolused, also depends on the absolute levels of radiocaesium activity relative to levels acceptable in human food. Excessively high levels of radiocaesium in lichens and reindeer can make it impossible to reach acceptable levels for food when using the sustained-release bolus in freely-grazing reindeer. For wild reindeer, the only method would be to alter the hunting season to fit in with the low-point activity of radiocaesium in reindeer. Placing salt blocks impregnated with AFCF on the range might also reduce the radiocaesium activity in wild reindeer.

4.4.11 BIOLOGICAL EFFECTS OF CHERNOBYL RADIOCAESILJMFALLOUT Disregarding the age of the animals, the pattern of chromosome abberation in reindeer in Norway shows no correlation with the level of Chernobyl radiocaesium fallout. However, calves from the most heavily contaminated area show a significantly greater number of chromosome aberrations, characterized by 2-break events, compared to calves from areas not contaminated by the Chernobyl accident. Furthermore, reindeer born during 1986 in the most heavily contaminated areas, show a greater amount of chromosome damage than do reindeer born before or after the Chemobyl accident (Roed et al., 1991; Roed, 1992). It has also been claimed that in the Rondane mountain range of Southern Norway (which received 40-60 kBq/m 2 of '37Cs)the wild reindeer replacement rate was reduced by about 25% in 1988 and 1987 (Skogland et al., 1991). However, differences of opinion concerning the accuracy of methods of estimating the age and sexual composition of reindeer populations have led to these conclusions being disputed (Reimers, 1993). In domestic reindeer herds living in Southern Norway in areas with similar levels of radiocaesium pollution, no effect on reproduction or replacement rate was observed after the Chernobyl accident (Pedersen et al., 1992). Experimental studies using yearling male calves that had high burdens of radiocaesium (<0.1, 70, 140, 420, and 1160 kBq/d) showed increasing chromosome aberration related to the total radiocaesium burden after 20 and 23 weeks of treatment (Figure 4.4.11). Although some chromosome damage can be attributed to radiocaesium fallout, it appears that any effects of Chernobyl radiocaesium on the biological performance of reindeer has not been conclusively documented.

325

Figure 4.4.11. Chromosome aberrations in reindeer fed daily high doses of 134Csfor 20 and 23 weeks. Redrawn from Raed 1992.

4.4.12 TRANSFER FROM REINDEER TO REINDEER-MEAT CONSUMERS There are few data on the radiocaesium burdens of predators following the Chernobyl accident. The seasonal variations in the radiocaesium activity of reindeer meat are reflected in the body burdens of those that consume them - wolves or man, Figures 4.4.12 and 4.4.13. The annual maximum in radiocaesium activity in predators is, however, delayed compared to that of the reindeer (Lindtn and Gustavsson, 1967; Hanson and Palmer, 1965; Holleman et al., 1990; Westerlund, 1985). Care should therefore be taken when comparing results from different areas and years. The Norwegian lynx population is monitored by the Norwegian Institute for Nature Research. From 1986 to1993 samples from about 150 animals, shot by hunters or found dead, were collected and measured for 13’Cs. Stomach contents are recorded and animals that have had access to reindeer consistently show a higher burden of I3’Cs than do those on a diet of a roe deer or hare (Gaare and Kvam, in prep). Maximal values seem to follow the pattern of fallout closely, with highest values found in the coastal district of the counties of southern Nordland and NordTrandelag, viz. 87000 Bq/kg in the winter of 1989. The biological halflife (after single injections) of radiocaesium in predators of 17.5-26 days and 40-64 days in man appears longer than the maximum biological halflife (20 days) found in reindeer, Table 4.4.8. In the arctic and subarctic, occupation (reindeer hunterherdsman) and consumption of reindeer meat are the major factors determining human radiocaesium burdens, Figure 4.4.13. Furthermore males always have higher burdens than females (Nevrestueva et al., 1967; Miettinen and Hlsanen, 1967; Rahola and Suomela, 1990; Westerlund, 1985; Hanson, 1967; Miettinen, 1967; Miettinen, 1977; Jokelainen, 1967; B@eet al., 1990; Anonymous, 1967; Linden and Gustafsson, 1967). In 1965 to 1983 the male/female ratio in Kautokeino was about 1.8 for radiocaesium burdens (Berthelsen, 1986). Based on Berthelsen’s data, a regression can be made

326

A

20 -I

Month

B

25' h

U J

i

s 2 0

-

15'

0 10-

8

5-

0

n l m

* 1

1 ' 1 1 1 1 1 '

' I

'"'1

1966

Figure 4.4.12. A: Absorbed radiocaesium intake for the wolf and the resulting radiocaesium concentrations i skeletal muscle of wolf. B: Redrawn from Holleman et al. 1990. C: Seasonal variations of the I3'Cs concentration in saami people from Funlsdalen, Sweden 1965. Circles and triangles show the results of whole body measurements at four occasions during 1965. Average values and the individual maximum values are indicated by open circles and triangles for men, by filled circles and triangles for women. Redrawn from Lindbn and Gustavsson 1967.

321

A

4Ooo

Rein1 le ?rmeat

3000 2000 1000

0 1960

1970

1980

1990

B 60 OW

Reindeer herders

8

yc

30000

20 OW

1oow

0

19Q

Figure 4.4.13. The concentrations of I3’Cs in fresh reindeer meat (winter animals) and male reindeer herders (whole body measurements) in the Enare district of Finland 1960-1990. Redrawn from Rahola and Suomela 1990.

328

between radiocaesium concentrations in meat (x) and body burdens (y) in humans (kBq/kg BM): y=9.65*x-0.21, (r2=0.89, n=18). Although there is an obvious relationship between the consumption of radiocaesium-contaminated reindeer meat and human radiocaesium burdens, less is known about the kinetics of absorption. It can be assumed that radiocaesium is highly bioavailable and in predators like the wolf (Holleman and Luick, 1976). Holleman et al. (1990) estimated an absorption factor of 1. This is consistent with an absorption factor of 1 for ionic radiocaesium in reindeer. Table 4.4.8. Biological halflive of radiocaesium in reindeer-meat consumers (slow component). Silver fox (male Arctic fox (female) Coyote (female) Coyote (female) Wolf (female Inuits, Anaktuvuk Pass** Children (4- 12 yr) Minors (15-20 yr) Adults (> 21 yr)

*

**

t 1/2 (days) 25.3* 17.5* 26.0* 22.0* 23.0*

Body mass (kg) 5.3 4.9 9.5 11.5 31.3

40 52 64

48-61 48-84

17-34

Single injection (Holleman and Luick 1976). People were given radiocaesium-free beef and chicken instead of their traditional caribou-meat diet (Hanson et al. 1967).

4.4.13 CONCLUSION The Chemobyl accident triggered intensified radioecological research in all Nordic countries. Institutions not previously active in this field also took part, e.g. The Agricultural University in Norway, The Norwegian Institute for Nature Reseach, and The Swedish Tame Reindeer Research. A wealth of data has been put on record, data that confirm experience from nuclear weapons tests. Like other animals reindeer are exposed to radioactive fallout through their food, and because they alone utilize lichens as forage, a high level of radiocaesium is found in their muscle. The lichen intake of reindeer varies from 70-80% of the diet in winter to 10-20% in summer. This factor, coupled with the short biological halflife of caesium in the reindeer body (10-20 days), leads to a strong seasonal variation: a late winter high that is about five times the late summer low of radiocaesium in the meat. Predictions based upon experience so far show that it will take about 20 years before radiocaesium burdens are the same as those prior to the accident. Estimates indicate that the effective halflife of radiocaesium in the meat of grazing reindeer is 3-4 years for the postChernobyl period, Table 4.4.7. The verification of the predictions made should be done by monitoring major tame and wild reindeer ranges. This must include sampling both of animals and pasture. This has justified early slaughter in the domestic reindeer industry and when hunting wild reindeer. Countermeasures based on fodder additives were tested successfully for domestic

3 29 animals, both bentonite and ammonium iron(II1)hexacyanoferrate may be given in several different ways prior to slaughter, in order to reduce the radiocaesium burden to below the critical level in the meat. The developing of countermeasures has been very successfull and should be continued. Indications of recycling of 13’Cs in alpine ecosystems has been found, but the evidence is not conclusive. It may, however, mean that I3’Cs is transported with litter from ridges grazed in winter to leeward sides grazed in summer and autumn just before hunt or slaughter. We suggest that more research is needed to clarify the extent of this.

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333 Pedersen, @., K. Hove and H. Staaland (1993). Seasonal variation and effective half-life of i37Cs in semidomestic reindeer grazing mountain pasture contaminated by deposition from the Chernobyl accident. In: Environmental radioactivity in the Arctic and Antarctic. Ed. by P. Strand and E. Holm. gsterls, Norway, 303-304. Pedersen, 0.,K. Hove, E. Reimers, K. RBed and H. Staaland (1992). Strilingsdoser og kalveddlighet i spr-norske reinstammer. Reindrifsnytt 26 3/4,34-39. content of the Rabon, E.W. (1968). Some seasonal and physiological effects on '37Csand s9T90Sr white-tailed deer, (Odocoileus virginianus). Health Physics 15,37-42. Rahola, T. and J.K. Miettinen (1977). Fallout levels of i37Csand some shortlife nuclides in Finnish Lapland during 1966-1976 in the food-chain lichen-reindeer-man. Progress report. Prepared for the U.S. Energy Research and Devlopment Adminstration under contract No Ch:E(ll1)-3011. Rahola, T. and M. Suomela (1990). Radiocaesium i renkot-okad straldos for samebefolkningen. Rangifer, Special Issue 4,38-40. Reimers, E. (1993). Tsjernobyl og kalved~dlighet.Jakt og Fiske 8,46-47. Ringberg, T.M., R.G. White, D.F. Holleman and J.R. Luick (1981). Prediction of carcass composition in reindeer (Rangifer tarandus tarundus L.) by use of selected indicator bones and muscles. Can. J . Zool. 59,583-588. Rissanen, K. and T. Rahola (1989). Cs-137 concentration in reindeer and its fodder plants. Science Total Envir. 85, 199-206. Rissanen, K. and T. Rahola (1990). Radiocesium in lichens and reindeer after the Chernobyl accident. Rangifer, Special Issue 3,55-61. Rissanen, K., T. Rahola and P. Aro (1990). Distribution of cesium-137in reindeer. Rangifer 2( lo), 57-66. Rissanen, K., T. Rahola, E. Illukka and A. Alfthan (1987). Radioactivity of reindeer, game and fish in Finnish Lapland after the Chernobyl accident in 1986. Annual Rep. STUK-A55. Supl.8:33pp. Rominger, E.M. and J.L. Oldemeyer (1990). Early-winter diet of woodland caribou in relation to snow accumulation, Sekirk Mountain, British Columbia, Canada. Can. J. Zool. 68, 26912694. Roos, P., C. Samuelsson and S. Mattsson (1991). 137Csin the lichen Cladina stellaris before and after the Chernobyl accident. In: The Chernobyl fallout in Sweden. Ed. by L. Moberg. The Swedish radiation protection institute, Stockholm, 389-400. Rusten, P. (1975). Krav og tilfredsstillelse av reinens energibalanse som en del av en simuleringsmodell for forvaltning av villreinstammen. Hovedoppgave Norges Tekniske C'P H~gskole,Trondheim. 65 pp + appendix A-E. Ryg, M. and E. Jacobsen (1982). Effects of castration of growth and food intake cycles in young male reindeer (Rangifer tarandus tarandus). Can. J . Zool. 60,942-945. R ~ e d K. , (1992). Genetiske skader hos rein etter Tsjernobylulykken. In: Radioaktivt nedfall fra Tsjernobyl-ulykken. Ed. by T.H. Garmo and T.B.Gunner0d. Norges Landbruksvitenskapelige Forskningsrad, 103-111. R ~ e d K.H., , I.M.H. Eiklemann, M. Jacobsen and 0.Pedersen (1991). Chromosome aberration in Norwegian reindeer calves exposed to fallout from the Chernobyl accident. Hereditas 115, 200-206. Skjenneberg, S., P. Fjellheim, E. Game and D. Lenvik (1975). Reindeer with esophageal fistula in range studies: a study of methods. In: Proc First Int. Reindeer Caribou Symp. Ed. by 3.R. Luick, P.L. Lent, D.R. Klein and R.G. White. Biol. papers Univ. Alaska, Spec. Rep. 1, 528545. Skogland, T. (1987). Radiocaesium concentrations in wild reindeer at Dovrefjell, Norway. Rangifer 7(2), 42-45. Skogland, T., 0. Strand and I. Espelien (1991). Den biologiske betydning av radiocesium i villrein. Tsjernobyl. Sluttrapport fra NINAs radioakologiske program 1986-1990.

334 Temahefte 2. Norsk Inst. for Naturforskning. (E. Gaare, B. Jonsson and T Skogland eds.) pp. 64-71. Sletten, H. (1987a). F8ringsfors~kfor il redusere radioaktivitet. Reindrifsnytt 21(2), 24-26. Sletten, H. (1987b). Slikkestein-forsoket i Elgil. Reindrifsnytt 21(2), 19-21. Sletten, H. and K. Hove (1990). Digestive studies with a feed developed for realimentation of starving reindeer. Rangifer 10(1), 31-37. Solem, J.O. and E. Gaare (1991). Radioaktivt caesium i invertebrater fra Dovrefjell, Norge, 19861989 etter Tsjemobylulykken. In: Tsjernobyl-sluttrapport fra NINAs radio@kologiske program 1986-1990. Ed. by E. Gaare, B. Jonsson and T. Skogland. NINA Temahefte 2,l-71. Staaland, H. (1984). On the quality of Svalbard reindeer pasture in the summer and autumn. Rangifer 4, 16-23. Staaland, H. and I.M.H. Eikelmann (1991). Status of the reindeer industry in Fennoscandia. In: Wildlife production: Conservation and sustainable developement. Ed. by L A . Renecker and R.J. Hudson. AFES misc. publ. 91-6 Uviv. Alaska, Fairbanks, 77-88. Staaland, H., K. Hove and 0.Pedersen (1990). Transport and recycling of radiocesium in the alimentary tract of reindeer. Rangifer, Special Issue 3,73-82. Staaland, H., H. Bjflrnstad, 0. Pedersen and K. Hove (1991). Radiostrontium, radiocesium and stable mineral composition of bones of domestic reindeer from Vilgil, Norway. Rangifer 11(l), 17-22. Staaland, H., R.G. White, J.R. Luick and D.F. Holleman (1980). Dietary influences on sodium and potassium metabolism of reindeer. Can. J . 2001.58, 1728-1734. Staaland, H., T.H. Garmo, K. Hove and 0.Pedersen (1994). Radiocesium intake in reindeer, sheep and goats grazing alpine summer habitats in southern Norway. In prep. Staaland, H., T.H. Garmo, 8. Pedersen and K. Hove (1991). Endring i innhaldet av radiocesium i plantemateriale og beitedyr pi3 fjellbeite 1986-90- F~rebelsresultat. NLVFs Forsknings-program om radioaktivt nedfall. Seminar 6.-7. November 1990. Inf. fra Statens Fagtjeneste for Landbruket. nr. 28, 1990, 84-95. Svensson, G.K. and K. Linddn (1965). The transition of I3’Cs from lichen to animal and man. Health Physics 11, 1393-1400. Svoboda, J. and H.W. Taylor (1979). Persistence of cesium-137 in arctic lichens, Dryas integrifolia, and lake sediments. Arctic and Alpine Res. 11.95-108. Tillander, M., T. Rahola, T. Jaakola and M. Suomela (1992). Radiocesium in Finnish Lapps after Chemobyl. Mimeographed 6 pp. Trudell, J. and R.G. White (1981). The effect of forage structure and availability on food intake, biting rate, bite size and daily eating time of reindeer. J. Appl. Ecol. 18,63-81. Westerlund, E-A. (1985). I3’Cs i norske samer. SIS Rapport 19851. National Inst. of Radiation Hygiene Plsterils, Norway. White, R.G. (1983). Foraging patterns and their multiplier effects on productivity of northern ungulates. Oikos 40,377-384. White, R.G., D.F. Holleman and A. Allaye-Chan (1986). Radiocesium concentrations in the lichen-reindeerkaribou food chain: Before and after Chemobyl. Rangifer 1: Appendix: 2429. White, R.G., B.R. Thomson, T. Skogland, S.J. Person, D.R. Russel, D.F. Holleman and J.R. Luick (1975). Ecology of Caribou at Prudhoe Bay, Alaska. In: Ecol. Inv. of the Tundra Biome in the Prudhoe Bay Region. Ed. by J. Brown. Alaska. Univ. Alaska, Special Rep. No.2 ,151201. Ahman, B. and G . Ahman (1994). Radiocaesium in Swedish reindeer after the Chernobyl fall-out: seasonal variations, transfer from fall-out and long-term changes. Health Physics 66, 503512. Ahman, B., S . Forberg and G. Ahman (1990b). Zeolite and bentonite as cesium binders in reindeer feed. Rangifer, Special Issue 3,73-82. Ahman, G., B. Ahman and A. Rydberg (1990a). Consequences of the Chernobyl accident for reindeer husbandry in Sweden. Rangifer, Special Issue 3,83-88.

335

4.5 THE DISTRIBUTION OF RADIOACTIVE CAESIUM IN BOREAL FOREST ECOSYSTEMS

RONNY BERGMAN National Defence Research Establishment, Department of NBC Defence, Umei, Sweden

SUMMARY Experience from current research in the Nordic countries concerning the behaviour of radioactive caesium (134Cs and 137Cs)in boreal forests is reviewed with emphasis on its distribution in various time perspectives. The analysis has thus been focused on data of relevance for both early and later phases after fallout over forest areas. Possibilities and limitations in using data from other time periods or regions, than that characterised by fallout over the boreal zones after the Chernobyl event are also discussed. This concerns extrapolations from information pertaining to neighbouring ecological areas - at higher altitudes (alpine, and sub-alpine regions) or below the southern limit (i.e. in the hemiboreal and nemoboreal zones), and to future time with respect to predictions of the behaviour of 137Cs, based on results for OLD (i.e. from atmospheric nuclear weapons tests - mainly in the sixties) versus CHERNOBYL caesium. Beside the principal terrestrial constituents of the soil-plant-animal system , the BOREAL POREST ECOSYSTEM will for the present purpose be considered to comprise the semi-aquatic and aquatic components pertaining to peat, open peat bog, and ground water. This implies that runoff from a catchment constitutes the main link between the terrestrial part considered here and the aquatic ecosystem proper. In boreal forests the humus layer usually retains a major fraction of the fallout of radioactive caesium, evidently even several decades after deposition. This notable feature, as well as a persistent high availability in important food-chains, emerges from the present Nordic radioecological research. Both constitute facets of a singularly conservative - although not at all static - situation prevailing for radiaoctive caesium in the boreal forest. The implication is that for 137Cs physical decay will be the major factor of loss from the boreal ecosystem in a long-term perspective, and that runoff, particularly from peat bogs, is expected to be the second in order of importance - even in areas subjected to logging during the period. There seems to be a gradual change from these conditions southwards of the boreal zone, where differences in the vertical distribution between nuclear weapons and Chernobyl caesium are manifest, and processes operate, leading to significant decrease in availability. Physical decay and decreasing availability constitute the dominant factors for the loss of radioactive caesium in these areas.

336

INTRODUCTION The boreal forest This review deals mainly with experience from current research in the Nordic countries concerning the behaviour of radioactive caesium (134 Cs and "'Cs) in boreal forests with emphasis on its distribution in various time perspectives. Different terminologies may be used to define the boreal vegetation or climatic zones. The possible choices for the classification, of course, all result more or less in the same overall pattern, but may differ in the zonal subdivisions. We have chosen the vegetational zonation according to Ahti et al. (1 968). The BOREALZONE PROPER, illustrated in fig. 1 by dark gray areas, will in the present case include the Northern, Middle and Southern Boreal zones. The last-named acts as a transitional area between these northern zones and the SOUTHERN CONIFER BELT (the Hemiboreal zone) and also constitutes a northern limit for the occurrence of the oak tree. The boreal and hemiboreal areas exhibit too different The vegetation 'Ones Of environmental conditions to allow, apriori, definite asFenno-Scandia: Norihern, Middle, (dark gray); and Hemi- sumptions about the level of similarity from a radioecological boreal (light dotted); Subalpine and point of view. Similar arguments distinguish the boreal zones Alpine zones (white); and Nemoral (black along the atlantic coast of Norway, which are influenced by a dotted). maritime climate and associated with especially high annual precipitation. These areas are labelled OCEANIC BOREAL, but have not been marked explicitly in the figure.The boreal zone includes large areas of peat-land particularly in the northern parts of Sweden and Finland. The forests are dominated by the coniferous trees: Scots pine (Pinus syivesfris) and Norwegian spruce (Picea abies), often interspersed with deciduous species, primarily birch (Betula sp), mountain ash (Sorbus aucuparia) and alder (Alnus incana). There exist - although rather sporadically- growing sites of a boreal type even outside the boreal zone proper. One such site in Denmark was investigated after the Chemobyl accident. However, natural boreal forests in the Nordic countries are, primarily for climatic reasons, confined to the areas covered by fig.1. The vegetational zones bordering the boreal are mainly deciduous, but are adapted to very different conditions at high altitudes or at northern as compared to southern latitudes. Close to the Fenno-Scandian mountain ridge birch dominates completely, establishing the subalpine region, which in turn is superseded above the tree limit by the alpine region. Denmark and the southernmost coastlines of Norway and Sweden belong to the beech region, the Nemoral zone. The climate in Fenno-Scandia is generally humid, especially along its western coastlines and in its northem parts. A considerable proportion of the precipitation percolates through the soils and forms groundwater. Thus leaching gives rise to a specific pattern of stratified layers in the soil profile of the PODSOLIC TYPE. This is the predominant soil-forming process in the Nordic boreal zones. The other main type of forest soil in Fenno-Scandia is the CAMBISOL, in which a mixed layer

331 of organic matter and mineral soil is present under the litter layer. The cambisols generally appear on more nutrient-rich parent rock materials than do podsols, and are found more frequently south of the boreal zones. Where the groundwater level lies rather deep in podsolic subsoil (the iron podsol type), the humus layer consists of very sour raw-humus, in general 2-8 cm in thickness. If the groundwater level lies near the surface of the soil, at least during certain seasons, profiles of the so-called humus podsol type have often developed, varying in thickness from about 10 to 50 cm. The humus layer usually retains a major fraction of the fallout of radioactive caesium, evidently even several decades after deposition. This notable feature, as well as a persistent high availability in important foodchains, emerges from the present Nordic radioecological research. Both constitute facets of a singularly conservative - although not at all static - situation prevailing for radiaoctive caesium in the boreal forest.

Post-Chernobyl studies The analysis has been focused on data of relevance for both early and later phases after fallout over forest areas. The possibilities and limitations of using data from other time periods or regions, than those characterised by fallout over the boreal zones after the Chernobyl event are also discussed. This concerns extrapolations : from information pertaining to neighbouring ecological areas - at higher altitudes (alpine, and sub-alpine regions) or below the southern limit (i.e. in the hemiboreal and nemoboreal zones), and to hture time with respect to predictions of the behaviour of (3-137, based on results for OLD (i.e. from atmospheric nuclear weapons tests - mainly in the sixties) versus CHERNOBYL caesium. Beside the principal terrestrial constituents of the soil-plant-animal system, the BOREAL FOREST ECOSYSTEM will, for the present purpose, be considered to comprise the semi-aquatic and aquatic components pertaining to peat, open peat bog, and ground water. This implies that run-off from a catchment constitutes the main link between the terrestrial part considered here and the aquatic ecosystem proper. The need for - but also the paucity of - radioecological knowledge concerning the boreal forest became particularly apparent after the nuclear power plant accident in Chernobyl in April 1986. As a consequence several new projects were initiated, and of necessity the majority of these have for several years been focused on descriptions of the dynamic changes of the content of radioactive caesium in the environment. Boreal forest radioecology deals as a rule with rather complex systems. The post-Chernobyl studies have gradually entered the stage where the specific causes of the caesium behaviour may be approached. However, this is a relatively recent development with only few reports available as yet. Distribution patterns and caesium behaviour are therefore in this review mostly discussed in terms of general concepts such as input,

338 redistribution, and loss based on actual activity concentrations, rather than with reference to specific processes and interactions. The presentation follows as far as possible a chronological order of the events from deposition of radioactive caesium and through subsequent phases of its turnover. The chain of events In this review the event sequence is studied in time periods covering situations during: (1) the phase of initial deposition; (2) the early phase after depositiona, i.e. the period from the end of the main wet deposition up to mid-July - generally corresponding to the end of the period of dynamic growth in large parts of the Fenno-Scandian forest areas; (3) the first five years after deposition mainly corresponding to the period for which experience of the Chernobyl fallout is available; and (4) the subsequent period of long-term turnover.

1 A L

Event Phases for Caesium Fallout in a Forest Ecosystem

I EARLY PHASE1 [ FIRST FIVE YEARS 1 IFUTURETIMEI hours-days-weeks-months

-

y e a r s

-

d

e

c

a

d

e

s

THE INITIAL FALLOUT AND EARLY PHASES AFTER DEPOSITION Late April -the time for the accidental release in Chemobyl, 1986 - implies late winter in large parts of the boreal forest of Fenno-Scandia, merging into early spring south of this region. In the boreal zone with the exception of its northern and alpine parts, snow melting had already become significant in the last week of April. The main wet deposition of radioactive matter from Chernobyl, lasting about 3 days, started between April 27th and 28th 1986 in the order: Finland, Sweden and central Norway (Arvela et a]. 1987, Persson et al. 1986, Saltbones 1986). In southern regions of Norway (Ssr- and Vestlandet) and to some extent also in neighbouring

Fig.4.5.2 Distribution of areas where fallout of radioactive caesium,. originating - from the Chemobyl accident, occurred mainly by wet deposition (dark dotted) and exceeded about 2 kBq/m*.

339 areas in south-western Sweden, relatively extensive wet deposition occurred during May 4 k 9 t h (Backe et al. 1986).The gray areas in fig. 4.5.2 indicating where the cumulative deposition exceeds about 2 k13q/mZ-based on gamma radiation or deposition measurements (Arvela et al. 1987, SIS 1987, Backe et al. 1986, LindCn and Melander 1987) - coincide with those where wet deposition was prominent during the period April 27th - May 10th . Outside these areas wet deposition was less although occasionally - e.g. over most of Denmark (Aarkrog 1988a, 1988b) -relatively important in comparison to the contribution from dry deposition. The snow-covered area during the main wet deposition period generally corresponded to the boreal zones (dark grey areas in fig. 4.5.1), while the hemiboreal areas typically were free from snow. If "early measurements" mean those performed within some days (to at most a week after the start) of the deposition on the ground, or of the vertical distribution between canopy and ground, then there are rather few results documented. However, there exist several observations based on y-measurements with portable, mostly simple "hand"-instruments, which indicate very variable deposition patterns. There are also results from alpine and subalpine areas that may give clues to the role of various phenomena in the initial redistribution effects, particularly the snow-cover. The southern (mostly hemiboreal) parts of Norway and Sweden affected by wet deposition, had already entered into the dynamic phase of early spring, and in addition suffered exposure about one week or more after the other boreal regions. This very likely influenced the initial fate of the fallout over the forest ecosystems, particularly with regard to the different interception capacity in winter and spring of both canopy and understorey vegetation. It may furthermore be of importance for the relative efficiency of conifers and deciduous trees. These factors must be kept in mind when comparing results from the different regions. Thus the initial distribution patterns are expected to result from the composite effect of several factors, of which some can only have been operative on the fallout after it entered the terrestrial environment. Factors of known or potential importance in this respect are:

- wet or dry deposition; - short or protracted period of precipitation, and amount of rainfall; - occurrence of snow-covered surface at time of deposition; - the physico-chemical characteristics of the particulate carrier of radioactive caesium. Vertical distribution Distribution in snow

Apparently the number of measurements performed on radioactive caesium in snow profiles after the Chernobyl accident was very limited in the Nordic countries, and still less has been reported. However, some results are available concerning radioactive caesium in snow profiles sampled within the first week after the deposition in Sweden in 1986 on a coastline (Gustavsson et al. 1987) and an inland site (Nylen and Grip, private comm.) in Vasterbotten, where the snow-depth at the

340 time of sampling was about 0.3 and 0.5 m respectively. These gradients indicate qualitatively a fast redistribution of the initial fallout from an early concentration (the day after the main deposition, according to Nylen and Grip) mainly on the superficial layer of snow to a more homogeneous distribution three days later (Gustavsson et al. 1987) varying within a factor two from the surface to the ground. The latter sample was taken after heavy rainfall, and the snow was very wet. Distribution in soil Measurements of the distribution of radioactive caesium in litter and soil in the period "days to a few weeks" after the main deposition are also very few. However, analyses of the subsequent changes in soil profiles from various sites with different soil characteristics, as well as comparisons with the depth distribution of "old" caesium in the same samples, have been used to estimate the initial distribution and probable causes of the behaviour of caesium. The vertical transport of caesium ions, of relevance also for the behaviour of its radionuclides when present in soil, may result from: -

transport of ions by means of percolating water or by ion diffusion in the soil water;

-

transport of ions bound to colloids;

-

redistribution by uptake and translocation in soil, as well as leakage and production of litter from plants and fungi.

mechanical mixing or bio-turbation e.g. caused by the presence of earthworms or other burrowing species in the soil;

As most of the radioactive caesium generally still remains in the upper organic horizons in podzol soil several years after the deposition, it seems primarily to be a question of how superficially the early distribution became established. This is of importance among other with regard to a possibly early contribution from uptake in the root zone, particularly for vegetation pertaining to the litter decomposers. An understanding of the early depth distribution is also essential for interpretation of y-spectrometry in situ (Finck 1992, Edvarson 1991a). And - conversely - the increase with time of the "soil samplelin situ" ratios, based on soil sampling and series of measurements over the same areas with airborne equipment (Edvarson 1991b) or ground-based (Gustafsson et al. 1987) y-spectrometry, give evidence of slow but generally significant redistribution processes. Starting from the earliest measurements performed by SCAB (Mellander 1988) and Gustafsson et al. (1987) in Sweden in May-June 1986, these analyses cover the subsequent period up to 1989. The findings indicate that the distributions obtained from soil profiles sampled in boreal forest ecosystems one or more years after the Chernobyl event, essentially describe the situation prevalent after the deposition from about a month at least. The behaviour of the caesium fallout during the first month in the boreal forests is not well understood. At the few boreal sites for which data for this period have been reported - or from which some notable features about the early redistribution may be inferred from observations at a later time - the specific soil characteristics or climatic conditions limit the potential for extrapolations to other forest regions. There exist, for instance, some results for the Vest-Agder

34 1 fylke in Norway from sampling in 1987 (Haugen and Uhlen 1992 ), from which have been deduced a rapid penetration of 40 percent or more of the activity to a depth below the upper 4 cm. Except in peat, this layer otherwise contains most of the deposition - frequently more than 90% - irrespective of soil type. This is manifest in the above findings from Srarlandet to the Helgeland coast (Rudjord and Haugen 1989), as well as in those of others from all over the boreal zones in the Nordic countries. Meteorological data give evidence of high precipitation in the Vest-Agder area during the main deposition phase, indicating the importance of heavy rainfall for the rapid penetration. Similar observations have been made in Bavaria in southern Germany (Block 1990, Bunzl et al. 1989). Already 13 days after the fallout, Bunzl found a considerable fraction of the radioactive caesium deposition at depths below 5 cm. This rather fast penetration was ascribed to the precipitation concomitant to heavy fallout during April 30th 1986. Experimental studies (Bunzl and Schimack 1992) give evidence of half-lives around 1 min for the sorption of I3’Cs in the organic horizons. This rather slow sorption rate may explain deep penetration if especially heavy rainfall occurs in connection with the deposition.

Interception It is reasonable to assume the interception efficiency of wet deposition to depend on the amount of precipitation. Typically trees may be exposed to about 2-3 mm of rainfall before they lose their capacity to retain the precipitation without extensive throughfall. This is close to the level obtained at the Vindeln experimental forest in Sweden (Nylen and Ericsson 1989), thereby indicating almost optimal conditions for interception. However, the protracted input mediated by a particularly light rainfall (lasting for about 14 hours) probably made the initial distribution relatively homogeneous within the forest in comparison to what would result from more intense rainfall. Supporting this assumption is the fact that N y l h and Ericsson (1989) found no significant difference in the concentration of radioactive caesium in needles from different heights in individual trees or between the average values from different sampling sites in a Scots pine stand. Results pertinent to other sites, where the rainfall was heavier but more short-lasting (Shaw and Smith, private communication, Steines and Njistad 1992), indicate rather inhomogeneous vertical distributions in coniferous trees. However, the vertical distribution in pine trees after the early phase of dry deposition reported by Roed (1988) is uniform and resembles the distribution obtained after protracted wet deposition under conditions of very light rainfall according to Nylkn and Ericsson (1989). It is probable that the interception calculated for the latter case of rather fog-like wet deposition - when the wind may carry very small droplets to the ground below trees without vertical penetration of the canopy - underestimates the efficiency attained under conditions when the same total precipitation is a consequence of more intense rainfall (cf. fig. 4.5.3). The vertical distribution between the aerial parts, particularly the tree canopy, and ground shortly after the fallout in the Nordic countries may be derived (a) for WET DEPOSITION from data on the distribution of radioactive caesium between tree-living lichen species (on birch, pine and Norwegian spruce), and lichen or moss on the adjacent ground’surface (Steinnes and Njdstad 1992) [sampling time July 19861, as well as from a few snow profiles gathered under Norwegian spruce and Scots pine, and from adjacent open areas (Bergman et al. 1991) [April 30th 19861; and (b) for

342 DRY DEPOSITION by the interception in several tree species including birch, Scots pine and

Norwegian spruce (Melin and Wallberg 1991) [mid-May 19861. Results concerning the distribution between canopy and ground early in May or June 1986 have also been reported for Norwegian spruce in Belgium (Sombre et al. 1990) and Southern Germany (Bunzl et al. 1989).

Fig. 4.5.3 Relative interception of 137Cs by forests of different types, ages and stem densities during wet and dry deposition in the transition from late winter to spring. The relative interception efficiency reported for dry deposition (Melin and Wallberg 1991), or estimated from data on wet deposition (Bergman et al. 1991, Bunzl et al. 1989, Sombre et al. 1990). is shown in fig. 4.5.3. The data from Steines and NjAstad do not directly concern interception, but are related to it, as they describe the I3'Cs contamination retained on tree-living lichen. The relatively high interception capacity of the needles on conifers in comparison to the capacity for unfoliated birch trees is reflected by a three to four times higher content of Chernobyl 13'Cs in lichen living on birch than in those living on Norwegian spruce or Scots pine trees. On the coniferous trees a significant fraction of the radioactive caesium was probably fixed to needles before having the opportunity to get in contact with the lichen. Interception by needles thus seems to have prevented extensive contamination of the lichen. If therefore the proportions of the initial deposition retained by the tree-living lichen in e.g. birch and Norwegian spruce correspond inversely to the respective interception capacity, a measure of the relative interception may be inferred. The three to four times higher interception for conifers in comparison to birch, derived from the results of Steines and NjAstad (1992), is similar to that found by Bunzl et al. (1989) and Sombre et al. (1990) for wet deposition and by Melin and Wallberg (1991) for dry deposition. Studies directly of throughfall (Bergman et al. 1988, Nylkn and Ericsson 1989), or implicitly by the increasing fraction of the total initial fallout appearing as ground deposition (Bunzl et al. 1989) - discussed later in this review (cf. Distribution between trees, understorey and soil and fig. 4.5.13),give evidence of a considerable loss of radioactive caesium by that route, especially during

343 early phases. The data of Melin and Wallberg based on sampling within about three weeks of the accident therefore represent lower levels for the fraction of radioactive caesium initially retained in tree canopies. The data of Nyltn (in Bergman et al. 1992) based on sampling on April 30th 1986, i.e. the day after the main wet deposition event in the district of Vasterbotten, constitute case studies based on just a few samples, thus associated with considerable uncertainty. Stem density and age of the forest stand, which differ between the sites considered here, also affect the interception capacity. In addition, all these estimates of interception are indirect (including extrapolations from the lichen data from Steines and Njistad). The deposition rate in a Norwegian spruce forest was 20% higher as compared to an adjacent, open, grass-covered field after the Chemobyl accident (Bunzl et al. 1989). Similarly the total activity of 137Cs deposited by the global fallout of earlier weapon tests is higher by about 30% in this forest as compared to the grassland, because of the interception between Norwegian spruce needles and the aerosols ( B w l and Kracke 1988). This means that differences between the activity on surfaces under trees and on open areas, used for some of the interception assessments, are not solely due to interception in the canopy. Nevertheless, the conclusion is that the initial retention by interception is high in Scots pine and Norwegian spruce, but relatively low in the deciduous trees lacking leaves at the time of the fallout. Evidently a large fraction (probably more than 50%) of the radioactive caesiwn in the fallout over coniferous forests will be initially retained in the canopy under both wet and dry deposition at least under the circumstances prevailing in the Chemobyl case. Therefore the possible effects of the caesium content in the canopy (operating as a secondary source, leading to contaminating of the understorey vegetation by subsequent throughfall and litterfall) is expected to be of the same relative importance for both wet and dry deposition. Thus this assumption means that even under conditions when sources generated by dry and wet deposition in the forest canopy cannot be neglected, they would lead to similar influence on, e.g., transfer factors for components of the understorey vegetation - a circumstance worth considering in intercomparisons of results from different boreal localities in the Nordic countries. Aspects of early vertical transfer by rainfall and internal translocation Provided the understorey vegetation is free from snow-cover, it effectively intercepts radioactive caesium both under wet and dry deposition, apparently even in late winter or spring. Also in areas where the snow-cover was about 50 cm thick at the time of the main deposition after the Chemobyl accident, the understorey became contaminated to a great extent. Beginning in the first week of June 1986 the activity concentrations in birch and bilbeny (Vaccinium myrtillus) - from such initially snow-covered sites - were followed by sampling at the same localities with 14-day frequency over a period of about three years (Bergman et al. 1988). These time series give evidence of a fast decline with a biological half-life of at most 14 days in bilbeny during the first half of June. Considering the actual activity concentration in bilbeny on the first sampling date, the aerial parts of the understorey vegetation (excluding the fractions present in lichen or moss carpets) may be estimated to have contained around 5% of the total deposition of radioactive caesium, assuming a biomass density of 0.3-0.5 kg d.m. per mz. If the sharp drop prevalent at the beginning of June

344 reflects the main trend also during the preceding month- as we consider likely -the content in the understorey vegetation should have been much higher during the early stage. In the perspective of potential contamination of the forest floor one should furthermore include the often more than 50% (cf. above Bunzl et al. 1989, Sombre et al. 1990, Melin and Wallberg 1991, Bergman et al. 1991, Steines et al. 1992) of the deposition intercepted by the forest canopy. A considerable fraction of this source was transferred to the ground, i.e. the understorey vegetation including litter and often a moss or lichen carpet, particularly within the first two months after the initial fallout. The transfer was primarily mediated by throughfalling rain (Bergman et al. 1988, Nylen and Ericsson 1989). The dominant role of rainfall does not necessarily imply that vertical transport from the aerial parts by means of internal translocation or litterfall can usually be disregarded. However, in the early period after deposition of radioactive caesium it appears to be of minor importance compared to the transfer effected by rainfall. Detailed studies of changes within the first two months in the concentration of radioactive caesium in different year classes of needles in Scots pine (NylCn and Ericsson 1989) and Norwegian spruce (Wyttenbach et al. 1986, Tobler et al. 1988, Wyttenbach and Tobler 1988) show that translocation should be considerable between compartments (different year classes of needles, small twigs etc.) on a single branch. By comparing the ratio of '34Cs to J3'Csin the deposition and in early needle samples from Norwegian spruce trees, Block (1990) and Bunzl et al. (199 1) conclude that contributions from uptake of "old" 137Csfrom the root-zone are negligible. Based both on pre-Chernobyl samples from 1984 and the ratio of 134 Cs to 137Cs after the accident, the results in 1986 for the different year classes of Scots pine needles (NylCn: in Bergman et al. 1988) show that the contribution from "old" I3'Cs in the current year shoots was less than 5% at the beginning of June 1986. When corrected for changes in the isotopic ratio due to decay, the influx of "old" 137Cs to the new needles appears to proceed at a relatively steady rate over the first year (with the exception of a period covering needlefall in autumn). At the same time the absolute level of radioactive caesium in the needles decreases fast - the composite effect of leakage, contamination from other parts of the canopy and internal translocation- resulting in a net increase of "old" cesium of on average about 1% per week in comparison to the remaining component of Chernobyl caesium. This means that in those needles not present at the time of the fallout, and which consequently obtained their content of radioactive caesium by external contamination or internal translocation, the ratio in early June 1986 of "old" and "Chernobyl" caesium is about half that obtained on average, when integrated over the soil column (Bergman et al. 1991). The effective translocation between the compartments within the same twig should be sufficient to explain this approximately equal contribution from internal and external sources. Similarly Raitio and Rantavaara (1993b) maintain that the proportion of radiocaesium of soil origin is still relatively low in the needles several years after the deposition, since no correlation was observed between the radiocaesium concentrations in conifer needles and the underlying soil. There is also other evidence from later phases - discussed further under Dynamics of translocation and feed-back in the forest of only a small contribution from root-uptake still several years after the deposition (see for

345 instance Ertel and Ziegler 1991). Direct translocation between the root zone and the conifer canopy is therefore not expected to play an important role during the early phase after the Chernobyl event. Nevertheless, the movement of caesium by internal translocation may be rapid under environmental conditions favoring transpiration. Witherspoon (1964) studying cycling of Cs- 134 in white oak trees found radiocaesium in all parts of the tree crown 1.5 hrs after trunk inoculation. Downward movement in the trunk was slower than upward movement. As in the case of the overstorey, the importance of rainfall compared to internal translocation has also been confirmed for transfer from smaller plants to litter and soil during the fust days to weeks after external contamination (Whitford 1968). Whitford found that only about 4% of the amount of (37Cs applied to the foliar parts of the understorey in a mesic forest was absorbed and translocated to stem and roots; most of the contamination was transferred to the litter by rainfall.

Thus the vertical distribution established early in soil may seemingly be described as resulting from two steps: one very rapid, operating on the initial deposition directly in litter and soil; and one belonging to the subsequent phase, and governed by the relatively intense release during precipitation over the first months from secondary sources in the aerial parts of the vegetation. That the distribution profile may be affected at least by heavy rainfall, is indicated by sometimes relatively deep penetration directly after wet deposition (Haugen and Uhlen 1992, Block 1990, Bum1 et al. 1989). Lateral distribution A characteristic feature of the lateral fallout distribution after the Chernobyl accident is the

considerable variability, which was manifested already from the first measurements performed during or shortly after the initial deposition phase. Such spatial variation of deposition in forests and forest soils has mainly been explained by the nature, structure and age of the forest stands and by the thickness and nature of the humus (Ernst and van Rooij 1987, Bunzl et al. 1989, Block and Pimp1 1990, Guillitte et al. 1990). This effect thus seems to be valid, but probably results from different causes with respect to the pattern appearing on a small and large scale. Typically it may be observed on each successive level in a stepwise increase of the linear dimensions of the area under study (e.g. diameters of about 0.1, I , 10, 100 m, 1 km etc). Several processes - operating directly during or subsequent to the deposition event - are supposed to contribute to this distribution pattern. Radiecological research in the Nordic countries has dealt with various factors of established or assumed importance in this respect. These are for example: type of deposition (wet or dry); bare or snow-covered surface; flat or sloping ground; dense or sparse vegetation (particularly the capacity to retain caesium by adsorption during surface runoff or snow melt); type and current condition of the vegetation (e.g. absorption capacity of a poikilohydric

346 lichen carpet depending on its moisture content at deposition); "ecological" redistribution (uptake and translocation processes leading to lateral redistribution). These factors often appear interrelated when considering the observed effects. Studies have been made of the presence or absence of snow on the ground at deposition in particular. In some cases analyses of the resulting patterns have been made with regard to probable interactions with other factors. Effects of snow cover and snow melt

In areas assumed to be covered by snow in April 1986 (either upper parts of the subalpine region leading to higher altitudes, or the still snow-covered leeward slopes lower down at the subalpine level), a characteristic relation is found between the distribution patterns of radioactive caesium on bare and adjacent snow-covered sites. On wind-exposed and hence snow-free ground, the concentration of radioactive caesium in the vegetation and lichen carpet was generally about three times higher than on sites in the vicinity still retaining snow (Gaare 1987a, Gaare 1987b, Bretten 1991). Dry lichens have a greater capacity to soak up water than moist ones. The higher activity concentration found in lichens on snow-free sites, where the water content generally is relatively low - except during rainfall and a brief period thereafter - has been ascribed to this absorption potential of relatively dry lichens (Gaare 1987a). According to Gaare(l991) the coefficient of variation (i.e. the quotient of standard deviation to the mean) with regard to the concentration of radioactive caesium in lichen was about twice as high on mostly snow-covered sites as compared to snow-free sites. The relatively high variability found in vegetation on snow-covered sites resembles that illustrated in fig.4.5.4, obtained from soil sampled on open fields from the coast as well as at alpine altitudes (Haugen and Bj~irnstad1990, Haugen 1992), or in a boreal forest (Bergman et al. 1991) and open grasscovered fields in the southern boreal (Finck 1992), as well as the nemoral (Isaksson and Erlandsson 1992) zone. The results reported by these authors have been used to calculate mean levels of the variability for certain categories ("open area or forests free from snow"; "snow cover in forest or on open areas"; snow cover on peat"). It is these mean values that appear in fig.4.5.4. The consistently higher variability in the presence of snow cover, despite the very different type of ground cover (in Fig. 4.5.4 The variability in the distribution of radioactive caesium over terms of type and open areas and in forests with and without snow cover during the deposition. Standard deviation in percent of the mean.

347 density of vegetation) at the growing sites, strongly favours the supposition that the common factor (i.e. snow and snowmelt) is the primary cause of this effect. Finck (1992) sampling on equivalent open grasscovered fields - one with, one lacking snow cover - obtained principally the same dependence of the variability on the presence of snow at deposition, as on the other sites referred to in figure 4.5.4. Thus, fallout on snow and transfer of caesium during snow melt seem to be the two primary causes of a comparatively high variability in the lateral distribution. Discharge areas, partly water-covered soon after the onset of snowmelt, offer a quite different environment for a recent fallout than that at recharge sites, where the ground water level is below the surface. The results from sampling in such discharges areas on peat ground exhibit a relatively high variability similar to that found for snow-covered podzol ground, cf. "snow-cover on open peat areas " in fig.4.5.4. It is expected that this variability largely reflects the presence at the flooded site of small areas projecting above the ground water level. This assumption conforms to the fact that the coefficient of variation is smaller (33%) over the fraction consisting of open peat bog , i.e. completely "flooded" during the whole period of snowmelt and intense runoff (Bergman et al. 1991). This variabiIity is closer to that found on open areas free from snow than to that on the snow-covered sites. Interaction between snowmelt and secondary processes acting on the water borne caesium, such as adsorption on vegetation and surfaces encountered during the transport or runoff from the site of deposition, appear to influence the distribution on both small scale (e.g. areas with a diameter less than lm) and a relatively large scale ( diameter of the order of 100 m). Growing site and topographical characterisitcs seem to be of importance with respect to the net effect on the lateral redistribution of these interactions. Effects on the distribution pattern of inhomogenities in the ground surface structure have been investigated by N y l h (private comm) at a boreal forest site, Svartberget in VSisterbotten, Sweden, covered by about 0.5 m snow at the time of deposition of radioactive caesium. In areadefined soil and vegetation samples taken from protuberances caused by overgrown old treestumps, small boulders, etc., the remaining 13'Cs activity concentration per m 2was on average three times lower than on the adjacent lower ground. Such a distribution may result from a relatively low retention capacity of vegetation soaked by melting snow (thereby increasing the mobility during the early redistribution), and also agrees well with the findings of Bretten and Gaare discussed above concerning the different retention capacity of bare and snow-covered sites on alpine and sub-alpine levels. It is also evident from previous studies of radioactive fallout originating from atmospheric nuclear weapons tests (Osburn 1963) that the vegetation density is of significance, as regards the capacity to scavenge and retain fallout on snow at the time of snowmelt. This interaction may in fact result in quite opposite distribution patterns within topographically similar catchments, although they were all covered by snow at the deposition event - in some cases yielding a relative increase by more than an order of magnitude in the valleys due to transfer of caesium from surrounding hill slopes; in others leading to a decrease, due to loss by runoff, from the low ground constituting discharge areas in the valleys. This is apparent from the result obtained in a study of an

348 alpine grass-covered site at Bstlandet in Norway (Haugen and Uhlen 1992 , Haugen and Bjerrnstad 1990), and a boreal forest catchment at Svartberget in Sweden (Nylen and Grip 1989, Bergman et a1.1991). At Bstlandet runoff from the surface occurs only in connection with the snowmelt during springtime, and the capacity for the vegetation to capture the waterborne caesium is considered to be high. The remaining fraction of the radioactive caesium deposition was at that site much higher in the valley than on the surrounding elevated parts. In contrast, the ridges and slopes confining the catchment at Svartberget retained most of the initial deposit of radioactive caesium, despite a very intense snowmelt. However, the lower ground in this catchment, consisting of a peat-covered discharge area, lost about two-thirds of its content of radioactive caesium within the first ten days after the fallout. The findings of Haugen and Uhlen are based on studies performed some years after the Chernobyl accident, and accordingly lack information about the fraction of radioactive caesium lost by runoff during the spring of 1986. Despite the fact that a large amount of the radioactive caesium still remained at the study site several years later, these results do not apparently contradict the notion that a large loss may have occurred also in this case from that part of the low ground which became flooded during snowmelt in 1986. The relatively small variability found after wet deposition on open areas free from snow (cf. fig. 4.5.4)indicates that radioactive caesium became deposited rather homogenously during this process. Even if this was the general situation, some findings point to the opposite. Gaare (1987b) found that lichens planted in pots in the open air and separated from one another by less than 1 m contained activity concentrationsof radioactive caesium that differed by more than an order of magnitude after exposure to the wet deposition of Chernobyl caesium. Inhomogeneous deposition is the plausible cause to this result according to Gaare.

In recent years the knowledge has considerably increased of the different types of rainfall originating from meteorological processes in cloud formation (Simkiss et al. 1993). Clearly, rain that originates from different phases is likely to scavenge radionuclides with different efficiencies and deposit them at different rates. ApSimon et al. (1990, 1991) developed a simple dynamic model of the life cycle of a convective storm of the type that led to much of the deposition of radionuclides over Europe. These new analyses indicate that high variability in the deposition resulting from wet precipitation is quite probable, and are in agreement with the observations reported by Gaare. The general contention from all these findings from different parts of the Fenno-Scandian forest environment is that the main feature of the lateral distribution pattern prevalent during several years after the Chernobyl accident appears to have become established already at the time of fallout - in the case of deposition over snow-free ground- or otherwise during the subsequent short-lasting period of snowmelt and intense runoff.

349 Expected effects on early distribution by a particulate carrier

The initial carrier A particulate carrier may affect the resulting vertical and lateral distribution after deposition of radioactive caesium in various ways, e.g. by modifying the retention capacity at interception in canopy or understorey, as well as the mobility and availability for uptake in soil. Both the physicochemical characteristics of the carrier and the "aggressiveness"of the receiving environment (i.e. its capability to release caesium from the carrier by weathering or due to microbial activity) will be of importance in this respect. This interaction between carrier and environment should largely determine the effect on the distribution in different phases after a deposition, and for how long the carrier will influence the behaviour of caesium. During the fallout period after the Chernobyl accident, more than 75% of the radioactive caesium in rainwater sampled in southern Norway was bound to colloids or particles (Salbu 1988b). This means that a major fraction of the radioactive caesium was initially present in rather inert forms. The hot particles found in the Nordic countries, however, only contained small amounts of radioactive caesium and consisted mainly of other nuclides (Devell 1988a, 1988b, Salbu 1988a, 1988b, Salbu et al. 1990, Raunemaa et al. 1988). Most of the radioactive caesium was evidently carried by particles with diameters below lpm (Person et al. 1986, Devell 1988a, Georgi et al. 1988). The "availability"of radiocaesium has been analysed in terms of the fraction of caesium extracted from soil based on sequential extractions (Riise and Salbu 1988, Oughton and Salbu 1990, Riise et al. 1990), as well as the effectiveness concerning uptake in plants (Haugen et al. 1990, Salbu et al. 1992), or animals (Hove and Hansen 1992). According to Riise et aL(1990) about the same ratio is obtained between stable and radioactive caesium in the successive fractions from soil sampled in 1989, which indicates that equilibrium was attained sometime within the first three years after the Chernobyl event. The availability to plants of radioactive caesium from the Chernobyl fallout in the litter and uppermost layer of soil was compared experimentally 1989-1991 with that of Cs-134 in ionic form (Haugen et al. 1990). They found a relatively low uptake of "Chernobyl"caesium on the first sampling occasion 1-2 months after the addition of Cs-134 in ionic form to the soil. Only small differences were apparent between these two components of radioactive caesium after one year. Based on these observations these authors conclude that plant availability may largely be described as that of caesium introduced in ionic form - at least from about 3 years after the Chemobyl fallout. Similarly, concerning the transfer from feed to goats it was observed that the bio-availability of Chernobyl caesium was low in 1986 as compared to in later years (Hove and Hansen 1992). The relative change in availability agreed well with that found for the exchangeable fraction in extractions performed on plant material (Salbu et al. 1992). These observations indicate that the particulate carrier probably influenced the behaviour of the caesium at an early stage after deposition. It might therefore also be responsible for the relatively deep penetration observed for sites where deposition was concomitant to intense rainfall.

350 The results (based on the distribution pattern of I37Cs and stable caesium in the sequence of extractions used in the analyses) also indicate that for the Chernobyl fallout a "memory" of its origin lasted for at least some months; after about one year only small differences were apparent in the availability (as compared to that of caesium supplied in ionic form); and within 3 years equilibrium was attained between "old" 137Cs and stable caesium. However, at three natural mountain sites in Central Norway receiving almost the same fallout after the Chernobyl accident, but representing very different nutrient states, Varskog et al. (1 990) found that in 1989 there still remained differences in the activity ratios for Chernobyl and stable caesium in dwarf birch, which appear related to the site specific productivity. Varskog et al. interpreted these differences in the plant material as reflecting either a corresponding difference in depth distribution in soil, or a not yet completely established exchange between the stable and radioactive isotopes.

This difference in availability for Chemobyl caesium, in comparison to e.g. stable caesium, maintained over a period of months to perhaps several years, has generally been ascribed to the physico-chemical character of the initial carrier. If this is the true and sole explanation of the phenomenon, it is quite likely that the dynamics of the adaptation process, tending towards (but not necessarily attaining) complete isotopic exchange between 137Cs from Chemobyl and stable caesium, will depend on site-specific environmental factors - especially as the high surface area of the sub-micronic particles carrying most of the radioactive caesium would easily subject them to weathering and decomposition. In areas where the fallout should encounter a rather aggressive chemical or biological environment, the radioactive caesium is therefore expected to attain free isotopic exchange with stable caesium relatively soon - and vice versa where environmental conditions allow a longer life for the carrier. The different findings of Varskog et al. compared to those of Salbu et al. with respect to the level of isotopic exchange obtained within the first three years after the fallout could well fit such circumstances. However, concerning processes attributed a causal role with regard to different behaviour or distributions of stable caesium and recent fallout, there may be other possible candidates than a particulate carrier. Findings that may be interpreted in this way concern retainment of the early caesium contamination in organical structures, which release their content very slowly into circulation. For example, the last year-class of Scots pine needles directly exposed to the Chernobyl fallout contained in 1990 a concentration of I37Cs and 134 Cs about ten times higher than that of the younger year classes (NylCn private comm.). Evidently, such trapping of radioactive caesium very early after deposition may prevent its release into circulation for a long time, and thereby prolong the processes leading towards isotopic equilibrium. See further under Dynamics of translocation and feed-back in the forest. Resuspension Wind and running water mediate resuspension to various degrees depending on topographic, climatic, vegetational and geological factors. It is therefore reasonable to expect very different conditions for resuspension, e.g. on open wind-exposed areas typical of the alpine environment and in forests belonging to the boreal zone. Such differences are also apparent from various studies in Fenno-Scandia.

351 As interception was generally high in the forest canopies (cf. fig. 4.5.3),much radioactive caesium remained initially as contamination above ground. A considerable fraction of this caesium was lost from the canopy mainly by throughfalling rain during the first months after the deposition. Provided activity in the overstoray contributes significantly to resuspension, this fast decrease of the content in the aerial parts of the vegetation should be reflected in a corresponding decrease in the amount of resuspended caesium. Studies (see discussion below) based on resuspension sampling at several sites in Finland (Aaltonen et al. 1990) reveal this dynamic change over the period 1986-1987 and, furthermore, indicate that the primary source for the resuspended radioactive caesium is the local forest; i.e. contributions associated with long range transport of resuspended matter were too low to be significant in these measurements. In those areas of Finland subjected to relatively high deposition, the proportion of 137Cs originating from Chernobyl and present in the resuspension in 1987 totalled around 3% of the corresponding level in 1986. Aaltonen et al. also observed that there were no major differences in the concentrations between summer and winter at a site (Nurmijhi) where the ground is covered by snow during winter. As regards the dependence on presence of snow-cover, similar observations were made by Vintersved et al. (1991) for sampling in forest environments at sites in Kiruna and Umeb. Hence, resuspension fiom the ground does not explain the winter values.The resuspension factor of 137Cs at the location in southern Finland was estimated to be 4.10-9 m-1 in 1987, when the annual average air concentration was used. A comparison of the total amounts of deposited radiocaesium in 1987 at different sampling stations with those obtained at the same sites in 1986 reveals that the amounts were still highest at the stations at which they were highest in 1986. This indicates that most of the 137Cs was of local origin, being resuspended from the nearby environment. The maximum figures for monthly 137C.s depositions at different stations do not correlate with the amounts of monthly precipitation. Resuspension seems therefore to be associated with some meteorological reason other than precipitation - maybe wind or storm. Therefore, coniferous trees in the vicinity probably constituted a significant source of radiocaesium in the deposition samples. This in turn means that such resuspension data are likely to overestimate the role for resuspension with respect to transport over longer distances. Comparisons with other data as those made below with results found by Aarkrog concerning measurements on arable areas need to consider this aspect. About four years after the accident the resuspension level has almost reached that obtained for nuclear weapons fallout deposited more than 15 years earlier (Vintersved et al. 1991). Provided the caesium activity deposited in Sweden after the Chernobyl accident behaves in the same way as that in fallout from nuclear weapons, any further decrease is therefore expected to be small. Uptake of scandium is very small in vegetation. To the extent vegetation may contain scandium, external contamination is therefore the probable cause. This fact was used by Oughton (1 990) in an analysis of vegetation sampled in seminatural areas in 1989. No measurable quantities of scandium were present in the samples, which means that the potential for contamination of the vegetation due to soil particles is relatively small in this environment three years after the Chernobyl accident.

352 The relatively low resuspension on an yearly basis in the boreal forest after 1986 (Aaltonen 1990, Vintersved et al. 1991) seems to be consistent with the negative findings by Oughton concerning actual contamination of the vegetation. The roughly 3 to 5 times higher resuspension factors obtained at about the same time in Denmark (Aarkrog 1998a, 1998b) mean a resuspension of about 3% during November 1986 - October 1987. This is relevant for agricultural areas with considerable influence of wind-erosion, and does not contradict the findings by Oughton. However, the percentage of resuspended deposition with rain from July 1986 to December 1988 in the Faroes was 30% at Torshavn and 24% at Klaksvig (Aarkrog 1992). The cause of this relatively high resuspension is not known with certainty, but its magnitude is similar to that of resuspension at the alpine levels according to findings by Gaare (cf. below). The change with time of the activity concentration of 137Csin humus and vegetation in windward and leeward sites of alpine and sub-alpine areas in Norway exhibits markedly different patterns (Gaare 1987, Gaare 1990, Bretten 1990). On windward sites the '37Cs concentration decreases during the period 1986-1990 from an initially relatively high level in the part of the vegetation that is assumed to have been free from snow during the fallout after the Chernobyl accident. These initial activity concentrationswere higher the more wind-exposed was the specific site, and the rate of decrease seems to have led to a convergence of the levels. The initially high levels on windward snow-free sites were interpreted as reflecting the capacity for absorption of water during the main deposition (Gaare 1987), e.g. dry lichens will have a higher capacity to soak up water than moist ones. The convergence is, on the other hand, not explained by this qualitative difference in absorption capacity, but fits an assumed role for resuspension - especially accentuated during the first year - (Gaare 1990). The marked increase with time in humus and in the vegetation on neighbouring leeward areas conforms with this assumption. This indicates possibly two resuspension processes induced by exposure to the wind: one operating directly on the particulate matter deposited on the vegetation; the other transferring debris from the vegetation carpet to leeward sites during subsequent years. Resuspension of this order is not expected to be a common feature in redistribution in the boreal forest.

Loss by runoff Loss of radioactive caesium by runoff from catchments was studied in the boreal forest (Middle boreal zone in Sweden, Nylen and Grip 1989 ; southern boreal zone in Sweden, Sundblad et a1.1991) and at sub-alpine altitudes (Haugen and Bjrarnstad 1990 at 0vre HeimdalenNorway). In the Middle boreal zone NylCn and Grip (1989) analysed the runoff from a catchment (Nyanget) in the district of Vasterbotten beginning about one week before the accident in Chernobyl. Based mainly on radioactive concentration in 1986, 1989 and 1990 as well as waterflow for the whole period 1986-1990, the amount of radioactive caesium leaving the terrestrial compartment was estimated (cf. fig.4.5.5). The integral ground deposition in the 0.5 kmzcatchment and the loss by runoff, measured by respectively soil and water sampling, means that about 7% of the initial fallout was transferred from the catchment by runoff during 1986. The main part was lost already

353 within the first four weeks after the accident. During subsequent years this leakage decreased to about 0.2% per year. The discharge area including an open peat bog covers about 20% of the catchment. Caesium is generally relatively mobile in soil high in organic matter. Under the assumption that mainly peat areas - where the organic component is especially prominent - are responsible for the release of activity to the stream, the fraction lost per year during 1987-1990 was 1-2% from such areas. A relatively high loss from peat is also apparent from the different fractions of "old" nuclear weapon caesium remaining in podzol soil and adjacent peat areas (Bergman et al. 1991). In the southern boreal zone Sundblad et al. (1991) studied runoff at Hillesjon and Sllgsjon from the beginning of June 1986 and continuing to 1991. Bog areas cover 2% and 32% respectively of these catchments. The decrease in concentration of '37Cs from their first measurements in comparison to that in the period 1990- 1991 is about one order of magnitude in the inlet streams at both lakes.

Discharge of Cs-137 10 [% of total deposition]

1-

1

0.1

The change with time in the activity 1986 1987 1988 1989 concentration at Salgsjon exhibits a fast Fig. 4.5.5 The discharge of 137Cs by runoff from Nyiinget in percent of the total deposition after the leveling off to a relatively stable activity accident in Chemobyl (NylBn and Grip 1989). concentration already after a year. The findings at Nyiinget (see above), where bog covers a similar fraction in the catchment, conform to this pattern. At Hillesjon, however, where the bog area is very small, the fall in activity concentration appears to proceed relatively rapidly over the whole period for which results are reported. This probably reflects a gradually decreasing

Cs-137 transfer by runoff

[m2/m3] 0.I

0.01

0Runoff]kBqlm'Jper unit deposition I kBq/m'J in the Catchment in 1989-90

0Runoff [m21ma]per fraction of bog area in the catchment

0.001 0.0001

SlllgsjSn

Nyllnget

Fig. 4.5 5 Estimated average W s concentration in runoff during 1989-90 from three catchments: inlets to respectively lakes Hillesjon and Salgsjon (Sundblad et al. 1991) in the Southern Boreal Zone and outlet from Nyiinget (NylBn and Grip 1989) in the Northern Zone. Levels per unit deposition of 137Cs(grey), and per unit deposition and fraction of bog area within the catchment (white).

capacity for leakage of 137Cs from podzol soil in contrast to the situation for peat. These different types of response thus agree well with the assumption that peat bogs constitute the main source for loss of 137Cs from catchments in the boreal zones - at least over a five-year period after the deposition. The analysis of the fraction of "old" 137Csresiding in peat more than two decades after deposition at Nyhget (Bergman et al. 1991) indicates that this state will probably prevail far into the future. Figure 4.5.6 illustrates the output or

354 transfer of W s by runoff from the respective boreal catchments in the period 1989-1990 under two conditions. Grey areas represent the quotient of the activity concentration of '37Cs in the runoff to the estimated average deposition at each site. White areas are the same estimates per fraction of bog area in the catchments. An estimated average is used to recalculate the transfer out of the system in terms of concentration in the runoff water per unit deposition. This operation may introduce additional uncertainty into the results. Nonetheless, it is evident that the one order of magnitude difference in concentration of 137Cs in runoff water per unit deposition between the two "high peat" and the "low peat" catchments almost disappears, when the results instead are recalculated to express runoff per fraction of bog area. The very similar transfer capacity obtained in this way emphasises the probable role of the peat areas concerning the loss by runoff from a catchment. At the sub-alpine level, surface runoff from a small watershed on a mountain pasture, Ovre Heimdalen, was studied during snow melt in spring 1989 (Haugen and Bjramstad 1990). The area was mainly snow-covered during the deposition in 1986. As the ground is frozen at this time, melt water mostly will be lost by runoff from the surface. There are no peat areas within the studied catchment. On sampling in 1989, 54-74% of the activity content in meltwater appeared in the high molecular weight fraction (> 2 nm). The particles present in the snow can plausibly adsorb a large part of the radioactive caesium present in ionic form in the snow. Direct measurements of the thickness of the snow-cover were lacking. Based on the amount of meltwater resulting from winter precipitation in a normal year, the loss of '37Cs by runoff during snow melt in 1989 was estimated to fall within the interval 0.01-0.1% (Haugen and Bjemstad 1990). This indicates that only a small amount is lost per year. Therefore, the low concentrationsof 137Cs found on ridges an slopes as compared to the lower parts of the catchment must be due to transport during previous years and 1986 in particular.

CAESIUM-137 IN A LONG-TERM PERSPECTIVE Distribution in soil during the first five years

The distribution of radioactive caesium in forest soils generally exhibits great variability. In studies south of the boreal zone this has mainly been considered to reflect the nature, structure and age of the forest stands and the thickness and nature of the humus (Ernst and van Rooij 1987, Bunzl et al. 1989, Block and Pimp1 1990, and Guillitte et al. 1990). The conservative behaviour of the forest ecosystem on the boreal latitudes, as regards retention of radioactive caesium, is apparent. The main features of the distibution usually become manifest within the first year. In particular, "penetration"or net changes by redistribution in soil is generally small or not significant in boreal forests after this early period, whereas protracted processes operating over the order of decades markedly affect the depth distribution south of this area. Concerning caesium, the dynamics and order of importance of the different redistribution processes are probably not the same under alpine, boreal or nemoboreal conditions.

355 Distribution patterns in podzol andpeat:

The presence of snow at the time of deposition evidently increases the variation in the lateral distribution prevailing in soil after the subsequent snowmelt (cf. fig. 4.5.4). Dependence of the spatial variation of deposition in the forest soil on vegetational characteristics, e.g. open or forested areas, type, age and stem density of a forest stand is expected to be important also in the boreal zone. However, in the Chemobyl event the deposition in the boreal zone was subject to the extreme dynamics associated with intense snowmelt, which may explain why dependence on the lateral distribution on snow generally dominates (Bergman et al. 1991, Raitio and Rantavaara 1993b). The distribution in various coniferous stands of different ages at the forest research station at Vindeln, Vasterbotten, (Bergman et al. 1991) according to table 1 and the distribution of 137Cs in all the soil samples from these sites (cf. fig. 4.5.7) illustrate a typical pattern in boreal podzol soil. In general, the sites show quite a uniform distribution. Density distribution However, there is a ten35 1 [%] dency towards particularly 30 .. high variability at sites (cf. 25 Podzol soil table 4.5.1) with sparce .20 understorey growth and a 15 .low stem density, often lo associated with relatively 5 c , , ? poor soil conditions. 0 5 10 15 20 25 30 35 40 45 50 55 60 65 70 115 I20 Skewed distributions simi0-137activity concentration (kBq/&] lar to those in fig. 4.5.7 Fig. 4.5.7 Density distribution of I T S in samples of podzol soil over are common at podzol an area of average activity 26 kBq/m2 ; r~ = 9 kBq/mz. sites studied all over the (Bergman et al. 1991). Nordic countries.

L

Caesium deposited on peat, on the other hand, encounters different environmental conditions resulting in a relatively high mobility - in turn leading to deeper penetration in the soil column, and potentially to considerable loss by runoff. The latter is evident in the results from discharge areas where ground water is level with or occasionally above the peat surface. It has become particularly pronounced at sites which were partly water-covered soon after the onset of snowmelt in the year of the Chernobyl accident. in peat in fig. 4.5.8 is based on results from sampling in the same 0.5 The distribution of kmzcatchment, and period, 1989-1990, as in the case illustrated in fig. 4.5.7. Results from this catchment concerning radiocaesium content in vegetation and soil, as well as detailed data on its loss by runoff - which are available for a period beginning one week prior to the Chemobyl accident - indicate that the deposition was of similar magnitude over both podzol and peat areas in the whole catchment (Bergman et al. 1991).Nevertheless, the average remaining activity in peat is about one third of that in podzol soil (cf. fig. 4.5.8). The main reason for this is thought to be a fast loss by runoff within the first weeks after deposition (Nylen and Grip 1989).

356 Table 4.5.1. The 137Csactivity concentration (kBq/m*)in local study sites at Svartberget estimated from measurements on soil sampled in 1989-1990 (corrected for physical decay from April 29 1986). The values include the components from both atmospheric nuclear weapons tests and the Chernobyl accident (Bergman et al. 1991).

Site description

Average activity

(kBq/mz) Young Scots pine forest Young Scots pine forest Young Scots pine forest Old Scots pine forest Old Scots pine forest Old Scots pine forest Norwegian spruce forest Mixed pine and spruce forest

28 21 37 ** 28 20 26 22 23

n*

std

(kBq/mz)

8

7 20** 12 4 7 6 7

14 19 17 20 20 9 20 19

............................................... All the podzol areas:

26

9

138

Forest covered peat aeas Open peat bog

11 6

6 2

16 15

All the peat areas:

9

6

31

*) n is number of measurements, each measurement is made on a homogeneous mixture of 5 samples. **) site with particularly sparse understorey growth, low stem density and thin organic soil layer. The amount of 13’Cs fallout from nuclear weapons testing remaining in the peat bog is relatively small (about 15% ) as compared to the levels in non-peat areas of the levels in non-peat of the catchment about areas 20 years after the catchment about 20 years after the is one of main deposition. This .30 peat ’Oil main deposition. Thissupporting is one of the several indications several indications supporting theto assumption that deposition due 20 -assumption that deposition due tests to atmospheric nuclear weapons atmospheric nuclear weapons behaves essentially similarlytests to 10 -behaves essentially to the fallout from thesimilarly Chernobyl the fallout from the Chernobyl accident - at least in the boreal accident - at leastetinal. the199 boreal 0 1, zone (Bergman Bergman et al. 1992). The estimated loss over long time periods, assuming a similar fate for the two types of fallout, is of the order of 1% per year from

357 peat, which equals the leakage rate found for the Chemobyl caesium over the whole period, except in the first year (cf. fig. 4.5.5).

Comparison of pre- and post-Chernobyl patterns "Old" versus "Chernobyl"caesium in soil

A comparison of the vertical distribution of "Chernobyl" and "old' fallout caesium based on samples from the same site should reveal cases where slow processes significantly affect redistribution over a time span of one to several decades, i.e. the time spent in the ecosystems by most of the "old" fallout of I3'Cs. Data on the vertical distribution satisfying this condition, and pertinent to soil of podzol type, are available from forest environments both in the Nordic countries and in central Europe. The depth of the organic layers differs both within and between sites. It is frequently not more than 5 cm in the boreal samples but generally increases in thickness south of this zone. In some cases sampling took place within a few weeks (Block and Pimp1 1990, Bunzl et al. 1989), about six months (Gustavsson et al. 1987), or one to several years after the Chernobyl accident. A few samples are based on pre-Chernobyl data (Melin and Wallberg 1991). The authors have not presented the depth distribution in exactly the same way. However, it is possible to get an approximate estimate of the amount residing in 1) the rootzone and humus; 2) about 0-3 cm of the mineral soil; and 3) mineral soil below about 3 cm. Distributions in podzol soil - in terms of the integrated amounts of "Chernobyl" and "old" fallout caesium over the organic layer and uppermost 3 cm of mineral soil are illustrated in fig. 4.5.9 for different sites arranged according to the vegetational zonation: Middle Boreal (Nyltn private comm., Gustavsson et al. 1987); Southern Boreal (Rantavaara and Raitio 1993); Hemiboreal (Rudfjord and Haugen 1989, Olsen private comm., Melin and Wallberg 1991); Oceanic boreal (Selnaes and Strand 1992); Nemoral Denmark (Strandberg 1992); Figure 4.5.9. Distribution in soil of radioactive caesium from nuclear weapons tests ("old" fallout) and from the Chernobyl release Nemoral Germany (Bunzl ("Chemobyl" fallout). Approximate content (percent of total in the et al. 1989,Block and whole soil column) in the upper layers (whole humus and root zone + about 3 cm of mineral soil) in coniferous forests on soil of podzol type Pimp1 1990). belonging to different vegetational zones.

Provided vertical transport in soil is a sufficiently dynamic process, the distribution obtained by sampling at different times after deposition should exhibit a significant time dependence. For caesium deposited over the boreal versus the nemoral regions, the long-term behaviour is evidently quite different (cf. fig.

358 4.5.9). In the boreal zone the distribution of 137Csremaining in the environment since about the middle of the sixties has not changed significantly from that established already within the first year. Furthermore, a high level of Chernobyl caesium in a single sample is frequently accompanied by a high level of old caesium, despite the considerable variability between different samples from the same site (Bergman et al. 1991). This exhibits a very conservative behaviour with an apparently static, or steady state, depth distribution of '37Cs in the boreal podzol soil, contrary to the situation in the coniferous forests on podzol type south of the boreal areas. The distribution in peat indicates a faster and deeper penetration than in podzol soil. Yet Gustavsson et al. (1987) have shown that more than 95% of the "Chernobyl " caesium frequently remains in the upper 5 cm six months after deposition. At this site, Finnsjon in the Middle Boreal Zone, the distribution in peat of "7Cs did not differ significantly from that of 134 Cs. Thus already within the first six months after deposition, the depth-distributionof the recent fallout appears to approach that of caesium residing in the peat for decades. Similar conditions prevail also in peat outside the boreal zone. For instance Cawse and Horrill(l986) found that 20 years after the cessation of nuclear weapons testing, 90%of the radiocaesium remained in the top 15 cm of two peat soils in the UK. A relatively high loss through leakage and runoff seems to occur over many years from peat (Bergman et al. 1991, see also fig. 4.5.6 and its text). This may agree with the findings of Gustavsson et al., if loss does not change the depth distribution - e.g. because the same proportion is removed from all depths, or the redistribution of the remaining fraction is fast and essentially completed during each subsequent growing season. Many observations confirm the prevalence of the latter factor.

Distribution between trees, understorey and soil Interception of the initial distribution in the tree canopy is evidently effective, although dependent on species, age, and - particularly for deciduous trees - season. The interception capacity of a forest, however, also depends on the stem density. Under the circumstances prevalent during wet and dry deposition early after the Chernobyl accident, often more than half of the radioactive caesium fallout was retained in the canopy of coniferous forests (cf. fig. 4.5.3).This is also reflected by the fact that the radionuclides from the Chernobyl fallout deposited per mz soil surface immediately after the deposition period were higher by a factor 2.5 in grassland as compared to forest soil (Schimack et al. 1989). The content in the canopy was subsequently removed from the needles and twigs as a result of weathering (rain, wind, litterfall) and transferred to the forest floor, but only rather slowly. Half-lives in a Norwegian spruce canopy (Bunzl et al. 1989) in southern Germany are estimated to be 90 days for the period 0-130 days, and 230 days for the period 130600 days. Tobler et al. (1 988) in Switzerland estimate the half-life to be 175 days for needles sprouted in 1985 and 1 15 days for twig-wood within the period 50 - 240 days after the Chernobyl accident. From data on throughfalling rain and litterfall in a Norwegian spruce stand (Bergman et al. 1988) a corresponding half-life of 140 days is estimated for needles over the period 35 - 180 days at a boreal latitude. In a Scots pine forest (Bergman et al. 1988) half-lives of 114 days are estimated for the period 40 - 200 days, and 250 days for the period 200 - 360 days.

359 The transfer to the forest floor by litterfall appears to be small compared to weathering by rain or wind. About 7% is lost by litterfall during the first year with respect to the total activity in the canopy in a Norwegian spruce forest (Bunzl et al. 1989). Similar levels are found in Scots pine forests (Bergman et al. 1991). These results imply that the transient phase of loss of the intercepted fallout in the canopy is essentially completed after one year, as the main part of the radioactive caesium transferred by the route of weathering has already entered the forest floor. The distribution between trees, understorey and soil after this phase ( i.e. from 1987 or later concerning Chernobyl caesium ) is illustrated in fig. 4.5.10. It is apparent that results concerning the behaviour of 13'Cs from atmospheric nuclear weapons tests ("old" caesium) - at least with regard to the site in Sweden (Melin and Wallberg: hemiboreal; pre-Chernobyl sampling in 1980) and that in Finland (Raitio and Rantavaara: Southern boreal; comparison of distributions of Cs and I3'Cs) - are very similar to those of "Chernobyl" caesium. Most of the activity resides in soil. The distribution between trees, understory and soil in the boreal zone reveals a notably unchanging pattern over several decades. South of this zone, however, the distribution between different horizons (cf. Fig. 4.5.10 The inventory of radioactive caesium in scots fig. 4.5.9) -and seemingly also the pine and Norwegian spruce forests distributed over content availability for the living matter - may in trees, understorey (including mosses and lichen) and differ, depending on the time that the soil. caesium has spent in the environment and on site-specific conditions. The generally high coverage of mosses and lichen in the boreal ecosystems captures a considerable part of the initial deposition and of the subsequent transfer to the ground by throughfalling rain. Mosses and lichens therefore constitute secondary sources. From these, radioactive caesium is released rather slowly to the environment with half-lives often in the range 2 - 10 years, depending on the biotope and the species constituting the ground carpet. Concerning the extensive ground coverage by mosses and lichen, the forest stand studied by Raitio and Rantavaara (1993a, 1993b) is representative of various boreal biotopes -on podzol type from low to rather high productivity. At this site the amount remaining at ground level -"understorey, mosses and lichen" in fig. 4.5.10 - is one order of magnitude higher than at the other sites in a hemiboreal (Melin and Wallberg) or

360 nemoral region (Strandberg, Block). However, 13% of the radioactive caesium remains in mosses and lichen according to Raitio and Rantavaara, while the understorey in this case contains about 1%. The levels found in the understorey at the other sites in fig. 4.5.10, where the coverage of mosses or lichens is sparse - if at all present - are close to that value.

Dynamics of translocation and feed-back in the forest The current distribution patterns of radioactive caesium on various growing sites (cf. figs 4.5.9 and 4.5.10) exhibit the composite and dynamic effect resulting from several interacting factors. Other indications of the influence from different processes may be derived from the appearance of similar or dissimilar distributions among the different species in plant communities at a growing site, and the ratios of "old" and "Chernobyl" caesium in the various components. Major factors in this respect are: direct contamination due to primary interception (and subsequent secondary transfer); translocation; uptake from the root zone; and growth. The order of importance for vertical transfer by rainfall and internal translocation with regard to the early phases was discussed in Aspects of early vertical transfer and internal translocation. The similar trends towards lower concentrations of I3?Csin Scots pine, birch, and bilbeny shown in fig. 4.5.1 1 are based on pooled data from regular sampling in Sweden at 10 different sites in the Vindeln experimental forest (Bergman et al. 1991). At the ten sampling sites covering about one hectare each, ten separate subsamples (for statistical assessment of the variability within and between sites) were taken regularly twice a year (July and October) in the period 1986- 1990. The sites comprise three in mature mesic Norwegian spruce forests, four in dry Scots pine forests, and four in young forest in areas that were clear felled within 5-10 years before the Chernobyl accident. In order to focus on the principal feature in the change over time, the level of 137Cs in the figure is normalized to initially ten units in 1986 for each plant species. Similarly illustrated is the concentration of I T S in milkweed (Epilobium angustifolium) during 1986-1989 based on pooled data from samples in July (for details cf. Bergman et al. 1991, Nelin and Nyltn 1993). Birch, Scots pine, bilbeny, and milkweed constitute perennial "key" plants with respect to their importance in food-chains to man (via moose or berries) and their abundance in the boreal forest ecosystem. They also represent specific categories of vegetation exposed under qualitatively different conditions during the wet deposition: The pine, because of effective scavenging by its needles, captures in its canopy much of the caesium transferred during wet deposition (cf. fig. 4.5.3). The birch, on the other hand, represents a category of directly exposed vegetation with a comparatively low capacity to retain the initial wet deposition. Apparently this reflects the small surface area of its canopy, because of the lack of leaves during the time, when the deposition occurred.

361 The dwarf shrubs were completely covered by a thick (> 50 cm) layer of snow at the time of deposition. Bilberry therefore represents a category of understorey vegetation not directly exposed during the initial wet deposition, although it became thoroughly contaminated during snow-melting in the subsequent week. The milkweed contains no aerial parts during the winter season (in contrast to the three species above). It belongs thus to the type of vegetation that became exposed indirectly during the growing season subsequent to the deposition. The level found in the aerial parts in July 1986 reveals a fast uptake from the root-zone to maximum concentrationalready the first summer. Consequently the fraction of Y k that entered into circulation (and not remaining as elative Cs-137 ctivity concentration a potential source after deposition per of d.m. on e.g. the lichen or moss carpet) appears to have a very fast turnover in the ecosystem.

As evident from fig. 4.5.1 1, Scots pine, birch, bilberry and milkweed exhibit a similar pattern with a slow decrease in I3'Cs concentration during the first 5 years after deposition over podzol-type areas. This is true despite the very different Pine [Year] 0 : exposure history, the different 1986 1987 1988 1989 1990 ecological conditions for big trees visavis dwarf shrubs in the Fig. 4.5.1 1. Concentration of l37Cs during 1986-1990 based on understorey, and the proportions pooled data from samples in July (milkweed, bilberry and birch twigs) and October (bilbeny, birch and pine twigs). For of 137Cs initially retained on and in their aerial parts. each species the level in summer 1986 is set at 10 units. Standard errors of the mean are illustrated by bars. The main trend towards decreasing levels in the vegetation is a common observation in the boreal zones, although the distribution over long time periods in soil is apparently unchanging (cf fig. 4.5.9). In Norway, for instance, Haugen (1989) and Staaland et al. (1990) reported a decrease of generally about 60-80% during 1986-1987, and after that period about 20% annually. Nonetheless, there are several exceptions exhibiting increasing concentrations with time in vegetation, particularly at sites with relatively low primary productivity, as on peat in discharge areas ( Rissanen and Rahola 1989, Naemann et al. 1990, Staaland et al. 1990). See further under Principalfeatures of an explanatory model for a borealforest.

-

Furthermore, there is certain evidence of partial retainment of the early caesium contamination in organic structures that release their content very slowly into circulation. For example, the

362 last year-class of Scots pine needles directly exposed to the Chernobyl fallout contained in 1990 a concentration of 137Csand n4Cs about ten times higher than the younger year classes (NylBn private comm.). A considerable fraction of the initial contaminationthus still remained trapped within the needles several years after the contamination. Despite this, translocation appears to have been considerable between the different year classes already at an early stage for that part of the contamination which had entered into internal circulation, i.e. constituting "bio-available" caesium (Bergman et al. 1988). This type of I' trapping" of radioactive caesium very early after deposition, may evidently prevent its release into circulation for a long time, and thereby prolong the processes towards isotopic equilibrium. Such an effect is expected to be particularly pronounced, when turnover by litter production and decompositionis slow, which is often the case under poor soil or climatic conditions in boreal and alpine areas. It is probable that differences in turnover rate andor such trapping phenomena have influenced the distribution and isotopic ratios found, for instance, at Dovrefjiill in Norway about three years after the Chernobyl accident. Varskog et aL(1990) found that there were no evident differences in the levels of stable caesium in dwarf birch and dwarf willow at two sites, while the level of 137Cs in these species decreased markedly in relation to the soil status, and was highest on the site with poor soil. Uptake in coniferous trees of radioactive caesium from the root-zone has been specifically studied - or information concerning this mechanism may be derived from various studies - in the Nordic countries (NylBn and Ericsson 1989, Rantavaara and Raitio 1993a, ibid. 199313, Strandberg 1992), as well as south of this region in Europe (Block and Pimp1 1990, Bunzl et al. 1989, Ertel and Ziegler 1991, Schimack et al. 1988). Some of these results have already been discussed (cf. Aspects of early vertical transfer by rainfall and internal translocation) in connection with the distribution of radioactive caesium in the early phase. Concerning Norwegian spruce, the data found by Schimack et al. (1988) show that Cs-134 was present in comparatively high concentrationsin needles and in twig-wood, which had already sprouted by the time of the radioactive fallout. Needles which sprouted in 1986, i.e after the fallout period, also contained 134 Cs, but at much lower concentrations.Because the 137CsP34Cs ratio in all needle samples had, within experimental error, the same value (1.75) characteristic of the Chernobyl-derivedfallout, they concluded that the 134 Cs in the needles from 1986 was not transferred to the needles as a result of root uptake, because this ratio is higher in the soil of the root zone. The presence of 134 Cs in the needles 1986 is either the result of Cs uptake by the older needles and subsequent transfer within the tree, or the result of the transfer of activity from the older, highly contaminated needles and twigs to the younger ones by weathering processes (rain wind); in fact, both mechanisms probably occur simultaneously. Similarly Ertel and Ziegler (1991) showed that only about 5% of the radiocaesium in the more highly contaminated branches of larch was derived from root-uptake 2.5 years afier the accident. Since no correlation was observed between the radiocaesium concentrations in conifer needles and the underlying soil, it can be presumed that the proportion of radiocaesium of soil origin is still relatively low. The observations from the boreal zones are less conclusive concerning the role of root-uptake in trees. In the middle (NylCn and Ericsson 1989) and southern (Rantavaara and Raitio 1993b)

363 boreal zones the distribution of Chernobyl caesium and fallout caesium is not significantly different in the soil horizons and needles sprouted after 1986 in mature Scots pine stands. Retranslocation of potassium is known to occur - and therefore probably takes place for caesium too - from older year classes to the new. Nonetheless, uptake of radioactive caesium from the root-zone seems after some years to contribute sufficiently to dilute the sources in the canopy and trunk resulting from direct contamination from the deposition of Chernobyl caesium. Furthermore, twigs sampled in 1984 from three different Scots pine stands with trees of age 58,9, or 5 years indicate that Scots pine trees in existence before the fifties (i.e. earlier than the start of fallout from nuclear weapons tests) have in the order of 30 times higher levels of Ij7Csin the twigs, as compared to plants of the same size, but less than 10 years of age in 1986 (Bergman et al. 1988). Whether the difference in concentration of I3’Cs mainly depends on the soil characteristics,on effective retention after deposition directly on old trees, or on cumulative retention as a function of the age of the tree is not clear. Nevertheless, it exemplifies the fact that, in neighbouring areas with almost the same deposition of 137Cs, the levels in vegetation might differ very much in future time. The studies by Strandberg (1992) in a Scots pine forest in Denmark reveal notable differences in the ratios of old fallout versus Chernobyl caesium in various compartments at the growing site. About 29% of the total inventory of 137Cs in this forest ecosystem in 1991 originated from the Chernobyl accident. Figure 4.5.12 Fractions of old and Chernobyi caesium [YO] illustrates the distributions for 137Cs IW occuring after about 5 years (Chernobyf)in 1s comparison to that present for about 25 so years (Old). The concentration of caesium 25 in end shoots of Scots pine at this site is 0 four times higher than in the leaves of the birch pine shrubs litter soil soil birch Strandberg supposes that this leaves end(0-5cm) ( e c m ) shoots indicates a higher rate of uptake for Scots 0 Chernobyl 0 Old pine than for birch. However, other

L

interpretations may also fit these findings. Fig. 4.5.12. The percentage of radioactive caesium As interception is lower in unfoliated birch present in various components of a forest ecosystem in 1991, originating from fallout after nuclear weapons than in ‘Onifen (cf. fig. 4’5’3)yuptake Of tests (old)and the Chernobyl accident (Strandberg radioactive caesium from the root-zone in 1992). comparison to the fraction adsorbed after direct contamination of branches and twigs may change the isotopic ratio more by dilution after influx to the birch leaves than to the pine needles. The higher initial retention of the deposition in the pine canopy and subsequent retranslocation from the directly contaminated pools - with a high WW37Cs ratio - in twigs and needles to the current year shoots would agree well with its known interception capacity, the findings illustrated in fig.4.5.12, and the role of internal translocation apparent in the results from other studies on Norwegian spruce (Wyttenbach et al. 1988, Tobler et al. 1988) and Scots pine (Nyl6n and Ericsson 1989).

364 In conclusion, root-uptake south of the boreal zones seems still to have minor or insignificant effects on the content of radioactive caesium in the tree canopy several years after deposition, in comparison to the contribution from the fraction retained at the initial contamination. In the Boreal zone uptake from the root-zone is manifest as a fast process in understorey vegetation not directly exposed to the radioactive deposition. Concerning recent deposition in mature Scots pine forests the influx of caesium from the roots to the needles is probably not the major cause of the observed isotopic ratios and their changes with time. Nonetheless, root-uptake is the probable candidate for the approach to equal 134 Cs/'37Cs ratios which has been found in humus, understorey vegetation and trees of the boreal zones within the first five years after the Chernobyl accident.

Theoretical presentation of caesium behaviour in the complex forest ecosystem The boreal forest ecosystem constitutes an ordered but complex entity. The behaviour of caesium, resulting from the interdependencies and effective interactions within the community, probably cannot be comprehended by constructing a theoretical model based on a detailed pattern of these interrelationships. Theoretical treatments are therefore frequently based on DESCRIPTIVE MODELS. These models attempt only to describe a set of observations in mathematical form, for example, by fitting a curve to a set of points. No explanatory mechanism is built into this model, although the model itself may be used to suggest possible mechanisms. The histograms, curves and estimates of various half-lives in the preceding text exemplify such applications. If the emphasis is on attempting to explain observed data in terms of more basic known mechanisms, and on showing the prinicipal trend in the dynamic behaviour e.g. of radioactive caesium in a forest ecosystem, use is made of the EXPLANATORY MODELS. An example of this category, focused on redistribution processes in a long-term perspective, is dealt with below. In this model (Bergman et al. 1992) the major regulators of energy flow, as well as of caesium turnover, are related to primary production and its constraints on the growth capacity. Certain fundamental physiological processes governing the metabolism of living matter in the biotope are also considered. Explanatory models are thus not primarily used to make precise predictions, in contrast to PREDICTIVE MODELS. Models of the latter type, simulating ecological systems, tend to be very complex and generally need to be based on a network including several compartments. However, the three functions of description, explanation and prediction are usually not clearcut and most modelling efforts comprise some aspects of all three. Principal features of an explanatory model for a boreal forest

An explanatory model was developed based on compartment theory and first-order kinetics for the turnover of caesium in the boreal forest (Bergman et al. 1992). The analysis is focused on testing the hypothesis that: "the 137Cs present in a boreal forest tends towards a homogeneous distribution among the living cells of that system". This hypothesis is mainly based on physiological

365 characteristics concerning transport of caesium over cell membranes and intracellular distribution, and the apparently conservative conditions prevailing for caesium in boreal ecosystems - e.g. the facts that very little of the radioactive caesium deposited over the forest area is lost from the system by run-off (cf. fig. 4 . 5 3 , about 90% of the total deposition of I T S occurs in the upper organic horizon in podzol areas (cf. fig. 4.5.10), and that the availability in the ecosystem, as can be seen from the 137Cs concentration in moose meat, was not significantly different in 1985 (i.e. prior to the Chernobyl accident) compared to the period 1986-1990 (Bergman et al. 1991). The primary purpose of applying this model was to elucidate qualitatively how predictions based on this hypothesis correspond with the main features of the time-dependent change of 137Cs activity according to measurements on perennial vegetation (cf. fig. 4.5.1 1). The model includes qualitatively effects of primary production and growth on turnover of caesium. The dependence on these factors is concluded from the following facts: primary production and its distribution over growth and litterfall constitute major regulators with regard to the dynamics of the redistribution processes of organic matter in the forest (Lundmark 1986). The same conditions should be true for redistribution effects on potassium due to its essential role in the living cell. Potassium and caesium are to a high degree exchangeable in active transport over cell membranes in living tissue (Guering and Wallon 1979). Evidently both elements may serve in the same vital processes. Accordingly, as primary production is of importance for the behaviour of potassium in the forest ecosystem, it should be so for caesium too. The principal compartment model structure is based on the actual results for the time-dependent T hrarghfall transfer of i3?Cs from secondary sources in a Scots pine canopy by throughfall and needlefall (cf. fig. 4.5.13), in addition to the release to the environment of 137Cs deposited over the moss and lichen carpet. Loss from the system by runoff, -. .... ......-....... ..................__ ........... Year from 1987 and onwards, is less than 0.001 I 1986 1987 1988 1989 1990 199 that due to physical decay, and therefore disregarded in the model. Fig. 4.5.13 Rate of 137Cs transfer per year from the canopy and by runoff. The fraction transferred is normalized to the The model also includes: a "competitor" compartment (i.e. initial deposition (Bergman et al. 1992) indicating the increase in biomass competing for the available caesium) to simulate influence on the redistribution processes of primary production and growth; perennial vegetation; litterfall from this compartment; decomposition in a litter compartment; and exchange of caesium between the vegetation compartments and soil. Transfer rate [% per year oftotal inventory]

o ( -

366

In fig. 4.5.14 the predicted I37Cscontent in the perennial understorey vegetation is shown as a function of time after deposition, where the moss and lichen carpet covers 90% of the ground surface. The estimated mean residence time for caesium in moss and litter is respectively 10 and 1 years. A number of other cases were also simulated to allow for different coverage by the moss carpet, and turnover times in litter and soil. The assumed early distribution after the fallout: 50% of the deposited 137Csinitially retained in the canopy and 50% on the understorey vegetation and ground floor, describes the distribution prevaling a few days after the wet deposition. Effects of competition for available caesium at the growing site, and of I T S mean residence time in secondary sources are illustrated for three cases:

(I) the net increase of biomass is negligible in comparison to the size of the existing perennial vegetation (“no competitor”); (11) the increase in biomass described by the size of the competitor compartment is 10 times that of the perennial vegetation (“competitor intermediate biomass”); and

(111) 100 times that of the perennial vegetation (“competitorhigh biomass“).

Relative Cs-137 level in the perennial vegetation l5 T

9 . ’

-

m i .

--Nocompetitor Competitor interow

. I

-

mediate biomass Competitor

5 --

I . Time after deposition [year]

0

5

0

I

10

15

20

25

Fig 4.5.14. Predicted changes of the 137Cs content in the perennial understorey vegetation with regard to growth of “competing”biomass. The mean residence time in moss and litter is assumed to be 10 and 1 year respectively, and mosses cover 90% of the ground surface. Model predictions andfactual results In a period of time in the order of the mean residence time of 13’Cs in mosses and lichen much of the original deposition still remains in these secondary sources, although continuous release occurs to the environment. The model predicts only minor dependence on growing site characteristics with respect to the relative change with time of the 137Cs content in the perennial vegetation (cf. fig.

367 4.5.14) during this period. The long term behaviour, however, is predicted to be governed mainly by the productivity and age of the forest stand. Because about 1% per year of the 137Cs content (cf. Distribution patterns inpodzol andpeat) is estimated to be lost specifically by run-off from peat in discharge areas, peat seems to constitute an exception with regard to the fact that removal of 137Cs from a catchment by run-off may generally be neglected in comparison to loss by physical decay. Therefore any potential increase, as in fig.4.5.14 for the case "no competitor", is expected to be partly counterbalancedby the continuous loss by run-off from such sites. The observed decrease in the perennial vegetation according to fig.4.5.11 conforms with all three cases during the time period 1986-1991. This time span is relatively short in comparison to the assumed mean residence time in the moss and lichen carpet. Consequently, the release of 137Cs from such secondary sources integrated over the elapsed period of time has been of comparatively minor importance. However, residence times in mosses are frequently shorter than the 10 years used in this example. In these cases - with faster release from mosses or lichen - differences in caesium concentration, due to productivity and growth, are expected to occur at a correspondingly earlier stage. Ratio of CS-137 conc. in plants on bog Venus dry soil The small primary production typical of peat soil conditions limits biomass growth. Such sites therefore Rissanen and Raholn 1989 belong to alternative (I), where the continous release of 137Cs from 0Cs-137 pre-Chemobyl secondary sources in the moss and I] Cs-137 : post-Chemobyl lichen carpet may partly or 1 completely compensate for losses by physical decay or transport of "7Cs out off the system. This appears to Horsetail Birch increase the 137Cs content in the Fig. 4.5.15 137Cs concentration in plants growing on peat bog perennnial vegetation after a time versus on "dry soil". Ratios for birch and horsetail are shown for sampling in 1979-1984("pre-Chernobyl";light gray) and period corresponding to the mean from 1986 ("post-Chernobyl";white). Ratios for potassium residence time for caesium in moss (dark gray). and litter. Several findings in the alpine and boreal zones of FennoScandia give evidence of this expected response related to productivity. Over the periode 19871990 the content of radioactive caesium almost doubled in bogbean (Menyanthes trifoliata) and increased fivefold in common cotton grass (Eriophorumangustijblium) (Staaland et al. 1990). Both plants grow on peat. The high levels of radioactive caesium established in vegetation within the first years after deposition on peat and wet bogs in comparison to those in plants growing in dry soil appear to be maintained over several decades. In the period 1979 - 84 ( i.e. about 15 - 20 years after the main deposition due to atmospheric nuclear weapons tests in the middle-sixties, but before the Chernobyl accident) birch and horsetail (Equisetumsylvaticum) growing on bogs (see fig. 4.5.15 ) had several times higher concentrationsof 17Csthan those growing on "dry soil" according

-

~

368 to Rissanen and Rahola (1989). These relations remained similar in 1986 after the Chernobyl accident. Potassium, on the contrary, occurs in about the same concentration in both soil types. A markedly different behaviour of radioactive caesium and potassium in boreal forest ecosystems is implicit in these results. The specific conditions in the boreal zones are emphasised by comparison with e.g. the study of different soil types in Belgium and Luxembourg (Andolina and Guillitte 1990), which indicates that radiocaesium availability is affected in particular by the potassium content. Even in the case of high net primary production, the increase in biomass may be small, because of the combined effects of growth and litter production. This is, for instance, a characteristic of the mature boreal forest, where growth and decomposition processes to a large extent neutralise each other, resulting in relatively small net contributions to the total biomass. Therefore a maturing forest lies in a category, whose response is expected to approach that of alternative (I) during ageing of the growing site.

In a young forest, or in a clear-felled area, a relatively large fraction of the primary production is directed into growth. As a consequence the 137Cs content in the perennial vegetation at such sites is predicted to decrease continuously in agreement with alternatives (11) and (111) in fig. 4.5.14. These two alternatives represent the effects of new growth at a level commonly found in the boreal forest depending on the stage of development and the limit to growth set by the local primary production. The uniform pattern of changes in I T S content exhibited by the perennial vegetation in CS-137 activity ratio content exhibited fig. 4.5.1 1 agrees well with the model predictions Bilberry in young versus mature forests that the caesium levels in the vegetation should primarily be a characteristic characteristic of the single not primarily species, but of the growing site. The information information species, P lost by pooling data over the different ecological ecological conditions conditions occurring at the 10 study sites concerns concerns 0 site-specific factors such as age of the forest stand and primary productivity. These factors are of Fig. 4.5.16 The ratio of 13’Csconcentration in particular importance according to the model. The Young versus mature coniferous forests. The comp&son in fig. 4.5.16, however, tests the vertical bars represent standard error. (Bergman model predictions against the actual results by et al. 1992). specifically focusing on the expected differences in the redistribution dynamics of IS7Csbetween young and mature forest stands.

0D

The increased competition for available caesium caused by the addition of new biomass affects the rate of decrease of the 137Cs content in the vegetation. This is indicated by the different dynamics in fig. 4.5.14 concerning the change of the IS7Cscontent in perennial vegetation in the cases “no competitor“ and “competitorintermediate (or high) biomass”. The categories (a) mature coniferous forests (i.e. comprising the Scots pine and mixed coniferous forests at the study sites), and (b) young forests or clear cuts, have been used to separate the material for bilbeny (the same as

369 in fig. 4.5.1 1) in populations belonging to the mature or young forest respectively. The ratio between the 137Cs concentration in bilberry from recently clear-felled areas and mature forests is shown in fig. 4.5.16 for three periods: 1) 1986; 2) 1987-1988; and 3) 1989-1990.On the first sampling occasion in July 1986 the "7Cs concentrationswere not significantly different in young and mature forest areas (Nelin and Nylen 1993). The ratio therefore starts in 1986 at a value close to 1, but decreases significantly over the follwing two periods down to about 113. Thus the change towards lower concentrations in bilberry is faster in the young forest or clear-felled areas, where the increase in biomass is relatively high, in comparison to that in the mature forest. Concerning the redistribution dynamics for caesium this qualitative dependence on competing biomass is exactly what is predicted by the model. Different dynamics depending on productivity and soil conditions also appear from the analysis of the content of radioactive caesium in dwarf birch (Nreumann et al. 1990). A faster decrease occurred for this species on good than on nutrient-poor soil. The ecological half-lives were estimated to 1.2 and 3.0 years respectively. The satisfactory agreement between the observed and the predicted behaviour of 137Cs in the boreal forest (A): in the early phase ( i.e. over the first five years after the Chemobyl accident),

and

(B): the distribution about two decades after deposition provides important clues to the expected general behaviour in the forest over a time span comparable to the physical half-life of 137Cs. It is contended that: - at the same primary production capacity the decrease in 1J7Csactivity concentration in the

perennial vegetalion will be faster in the young than in the matureforest;

- at a particular growing site the rate of decrease of the 13'Cs content in perennial vegetation will be positively related to the local primary production capacity (at sufficiently low productivity, e.g.peat soil conditions, decrease may not occur or be extremely low for several decades).

CAESIUM DISTRIBUTION AND "AVAILABILITY"IN FOOD-CHAINS: AN EXTRAPOLATIONTO FUTURE SITUATIONS The change in availability with time after deposition, as regards observed ratios (OR) between the content in vegetation and that in the total soil column, is presumably at least partly related to chemical processes leading to the incorporation of caesium in clay matrices - a phenomenon well known in agricultural radioecology. Cations that enter into mineral soils tend to be removed rapidly from solution on to exchange sites which occur on both clay minerals and organic molecules. Cations having a small ionic radius, (K,Cs,Rb), can also enter the interlayer space of micaceous

370 clay minerals where they are held until the mineral is weathered. Thus, the immobilization of radiocaesium in mineral soils is rapid. An often accompanying vertical transport in undisturbed soil may gradually transfer radioactive caesium out of the horizons where effective recirculation occurs in the living matter. However, the distributions illustrated in fig. 4.5.9 indicate a distribution pattern in soil that is essentially the same over several decades in the boreal zone. Furthermore, in this zone no significant decrease in availability is apparent for the distribution of "old" and "Chernobyl" caesium in vegetation and soil (Raitio and Rantavaara 1993b) and in certain forest food-chains, as indicated by the OR for moose (Bergman et al. 1989, Bergman et al. 1991, Johansson 1993), as well as by analysis of various foodstuffs of "wild" origin (Rantavaara 1990). However, south of that area changes are I apparent in the vertical distribution (cf. fig. 4.5.9), and combined with effects ysical decay modifying the availability over long time periods. For a growing site of boreal type on podzol soil in Denmark input (Strandberg 1992), a decrease in OR to a third or half is obtained for "old" caesium in comparison to that found at the beginning of the nineties for "Chernobyl"caesium. There seems thus to be a gradual change from the Fig. 4.5.17 The principal natural processes affecting the conditions governing the rather budget of radioactive caesium in a forest ecosystem. conservative behaviow of caesium in the

The caesium budget in a forest ecosystem

-t

+-

boreal zones towards those in the nemoral zone in Denmark and Germany. The main natural causes of changes in the inventory of radioactive caesium in a forest ecosystem are indicated in fig. 4.5.17. Recirculation concerns internal translocation within plants, as well as redistribution between different compartments of the ecosystem; e.g. by litterfall , herbivory (Bergman et al. 1992a), decompositionprocesses and uptake. Vertical transport means the downward penetration in soil, and concomitant physico-chemical loss by leakage to groundwater or decreasing availability for living cells. Runoff of 137Cs, although generally very small with regard to loss from podzol soil in comparison to loss by physical decay, appears to be particularly important when deposition occurs during winter or early spring as in the case of the Chernobyl accident. Furthermore, the release from peat in discharge areas seems to be relatively high during several decades in the boreal zones. Resuspension - potentially contributing both as an input to and loss from a forest ecosystem - is not shown in the figure as its effects appear to be negligible (Oughton 1989) compared to that of the processes identified in fig. 4.5.17. Nonetheless, it may be relatively important for short and long-term redistribution at the alpine level (Game

37 I

1989). Besides the natural processes, removal associated with human impact on the system, e.g. forest practice - especially logging - should also be considered. The new knowledge gained from Nordic radioecological research since the Chernobyl accident fits a pattern, where primarily recirculation processes and physical decay affect the distribution and loss in the boreal forest ecosystems. Our interpretation is that during the successional stages of development - i.e. the ageing and redistribution will comply with these population changes in biotopes - the dynamics of general characteristics, not only in the five-year period in which direct observations are available concerning the Chernobyl fallout, but practically for as long as the ceasium isotope still remains in the ecosystem. If this interpretation is true, the pattern of redistribution will be repeated at sites where, for instance, a mature forest stand is clear-felled - even far into the future. Depending on forest practice, we expect a transient phase to occur initially, during which the prerequisites of a sufficiently intact feed-back network may not be satisfied. The re-establishment of a functional complex network, which mainly relies on the recovery of the microbial, mycelial and fine root systems, will probably be fast. This means that the 137Cscontent of the perennial vegetation, after a short-lasting transient phase, is expected to undergo the same dynamic stages of redistribution as described for the forest exposed to direct deposition - but, of course, only with regard to the recirculated fraction.This assumption is supported by the fact that 3-1 8 % of the total deposition over a mature forest is contained in the tree biomass over the period from some years to several decades after fallout (see fig. 4.5.10 and Holm 1993). About two-thirds of this will be removed with the tree trunks, as can be estimated from the internal distribution. Consequently only about 212%of the total deposition is lost by that route. This in turn implies that the total inventory of 137Cs will not change considerably as a result of logging. There is, however, an apparent paradox in the occurrence of a generally rather uniform and longtime decrease of radioactive caesium in vegetation on podzol soil, whereas the concentrations in certain food-chains - as manifest by the OR in moose meat, or from decay-corrected samples of “wild” foodstuffs over large areas - are not significantlytime-dependent. As long as the large-scale pattern (e.g. on a regional basis) of clear-felled areas and forest stands of various ages is not radically changed by forest practice or extremely extensive forest fires, the distribution of biotopes prevailing at the time, when the deposition of radioactive caesium occurred, will persist or at least be similar over a long time - despite ageing and a continuing successional process for each individual growing site. The explanatory model predictions ,and the small losses from the system, indicate that in this time span a regional average for the ‘37Cs concentration in the key-plants for moose and in bemes used by man will be essentially unchanged over the different growing sites. The average 137Cs content in the moose population should thus also remain on a similar level under these conditions.This supports the assumption that despite the dynamic behaviour of redistribution observed for 137Csat a particular growing site, the cycle provided by ageing of the forest stands and the changes (through succession and forest practice) to young forest or clear-felled areas will preserve the 137Cscontent in the boreal forest vegetation on a

372 relatively stable level, which is essentially only subjected to a decrease as a result of physical decay. There seems to be a gradual change from the conditions governing the rather conservative boreal behaviour of caesium to those of the nemoral zone in Denmark and Germany, where there are differences in the vertical distribution between nuclear weapons and Chemobyl caesium, and processes operate that lead to a significant decrease in availability. Although the causal relationships between these phenomena and the underlying processes are still insufficiently elucidated, the scarcity of earthworms ( i.e. effective bioturbators ); relatively slow rate of litter decomposition; and low pH in podzol soils of boreal coniferous areas, as compared to the conditions at growing sites of predominantly decidous types - or for the more favourable climatic conditions to the south - are possible and even probable causes of the notable differences. The relative contributions of the various processes are indicated semi-qualitatively in fig. 4.5.18, as averages of the loss over a decade within and south of the boreal zone. The implication is that physical decay will be the major factor of loss from the boreal ecosystem in a long-term perspective, and that runoff is expected to be the second in order of importance - even in areas subjected to logging during this time. South of the boreal zone decreasing availability appears to be of progressively greater importance in passing from the hemiboreal to the nemoral vegetation zones. Physical decay and decreasing availability constitute the dominant factors for the loss of radioactive caesium in these areas.

IMPLICATIONS FOR ASSESSMENTS OF DOSE TO MAN Usually the forest may be considered as an undisturbed system in the sense of a comprehensive intact network, although forest practice will temporarily more or less extensively disrupt its system of ecological interactions. Agricultural practices, on the contrary, often focused on optimal growth, imply radical disturbances and extensive disruption of the network that would otherwise be attained

313

in a natural undisturbed ecosystem. It therefore in general remains a reduced number of feed-back pathways operative in the system with respect to turnover of caesium. A successful explanatory or predictive model of transfer between soil and plant may, under such circumstances, be based on a few major factors and pathways, e.g. clay content in soil and uptake by the root. We believe that problems encountered in extrapolationsof strategies useful for radioecological studies of caesium behaviour in agricultural ecosystems to studies of the forest environment, may be traced back to these qualitatively different conditions. It appears to be due to - rather than despite - the complex network in the forest that an approach based on growth and an assumed effective feed-back, a "holistic" model principle, may be successful in describing caesium behaviour in long-term perspectives. A holistic approach to the behaviour of caesium in the boreal forest ecosystem offers particular advantages concerning assessments of the dose to man from transport of 137Cs through forest food-chains. The explanatory model approach, which we consider successful for prediction of the main features concerning 13'Cs distribution in the boreal forest in different time perspectives, indicates that the overall distribution pattern is essentially conserved, and in the boreal zone primarily subjected to changes because of physical decay. This conclusion is valid despite the effects of dynamic redistribution processes operating on the scale of the local growing site, leading to high variability between samples and often pronounced long-term site-specific changes in the concentration of radioactive caesium . Therefore assessments are expected to become quite straightforwardwith regard to the activity concentration in boreal forest products such as moose meat and certain berries, as well as the subsequent internal exposure of humans because of their consumption of these forest products. The transfer factor (UNSCEAR 1977), or transfer coefficient (UNSCEAR 1982), has been defined as the quotient of the infinite time integral of the appropriate quantity in one compartment to the infinite time integral of the appropriate quantity in the preceding compartment. The ratio for 137Cs concentration in e.g. compartment moose to that in compartment ground over the time interval 1985-1990does not comply with this definition of a transfer factor. However, the conservative role of the boreal forest ecosystem regarding the inventory of 137Csfacilitates the translation of the properly used "observed ratio" for short time periods to a transfer factor. It also offers a comparatively safe way of estimating the precision of the results. This advantage Concerning dose predictions in a long-term perspective for forest ecosystems seemingly lacks a counterpart outside the boreal areas. Instead, the availability there of radioactive caesium for plants decreases considerably over a time corresponding to the physical half-life of 137Cs. The quantitative expression for this time dependence is, however, seldom known with sufficient precision. Improper use of transfer factors, i.e. based on OR over a too limited period, is therefore likely to be misleading when assessing dose commitments from forest products. This particularly concerns the Nemoral zones. The relatively small OR observed in vegetation for fallout of nuclear weapons caesium about 20 years after deposition, in comparison to the OR based on the recent deposition after the Chernobyl accident, leads to a notable overestimation of internal exposure or dose, if "Chernobyl data" are used uncorrected for this effect.

374

ACKNOWLEDGEMENTS This work was supported by the organisation for Nordic Nuclear Safety Research, the National Radiation Protection Institute in Sweden, and the Organisation for Radiation Protection Research under the Commission of the European Communities.

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375 Environments.(ed Desmet E, Nassimbeni P, Belli M) Elsevier Applied Science London & New York. 450-458. Bretten S. 1991. Radioaktivt Cs-137 etter Tsjemobylnedfalleti alpine p l a n t e s a m h p i Dovrefiell. In: TSJERNOBYL Slutrapportfra NINA’S radioakologiprogram 1986-1990. Temahefte 2.2835, (in Norwegian) Bunzl K, Schimmack W, Kreutzer K and Schierl R. 1989. The migration of fallout 134Cs,V3 and ‘06Rufrom Chemobyl and of 137Csfrom Weapons testing in a forest soil. Z P’amenniihr. Bodenk. 152. 39-44. Bunzl K, Schimmack W, Kreutzer K and Schierl R. 1989. Interception and retention of Chernobylderived 134 Cs, 137Cs and lo6Ru in a spruce stand. The Science of the total environment. 78. 77-87. Cawse P A, and Horrill A D (1 986). A survey of caesium 137 and plutonium in British soils in 1977. UKAEA, Hanvell55pp. DeGeer L-E, Amsting R, Vintersved I, Sisefsky J, Jakobsson S and Engstrom J-A. 1987. Particulate radioactivity, mainly from nuclear explosions in air and precipitation in Sweden mid-year 1975 to mid-year 1977. FOA rapport C 40089-T2. 1987. Devell L. 1988a. Characteristicsof the Chemobyl release and fallout that affect the transport and behaviour of radioactive substances in the environment. STUDSVIK/NP-88/1. Ibid. 1988b. Effects of atmospheric deposition processes and surface contamination on accident consequences, their mitigation and emergency response planning. Studsvik Nuclear - Technical Note NP-88/109. Ibid. 1988c. Nuclide composition of Chemobyl hot particles. Proc.Joint CEC/OECD(NEA) Workshop on Recent Advances in Reactor Accident Consequences Assessment. Rome, Italy, January 1988.CSNI Report 1-45, 1,23-34, 1988. Ibid. 1989. Characteristics of the Chemobyl releaseand fallout. Studsvik Nuclear - Technical Note NP-89/34. Degerrnark C. 1987. Climate and Chemistry of Water at Svartberget. Reference measurements 1986. (In Swedish) Swedish University of Agricultural Sciences. Ibid. 1988,1989,1990 and 1991. Edvarson K. 1991a. Fallout over Sweden from the Chemobyl accident. 47-66. In: The Chemobyl fallout in Sweden: Results from a research programme on environmentalradiology (ed L Moberg th Swedish radiation protection institute). Ibid. 1991b. External doses in Sweden from the Chemobyl fallout.527-246 In: The Chemobyl fallout in Sweden: Results from a research programme on environmentalradiology (ed L Moberg the Swedish radiation protection institute). Eriksson 0, Jones B and Raunistola T. 1991. Radiocesium contamination and the reindeer. In The Chemobyl fallout in Sweden: Results from a research programme on environmental radiology (ed L Moberg the Swedish radiation protection institute). Emst W H 0 and van Rooij L F. 1987. Cs-134/137 fall-out from Chemobyl in Dutch forest. In: Lindberg S E and Hutchinsson T C (Eds). Heavy metals in the environment. Proceedings of Intemational Conference in New Orleans - September 1987.284-296. Ertel J and Ziegler H. 1991. Cs-l34/137 contamination and root uptake of different forest trees before and after the Chemobyl accident. Radiation and Environmental Biophysics 30:147-157. Finck R. 1992. Thesis: High resolution field gamma spectrometry and its application to problems in environmental radiology. LUNFD6/(NFRF-1004)/1-138/(1992). Department of Radiation Physics, Malmo Gemeral Hospital, S-21401 Malmo, Sweden.

376 Fujita M, Iwamoto J, and Kondo M. 1966. Comparative metabolism of cesium and potassium in mammals interspecies correlation between body weight and equilibrium level. Health Phys. 12, 1237-1247. Gaare E. 1987a. Radioaktivt cesium i noen reinbeite-lav fra sentralnorske fjellstr~k.- Third Nordic Workshop on Reindeer Research, Rovaniemi, Finland 15-17 October 1986. Rangifer 1 Appemdix:45-47.(in Norwegian). Ibid. 1987b. Hvofor varierer innholdet av radiocesium i lav s i sterkt over korte avstander? - S, 5357 i Jensen B.M. (red.) Radioakologisk forskningsprogram:Resultater fra undersakelser i 1986. Foredrag holdt p i seminar i DN 22. April 1987. (in Norwegian) Ibid. 1991. Virkningen p i reinens beite i traktene fra Dovrefjell ti1 Rondande av ulykken i Tsjemobyl, april 1986. In Tsjernobyl: slutrapport fra NINA'S radio6kologiske program 19861990. NINA Temahefte 2: 1-71 -8ed Gaare, Jonsson of Skogland) Trondheim 1991. (in Norwegian) Ibid. 1992. Monitoring of radiocesium in Norwegian natural areas. Det sjette Nordiske Radio$kologi Seminar. 14-18juni 1992. Torshavn Faeroyar. Georgi B, Helmeke H-J, Hietel B and Tschiersch J. 1988. Particle size distribution measurements after the Chernobyl accident. Proc.Joint CEC/OECD(NEA)Workshop on Recent Advances in Reactor Accident Consequences Assessment. Rome, Italy, January 1988.CSNI Report 1-45,1,39-52,1988. Guerin M. and Wallon G. 1979. The reversible replacement of internal potassium by caesium in isolated turtle heart. J Physiol p 525-537. Guillitte 0, Koziol M, Debauche A, and Andolina J. 1990. Plant-cover influence on the spatial distribution of radiocaesium deposits in forest ecosystems. In: Desmet G, Nassirnbeni P, and Belli M (Eds). Transfer of radionuclides in natural and semi-natural environments. Elsevier Applied Science. London and New York. pp. 441-449. Gustafsson E, SkAlberg M, Sundblad B, Karlberg 0, Tullborg E-L, Ittner T, Carbol P, Eriksson N and Lampe S. 1987. Radionuclide deposition and migration within the Gidei and Finnsjon study sites, Sweden: A study of the fallout after the Chernobyl accident. SKB Technical report 87-28. Swedish Nuclear Fuel and Wast Management Co. Haugen L.E. 1989. Transportmekanismer og plantetilgjengelighetav radionuklider i dike jordtyper. Informasjon fra Statens fagtjeneste for lantbruket. 1989 (1): 65-74. Haugen, L.E. and H.E. Bjerrnstad. 1990. Transport of radiocaesium during snowmelting on a mountain pasture in Norway, spring 1989. Paper presented on: IUR Working group on Soil-Plant transfer. Workshop on The Contamination of Crops because of Soil Adhesion; Uppsala, Sweden, September 27-28, 1990. p 139-142. Holm E. 1993. Flux of radionuclides and absorbed doses in the forest industry.(submitted for publication, Seminar on The Dynamic Behaviour of Radionuclides in Forests, May 18-22 1992) Hove K, Pedersen 0, Garmo T, Hansen H.S. and H Staaland. 1990. Fungi: A major source of radiocesium contamination of grazing ruminants in Norway. Health Phys vol. 59, no2, pp 189192. Johansson K-J .1993. Radiocaesium in game animals in the Nordic countries. (submitted for publication, Seminar on The Dynamic Behaviour or Radionuclides in Forests, Stockholm, Sweden, May 18-22 1992). LindCn A and H. Mellander. 1986. Airborne measurements in Sweden of the radioactive fallout after the nuclear reactor accident in Chernobyl, USSR.SGAB report TG 8606. Livens F.R., Horril A.D. and Singleton D.L. 1991. Distribution of radionuclides in the soilplant systems of upland areas of Europe. Health Phys.60.539-545

377 Lundmark J-E,. 1986. Skogsmarkensekologi- stindortsanpassat skogsbruk. Del 1 -Grunder. Skogsstyrelsen,JonkBping. Melin J and Wallberg L. 1991. Distribution and retention of Cesium in Swedish boreal forest ecosystems. In The Chernobyl fallout in Sweden (ed. L Moberg) The Swedish Radiation Protection Institute 1991. Mattsson S. 1972. Radionuclides in Lichen, Reindeer and Man: Long-term variations and internal distribution studied by gamma-spectrometric methods. Thesis from the Radiation Physics Department,Universityof Lund, Sweden . 1972. Ibid. 1975. Deposition, retention and internal distribution of Eu-155, Ce-144, Sb-125, Ru-106, Zr95, Mn-54 and Be-7 in the reindeer lichen Cladonia Alpestris, 1961-1970.Health Phys. 29 27. Naxmann R., E. Steines and P. Varskog. 1990. Mobilitet og plantetilgjengelighetav radiaktivt cesium i naturlig jord. Informasjon fra Statens fagtjeneste for lantbruket. 1990 (1): 61-65. N y l h T and Ericsson A. 1989. Uptake and retention of Cs-137 in Scots Pine. Proc. XVth Congress of IRPA. Progress in radiation protection (ed W Feldt). 1989. Nylkn T and Grip H. 1989. Transport of Caesium-137 in a forest catchment. Proc. XVth congress of IRPA. Progress in radiation protection (ed W Feldt). 1989. N y l h T and Nelin P. 1992 . The time related distribution of radioactive cesium in boreal forests. (submitted for publication, Seminar on The Dynamic Behaviour of Radionuclides in Forests, Stockholm, May 18-22 1992). Osbum W.S. 1963 The dynamics of fallout distributionin a Colorado alpine tundra snow accumulation ecosystem. In: V. Schultz & A.W. Klemens (eds.) Radioecology. Proceedings of the First National Symposium on Radioecology held at Colorado State University, fort Collins. Colorado, September 10-15, 1961. p 50-71. Oughton D.H. 1990. Radiocaesiurn association with soil components: The application of a sequential extraction technique. Paper presented on : IUR Working group on Soil-Plant transfer. Workshop on The Contaminationof Crops because of Soil adhesion; Uppsala, Sweden, September 27-28, 1990. p 182-189. Oughton D.H. og Salbu B., 1990, Estimation of the mobility of radionuclides in soil with regard to mobility factors. NLVFs research programme about radioactive fallout. Seminar 6-7 november 1990. Informasjon fra Statens Fagtjeneste for lantbruket Nr. 28 1990 (in Norwegian). Palo T, Nelin P and Lindstrom E. 1989. The Chernobyl aftermath. Uptake of Caesium-137 in vegetation and wildlife in Northern Sweden. IUGB Congress Trondheim 1989. Persson C, Rodhe H and De Geer L-E, 1986. The Chemobyl accident - A meteorological analysis of how radionuclides reached Sweden. SMHI ReportsNr 55, December 1986. Raitio H and Rantavaara A. 1993a.Airbome radiocesium in Scots pine and Norway spruce needles. Submitted for publication Sci. Total Environment. Raitio H and Rantavaara A. 1993b. Radiocesium budget in young Scots pine stand. Preliminary results at seminar, Aas Norway. January 27-29 1993. Rantavaara A, Nygren T, Nygren K and Hyvonen T. 1987. Radioactivity of game meat in Finland after the Chernobyl accident in 1986. STUK-A62. Supplement 7 to Annual Report STUK-A65.1987. Rantavaara A. 1990. Transfer of radiocesium through natural ecosystems to foodstuffs of terrestrial origin in Finland. The 7th regular meeting of the Nordic Radiation Protection Society, Ronneby, Sweden. 26-29 August 1990.

378 Raunemaa R. Saari H. Luokkanen S and Lehtinen S. 1988. Hot particles in the Fallout of Chernobyl in Fin1and.h: Hot particles from the Chernobyl fallout, B. 16. H. von Phillipsbom, F. Steinhausler (eds.): Schriftenreihedes Bergbau- und Industrimuseums Ostbayern Theuer, 1988. Riise G., H.E. Bj~rnstadH.N. T.Krekling, H. Lien, G. Riise and G. 0stby. 1990. A study on radionclide association with soil components using a sequential extraction procedure. Journal of Radioanalytical and Nuclear Chemistry, articles, 142 (2): 53 1-538. Riise G, Bjmnstad H.E., Lien H.N., Oughton D.H. and Salbu B. 1990. A study on radionuclide association with soil components using a sequential extraction procedure. J of Radioanalytical and Nuclear chemistry, Articles 142. no 2. 1990,531-538 Riise G. and Salbu B. 1989. Sekvensiell extraktsjonsteknikfor bestemmelse av radioaktive Csisotopers assosiasjon ti1 dike jordfraksjoner. In: Forskningsprogramom radioaktivt nedfall. seminar 22-23 november 1988. Informasjon fra Statens Fagtjeneste for Lantbruket. Nr 1 1989. Rissanen K and Rahola T. 1988. Cesiumhalter i renlav fire och efter Tjernobylolyckan. Det femte Nordiska Radioekologiseminariet22-25 augusti 1988, Riittvik, Sverige. (in Swedish). Ibid. 1989. Cs-137 concentration in reindeer and its fodder plants. The Science of the Total Environment, 85.199-206. Rissanen K. Rahola T. and Illukka E. 1987. Radioactivity in plants and foodstuffs in Lapland 1979-1986, Studies on environmental radioactivity in Finland in 1986, STUK-A55, Annual Report, 1987, Helsinki. Roed J. 1988. Dry deposition on trees and grass. The 5thNordic seminar in Radioecology, Rattvik, Sweden. August 22-24 1988. Rudjord A.L. and L.E. Haugen. 1989. Distribution of radiocesium in soil profiles 1986-88. NJF-seminar "Deposition and transfer of radionculides in Nordic terrestric environment". Beitostalen 2 1-23 August, 1989. Nordisk JordbruksforskeresForening, NJF-utredningedRapport Nr. 59 .13-19. Salbu B., 1988a, Radionuclides associated with Colloids and Particles in the Chemobyl Fallout. Proc.Joint CEC/OECD(NEA)Workshop on Recent Advances in Reactor Accident Consequences Assessment. Rome, Italy, January 1988.CSNI Report 1-45, 1,53-68, 1988. Salbu B. 1988b, Salbu B, Bjarnstad H.E. Lien H.N. Riise G. and 0stby G.1990. Determination of physicochemical forms of radionuclides deposited after the Chernobyl accident. IAEA. -172. SM-306/35P. V O ~1,171 Salby B. 0stby G. Garmo T and Hove K. 1992. Availability of Cs-isotopes in vegetation estimated from incubation and extraction experiments. The Analyst (in press). Saltbones J. 1986. Kjernekraft-ulykken i Chemobyl: Atmosfzrisk transport og spredning av radioaktivt materiale. Det Norske Meteorologiska Institutt 1986. (in Norwegian). Selnaes T. D. and Strand P. 1992. Comparison of the Uptake of Radiocaesium From Soil to Grass After Nuclear Weapons Tests and the Chemobyl Accident. The Analyst. 117.493496. Schimmack W, Bunzl K and Zelles L. 1989. Initial Rates of Migration of Radionuclides from Chernobyl Fallout in Undisturbed Soils. Geoderma, 44,211-21 8. Shaw G. and Smith J., private communication(Progress report at radioecological meeting in Barcelona 12-15 May 1991). Simkiss K et al. 1993,( Environmental Radioactivity Special Steering Committee. National Environment Research Council, UK). 1993. Radiocaesium in Natural Systems - A UK Coordinated study. J Environ. Radioactivity 18.133-149.

379 SIS. 1987. Radioaktivt nedfall. Nedfallsm~nster for cesium-134 of cesium-137 etter kjernekraftulykkeni Tsjernobyl. Kart i maestokk 1:5 000 000. Skhlberg M. 1992. The Gidei Study - Area description and Radionuclide Deposition PatternSeminar on The Dynamic Behaviour or Radionuclides in Forests, Stockholm, Sweden,May 18-22 1992). Slutrapport fra NLVFs forskningsprogram om radioaktivt nedfall 1988-1991. Radioaktivt nedfall fra Tsjernobyl-ulykken. Garmo T.H. and T.B. Gunnerad (eds.) Norges Landbruksvitenskapeligeforskningsrid. 1992. Sombre L, Vanhouche M, Thiry Y, Ronneau C, Lambotte J M, and Myttenaere C. 1990. Transfer of radiocesium in forest ecosystems resulting from a nuclear accident. In Transfer of Radionuclides in Natural and Semi-Natural Environments.(ed Desmet E, Nassimbeni P, Belli M) Elsevier Applied Science london & New York. Staaland H., T.H. Garmo, 0 Pedersen and K. Hove. 1990. endring i inhaldet av radiocesium i plantemateriale of beitedyr p i fielbeite 1986-1990. Informasjon fra Statens fagtjeneste for lantbruket. 1990 (28): 84-95. Steinnes E. and Njhstad 0. 1992. Uses of mosses and lichens for regional mapping of Cs-137 fallout from the Chernobyl accident. Det Sjette Nordiske Radio6kologi Seminar, Torshavn, Faeroyar, 14-18juni 1992. Strandberg M. Radiocesium in a Danish boreal pineforest ecosystem. The VI:th Nordic seminar on radioecology, Torshavn, the Faroes, June 14-18 1992. Tobler L., Bajo S. and Wyttenbach A. 1988. Deposition of Cs-134/-137 from Chernobyl fallout on Norway spruce and forest soil and its incorporation into spruce twigs. J. Environ. Radioactivity, 6 : 225-245. UNSCEAR. 1977. Sources and effects of ionizing radiation. United Nations Scientific Committee on the Effects of Atomic Radiation, Report to the General Assembly. United Nations. UNSCEAR. 1982. Ionizing radiation. Sources and biologicasl effects., Ibid. United Nations. Varskog P, Naumann R and Steiness E. 1990. Opptak av radioaktivitet og stabilt cesium i naturlig Qellvegetasjon fra jordsmonn med varierende nmingsstatus. In NLVFs forskningsprogram om radioaktivt nedfall. Informasjon fra Statens fagtjenste for lantbruket. Nr 28 1990 (ed. Gunnerfd og Garmo) (in Norwegian). Vintersved I, Arntsing R, Bjurman B, De Geer L-E, Jakobsson S. 1991. Resuspension of radioactive caesium from the Chemobyl fallout. pp 85-106. In The Chernobyl fallout in Sweden: Results from a research programme on environmental radiology (ed L Moberg the Swedish radiation protection institute). Whitford P. B. 1968. Foliar application of 137Cs on understory species of mesic forest. Radiation Botany.8. 509 -5 13. Winteringham F.P. 1989. Radioactive fallout in soils, crops and food. a background review prepared by F.P.W. Winteringham for the FA0 Standing Committee on Radiation Effects, the FA0 Land and Water Development Division and the Joint FAOLAEA Division on Nuclear Techniques in Food and Agriculture. FA0 Soils Bulletin 61. 84. Witherspoon J.P. 1964. Cycling of Cesium-134 in white oak trees. Ecol Monogr. 34.403-420 Wyttenbach A. and Tobler L. 1988. The seasonal variation of 20 elements in 1st and 2nd year needles of Norway spruce, Picea abies (L.) Karst. Trees 2. 52-64.

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383

5.1. INTRODUCTION TO INTECALIBRATION/ANALYTICAL QUALITY CONTROL AND DOSES

ELIS HOLM Department of Radiation Physics, Lund University, Sweden.

The purpose of the Analytical Quality Control programmes was to assist laboratories engaged in the analysis of samples of environmental origin for radionuclide determination and to check the quality of their work. Such control is necessary as results of analytical activities may be the basis upon which economic, administrative, legal or medical decisions are taken; they must, therefore, be documented to be sufficiently reliable. Reliability of results is a function of precision (reproducibility) and of accuracy (true value). The precision of results can easily be determined by internal measures. The determination of accuracy, however, in most cases requires more detailed procedures such as: -Analysis to be carried out by as many different, independent methods, analysts and techniques as possible. In cases when agreement is good, results can be assumed to be accurate. -Control analysis with Reference Materials which are as similar as possible to the materials to be analysed. Agreement between certified and observed values then allows a direct inference of accuracy for that particular determination -Participation in interlaboratory comparison studies. Samples used in such studies should, as far as possible, be similar in composition and concentration to the samples to be analysed on a routine basis. The agreement of their results received from a particular laboratory with the most probable mean value obtained from statistical evaluation of all the results will be a measure of the accuracy for that particular determination. For practical reasons, most analytical laboratories are not

in a

position to check accuracy internally, without an external source of Reference Material. Frequently resources are available for only one method

384 and/or technique. Only in exceptional cases Reference Materials, particularly for trace analysis, are normally prepared by the institutes themselves. Intercomparison Runs are organised on a rather limited basis and many important materials and/or analytes have not been covered so far. The International Atomic Energy Authority runs a number of intercomparison exercises annually under their AQCS (Analytical Quality Control Services) programme. For controlled and accidental releases from nuclear facilities experience shows that assessment of gamma-emitting radionuclides is the most important tool to quantify general releases and environmental transport. On the basis of results for gamma-emitting radionuclides, the impact of beta emitters can also be assessed, using calculated/known ratios. In addition gamma-spectrometric methods are far more rapid and need less manpower than the analysis of beta and alpha emitters. Following the Chernobyl accident it was soon understood that the most important radionuclides, from a radiological point of view, were the radiocaesiums, 137Cs and 13'Cs. Analytical control was therefore carried out for gamma spectrometry by sending homogenized samples (tree bark from coniferous and deciduous forest) to 30 Nordic laboratories for the determination of radiocaesium. Often neglected but of great importance for comparison of environmental measurements, are the sampling methods f o r air, water sediments and biological material and possible variation in techniques at different laboratories. In the case of a nuclear accident air concentrations are of great interest in order to follow the initial movement of a radioactive cloud. Therefore different air filtration set-upsin the Nordic countries were calibrated against one another by sending two sampling devices to various laboratories. Measurement of the actual concentration of radioactivity in man is the only true tool for estimating the real dose received . The results can then be compared with different models and estimations. A calibration of whole-body measurement was performed by sending "phantoms" simulating the human body and containing known activities of 134Cs and 137Cs around the Nordic countries. The dose to the Nordic population was also calculated on the basis of the intake of foodstuff using data for activity concentrations produced under the radioecology programme of NKS. These results were compared with those obtained by whole-body counting.

385

5.2 INTERCOMPARISON OF LARGE STATIONARY AIR SAMPLERS

INGEMAR VINTERSVED' National Defence Research Establishment, Sundbyberg, Sweden

SUMMARY The performance of large stationary air samplers at nine different laboratories was tested by employing two transportable high-volume air samplers, which were used in parallel with the stationary sampler at each laboratory for periods lasting between 2 and 6 months. Comparisons were made for 7Be at each laboratory and for '"Cs and ""Pb at some of the laboratories. The last two nuclides proved difficult to use for this purpose whereas 7Be was very useful.

INTRODUCTION Surveillance of airborne radionuclides using high-volume samplers is one of the most sensitive ways of detecting fresh fallout. In emergency situations concentrations of radioactive nuclides are quickly obtained. This is a very important parameter to consider in the evaluation of radiological consequences. Therefore an intercomparison test of large air samplers is of great interest. Most large air samplers are stationary and not easily transportable to a common site for an intercomparison test. To be able to compare results from different stationary samplers, one must employ one or two high-volume samplers which are moved sequentially to all participating sites. Several transportable samplers suitable for such an exercise were constructed at the National Defence Research Establishment (FOA) after the Chernobyl accident, primarily for resuspension studies. They have now been used at nine different laboratories to compare the results from the stationary samplers. 7Be was used in this intercomparison and at some sites it was also possible to compare '37Cs and 'l"Pb results.

THE FOA SAMPLER The transportable FOA sampler was constructed with the intention of having an easily movable sampler with high air sampling capacity and reasonably low power consumption. A sketch of the sampler together with its specifications is shown in figure 1. Of the two principal types of pump most frequently used for aerosol collection from air - positive displacement pumps and centrifugal pumps - the FOA sampler uses a centrifugal pump. Contrary to the positive 'The participants in this project ace presented in table 5.2.1

386

Figure 5.2.1 Sketch of FOA transportable hi h volume air sampler. The sampler is shown with lid open. During sampling the lid is closed feaving a slit around to allow the air to enter the filter. displacement pumps there is no linear relationship between suction pressure and capacity of a centrifugal pump. Centrifugal blowers draw air through the filter at a fairly constant linear velocity although clogging of the filter will reduce the flow rate. Therefore it is very important to measure the flow rate during the whole sampling period. In the FOA sampler this is done by using a gas flow meter with high accuracy and reliability. It is made by Fluid Inventor

AB, Stockholm, Sweden, and it uses a patented fluidic oscillator. The characteristic of this fluidic device is that its oscillating frequency is proportional to the velocity of the gas passing through. This flowmeter has also been used in FOA's permanent sampling stations since 1976. The calibration of the electronic converter, which is adjusted to give one pulse per cubic meter of air, is checked once a year against standard Pitot tube measurements. The adjustment required by any of our 8 transportable samplers during the last 5 years has never exceeded 2%.

RADIONUCLIDES USED Intercomparison testing requires measurable quantities of a radionuclide in air. A suitable radionuclide to use is 7Be, which is produced by cosmic rays and occurs in concentrations of one to several mBq/m3. The concentration of 13'Cs is at present very low (<5pBq/m3) but i t has been possible to compare results at some sites. To illustrate the fact that the concentration of 7Be in ground level air does not vary within short distances, we studied the 7Be concentration measured by three FOA samplers during 1987 to 1990 in the Gavle area. The samplers were placed in the city of Givle (SI), 5.5 km west of (S2) and 4.5 km north of the centre (S3). For samplers S1 and S2 there are 91 pairs of weekly samples. The mean value of the ratio between the 7Be concentration measured by samplers S1 and S2 is 0.993 with a standard deviation of 0.028. For samplers S1 and S3 there are 104 simultaneous samples with a mean ratio of 1.005 and standard deviation of 0.037. In these cases the samplers were kilometers apart while they

387 were only meters apart in the intercomparison test. This means that one should not expect to find any difference in the 'Be concentrations measured at the respective sites. Making the same comparison for the I3'Cs concentrations in the Gavle area we find that the ratio between samplers S1 and S2 is 1.35 with a standard deviation of 0.48 and the ratio between samplers S1 and S3 is 1.65 with a standard deviation of 0.87. This indicates that, in a heavily contaminated area like Gavle (>lo0 kBq/m2), the concentration of '"Cs in air depends on local resuspension, so one cannot expect to find the same concentration in samplers placed kilometers apart. In

the intercomparison test the samplers were placed only 10 to 50 m apart and none of the stationary samplers was located in a heavily contaminated area. Therefore one should expect to obtain agreement between the results from the different samplers at the sites participating in the intercomparison. The long-lived radon daughter 'loPb was also used in a couple of cases. This nuclide has the disadvantage of being difficult to measure by y-ray spectroscopy because of the low energy of its y-ray (46.5 keV).

STATIONARY SAMPLERS The laboratories taking part in the intercomparison test are presented in table 5.2.1. The characteristics of the stationary samplers are presented in table 5.2.2. As can be seen from table 5.2.2 six different types of filter were used, the most common being glass fibre. The FOA glass fibre filter is made by FOA and has qualities very similar to the commercially available Whatman GF/A glass fibre filter. Information on the physical characteristics and efficiencies of the FOA and Whatman GF/A glass fibre filters can be found in Lockhard et al. (1964). The Petrianov filter is described in Krehiak et al. (1981). The cellulose filter used by SMSR is of type C 135 made by BERNARD DUMAS S.A., Creysse, 24100 BERGERAC, FRANCE, and the polypropylene filter used by LMRE is made by JONELL INC., HOUSTON, TEXAS

77014, USA (Product number JP-3710).

MEASUREMENTS AND RESULTS Table 5.2.3 gives a summary of the intercomparisons performed.

Two transportable FOA

samplers were used at each site, except in the Faroe Islands, Iceland and at Lund, Sweden, where only one FOA sampler was used. In the stationary samplers some samples were taken using FOA glass fibre filters instead of the filters normally used. In the transportable FOA samplers FOA glass fibre filters were always used except on the icebreaker Oden where Microsorban filters were used. All samples were measured by the participating laboratory except at Montlhkry, France, where the filters from the FOA samplers were measured only by FOA. For comparison some of the samples from most of the participating laboratories were sent to FOA for measurement.

In the tables results are presented for each sample and the ratios between the concentration

388 TABLE 5.2.1 Participants in the intercomparison Laboratory Investigator Address FOA I. Vintersved National Defence Research Establishment Sundbyberg R i s l National Laboratory S. P. Nielsen RIS0 Roskilde Norwegian Radiation Protection Authority F. Ugletveit NRPA ijsteris Finnish Centre for Radiation and Nuclear Safety A. Leppanen STUK Helsinki Service Mixte de SCcuritC Radiologique Y. Bourlat SMSR MontlhCry Laboratoire de mesure de la radioactkite D. Calmet LMRE de I'environnement Orsay National Institute of Radiation Protection S. E. Palsson NIRP Reykjavik University of Faroe Islands T. Vestergaard N DV Thorshavn Institute of Radiation Physics IORP P. Roos Universitv of Lund

Country Sweden Denmark Norway Finland France France

Iceland Faroe Islands Sweden

TABLE 5.2.2 Stationary samplers used in the intercomparison. In the type of pump column, C stands for centrifugal and D for displacement. Laboratory Type Type of filter Filter size Capacity Flow rate of

FOA RIS0 N RPA STUK SMSR LMRE NlRP NDV IORP

Pump C D D D D

C C D D

FOA Glass fibre WhatmanGF/A WhatmanGF/A WhatmanGF/A Cellulose Polypropylene Petrianov WhatmanGF/A Membrane

Number 1 6 1 1 1 1 1 1 1

mxm

0.56~0.56 0.56~0.48 0.56~0.48 0.53x0.42 0= 0.11 0.47~0.28 0.42x0.40

m3/h 1100 1600

750 860 100 800 550

m/s

1.0 0.3 0.8 1.07

3 6

0.57~0.46

750

1.5 0.8

0.25x0.20

100

0.6

389 TABLE 5.2.3 Summarv of intercomDarisons performed. Laboratory Sampling period lntercomparison of 7Be 137Cs 'l0Pb FOA Apr-May 1988 Yes Yes No No Nov-Dec 1990 Yes Yes No Apr-May 1991 Yes Yes RIS0 No Jul-Dec 1991 Yes Yes N RPA No Feb-May 1992 Yes Yes STU K No No NlRP May-Sep 1992 Yes No No Aug-Sep 1991 Yes IORP No No Mar-May 1993 Yes No No Yes Jul-Oct 1992 N DV Yes Feb-May 1993 Yes Yes SMSR Yes No LMRE May-Aug 1993 Yes

measured by the stationary sampler and the mean values of the concentrations measured by the FOA transportable samplers. The mean value of the ratios is also presented together with the standard deviation and the standard error of the mean. Finally a t-test is performed to see if the ratio deviates significantly from the expected ratio of 1.0. The first tests were made

at FOA in 1988 and 1990. In table 5.2.4 and 5.2.5 the results are given for 'Be and 137Cs respectively. The t-test of the ratio between the stationary FOA sampler and the transportable ones implies a significant deviation at the 95% significance level, but the deviation is only less than 3%. For 137Cs the deviation is larger (8%), as can be seen in table 5.2.5. The test continued in April 1991 with two samplers at Risa. The 7Be comparison is presented in tables 5.2.6 and 5.2.7. When using Whatman GF/A in the stationary sampler there is a significant discrepancy between the FOA samplers and the Ris0 sampler, showing less 7Be in the R i s ~sampler than in the FOA ones. The three samples that were taken using the FOA filter in the Riser sampler do not show any significant deviation. Table 5.2.8 presents the samples where '"Cs

could be measured. In this case a mean value has been calculated

for the four samples ignoring the fact that one sample was taken with FOA filters in the Ris@ sampler and the other three with WhatmanGF/A filters. The mean value shows that with 95% certainty there is a deviation between the results with less "j7Cs in the Ris@sampler than in the FOA samplers. The next test was performed at NRPA in Bsteris. Tables 5.2.9 and 5.2.10 show the results from the 7Be intercomparison. There seems here to be a discrepancy similar to the Ris6 one, with less 7Be in the NRPA sampler when using Whatman GF/A. When using FOA filters in the NRPA sampler there is no significant deviation but there is a large uncertainty here because of only four measurements. Tables 5.2.11 and 5.2.12 show that, for 13'Cs, the concentration

in the NRPA sampler is significantly higher than in the FOA samplers. This is true for both

390 TABLE 5.2.4 Intercomparison, sampling of ‘Be at FOA, Sweden (pBq/m3) FOA FOA Ratio Mean FOA large Week number sampler 6 sampler 7 6/7 of 6 and 7 sampler

8815 8816 8818 8819 8820 8821 9045 9046 9047 9048 9049

3090 3330 3160 5660 3710 5360 1020 893 1450 1140 1220

1020 899 1460 1140 1250

Mean value Standard deviation Standard error of the mean t-test of mean = 1.0 Significance of deviation from 1.0

0.993 0.993 0.998 1.002 0.976 0.992 0.010 0.004 -1.533 ns

3090 3330 3160 5660 3710 5360 1020 896 1450 1140 1230

3080 3310 3210 5550 3560 5500 978 846 1360 1080 1140

Mean value Standard deviation Standard error o f the mean t-test o f mean = 1.0 Significance of deviation from 1.0

TABLE 5.2.5 Intercomparison, sampling o f I3’Cs at FOA, Sweden (pBq/m3) FOA FOA Ratio Mean FOA large Week number sampler 6 sampler 7 6/7 of 6 and 7 sampler

8815 8816 8818 8819 8820 8821 9045 9046 9047 9048 9049

12.3 10.6 7.56 9.70 6.52 16.7 2.8 1.67 2.67 3.99 1.76

2.23 2.01 2.72 3.77 1.67

12.3 10.6 7.56 9.70 6.52 16.7 2.52 1.84 2.70 3.88 1.72

9.11 11.3 7.61 8.78 5.86 15.9 2.29 1.38 2.47 3.61 1.79

1.26 0.83 0.98 1.06 1.05 Mean value 1.036 Mean value Standard deviation 0.155 Standard deviation Standard error of the mean 0.069 Standard error of the mean t-test of mean = 1.0 0.464 t-test of mean = 1.0 Significance of deviation from 1.0 ns Significance of deviation from 1.0

Ratio Large/Mean

0.997 0.996 1.017 0.981 0.959 1.026 0.959 0.944 0.934 0.944 0.926 0.971 0.034 0.010

-2.666 >95%

Ratio Large/Mean

0.738 1.063 1.007 0.905 0.899 0.953 0.909 0.750 0.917 0.930 1.044 0.920 0.104 0.031 -2.455 >95%

39 1 TABLE 5.2.6 Intercomparison, sampling o f 7Be at RIS0, Denmark (pBq/m3) FOA FOA Ratio Mean RIS0 Week number

sampler

6

sampler

7

6/7 6

of and

7

sampler with Whatman

Ratio RISB/Mean

GF/A

9115 9116 9117 9118 9121 9122 9123 9125

2600 2780 3200 1830 2160 2500 2640 1290

2580 2640 3350 1870 2160 2440 2700 1280

1.008 1.051 0.956 0.977 1.002 1.026 0.980 1.013 Mean value 1.002 Standard deviation 0.030 Standard error of the mean 0.011 t-test o f mean = 1.0 0.176 ns Significance o f deviation from 1.0

2590 2710 3280 1850 2160 2470 2670 1280

2290 2340 2780 1390 1900 2090 2190 1190

Mean value Standard deviation Standard error o f the mean t-test o f mean = 1.0 Significance of deviation from 1.0

TABLE 5.2.7 Intercomparison, sampling of 7Be at R I S 0 , Denmark (pBq/m3) FOA filters in all samplers. FOA FOA Ratio Mean RIS0 Week 6/7 of sampler number sampler 6 sampler 7 6 and 7 with

0.884 0.863 0.847 0.750 0.877 0.846 0.819 0.923 0.851 0.051 0.018 -7.691 >99.9%

Ratio RISB/Mean

FOA filter

0.967 1.057 1.037 Mean value 1.020 Standard deviation 0.047 Standard error of the mean 0.010 t-test of mean = 1.0 0.599 Significance of deviation from 1.0 ns 9119 9120 9124

2630 1720 1810

2720 1620 1750

2670 1670 1780

2540 1830 1570

Mean value Standard deviation Standard error o f the mean t-test o f mean = 1.0 Significance of deviation from 1.0

0.948 1.097 0.882 0.976 0.110 0.064 -0.312 ns

TABLE 5.2.8Intercomparison, sampling of 137Cs a t RIS0, Denmark (pBq/m3) FOA FOA Ratio Mean RIS0 Ratio Week number sampler 6 sampler 7 6/7 of sampler RIS0/Mean

6 and 7 3.60 1.62 3.79 0.90 3.60 3.07 2.04 0.69 1.73 1.00 2.09 0.79 1.88 1.07 Mean value 0.79 Mean value Standard deviation 0.11 Standard deviation Standard error of the mean 0.06 Standard error o f the mean t-test o f mean = 1.0 -2.70 t-test o f mean = 1.0 Significance of deviation from 1.0 ns Significance of deviation from 1.0 9115 9117 9118 9119

3.60 3.40 1.41 1.66

0.45 0.85 0.58 0.57 0.61 0.17 0.08 -3.97 >95%

392 TABLE 5.2.9 Intercomparison, sampling of 7Be a t NRPA, Norway (pBq/m3) FOA FOA Ratio Mean N RPA Week 6/7 o f 6 and 7 sampler number sampler 6 sampler 7

Ratio NRPA/Mean

with Whatman GF/A filter

1.109 1.016 1.056 1.060 0.984 1.054 1.063 0.967 1.018 0.984 0.991 0.994 0.976 0.971 1.015 Mean value 1.017 Standard deviation 0.042 Standard error of the mean 0.011 t-test o f mean = 1.0 1.505 Significance o f deviation from 1.0 ns 9130 9131 9132 9133 9134 9135 9141 91431 91432 91441 91442 91471 91472 9148 9149

2130 3840 1890 1060 1840 2350 2020 1470 1160 2530 1140 1570 830 1000 1350

1920 3780 1790 1000 1870 2230 1900 1520 1140 2570 1150 1580 850 1030 1330

2025 3810 1840 1030 1855 2290 1960 1495 1150 2550 1145 1575 840 1015 1340

1750 3430 1650 960 1800 2200 1700 1330 1000 2490 990 1390 740 720 1180

Mean value Standard deviation Standard error o f the mean t-test o f mean = 1.0 Significance o f deviation from 1.0

TABLE 5.2.10 Intercomparison, sampling o f 7Be at NRPA, Norway (pBq/m3) FOA filters in all samplers. FOA FOA Ratio Mean N RPA Week sampler 7 6/7 o f 6 and 7 sampler number sampler 6

0.864 0.900 0.897 0.932 0.970 0.961 0.867 0.890 0.870 0.976 0.865 0.883 0.881 0.709 0.881 0.890 0.063 0.016 -6.533 >99.9%

Ratio NRPA/Mean

with

FOA filter

1.022 1.192 0.860 1.077 Mean value 1.038 Standard deviation 0.138 Standard error of the mean 0.069 t-test of mean = 1.0 0.477 Significance of deviation from 1.0 ns 9137 9138 9139 9140

1410 1490 860 1120

1380 1250 1000 1040

1395 1370 930 1080

1390 1300 830 970

Mean value Standard deviation Standard error o f the mean t-test o f mean = 1.0 Significance of deviation from 1.0

0.996 0.949 0.892 0.898 0.934 0.049 0.024 -2.333 ns

393 TABLE 5.2.11 Intercomparison, sampling of Week FOA FOA Ratio number

sampler

9130 9131 9132 9133 9134 9135 9141 91431 91432 91441 91442 91471 91472 9148 9149

1.4 1.2 1.5 2.1 1.3 2.2 3.3 5.1 4.2 4.7

6

sampler

7

1.1 1.4 1.6 1.7 1.4 2.7 2.8 5.2 4.4 4.6 0.9 4.9 5.5 2.0 7.6

4.9 5.5 2.4 7.9

6/7

0.141 0.038 0.769 ns

TABLE 5.2.12 Intercomparison, sampling of FOA filters in all samplers. Week FOA FOA Ratio sampler

6

sampler

7

6/7

NRPA, Norway (pBq/m3) NRPA

7

1.25 1.3 1.55 1.9 1.35 2.45 3.05 5,15 4.3 4.65 0.9 4.9

1.000 1.000 1.200 1.039 1.030

at

Mean of 6 and

1.273 0.857 0.938 1.235 0.929 0.815 1.179 0.981 0.955 1.022

Mean value Standard deviation Standard error of the mean t-test of mean = 1.0 Significance of deviation from 1.0

number

13'Cs

sampler with Whatman GF/A filter

1.6 2.9 2.3 2.5 2.4 2.7 2.7 5.6 5.7 5.1 2.3 4.7 11.3 2.9 7.1

5.5 2.2

7.75

Mean value Standard deviation Standard error of the mean t-test of mean = 1.0 Significance of deviation from 1.0

""Cs of

Ratio

N RPA/ Mean

1.28 2.23 1.48 1.32 1.78 1.10 0.89 1.09 1.33 1.10 2.56 0.96 2.06 1.32 0.92 1.428 0.507 0.131 3.159 >99%

at N RPA, Norway (pBq/m3)

Mean 6 and

NRPA

7

sampler with

Ratio NRPA/Mean

FOA filter

9137 9138 9139 9140

1.063 1.357 0.700 0.952 Mean value 1.018 Standard deviation 0.272 Standard error of the mean 0.136 t-test of mean = 1.0 0.114 1.7 1.9 1.4 2.0

1.6 1.4 2.0 2.1

Significance of deviation from 1.0

ns

1.65 1.65 1.7 2.05

2.8 2.6 2.2 3.4

Mean value Standard deviation Standard error of the mean t-test o f mean = 1.0 Significance of deviation from 1.0

1.70 1.58 1.29 1.66 1.558 0.185 0.093 5.224 >98%

394

TABLE 5.2.13 Intercomparison, sampling of 7Be at STUK, Finland (pBq/m3) Week FOA number sampler 6

9209 9210 9211 9212 92131 92132 9214 9215

2050 1520 2460 2080 1620 3000 3030 2480

FOA sampler 7

2080 1470 2430 2150 1780 3020 2970 1080

Mean value Standard deviation Standard error of the mean t-test of mean = 1.0 Sianificance of deviation from 1.0

Ratio

6/7

0.986 1.033 1.014 0.965 0.910 0.994 1.021 2.306 1.154 0.467 0.165 0.872 ns

Mean of 6 a n d 7

2070 1500 2460 2110 1700 3010 3000 1780

STUK sampler with Whatrnan G F / A filter

Ratio STU K

/

Mean

2050 1480 2310 2110 1790 2920 2940 2500

0.993 0.988 0.946 0.999 1.053 0.970 0.981 1.404 Mean value 1.042 Standard deviation 0.149 Standard error of the mean 0.053 t-test of mean = 1.0 0.739 Sianificance of deviation from 1.0 ns

types of filter. The results from STUK in Helsinki are presented in tables 5.2.13, 5.2.14, 5.2.15 and

5.2.16. Here the agreement is good in all cases. During the intercomparison at SMSR, Montlhbry, the flow meter in the FOA sampler no.

7 failed in the first part of the test period. After replacement of the flow meter both samplers were used. As can be seen in table 5.2.17 there is a significant difference between the two FOA samplers. Although only 4 % it is still more than expected. It was probably the result of poor alignment of the new flow meter, as the newer models of the Fluid Inventor flow meters are very sensitive to the alignment of the measuring tube. Because the SMSR laboratory uses cellulose filters, their measuring procedure is not suitable for measuring the radioactivity on the glass fibre filters used in the FOA samplers. Therefore all samples from the FOA samplers were only measured by FOA, and the samples from the SMSR samplers were only measured by SMSR. For 'Be the agreement between the two stationary samplers used by SMSR is very good and there is excellent agreement between the mean values of the FOA samplers and the SMSR samplers. Table 5.2.18 presents the intercomparison of I3'Cs at SMSR. Poor statistics make it impossible to draw any conclusions from the data. At SMSR as well as at FOA the zl"Pb concentration was also measured routinely on the samples. That is why an intercomparison could be made on this nuclide too and the results are presented in table 5.2.19. There is a significant difference of 14 % between the FOA and SMSR results, but taking into account the fact that the *'OPb concentration is based on measuring the 46.5 keV y r a y , an energy in a region where the efficiency of the detector is very hard to determine due to self-absorbtion in the sample, the result is satisfactory. At SMSR use was made of a third stationary sampler

395 TABLE 5.2.14 Intercomparison, sampling o f FOA filters in all samplers. Week FOA FOA Ratio number

sampler

6

sampler

7

6/7

'Be a t

of

STUK, Finland (pBq/m3)

Mean 6 and

7

STUK

Ratio

sampler

STUK

with

FOA

i Mean

filter

9216 9217 9218 9219 9220

1930 2630 1960 2170 2300

2850 2690 1980 1940 2410

0.678 0.978 0.989 1.114 0.954

2390 2660 1970 2060 2360

1890 2620 1900 2100 2390

0.790 0.986 0.966 1.022 1.015 Mean value 0.943 Mean value 0.956 Standard deviation 0.160 Standard deviation 0.095 Standard error of the mean 0.072 Standard error o f the mean 0.043 t-test o f mean = 1.0 -0.711 t-test of mean = 1.0 -0.927 Significance o f deviation from 1.0 ns Significance of deviation from 1.0 ns

TABLE 5.2.15 Intercomparison, sampling of 137Csat STUK, Finland (pBq/m3) Week FOA FOA Ratio Mean STUK number

9209 9210 9211 9212 92131 92132 9214 9215

sampler

1.74 1.96 2.04 3.07 8.78 7.11 4.23 4.44

6

sampler

1.88 2.59 1.99 3.53 8.10 6.57 3.80 2.66

7

6/7

0.924 0.758 1.027 0.870 1.083 1.082 1.113 1.670 Mean value 1.066 Standard deviation 0.273 Standard error of the mean 0.097 t-test o f mean = 1.0 0.639 ns Significance o f deviation from 1.0

of

6 and 7

1.81 2.28 2.02 3.30 8.44 6.84 4.01 3.55

sampler with Whatman GF/A filter

1.70 2.03 1.91 2.82 8.53 6.65 3.77 5.19

Ratio

STUK Mean

0.939 0.891 0.948 0.856 1.010 0.975 0.939 1.461 Mean value 1.002 Standard deviation 0.191 Standard error o f the mean 0.068 t-test o f mean = 1.0 0.033 Significance of deviation f r o m 1.0 ns

396 TABLE 5.2.16 Intercomparison, sampling of 13’Cs at STUK, Finland (pBq/m3) FOA filters in all samplers. Week FOA FOA Ratio Mean STUK number sampler 6 sampler 7 6/7 o f 6 and 7 sampler

Ratio

STUK

i

with

FOA

Mean

filter

9216 9217 9218 9219 9220

2.33 7.08 6.15 1.41 3.43

4.41 7.32 6.28 1.03 4.45

0.528 0.967 0.980 1.367 0.771 Mean value 0.923 Standard deviation 0.309 Standard error of the mean 0.138 t-test o f mean = 1.0 -0.499 Significance o f deviation from 1.0 ns

3.37 7.20 6.21 1.22 3.94

3.30 7.79 6.88 1.35 4.03

0.978 1.082 1.108 1.104 1.022 Mean value 1.059 Standard deviation 0.057 Standard error o f the mean 0.025 t-test o f mean = 1.0 2.072 Significance o f deviation f r o m 1.0 ns

TABLE 5.2.17 Intercomparison, sampling o f 7Be at SMSR, France (pBq/m3) FOA FOA Ratio Mean SMSR SMSR Ratio Mean Week number smpl 6 smpl 7 6/7 6 7 smpl 1 smpl 2 1/2 12

9305 9306 9307 9308 9309 9310 9310 9312 9313 9314 9315 9316 9317 9318

3820 2020 2240 2020 2460 3430 2940 3520 3450 2680 3580 3890 3980 4500

3250 2620 3450 3850 3730 4320

Mean value Standard deviation Standard error o f the mean t-test of mean = 1.0 Significance o f deviation from 1.0

3820 3470 3620 2020 2030 2040 2240 2300 2220 2020 1980 2050 2460 2340 2390 3430 3590 3710 2940 3210 2820 3520 3480 3570 3310 1.062 3350 3210 1.025 2650 2580 2500 1.039 3520 3550 3730 1.011 3870 3900 3990 3860 1.068 3860 3790 1.044 4410 4280 4360 Mean value 1.042 Standard deviation 0.022 0.009 Standard error o f the mean 4.269 t-test o f mean = 1.0 >99%

Significance o f deviation from 1.0

0.959 0.995 1.036 0.966 0.979 0.968 1.138 0.975 0.970 1.032 0.952 0.977 0.982 0.982 0.994 0.048 0.013 -0.451 ns

3545 2035 2260 2015 2365 3650 3015 3525 3260 2540 3640 3945 3825 4320

Ratio

SMSR /FOA

0.928 1.007 1.009 0.998 0.961 1.064 1.026 1.003 0.973 0.959 1.035 1.019 0.991 0.980 0.997 0.035 0.009 -0.309 ns

397

TABLE 5.2.18 Intercomparison, sampling of 137Cs at SMSR, France (pBq/m3) Week FOA FOA Ratio Mean SMSR SMSR Ratio Mean Ratio number smpl 6 smpl7 6/7 6 7 smpl 1 smpl 2 1/2 1 2 SMSR /FOA 9305 1.9 1.9 3.02 1.58 1.911 2.30 1.211 1.8 2.50 2.00 1.250 2.25 1.250 9306 1.8 1.2 0.76 0.633 0.76 9307 1.2 0.7 1.28 1.63 0.785 1.46 2.079 9308 0.7 1.9 0.942 1.72 1.86 0.925 1.79 9309 1.9 1.483 1.2 2.28 1.28 1.781 1.78 9310 1.2 2.483 0.6 1.67 1.31 1.275 1.49 9310 0.6 0.9 1.11 1.11 1.233 9312 0.9 1.000 0.909 1.05 1.05 1.05 9313 1.0 1.1 1.200 0.55 9314 0.6 0.5 1.714 0.95 9315 1.2 0.7 1.522 0.800 0.9 1.37 1.37 9316 0.8 1.0 1.833 1.7 1.56 1.37 1.139 1.46 0.862 9317 2.2 1.2 1.094 1.286 0.8 0.83 0.92 0.902 0.88 9318 0.9 0.7 Mean value 1.246 1.316 Mean value 1.290 Standard deviation 0.410 0.522 Standard deviation 0.417 Standard error of the mean 0.145 0.151 Standard error of the mean 0.170 t-test o f mean = 1.0 1.587 2.008 t-test of mean = 1.0 1.555

-

SiRnificance o f deviation from 1.0

Significance of deviation from 1.0

ns

ns

ns

TABLE 5.2.19 Intercomparison, sampling of '"Pb at SMSR, France (pBq/m3) Week FOA FOA Ratio Mean SMSR SMSR Ratio Mean number smpl 6 smpl 7 6/7 6 7 smpl 1 smpl 2 1/2 12

9305 9306 9307 9308 9309 9310 9310 9312 9313 9314 9315 9316 9317 9318

1090 980 400 210 954 780 600 280 257 146 346 395 530 380

1.071 1.081 0.989 0.988 1.027 1.041 Mean value 1.033 Standard deviation 0.040 Standard error of the mean 0.016 t-test o f mean = 1.0 1.845 240 135 350 400 516 365

Significance of deviation from 1.0

ns

1090 980 400 210 9 54 780 600 280 248 140 348 398 523 372

1250 1140 311 244 840 860 750 324 286 178 413 497 647 441

1240 1310 321 271 960 980 660 339 256 133 435 497 642 391

1.008 0.870 0.969 0.900 0.875 0.878 1.136 0.956 1.117 1.338 0.949 1.000 1.008 1.128 Mean value 1.009 Standard deviation 0.131 Standard error o f the mean 0.035 t-test of mean = 1.0 0.248 Significance of deviation from 1.0

ns

1245 1225 316 258 900 920 705 332 271 156 424 497 644 416

Ratio

SMSR /FOA 1.142 1.250 0.790 1.226 0.943 1.179 1.175 1.184 1.093 1.107 1.218 1.250 1.232 1.117 1.136 0.129 0.034 3.801 >99%

398 equipped with FOA glass fibre filters. The results from this intercomparison are presented in tables 5.2.20 and 5.2.21. There is a significant difference indicating less activity of 'Be in the SMSR sampler than in the FOA samplers when using FOA filters in the SMSR sampler. But there is no difference in the results for 'lOPb. The intercomparisons made at LMRE are presented in tables 5.2.22 and 5.2.23. At LMRE the normal procedure is to run for 12 hours per day, from 8 pm until 8 am, for a 10 days period.

In the intercomparison test FOA sampler 6 ran at the same time as the LMRE sampler, while FOA sampler 7 ran continuously during the sampling period. As can be seen in table 5.2.22 there is excellent agreement between sampler 6 and the LMRE sampler. It is also interesting to notice that the two FOA samplers give the same result, even though one was sampling only

50% of the time. This indicates that the 7Be concentration varies smoothly with time. At LMRE it was also possible to compare the sampling of 21"Pb. Table 5.2.23 shows that the LMRE sampler retains much more ""Pb in the filters than do the FOA samplers. This large difference cannot be explained by self-absorption effects. The intercomparison at NIRP, Iceland, was made with only one FOA sampler. In table 5.2.24 the results are presented for 7Be which is the only radionuclide that could be used as the

137Csconcentration was too low.

On the Faroe Islands one sampler was used. The intercomparison took place in two very long sampling periods. As can be seen from table 5.2.25 the 7Be concentration found in the stationary sampler is only one fourth of that found in the FOA sampler. This is explained by the fact that the stationary sampler is located indoors, which of course reduces the amount of airborne particles that can reach the filters. The Institute of Radiation Physics in Lund participated in this intercomparison with an Andersen sampler on board the icebreaker Oden during the 1991 North Pole expedition August - October 1991. The FOA sampler was close to the Andersen sampler on the same deck. The

results are shown in table 5.2.26. The difference between the samplers is very large and the reason for this must be the difficulty in determining the flow rate in the Andersen sampler. An extra test was therefore made in Lund where the Andersen sampler was equipped with a Fluid Inventor flow meter on the exhaust pipe. Technical problems meant that it was only possible to get two weekly samples which were compared with the results from a FOA sampler placed close to the Andersen sampler. The results are shown in table 5.2.27. In this case the concentrations found in the FOA sampler are higher than in the Lund sampler. This can be explained by the fact that the flow rate in the Andersen sampler was much lower than expected and therefore was below the working range of the Fluid Inventor flow meter. After being measured at the participating laboratory some of the samples were sent to FOA for measurements and then returned to the laboratory for a second measurement. The results of this intercomparison are presented in tables 5.2.28 to 5.2.30. As the samples were

399 TABLE 5.2.20 Intercomparison, sampling of 7Be at SMSR, France (pBq/m3) FOA filters in SMSR sampler. Week Mean of SMSR Ratio numFOA sampler 3 SMSR/Mean ber sampler 6 and 7 with FOA filter

9309 9310 9310 9312 9313 9314 9315 9316 9317

2460 3430 2940 3520 3350 2650 3520 3870 3860

2120 3240 2780 3270 3010 2260 3420 3460 3650

Mean value Standard deviation Standard error of the mean t-test of mean = 1.0 Significance o f deviation from 1.0

0.862 0.945 0.946 0.930 0.899 0.853 0.972 0.894 0.946 0.916 0.041 0.014 -5.764 >99.9

TABLE 5.2.21 Intercomparison, sampling of ""Pb at SMSR, France (pBq/m3) FOA filters in SMSR sampler. Week Mean o f SMSR Ratio numFOA sampler 3 S MSR/ Mean ber samplers 6 and 7 with FOA filter

9309 9310 9310 9312 9313 9314 9315 9316 9317

954 780 600 280 248 140 348 398 523

750 770 570 317 210 152 437 429 597

Mean value Standard deviation Standard error of the mean t-test of mean = 1.0 Significance o f deviation from 1.0

0.768 0.987 0.950 1.132 0.847 1.086 1.256 1.078 1.141 1.029 0.150 0.050 0.546 ns

400 TABLE 5.2.22 Intercomparison, sampling of 'Be at LMRE, France (kBq/m3) Sampling FOA FOA Ratio LMRE 6/7 sampler period sampler 6 sampler 7

Ratio LMRE

i 6 0.788 0.987 1.085 1.080 1.000 1.173 0.903 1.053 1.097 0.975 1.027 1.015 0.104 0.031 0.456

FOA

1 2 3 4 5 6 7

a 9 10 11

4060 2390 3870 3520 2800 3410 2770 4750 2920 2000

3660 2470 3890 3440 2780 3340 2700 4430 2710

1.109 0.968 0.995 1.023 1.007 1.019 1.026 1.073 1.076

3200 2360 4200 3800 2800 4000 2500 5000 3200 1950 3000

1.033 0.045 0.015 2.074

Mean value Standard deviation Standard error of the mean t-test of mean = 1.0 Significance o f deviation from 1.0

3080

Mean value Standard deviation Standard error o f the mean t-test of mean = 1.0 Significance o f deviation from 1.0

ns

TABLE 5.2.23 Intercomparison, sampling of '"'Pb Sampling FOA FOA Ratio period sampler 6 sampler 7 6/7

at LMRE, France (pBq/m3) LMRE sampler

ns

Ratio LMRE

6 1.556 1.706 1.920 1.848 1.435 1.840 1.577 1.655 1.596 1.479 1.876 Mean value 1.681 Standard deviation 0.169 Standard error of the mean 0.051 t-test o f mean = 1.0 12.743 Significance o f deviation from 1.0 >99.9% FOA

1 2 3 4 5 6 7

a 9 10 11

540 170 250 330 230 250 260 390 245 285

520 190 220 350 230 225 250 370 205

0.963 0.895 1.136 0.943 1.000 1.111 1.040 1.040 1.201

315

Mean value Standard deviation Standard error of the mean t-test o f mean = 1.0 Significance of deviation from 1.0

1.037 0.099 0.033 1.057 ns

a40 290 480 610 330 460 410 640 390 420 590

40 1 TABLE 5.2.24 Intercomparison, sampling of 'Be a t NIRP, Iceland (pBq/m3) Sampling FOA NVD Ratio Type of period sampler sampler NVD/FOA filter in sampler 1580 1.07 Petrianov 92-06-04-06-26 1480 890 0.89 FOA filter 92-09-01-09-10 1000 1850 0.94 Petrianov 92-09-21-10-23 1970

TABLE 5.2.25 Intercomparison, sampling of 7Be at NDV, Faroe Islands (uBq/m3) Sampling FOA NVD Ratio period sampler sampler NVD/FOA 92-07-07-08-17 1120 253 0.23 92-08-20-09-28 881 219 0.25

TABLE 5.2.26 Intercomparison, sampling of 'Be on the icebreaker ODEN (mBq/m3) Sample FOA IORP Ratio number sampler sampler IORP/FOA with with Microsorban membrane filter filter LF129 1.05 1.54 1.47 1.94 2.26 LF130 1.16 1.08 1.69 LF131 1.56 1.04 1.78 LF132 1.71 0.51 0.83 LF133 1.63 0.61 1.06 1.74 LF134 0.23 0.40 1.74 LF136 LF137 0.27 0.39 1.44 Mean value 1.558 Standard deviation 0.196 Standard error of the mean 0.069 t-test of mean = 1.0 7.529 Significance of deviation from 1.0 >99.9%

TABLE 5.2.27 Test Week FOA number sampler pBqJm3 9320 5750 9321 4760

of Andersen sampler at IORP, Sweden Andersen Ratio sampler Andersen / FOA pBq)m3 5170 0.90 3800 0.80

402 TABLE 5.2.28 Intercomparison, 7-ray measurements of 'Be at RIS0, Denmark (pBq/m3) FOA sampler 6 Ratio RIS0 Ratio FOA Week RIS0 measuresecond Rl/FOA number first R2/FOA measurem. ment measurem.

1.068 1.082 FOA sampler 7

9117 9118

3200 1830

3000 1690

2930 1660

Week number

RIS0 first measurem.

FOA measurement

second measurem.

9117 9118

3350 1870

2940 1650

3020 1730

RIS0

Mean value Standard deviation Standard error of the mean t-test of mean = 1.0 Sirrnificance of deviation from 1.0

0.978 0.979

Ratio Rl/FOA

Ratio R2/FOA

1.140 1.135 1.106 0.036 0.018 5.029 >98%

1.028 1.048 1.008 0.035 0.018 0.405 ns

TABLE 5.2.29 Intercomparison, 7-ray measurements of 7Be at NRPA, Norway (pBq/m3) NRPA sampler Week number

NRPA measurement

FOA measurement

Ratio NRPA/FOA

9131 9140 9148

3430 970 720

3180 978 700

1.079 0.992 1.029

Week number

NRPA measurement

FOA measurement

N RPA / FOA

9131 9140 9148

3840 1120 1000

3380 978 933

1.136 1.145 1.072

Week number

NRPA measurement

FOA measurement

9131 9140 9148

3780 1040 1030

3420 984 921

FOA sampler 6 Ratio

FOA sampler 7

Mean value Standard deviation Standard error o f the mean t-test of mean = 1.0 Significance of deviation from 1.0

Ratio

NRPA/FOA 1.105 1.057 1.118 1.081 0.050 0.017 4.582 >99%

403 TABLE 5.2.30 Intercomparison, y r a y measurements of 7Be at STUK, Finland (pBq/rn3) FOA sampler 6 Ratio Ratio STUK STUK FOA Week Sl/FOA S2/FOA measuresecond first number measurem. ment measurem. 9210 1520 1540 1520 0.989 0.987 9213 3000 2700 2690 1.112 1.098 9216 1900 1930 1.017 FOA sampler 7 Ratio STUK Ratio STUK FOA Week Sl/FOA S2/FOA measuresecond number first measurem. ment measurem. 1470 1510 1.004 1.026 9210 1470 2910 1.100 1.059 3020 2750 9213 9216 2850 2860 2840 0.997 0.991 Mean value 1.040 1.030 Standard deviation 0.060 0.042 Standard error of the mean 0.027 0.017 t-test of mean = 1.0 1.341 1.562 Significance of deviation from 1.0 ns ns

measured at FOA in FOA’s measuring geometry, they had to be transferred from the original container to a FOA container. This procedure can have caused some loss of activity so one should not expect perfect agreement between the measurements. As can be seen from the tables all measurements agree within 15%.

CONCLUSIONS A summary of the results is presented in figure 5.2.2. There is less 7Be in the Rise and NRPA samplers than in the FOA sampler when WhatmanGF/A is used. The difference is not very large (<15%) but still significant. When using FOA filters in all samplers there is no significant difference apart from the stationary samplers at FOA and SMSR. One can only speculate whether this is an indication that the WhatmanGF/A filter is somewhat less efficient at lower flow rates than the FOA filter for the 7Be particles. Nevertheless taking into account all uncertainties involved in these measurements, an agreement within 15% is still satisfactory. For 137Cs the deviations between the results are much larger and no definite trend can be found. The STUK sampler shows excellent agreement in all cases. In the NRPA sampler there is a 50% deviation from the expected result with too much 137Cs. A possible explanation for this difference could be that, in the NRPA sampler, the opening into the filter compartment measures only 0.064 mz while the filter area is 0.269 m2. This means that even though the flow rate through the filter is 0.8 m/s the air is sucked into the sampler with a speed at the inlet of 3.3 m/s. So there is no laminar flow into this sampler and that, together with the fact that the 137Csparticles have a size distribution that differs from that of the 7Be particles with

404

m

RlSO ' NRPA . STUK ' SMSR . LMRE NlRP .

YI I+

It IH

'

NDV.

IORP . FOA * RlSO '* NRPA STUK SMSR

'

'

-

RlSO * NRPA STUK . SMSR *

7

Be

4

I I-H

m It

7

-4

m

-H 4 HI

I-4

Be FOA filters

137

cs

m

FOA . NRPA . STUK *

m

137

m

SMSR . LMRE .

m

m

SMSR .

*"Pb 210

Pb FOA filters

k H

0.5

1

cs

FOA filters

1.5

Ratio Laboratory sampler / FOA sampler Figure 5.2.2 Summary of the intercomparison. Error bars indicate the standard error of the mean at a confidence level of 95%.

405

much larger particles for Cs, might be the reason for the discrepancy. However these differences require further study. The conclusion must be that ':'7Cs in such low concentrations as found today in northern and western Europe is not suitable for use in an intercomparison. Moreover the 137Cswhich results from resuspension may be more sensitive to the micro weather around the stations, so that very different results can be obtained even from very close-lying samplers. The few results on "'Pb

also indicate unexplained differences. On the other hand 7Be proved

to be an excellent nuclide for intercomparison tests.

ACKNOWLEDGEMENTS All the participants and laboratories that contributed their time and facilities to the present work are hereby acknowledged. The French participation has been possible thanks to financial support from LMRE and SMSR.

REFERENCES Kreelniak J. W. and J . Porstend6rfer (1981). Experimental Determination of the Collection Efficiency of the Petrianov Type F P P 15-1.7 Filter for Sampling Submicron Atmospheric Aerosols. Nukleonika 20 669-679. Lockhard, L. B. and R. L. Patterson (1964). Characteristics of air filter media used for monitoring airborne radioactivity. Report NRL 6054. Also in: 0. Suschny, Technical Note No. 94, The Measurement of Atmospheric Radioactivity, W M O - No. 231.TP.124, 1968.

This Page Intentionally Left Blank

407

5.3. INTERCALIBRATION OF WHOLEBODY COUNTING SYSTEMS TUA RAHOLA’, ROLF FALK’ AND MICHAEL TILLANDER3 I Finnish Centre for Radiation and Nuclear Safety, P.O. Box 14, FIN-00881 Helsinki, Finland. Swedish Radiation Protection Institute, Box 60204, S-104 01 Stockholm, Sweden. Department of Radiochemistry, P.O. Box 5, FIN-00014 University of Helsinki, Finland.

SUMMARY Whole-body counting is an established method for the determination of radionuclides that emit X or gamma radiation in the human body. Quantitative measurement is necessary for radiation protection purposes, but requires reliable calibration. Whole body counters are generally calibrated by means of phantoms, i.e. containers resembling the human body and containing calibrated solutions of the radionuclides to be measured. After the Chernobyl accident it was found necessary to be able to measure any member of the general population, including children. In order to extend the calibration of existing systems a modular phantom was developed, which can be stacked to resemble humans of varying weigth and height. Twenty whole-body counting installations in the Nordic countries participated in the intercalibration project. The results were better than expected; systematic errors were usually less than +lo%. INTRODUCTION ‘Whole-body counting’ is a colloquial term for the measurements of X and gamma radiation emitted from radionuclides in the human body. A typical whole-body counting facility consists of

shielding, detectors and analyzers. The geometry of the measuring system, i.e. the position of the subject’s body in relation to the detector@),is also an important parameter of the installation.

Shielding must be used to reduce the influence of cosmic radiation and radionuclides present in the environment. Low-activity concrete, old steel, and lead are typical shielding materials. The shielding may enclose the system completely (measuring rooms) or be partly open (shadowshielding). Two types of detector are in common use: NaI(T1) scintillation crystals and semiconductor

detectors. Scintillation crystals have higher detection efficiencies but poorer energy resolution than semiconductor detectors. The purpose of the analyzing equipment is to provide an energy-dispersive measurement, i.e. a gamma spectrum of the radiation emitted from the person or object being measured.

408 Personal computers are generally used to store spectra and to compute the activities present, using either commercial or in-house software. The two main types of subject geometry are the bed geometry, usually with longitudinally scanning detectors, and the chair geometry. Chair geometries usually give higher counting efficiencies because the subject is closer to the detector. The penalty for this is a higher sensitivity to anthropomorphic parameters such as weight and length, which means difficulties in calibrating the system for subjects of varying size and shape. Comprehensive reviews of the method can be found in Toohey et al. (1991) and (IAEA, in press) The need for surveillance after the Chernobyl accident provided a fresh impetus to the field of whole-body counting. New installations were constructed and older systems refurbished. There has also been considerable scientific progress. The Nordic countries (Denmark, Finland, Norway and Sweden) have a total of some 20 whole-body counters in operation. These are used for a variety of purposes, including:

* surveillance of the general population or specially selected population groups; * surveillance at special sites, e.g. nuclear power plants; * medical or other scientific research; and * clinical measurements. EFFICIENCY CALIBRATION OF WHOLEBODY COUNTERS

The efficiency calibration of whole-body counters present special problems. The human body is not a very amenable 'sample' for activity measurement. Humans differ considerably in size and shape. The absorption of electromagnetic radiation is not uniform for all tissues. The distribution of the radionuclides to be measured is rarely homogeneous within the body. There are four different methods for efficiency calibration:

Phantom calibration An anthropomorphic phantom is a volume of material that in all respects (shape, size, density and

atomic number) represents the subject or group of subjects to be measured. In practice a phantom is an approximation of the ideal and usually consists of containers that are filled with a solution of known activity of the radionuclide in question. In many cases a homogeneous distribution of the radioactivity gives a sufficient accuracy. Depending on the physiological behaviour of the radioactive element or compound, a heterogeneous distribution, i.e. concentration in specific organs or tissues, may have to be simulated by using special assemblies.

409

Point source calibration Small solid sources are placed in cavities in an anthropomorphic mannequin. Single solid sources may also be used for routine checks on the efficiency. The latter method is commonly used to check the validity of a factory calibration of an installation.

Calibration by computer simulation Sophisticated computer programs may be used to calculate the measuring efficiencies for different energies and different subject geometries. Use of this method has so far been restricted to extending the validity of other calibration methods.

In-vivo calibration Physical or computational phantoms do not simulate the human body accurately in every respect. For example, the absorption of gamma radiation in a phantom filled with an aqueous solution is only 85% of that in the human body (Kramer et al., 1991). The best estimate of the measuring efficiency can be obtained by measuring a person containing a known amount of the radionuclide in question. Administering radionuclides to humans solely for this purpose is not, however, justified even if small amounts of activity are used.

AIMS Emergency planning for situations with large-scale radioactive contamination did not in the past include the use of whole-body counters for surveillance. In the early 1980’s this attitude was revised, and subsequent events have demonstrated the value of the method. A prerequisite for the utilization of all available whole-body counting facilities, including those in neighbouring countries, is a common calibration standard. Commercially available phantoms are usually fabricated for a specific geometry and thus fit poorly into other measuring arrangements. A project with funding from the Nordic Liaison Committee for Atomic Energy ( N U ) was initiated in 1984 to develop a modular arrangement, i.e. the phantoms were constructed of identical modular blocks, which could be arranged to represent persons of different weight and length in the measuring geometries used by the participating laboratories. Other requirements were that the possibility of leakage from the phantoms should be minimized, and that the total activity and the activity concentration should not exceed the exemption limits for sea and road transport. The final report of this project is included as Annex A.

410 The present study continues the NKA project. Homogeneous, modular phantoms were fabricated for the caesium radioisotopes '"Cs and '"Cs. The participating laboratories were requested to construct phantoms representing subjects of different weight and length, measure the phantoms and report their results for this report.

MATERIALS AND METHODS Phantom modules Flat plastic containers with a capacity of 550 ml were selected for the modular phantoms. They were manufactured by a Finnish company (Plastex Oy, FIN-08700 LOHJA) for use as ice packs and have the dimensions 190 mm

*

105 mm

* 35 mm.

Calibrated solutions of the caesium radioisotopes were obtained from Amersham International plc, Amersham, UK. Caesium carrier, hydrochloric acid and water, were added, and the tared bottles were filled with this solution, stoppered, sealed with silicon sealant and weighed. Table 1 lists the main characteristics of the phantoms. Table 5.3.1. Modular phantoms

Nuclide

(3-137

CS-134

Number Name

Reference date

Content of one module Activity

Net weight

Total weight

NKS-I Glader

15 FEB 1991

115 Bq

538 g

6oog

NKS-I1 Glader

15 FEB 1991

135 Bq

533 g

595 g

NKS-111 Butter

01 MAR 1992

198 Bq

550 g

612 g

NKS-IV Butter

01 MAR 1992

193 Bq

550 g

612 g

Geometries It was not possible to provide strict instructions for stacking the modules into phantoms, since the measuring geometries vary widely. The basic arrangement for a subject weighing 70 kilograms (the primary calibration point for most systems) was obtained by stacking the modules beside a 70 kg BOMAB (Bush, 1946) phantom on a flat surface (Figure 5.3.1).

41 1

Bed geometries are much easier to standardize than chair geometries. Auxiliary constructions were in some cases necessary to support the phantoms in the chairs. Making up bundles of 2, 3 and 4 modules using adhesive tape aided in stacking phantoms.

Figure 5.3.1. The phantom to the left is constructed of plastic containers, that to the right is a commercial specimen from Gray & Martin Ltd. Participating installations The participating laboratories are listed in Table 5.3.2 with the principal characteristics of the

installations in Tables 5.3.3a (bed geometries), 5.3.3b (chair geometries with scintillation detectors) and 5.3.3~(chair geometries with semiconductor detectors).

412 Table 5.3.2. Participating laboratories.

Number

Contact person, name, country

1

Klaus Ennow, Institute for Radiation Hygiene, Danmark

2

Bente Lauridsen, Ris0 National Laboratory, Danmark

3

Michael Tillander, University of Helsinki, Department of Radiochemistry, Finland

4, 5

Tua Rahola, Finnish Centre for Radiation and Nuclear Safety, Finland

6, 7

Torolf Berthelsen, Norwegian Radiation Protection Authority, Norway

8

Oyvind Bergene, Institutt for Energiteknikk, Norway

9

Trygve Bjerk, Institutt for Energiteknikk, Norway

10

Rolf Falk, Swedish Radiation Protection Institute, Sweden

11

Orvar Sundgren, Studsvik Nuclear AB, Sweden

12

Hans Tovedal, The National Defence Research Establishment, Sweden

13

Kerstin Muntzig , University of Goteborg, Department of Radiation Physics, Sahlgrenska sjukhuset, Sweden

14

Lennart Johansson, University of U m d , Department of Radiation Physics, Sweden

15

Lennart Lindgren, Barseback Nuclear Power Plant, Sweden

16

Monica Eklof, Vattenfall AB, Forsmark Nuclear Power Plant, Sweden

17

Christer Solstrand, OKG Aktiebolag, Sweden

18

Annete Lijvefors, Vattenfall AB, Ringhals Nuclear Power Plant, Sweden

19

Bengt Hemdal, Malmo General Hospital, Department of Radiation Physics, Sweden

20

Bengt Hemdal, University of Lund, Department of Radiation Physics, Sweden

413 Note the following abbreviations in Tables 5.3.3a, 5.3.3b and 5.3.3c, column 'Calibration': BOMAB: Bottle Manikin ABsorption, phantom bottles fabricated for this purpose. Bottle; Similar to BOMAB, but using commercially available plastic containers. Modular; Flat plastic containers (ice packs), about 0.5 litre each. Livermore: Complex phantoms constructed from solid tissue-equivalent material. BOMAB, bottle and modular phantoms are filled with calibrated solutions of the appropriate radionuclides. Solid sources are inserted into the Livermore phantoms. Details of the measuring systems and calibration procedures can be obtained from the contact persons listed in Table 5.3.2. RESULTS Standard geometries It is common practice to calibrate each system for the body size and geometry for which the system is designed, usually a body mass around 70 kg and a length of 1.8 m. Results for phantoms conforming to this 'standard' or 'design geometry' are listed in Table 5.3.4 for both caesium isotopes: total phantom weight in kg, photo-peak efficiency in s-'/kBq (for 134Csthe 796 keV photo-peak is used) and the quotient measured activity/expected activity.

Weight dependency The counting efficiency in s-'/kBq is shown as a function of total phantom weight in Figures 5.3.2-5.3.7. Scanning bed geometries are presented in Figures 5.3.2 and 5.3.3. Note that laboratories 4, 10 and 13 use longitudinal scan. In laboratory 1 the detectors were stationary during measurement. Figures 5.3.4 and 5.3.5 show the chair geometries using semiconductor detectors. For the chair geometries with NaI(T1) scintillation detectors (Figures 5.3.6 and 5.3.7) the counting efficiencies have been normalized by dividing them by the detector volumes, in cubic decimeters. CONCLUSIONS The mechanical performance of the modular phantoms was satisfactory. The packs were relatively easy to assemble and withstood moderate mechanical stress. Filling the modules must be done with care, avoiding overfill and excluding air bubbles, otherwise the modules will be difficult to stack because of bulging. Leakages occurred very rarely. The reproducibility of arranging the phantoms in the standard geometry was tested by only one laboratory. The standard deviation for five results, the phantom having been rearranged for

Table 5.3.3a: Installations. Scanning bed geometry with NaIW) detector(s)

Laboratory Number 1

Number and size of detectors

Measuring chamber

Calibration method (see text)

Background

Software

4 detectors

Low-activity concrete t steel 150 mm lead 2 m m t cadmium 1 mm

2-litre bottles

Empty chamber

Nuclear Data AccuSpec

Modular phantoms of variable size

BOMAB or variablesue sugar phantom

In-house program for preset energy windows

Empty chamber

Manual calculation

B O M B (4 sizes)

Empty chamber

In-house program for least-squares fit of reference spectra

5-litre bottles and SSI phantom

Empty chamber

Manual calculation

BOMAB

Empty chamber

In-house program for preset energy windows

150 mm dia * 100 mm

+

4 detectors

Steel 150 mm t lead 3 mm t copper 1 mm

6

1 detector 100 mm dia * 150 mm

Low-activity concrete 0.6 m t olivine gravel 0.4 m + lead 30 mm

10

3 detectors 125 mm dia 100 m m

Low-activity concrete 0.6 m 40 mm

4

-

125 mm dia * 100 mm

*

+ lead

13

2 detectors 125 mm dia * 100 mm

Magnetite concrete 0.6-0.9 m + steel 150 mm + lead 5 mm

19

2 detectors 125 mm dia * 100 mm

Steel 150 mm lead 3 mm

+

Table 5.3.3b: Installations. Chair geometry with single NaI(T1) detector

Calibration method

Background

Software

Shadow shield: lead 40 mm + copper 1 mm

Modular phantoms of variable size

Modular phantoms of variable size

MicroSAMPO peak fitting program

76 mm dia * 76 mm

Shadow shield: lead 4-10 mm

Comparison with scanning bed results (N:o 6)

Empty chair

In-house program

8

76 mm dia * 76 mrn

Shadow shield, lead

Bottle phantom, 50 kg

9

150 mm dia * 225 m m

Steel

11

200 mm dia 100 mm

Steel 160 mm lead 3 mm

12

125 mm dia * 125 mm

Shadow shield: lead 50 mm

20

200 mm dia * 100 mm

Steel 200 mm lead 3 mm

Size of detector

Measuring chamber

3

200 mm dia 100 mm

7

Laboratory Number

*

*

(see text)

Maestro (Ortec)

+

+

BOMAB

Empty chamber

Nuclear Data peak finding program

BOMAB

Sugar phantom

In-house program for preset energy windows

BOMB

Sugar phantom

In-house program for preset energy windows

Table 5.3.3~:Installations. Chair geometry with semiconductor detector(s)

Laboratory Number

Number and relative efficiency of detector(s)

Measuring chamber

Calibration method (see text)

Background

Software

2

One, 54%

Steel 300 mm

Bottle phantom

Empty chamber

In-house software

5

One, 56%

Shadow shield lead >70 mm

Modular phantoms, variable size

Sugar phantom

In-house with peak search

12b

One, 50%

Shadow shield lead 4-10 mm

Modular phantoms of variable size

Sugar phantom

14

One, 36%

Shadow shield lead 50 mm

B O W

Bottle phantom 70 kg

Oaec library

15

One, 55%

Steel 120 mm lead 3 mm

+

BOMB

Empty chamber

Omnigam

16

One, 21%

Steel 130 mm lead 20 mm

+

BOMAB

Empty chamber

Nuclear Data peakfinding program

17

Two, 23%

Concrete room, detectors shielded with: copper 2 mm cadmium 1 mm + lead 100-300 mm

Livennore phantom

Phantom

Canberra Packard WBC-6000

18

One, 18%

Steel 150 mm

+ Empty chamber

417

Table 5.3.4 Results for standard geometries

CS-137

CS-134

N:0

Weight

Effic.

M/E

Weight

Effic.

1

77

3.69

0.83

72

4.41

2

61

0.194

0.97

63

0.136

0.95

3

65

4.69

1.05

65

4.3

1.oo

4

73

5.41

1.04

75

5.16

1.oo

5

71

0.242

1.05

75

0.195

0.95

7

73

0.85

0.78

75

0.68

0.63

8

0.508 2.0

1.01 -

78 83

0.497 2.75

-

9

78 73

10

68

2.06

1.10

70

1.87

1.02

M/E

6

-

11

72

4.93

0.93

72

4.36

0.84

12 12b

78 70

0.616

1.15

-

0.59 0.104

1.07

0.115

72 72

-

13

70

2.14

1.14

72

1.41

1.10

14

71

0.24

1.01

72

0.21

0.97

15

71

0.45

1.43

65

-

1.06

16

74

0.09

1.08

74

0.128

1.16

17

75

0.14

0.94

75

0.122

1.04

18

71

0.103

1.42

71

0.079

1.23

19

77

3.41

1.11

79

3.03

1.11

20

77

3.92

1.02

-

-

-

Weight: total weight of solution and bottles, in kg Effic.: measuring efficiency, i.e. photopeak pulse rate divided by activity, in s-'/kBq. For IMCsthe 796 keV photopeak is used. M / E measured activity divided by expected activity

418

Figure 5.3.2. Normalized counting efficiency for homogeneously distributed 137Cs as a function of total phantom weight, bed geometry using NaI (Tl) detector(s).

Figure 5.3.3.Normalized counting efficiency for homogeneously distributed IMCsas a function of total phantom weight for bed geometry, using NaI (Tl) detector(s).

419

Figure 5.3.4.Normalized counting efficiency for homogeneously distributed 137Csas a function of total phantom weight, chair geometry using a semiconductor detector.

Figure 5.3.5. Normalized counting efficiency for homogeneously distributed 134Csas a function of total phantom weight for chair geometry using a semiconductor detector.

420

Figure 5.3.6. Normalized counting efficiency for homogeneously distributed I3’Cs as a function of total phantom weight, for chair geometry using a NaI (Tl) dctector.

Figure 5.3.7. Normalized counting efficiency for homogeneously distributed 134Csas a function

of total phantom weight for chair geometry using a NaI (Ti) detector.

42 1 each measurement, was one per cent of the counting rate. The statistical uncertainty of the counting rate is in most cases insignificant. The quotient measured activity/expected activity (Figure 5.3.8) for the standard geometries was between 0.8 and 1.2 for most laboratories. This result is better than expected and indicates a measuring accuracy quite sufficient for surveillance and radiation protection purposes.

2.0

MeasuredJExpexted -

-

1.0

0 CS-137 CS-134

n

'

0.5

1

3 2

5 4

7 6

9 8

11 10

13 12

15 14

17 16

19 18

20

Figure 5.3.8. The quotient measured activity/expected activity for eighteen laboratories. The influence of phantom weight on the counting efficiency is smallest in the scanning bed geometries. This advantage is lost if the scan is disabled, i.e. the detectors kept stationary (Laboratory 1). Chair geometries using scintillation detectors exhibit considerable dependency on body weight; this can be reduced by sacrificing efficiency in an arc-shaped geometry (Laboratory 12). Semiconductor detectors generally are less sensitive to changes in weight, but unexpected divergences at small weights can be seen for some systems. Many whole-body counting systems are designed for measurement of specific target groups, e.g. radiation workers. This study shows that their range of subjects can be extended with proper calibration. Modular phantoms are useful for calibrating systems with varying geometries. The accuracy of the whole-body counting technique was demonstrated to be very satisfactory.

REFERENCES F. Bush, Br. J. Radiol. 19 (1946) 14-21 Direct methods for measuring radionuclides in man, L4EA Safety Practices, in press G.H. Kramer, Linda Burns and L. Noel, Health Phys. 61 (1991) 895-902 R. Toohey, E. Palmer, L. Anderson, Carol Berger, N. Cohen, G. Eisele, B. Waccholz and W. Burr, Jr., Health Phys. 60: Suppl. 1 (1991) 7-41

422 Annex A

NORDIC INTERCOMPARISON OF WHOLEBODY COUNTERS 1984.1985

In order to improve possibilities of meeting the requirements of the bilateral agreements on assistance in emergency situations made between the Nordic countries, in 1984 the Nordic Liaison Committee for Atomic Energy decided to start an intercalibration project on wholebody counting. Whole-body counting results can be used for assessment of internal doses. The counting procedures and dose assessment of internal doses. The counting procedures and dose assessment methods have to be comparable. The Finnish Centre for Radiation Safety was responsible for this intercalibration programme for Nordic whole-body counters. The task for each participating whole-body counting laboratory was to measure a phantom made of plastic bottles (Fig. 5.3.1 in main text) homogeneously filled with an unknown aqueous solution and to determine what radionuclides were contained in the phantom and their concentrations.

The two radionuclides chosen for this purpose were 137Csand %o, 137Csbecause of the high probability of it being the contaminant causing the largest dose after a nuclear accident, and because it is the radionuclide causing most concern among those detected in nuclear power plant workers.

Table 5.3.A.1 gives the type of detector and the measuring geometry used by the 16 participating laboratories.

423 Table 5.3.A. 1.Type of detector and geometry used in the Nordic whole-body counting intercalibration project.

Lab.no

Country

Geometry

Detector(s)

1

Sweden

chair

1 Ge

2

Sweden

bed

4 planar Ge

3

Norway

bed

4

Sweden

scanning bed

1Na-m 3 NaI(T1)

5

Norway

chair

1 NaI(T1)

6

Finland

chair

1 HPGe

7

Sweden

chair

2 x HPGe

8

Sweden

scanning bed

1 NaI(T1)

9

Sweden

chair

1 NaI(T1)

10

Sweden

chair

1 NaI(T1)

11

Sweden

chair

1 NaI(T1)

12

Sweden

scanning bed

1 NaI(T1)

13

Norway

chair

1 NaI(T1)

14

Finland

scanning bed

4 NaI(T1)

15

Denmark

chair

1 NaI(T1)

16

Sweden

chair

1 NaI(T1)

The time schedule for the work was tight. Therefore it was decided that two experts from Finland would travel with the phantom to the whole-body counting laboratories. These experts also loaded the phantom into the measuring position at each laboratory. The measurements and calculation of the results were carried out by the local staff. All laboratories identified the nuclides '37Csand 6oCo.The results for phantom contents are given in Table 5.3.A.2. It appeared that bed geometries give results in good agreement with the "true" values, except for 6oCo measured in the system with a planar germanium detector designed for measurements of uranium in lungs and not for whole-body counting. Some systems were not routinely used, and some were only intended for qualitative checking of possible internal contamination.

424

Table 5.3.A.2

Contents of 6oCoand '37Cs(kBq) in intercomparison phantom "Sleepy" as

reported by the participants in the Nordic projekt (NKA) in 1984-85. The contents of "Sleepy" were 22.3 kBq %o and 17.3 kBq 137Cs.

Lab. no

1 2 3 4 5 6 7" 8 9 10 11 12 13 14 15 16

34.3 k0.8 40 & 7.5 18.9 20 2 19.8 53.0 18.2 2 1.8 55

23.5 k0.6 30 5 18.5 15 5 1.5 16.3 k2.4 14.9 k 1.5 43

18.4 k 0.08 8.7 16.7 18.0 8.9 20.5 k 2.0 35 5 5 17.1

13.8 0.08 9.6 10.4 13.2 7.0 15.1 1.5 28 k 4 12.5

Mean (min - max)

23 5 12 (8.7-55)

18.1 & 9.4 (7.0-43)

*

*

'Preliminary calibration used

Many of the participants had no chance of checking their calibration factors before this intercomparison project. After the results were collected and presented at a meeting of the Nordic Liaison Committee for Atomic Energy in 1985, all participants could compare their own results with those considered "true" values. For radiation protection purposes, and especially in accident situations, the performance of the whole-body counting procedures were found to be statisfactory. For more demanding internal dose calculations, improved procedures were needed.

425

5.4 INTERCALIBRATION OF GAMMA-SPECTROMETRIC EQUIPMENT Elis Holm Department of Radiation Physics, Lund University, Sweden

SUMMARY The results are reported of an intercomparison exercise on samples of t e r r e s t r i a l origin (bark from deciduous and coniferous trees) designed f o r the determination of radiocaesium. Data have been evaluated from 26 laboratories representing all t h e Nordic countries. The mean values f o r I3'Cs were 28.9k5.3 Bq kg-' and 47.528.4 B$ kg-' 134 respectively and the corresponding values f o r Cs were 2.9t1.1 Bq kg and 2.8k1.6 Bq kg-' respectively. The results show that most laboratories produced d a t a within acceptable ranges.

INTRODUCTION The most common method f o r assessment of the radioactive contamination of our environment is gamma-spectrometry. This limited sample preparation,

provides

one measurement and,

at

not

method takes little time , requires results f o r

least,

most

of

several radionuclides in

our

radiologically important

fission and activation products from controlled or accidental

radionuclides,

releases from nuclear power plants a r e gamma emitting radionuclides. In

1991 t h e

Department

of

Radiation

Physics,

Lund,

prepared

and

distributed two samples of terrestrial origin (bark from t r e e s in deciduous

or coniferous forest). I t was anticipated t h a t these samples would contain moderate levels of radiocaesium from nuclear weapons testing but mainly from the

Chernobyl

accident.

The

samples were distributed

to

a

total

of

39

laboratories. Within

t h e different

NKS programmmes,

radiocaesium was

t h e major

radionuclide t o be assessed with respect t o radioecology and doses t o man. The

participating

laboratories

(134Cs, 137Cs) and 40K these

laboratories

were

instrumental

to

determine

by using gamma-spectrometric

also

participated

samplers. This gamma-spectrometric for

requested

correlation

radioactivity measurement.

for

in

an

radiocaesium

technique.

intercalibration

of

Several of large

air

intercalibration thus constitutes a basis possible

deviations

in

results

of

the

426 MATERIALS Samples of deciduous and coniferous bark were collected from a pulp factory in southern Sweden (about 100 kg each). The r a w material supplied t o t h e factory originates, f o r both types of wood, from an area within a radius of 100 km. A t the factory the t r e e trunks a r e washed with water and t h e bark then peeled off mechanically and transported directly t o a burner. The bark was collected before the burner, air-dried and then ground in a "garden mill" t o pieces of about 1 cm in size. Further grinding was done in a laboratory mill, whereafter the samples were homogenized by mechanical

mixing.

The

samples were placed in consecutively numbered (0-50 and 51-100 respectively) plastic

bottles containing about 70 g each.

Ten bottles

of

each type

of

sample were randomly selected and gamma-spectrometry was carried out on 60 ml.

137Cs the

For

maximal deviation from the mean value was 3.6% f o r

decidious bark and 2.4% f o r coniferous barks. On this bases the samples were considered t o be sufficiently homogeneous f o r performing an intercalibration on 60 ml or larger volumes. The samples were sent t o the laboratories t h a t had expressed an interest inparticipating

in

the

intercalibration

programme,

and

to

laboratories

participating in t h e general radioecological programmmes of NKS as well as t o laboratories known t o perform continuously gamma-spectrometric measurement on environmental samples f o r different purposes such as environmental monitoring and the

control

of

radioactivity levels

in foodstuffs.

In total 3 9 Nordic

laboratories received the samples. Several laboratories asked f o r additional samples in order t o increase volumes subject f o r analysis. A complete list of the participating laboratories is provided in Appendix 5.4.1. The laboratories were also asked to provide information on type of detector,

detector

volume/relative

efficiency,

amounts

analysed,

methods

of evaluating the results, and water content of the samples. The results were t o be presented as Bq per kg dry weight on 1991 07 01 as reference date. Of the 3 9 laboratories twenty-six

laboratories reported results more o r

less before the deadlines; 14 had t o be reminded which resulted in a f e w additional reports analysis. resolve

or explanations t h a t they were unable to perform

Only 2 laboratories, using

134

the

Nal detectors, were either unable t o

Cs f r o m 137Cs o r t o produce reliable results.

RESULTS Analytical methods Most laboratories used Ge or HpGe gamma spectrometry, having detectors with

421 relative efficiencies of 12-55 %. For the evaluation of the results, PC-based evaluation programme, provided by the companies selling gamma-spectrometric

or

equipment, indicated

that

"home-made'' they

programmes,

had taken

were

used.

coincidence effects

Several

laboratories

134

for

Cs

into

account

either directly in t h e programme or "by hand" afterwards. The water content of t h e samples was between 3 and 12 Z f o r bark from trees in deciduous f o r e s t and generally slightly higher 4-14 % f o r coniferous bark forest. The amounts analyzed were between, 10 and 430 g dry weight, but generally around 20-50 g. Caesium-137 Results from t h e different laboratories a r e given in Table 5.4.1. Twenty-six 137

laboratories provided results f o r

Cs. The arithmetic mean f o r deciduous

bark, was f o r those using Ge detectors, was 28.9 f 5.3 Bq kg-' (n=24, 1 S.D) with a geometric mean of 28.5 Bq kg-'.

For coniferous bark the r e s d t was

47.5 f 8.4 Bq kg-' with a geometric mean of 46.8 Bq kg-'.

I t is obvious t h a t

most laboratories reporting results and working within the NKS programme are capable of performing an analysis of 137Cs in environmental samples at these levels. Caesium-134 Twenty-two

laboratories

coniferous bark

provided

respectively.

134

results

for

into

account

Taking

Cs

in

results

deciduous obtained

and

by

Ge

gamma spectrometry the arithmetic mean f o r decidouus bark was 2.9 ? 1.1 (1.S.D.) Bq kg-'

with a geometric mean of 2.8 Bq kg-'.

For coniferous bark

the corresponding result was 2.8 f 1.6 Bq kg-' with a geometric mean of Bq kg-'.

3.1

I t is again satisfying t o see t h a t most laboratories a r e capable of

carrying out an analysis f o r

Cs a t these low levels, although t h e s c a t t e r

137

is larger than f o r difficulties,

134

Cs. The larger scatter can partly be the results of 134

or neglect of coincidence effects in the analysis of

Cs but,

of course, i t could also be connected with poorer counting statistics.

It

is

coniferous

interesting bark

concentrations of higher related

to

than

it

note is

that in

the

134Cs/'37Cs r a t i o

deciduous

bark

is

although

134

Cs a r e the same. The total concentrations of

in coniferous bark, radiocaesium

and

which

indicates a lower fraction of

a higher

contribution from nuclear

related radiocesium in coniferous than in deciduous bark.

lower the

in

total

137

Cs a r e

Chernobyl

test

fallout

428 Table 5.4.1, Results from the different laboratories participating in the intercalibration exercise. Bark, deciduous

1

2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26

27 26.9 26.851.1 32f3 34.2f1.5 26.2f1.6 27.0f1.4 23.5 25 26.7f28 26.lf2.1 30.2f0.5 27.5+3% 25fl 26.7fO. 7 21f2 34.8f2.1 29+4 28.1f1.2 38f3 26.3f3.1 35.5*4% 27.7 27 t2 23

Bark, coniferous

I

nd

3.3 2.2fO.3 2.6f0.6 3.1f0.3 2.450.4 1.97f0.2 4 3.1 2.9f15% 3.1f0.6 2.3fl.1 3.Of40% 2fl 2.3f0.12 1.9f0.9 2.9f0.7

31 45fll

2f8%

2.5f0.3 nd t3 4.3*41% 19.2 26 <2 included

24f4 55fll 46.9f3.7 60 26 63f15% 61f20 55f26 52.6*30% 57f16 46.4 nd 74.6f1.5 104fl% 72f13 9f3 151f86 320+4.6% 48.6 39 t2

-

42 42.4 59f2 47f4 56.5f2.3 43.lfl.3 46.8f2.3 38 43 45.5fl% 50f2.8 45.9f3.9 48.2f3% 45f2 45.8f0.8 35f3 63.5f3.8 45f4.4% 47.2f2.2 65f3 41.5f2.2 53.lf3% 45.8 57 t2 19

7 1.7 4.5f0.3 3.5f0.6 4.2f0.4 2.5f0.6 2.3f0.2 3.4 3.2 3.5+14% 3.6fl. 1 2.6f0.7 2.4f40% 2fl 2.9f0.4 2.8fl. 1 4.4fl. 1 3f8% 2.8f0.3 nd (3

3.1f44% 2.8 3.2 t2 included

nd

65 64f13

80f6 62fll 74f4 70 47 69f148 49f33 75f10 90f15% 98f19 16.4 0.8f0.6 94f9.4 243f1% 86f13 7f3 217f86 360f4% 73.1 48 t2

-

429 Potassium-40 Twenty-one

and 22 laboratories,

respectively, reported results f o r

decidious and coniferous bark respectively. Some of

as

considered

"outliers".

The

results into account was 76 outliers 51 f 14 Bq kg-'

arithmetic

+

mean

75 Bq kg-'(l.

these must

however

concentration

in

40K be

taking

all

S . D . , n = 22) and deleting

(n = 171 f o r deciduous bark.

The corresponding

values f o r bark from coniferous trees were 87 f 73 Bq kg-'[n = 22) and 74 f 17 Bq kg-'

Bq kg-'

(n = 18) respectively. The geometric means were 56 Bq kg-' and 60

respectively taking all results into account and 49 Bq kg-' and 72 Bq

kg-' without outliers. It

is surprising t o find t h a t several laboratories have difficulties in

analyzing

40

K. This must mainly be

due t o high and varying backgrounds in

t h e laboratories, Potassium-40

is

not

of

importance

accidental or controlled releases of

as

an

artificial

radionuclide

anthropogenic radionuclides, but

used in radioecology as a tool f o r food chain transfer using t h e Cs/K on which t o base basis scientific conclusions.

it

in is

ratio

430 TABLE 5.4.2 Laboratories participating in the gamma spectrometry intercalibration exercise. Institute

Responsible investigator

1.

Valtion Teknillinen Tutkimuskeskus Reaktorlaboratorio, Espoo, Finland

P. Manninen/E. Hasanen

2.

Maatalouden Tutkimuskeikos Jokioinen. Finland

A. Paasikallio

3.

Puolustusvoimien Tutkimuskeskus Lakiala, Finland

M. Kettunen

4.

Jayakyln Yliopisto, Fysikan Laitos Jyvaskya, Finland

S. Justinen

5. Helsinki Yliopisto,Radiokemian Laitos

T. Jaakkola

Helsinki, Finland 6.

Sateilyturvakeskus Helsinki, Finland

7. Risa Nat. Lab., MIL ECO Roskilde, Denmark 8.

Statens Stralskyddsinstitut Stockholm. Sweden

9. Forsvarets Forskningsanstalt Umeh, Sweden

H. Aaaltonen/S. Klemola

H. Dahlgaard

J. Melin/M. Elbe R. Bergmann. NylCn

10.

Geislavarnir Rikisins Reykjavik, Iceland

S. Magnusson/S.E. Palsson

11.

Studsvik Nuclear Nykoping, Sweden

R. HedvalVY. Sandell

12.

Forsvarets Forskningsanstalt Sundbyberg, Sweden

I. Vintersved

13.

Statens Naturvlrdsverk Drottningholm, Sweden'

M. Notter/G.Neumann

14.

Radiofysiska Inst., Umei Univeritet Umeh., Sweden

G. Wickman&.

15.

Institutionen f o r Radiofysik, Lunds Universitet, Lund, Sweden

E. Holm

16.

Forsvarets Forskningsinstitutt Kjeller, Norway

A. Lillegraven

17.

Institutt f o r Energiteknikk Kjeller, Norway

G. Christensen

Agren

43 1 TABLE 5.4.2 continued

18.

Norges Lantbruksh~gskole,Isotoplaboratoriet B. Salbu/G. Osley As, Norway

19.

Statens Institute f o r Strklehygiene Osterks, Norway

20. Universitetet i Oslo,Avd. f . Karnkemi Blindern, Oslo, Norway

F. Ugletveit P. Hoff

21.

Universitetet i Trondheim Inst. f Uorg. Kemi-NTH, Trondheim, Norway

R. Nreumann

22.

Bareslcksverket Loddekopnge, Sweden

T. Aberg

23.

Ringhalsverket Varobacka, Sweden

L. Helgesson

24.

Forsrnarks Karnkraftverk asthamrnar, Sweden

F. Keffner

25.

Milo och Halsoskyddsforvaltningen Gavle, Sweden

H. Sjolund

26.

Mil j o och Halsoskyddsforvaltningen Uppsala, Sweden

C. Kihlberg

n This Department (lab No 13) under t h e Swedish Environmental Protection Board has been closed.

This Page Intentionally Left Blank

43 3

5.5 DOSES FROM THE CHERNOBYL ACCIDENT TO THE NORDIC POPULATIONS VIA DIET INTAKE

ASKER AARKROG Rise National Laboratory, DK-4000 Roskilde, Denmark. SUMMARY The individual mean doses from radiocaesium intake with diet since the Chernobyl accident in 1986 were determined for the five Nordic countries: Denmark, Finland, Iceland, Norway and Sweden. The estimates were obtained by two methods. The first used consumption data, i.e. information on the amounts of food eaten by an average individual in each of the five countries. The other method applied food production in the Nordic countries, ignoring the export and import of food but taking into account the amounts actually eaten. The consumption method gave an individual mean dose commitment of 1.3 mSv and the production method gave 1.0 mSv. In comparison the external mean dose for the Nordic countries was 0.8 mSv. The study emphasizes the importance of wild produce for the internal doses from radiocaesium. More than 50% of the total 137Csintake with the Nordic diet came from natural and seminatural ecosystems. In this context it is unfortunate that information on the consumption of and radiocaesium concentration in wild produce is relatively scarce. This makes the dose estimates imprecise, probably with a tendency to overestimation. INTRODUCTION One of the chief objectives of the NKS radioecology programme (RAD) was to assess the doses to the Nordic countries from the Chernobyl accident in 1986. This assessment should comprise collective effective doses received from food produced in the Nordic countries as well as the calculation of individual mean effective doses to an "average" inhabitant in each of the five Nordic countries. A further aim was to estimate the doses to critical groups. Finally a comparison was to be made between the internal doses from radiocaesium based on diet and wholebody measurements respectively. As this NKS project deals with radioecology, the main efforts focus on internal doses received from food consumption. These doses are, however, compared with the external doses reported from the various countries. Inhalation doses are not dealt with as they are of relatively minor importance (UNSCEAR 88). The very patchy deposition of the Chernobyl fallout, together with the significant differences in agricultural patterns both between and within the Nordic countries, make dose assessment

434 difficult, e.g. not all relevant activity data exist. Hence it was necessary to use data based on information available from neighbouring countries when national data were lacking. Experience from the various RAD projects described in this book was also applied whenever possible.

FOOD PRODUCTION IN THE NORDIC COUNTRIES General remarks Table 5.5.1. gives the mean amounts of nationally-produced food for human consumption in the five Nordic countries. The data are based on statistics for agriculture, fisheries and hunting for the time around 1990: Denmark: Danmarks Statistik (1990-1993), Finland: Statistical Yearbook of Finland (1991), Yearbook of farm statistics (1991), Tuumaninen ef al. (1992), Finnish Game and Fisheries Research Institute (1992), Iceland: The Icelandic Agricultural Informationservice(1992), Institute of Freshwater-fisheries (1992), Fiskifelag Islands (1992), Icelandic Museum of Natural History (1992), Norway: Statens Ernaeringsrad (1991), Statistisk Sentralbyrk (1992) and Sweden: Tradgardsnaeringens riksforbund (1992); Statistiska Centralbyrkn (1988, 1991); Skogstyrelsen (1993), Haggstrom (1993), Jordbrugsverket (1992), Moberg (1991). Table 5.5.1. is used for the calculation of the collective effective doses from the food for humans produced in the five Nordic countries. It is understood that to some extent this dose will be received by people living outside Scandinavia. This applies in particular to foodstuffs produced in Denmark, which exports most of its dairy and meat production. However fish, e.g. from Norway, Iceland and the Faroes, are also exported to a large extent. In some cases the figures in table 5.5.1 comprise minor amounts not used directly for human consumption. In the calculation of the collective doses (seetable 5 . 5 . 5 . ) it is assumed that 3/4 of the grain, potatoes, vegetables and fruit, 213 of the meat and 1/2 of the fish, game and mushrooms are consumed. The above corrections are also supposed to include the losses of radiocaesium resulting from cooking. In the dose calculations it is assumed that the production figures given in table 5.5.1. are valid for the next 50 years.

Comments on the production data (Table 5.5.1.) General The amounts are rounded off to two significant figures. Some figures are mean values for a number of years (- 1986 - 1991). The relative standard deviation of these means is usually less than 10%. However, for wild produce they are often of the order of 50% or more.

435

Table 5.5.1. Mean amounts of nationallyproduced food for human consumption in Gg (thousand tons or million kg) DEN

FIN

ICE

NOR

SWE

5.1

5.0

0.25

4.2

8.4

4600

2800

105

740

3400

Grain

400

570

0

140

1900

Potatoes

250

300

15

330

770

Vegetables & fruit (also greenhouse)

340

190

3

190

370

beef & pork Meat lamb

1480

290

6

160

430

9

25

4.8

Marine fish

330

105

830

900

230

Freshwater fish

30

15

3

115

2

Deer

2

0.3

0

1.2

2

Elk

0

6

0

4

14

Reindeer

0

3.3

0.01

2.8

2

Freshwater fish

0.5

38

0.15

5

35

Berries

0

7

0

11

40

0.25

0.4

0

1

24

1

1.3

0.13

1

0.8

Population in millions Cows’ milk

2

1.2

Wild produce:

Mushrooms Birds and other game

Cows’ milk

These figures represent the production of wholemilk in the various countries. The figures may contain minor contributions not used for human consumption.

436

Grain The figures should comprise rye and wheat grain for human consumption only. However, the Swedish figure may also include some fodder and seed grain. The Finnish figures include grain stored in granaries, which for the period under review resulted in excessively high figures because the granaries were nearly full.

Potatoes To some extent the figures (Norway & Sweden) may also include potatoes fed to livestock. The total amount used for human consumption may be obtained from table 2 by multiplying by the population number. In Finland fodder potatoes are estimated at 500 G g y-'.

Vegetables and fruit Figures include field-grown and greenhouse produce. The Finnish data suggest that greenhouse produce is of the order of 25-30% in Finland.

Meat The data represent slaughter weights, i.e.including by-products not eaten by man (e.g. bones). About two-thirds of the total live weight is available for human consumption and of this two thirds again is actually consumed by man.

Marine fish The data comprise whole fish used for human consumption only. Fish for industry (e.g. fishmeal) is thus excluded. Most of the Finnish catch comes from brackish waters. Fish farms are included and so is the catch of non-professional fishermen.

Freshwater fish The-non professional catch is excluded and entered under wild produce. The Norwegian figure is remarkably high. The Danish figure is a best estimate.

Wild produce The Finnish data for berries and mushrooms only include products offered for sale. Hence the actual figures are higher (cf. the Swedish and Finnish figures obtained from table 5.5.2. by multiplying by the respective population numbers). Reindeer comprises both domestic and wild animals. The Swedish catch of freshwater fish was estimated to 24 Gg in 1990 and to 47 Gg in 1989. (Haggstrom (1993)). Table 5.5.1. uses an average of these two figures.

437 INDIVIDUAL MEAN CONSUMPTION IN THE NORDIC COUNTRIES General remarks The figures in table 2 are based upon diet surveys in the Nordic countries, but also to some extent on estimates achieved by dividing the total net amount of a food product by the total population number of the country in question. Denmark: (Rise National Laboratory (1986-1992). Faroes: Rise National Laboratory (1986-1993). Finland: STUK-A78 (1991), STUK-A62 (1987), STUK-A94 (1992), Statistics Finland (1986-1990). Iceland: Manneldisrad Islands (1992), Steingrimsdottir et a1 (1991). Norway: Strand et a1 (1987) Statens Ernaeringsrad (1991). Sweden: Jordbruksverket

(1992), Skogstyrelsen (1993), Johanson et a1 (1993), Moberg (1991). Net amount is understood as [production

+ import - export] multiplied by the fraction actually consumed.

Comments on the consumption data (Table 5.5.2.) General The diet intake of a population changes with time. With an increasing standard of living the lowpriced foodstuffs such as grain products (and potatoes) are exchanged for more expensive ones, viz. meat, fish (and fruit). The total mean intake of energy in the five Nordic countries is 4.5f0.2 GJ year-'cap-' assuming equal mean intakes in the five countries of sugar, fats, beer & wine, and eggs, which together contribute 2GJ year'kap-'. Table 5.5.2. also comprises the consumption data for the Faroe Island, but these data were not included in the calculation of mean values for the Nordic countries. Milk products The annual mean intake of milk varies from 158 kg in Sweden to 262 kg in Finland, corresponding to a relative standard deviation between the five countries of 22%, which may be compared with the rel. SD for cheese of 15%. Grain products The mean consumption of bread and other grain products is almost similar in the Nordic countries. The rel.SD of the annual amounts is 15% only. However, various types of bread are eaten. For example Denmark has a relatively high consumption of wholemeal bread (which influences the intake of radionuclides significantly). Vegetables, potatoes and fruit In Iceland the total mean intake of these products is 133 kg year' cap-', which is significantly less than in the other Nordic countries. The Nordic mean was 176 kg vear-l cap-' with a relative SD of 17%.

438

Table 5.5.2.

Mean amounts of food consumed by an average person in kg year-' cap" (% imported) DEN

FAR (not included in

FIN

ICE

NOR

SWE

Nordic rel. mean

SD %

means)

Milk

164

146

262

176

206

158

cow Cheese goat

9.1

7.3 (100%)

12

14

9.5

12.7

3.3

0.4

Grain products as flour Potatoes

80

73

80 (100%) 91

73 57 59 (23%)(100%) (67%) 66

50

60

193

22

12

15

60

66

15

84

71

20

Vegetables

44 (34%)

20 (100%)

49 26 39 55 (20%) (60%) (68%) (-60%)

43

26

Fruit

51 (34 %)

18 (100%)

94 57 65 55 (93%)100(%) (68%) (-60%)

64

27

pork, beef Meat lamb

55

18.5 (100%) 18.5

} 45

18

Marine fish*

11

17

36

4.5

4

Deer

1

91

35

15

0.2

24

40 (4%) 6

14

27

20

48 1 (23 %) 15

0.14

0.07

-

0.3

0.2

Elk

0

1.14

-

0.94

1.6

Reindeer

0

0.5

0.04

0.67

0.25

Freshwater fish (wild)

-0

4.1

0.6

1.2

Berries (wild)

0

3.3

Mushrooms (wild)

0.05

1.4

Various game

0.46

GJ year-' cap-'

4.7

1.1

-

2.6

4.5

-

0.24

2.8

0.2

1.1

0.23

0.05

4.7

4.3

4.4

4.4

* / freshwater fish from fish farms are included,

439 Meat & fish An "average" Finn consumes 55 kg year-'cap-' of the meat and fish group (wild produce inchded).

As the other extreme an "average" Norwegian consumes 69 kg. The relative SD between the Nordic countries is only 9 % , in this group, and the Nordic mean is 65 kg year.' cap-'. Wild produce The mean consumption of terrestrial animals (deer, elk, reindeer and various game) is 1.5k0.7 kg y-' cap-', of freshwater fish it is 1.4f1.6 kg y-' cap-', and of vegetable produce (berries and mushrooms) 3+3 kg y-' cap-'. Finland and Sweden have the highest individual mean consumption of wild produce.

RADIOCAESIUM CONCENTRATIONS IN NORDIC DIET COMPONENTS General remarks Systematic countrywide diet surveys of radiocaesium have not been carried out in all Nordic countries. The measurements were most numerous in 1986-1987just after the Chernobyl accident. Hence it is to a large extent necessary to estimate the concentrations of 137Csin the Nordic diet. This can be done in several ways: e.g. milk data can be used for estimating levels in cheese and meat. Data from one country can be applied when estimating the data of a neighbouring country, and the data missing for a year can be estimated from adjacent years. Table 5.5.3. shows the measured and estimated (in brackets) I3'Cs in Nordic diet components. As appears from table 5.5.2., a substantial part of the vegetable diet (grain, potatoes, vegetables and fruits) consumed in the Nordic countries is imported; this may either lower or enhance the radiocaesium intakes from this diet group. Products imported e.g., from Northern Italy, may have shown higher levels than the locally-produced products in the Nordic countries, whereas products from Spain and Portugal were lower in radiocaesium. To overcome this problem, the local Nordic concentrations are used whenever available. In Iceland, more than half of all vegetables products are imported, the Danish concentrations have been used in this case. The references used for the radiocaesium concentrations in the Nordic diet are summarized below: Denmark: Rise National Laboratory (1985-1992). Faroes: Rise National Laboratory (19851993). Finland: STUK-A58 (1987), STUK-A55 (1987), STUK-A78 (1991), Rantavara (1993), STUK-A66 (1987), STUK-A82 (1988), STUK-A90 (1990), Rissanen (1993), STUK-A59 (1987), STUK-A61 (1987), STUK-A77 (1990), STUK-A94 (1992). Iceland: PNsson (1993), Magnusson

et al. (1992), Kershaw et al. (1992). Norway: Strand et al. (1987). Sweden: Johanson et al. (1993), Howard et al. (1993), Moberg (1991). Besides these references, the information given in

P P

0

Table 5.5.3. Caesium-137 concentrations in the Nordic diet. Countrywide means in Bq kg-’ freshweight

COUNTRY

Grain (as

(Chemobyl

flour)

Year

Milk

Cheese

Potatoes Vegetables Fruit

Pork Beef

Lamb

‘37Cs)

86 87 Denmark 88 (1.2kBqm.389 90

86 87

Finland (16kBq m-’)

88 89 90

91 86 87

Iceland (0 kBq m-’)

88

89 90 91

1.06

0.60 0.28 0.18

0.12 0.09 19.0 17 8.1 5.6 3.1 [ 21 2.8 2.9 2.3 2.3 2.5 2.3

0.77 0.43 0.14 0.15 0.09 0.06

[I31 [ 61 [ 61 [ 41 [ 21 [ I] [ 21 [ 21 [ 21 [ 21 [ 21 [ 21

1.11

3.2 0.11 0.11 0.11 0.12 1.2 3.6 0.64 0.75 0.61 0.21 11.11 [ 31 [0.1]

l0.11 r0.11 r0.11

0.53 0.12 0.08 0.09 0.07 0.06 5.0 2.1 1.0 0.8 0.8

1.16 2.5 0.72 0.73 1.3 0.47 33 45 22 16 10

0.5 10.51 r0.11 [0.11 10.11 r0.11 [0.1]

[I01 [I01 [ 101 r101 [lo] [lo]

[lo1

[lo] 5.6 131 r1.41 [1.4] 1.22 200 100

50 [SO] 1401

“I

r601 r601 I501

[501 56 49

Fish (Marine and freshwater fish from fish farms) 4.8 4.7 3.6 4.6 4.5 5. I 32 50 40 35 30 25 ~0.31 r0.31 ro.31 ro.31 ro.31 ro.31

Non-farmed

Wild plant produce

a n i d S

(Rein)*

(Elk)t

Non-farmed freshwater fish

(Bemes) (Mushrooms)

[I21

[lo1 r81 r71 r61 5

510*

680* 610* 550* 490* 430* u51* r151* [151*

[200]7 [200]? [200]7 [200]5 [200]7 [200]7

~51*

WI*

r151*

To be continued

Table 5.5.3. Caesium-137 concentrations in the Nordic diet. Countrywide means in Bq kg-’ freshweight (continued) Year

Milk

Cheese

COUNTRY (Chernobyl ‘3’Cs)

86 Norway (7 kBqm-”

87 88 89 90

91 Sweden (10kBq m-’)

86 87 88 89 90

91 86 Faroe Isl. (2 kJ3q m-’)

87 88 89 90 91

21 11.3 7.8 7.7 5.8 3.6 6.0 3.7 1.6 1.1 0.7 [0.5] 5.8 7.6 3.6 2.5 2.1 1.67

[I51 [ 81 [ 51 [ 51 [ 41 [ 31 [ 41 [ 31 [ I] [ 11

[0.5] [0.4] [0.8] [0.4] [0.1] [0.2]

[0.1] [0.1]

Grain (as flour)

0.94 3.0 0.78 [0.8] 0.6 0.2 [1.1] [3.3] [0.7] [0.8] [0.6] [0.2] [1.1] [3] [0.1] [0.1] [0.1] [0.1]

Potatoes Pork Vegetables Beef Fruit

5.0 “2.11 [1.0] [0.8] P.81 [0.5] [5.0] [2.1] [1.0] [0.8] [0.8] [0.5] 10.9 2.6 6.0 4.9 2.3 2.8

[48] [50] [30] [30] [201 [20] [30] [20] [lo] [ 51 [ 31 [ 21 11.21 12.51 [0.7] [0.7] [1.3] [OS]

Lamb

Fish Non-farmed (Marine and animals freshwater fish from fish farms) (Rein)* (Elk)?

Wild plant produce

Non-farmed freshwater

fish (Bemes)

(Mushrooms)

442

chapters 3 and 4 of the present book, on the radioecology of agricultural (RAD 3 ) and seminatural ecosystems (RAD 4), has been used. Table 5.5.3A. Applied effective ecological halflives for 137Csin Nordic components since 1992 in years ~

Milk

Grain

~~

Potatoes

Beef

Lamb

Vegetables & & Fruit

Mar.

Non-

Fish

farmed

Pork

Freshw. Mushrooms, fish

wild berries

animals ter.

~~~~

~

~

5

5

10

10

5

20

20

5

10

10

10

5

20

10

10

10

10

5

20

Denmark

5

10

20

Finland

5

10

Iceland

20

(10)

20

Norway

5

10

20

5

10

10

10

5

20

Sweden

5

10

20

5

10

10

10

5

20

Faroes

5

(10)

10

5

5

10

5

Comments on the IJ7Csdata (Table 5.5.3.) General

Table 5.5.3. shows the 137Csconcentrations in the various diet components in the Nordic countries, the Faroe Islands included. The mean depositions in kBq 137Csrn-' from the Chernobyl accident are also indicated; for Finland this applies the mean deposition in southern Finland, where agricultural production takes place. The average fallout in Finland after Chernobyl was 10.7 kI3q 1 3 7 m-2 ~ ~

It is a general problem that production of the various food products in the Nordic countries, except Denmark, is very unevenly distributed. This made it questionable to relate the concentrations of Chernobyl 137Csto the mean deposition of '37Cs from Chernobyl, because the fallout pattern over Scandinavia differed significantly. For example, reindeer in Finnish Lapland showed lower 137Csconcentrations than those from Norway and Sweden because most Chernobyl fallout in Finland was deposited over the southern part of the country, whereas it was the central area of Norway and Sweden, where reindeer are grazing, which received most Chernobyl fallout in these two countries. For milk, the Finnish concentrations were higher than the Swedish ones because most milk in Sweden comes from the southern part of the country, which was relatively uncontaminated by Chernobyl, whereas the opposite was the case in Finland.

443 In order to estimate the concentrations from 1992 and onwards, it is necessary to know the effective ecological halflife of '37Csin the various diet components for all the Nordic countries. This information is only partly available. The halflives used are summarized in table 5.5.3A. These halflives should be considered as a best estimate only.

Milk The values are taken from the references given above (cf. also 3.5 in the present book). Cheese

Finnish, Icelandic, Norwegian and Swedish figures were calculated from the milk levels assuming the same ratio between cheese and milk as observed in Denmark, i.e. 0.7. The 137Csconcentration

in goat cheese was assumed to be 2/3 of the concentration in lamb meat, because this was the ratio found in Norway in 1986-1987. Grain

The 137Csconcentrations in grain were based on the barley data collected in RAD-3 (cf. 3.4), except for the Danish and Finnish data, that were actual measurements of wheat and rye. Icelandic and Faroese figures were assumed equal to the Danish as all grain in these countries is imported. Swedish figures were calculated as the mean of the Finnish and Norwegian measurements, because the RAD-3 data from Sweden did not represent the mean levels in Swedish grain; for 1989-1991

use was made of the Finnish data. The Norwegian figures for 1989-1991 which were unavailable, were assumed identical to the Finnish grain levels because this was approximately the case in 19861988. In the calculation of 137Csin grain as flour (figures shown in table 5.5.3.) we assume that 1/4 of the harvest in year (i) is used in year (i) and 3/4 in year (i+l) a) further assumption is that the flour contains 50% of the 137Csfound in barley grain. (The level in grain from the 1985 harvest was assumed equal to the Danish, i.e. 0.08 Bq 137Cskg-I wholegrain). The Danish data take into account the amounts of wholemeal bread and white bread consumed. Potatoes, vegetables and fruit

The Danish and Finnish data are weighted means of the concentrations in potatoes, vegetables and fruit. Otherwise we assume that the mean content in these products (as obtained from RAD-3 cf. 3.4) equals the mean found in potatoes, cabbage, carrots and peas from the various countries. If no data are available from Norway or Sweden, the values are assumed equal to the Finnish. For Iceland, the Danish figures were used.

444

Pork and Beef Levels that were lacking were calculated from the milk concentrations by multiplyling by 4.2, i.e. the median ratio between Danish meat and Danish milk.

Lamb The 137Csconcentrations in lamb (cf. also 3.6) were obtained as follows: Denmark: The mean ratio between the levels in lamb and deposited activity in soil was 1.3kO.42 Bq 137Cskg-I pr. kBq 137Csm-' (0-5 cm soil layer) (k ISE, n=4) in 1990-1991. For a Danish mean deposition of 1.2 kBq 137Csm-z in the 0-5cm soil layer the calculated

mean level in Danish lamb for 1990 and 1991 becomes 1.24 Bq kg''. This level may be compared with the 137Csconcentrations in milk for 1990 and 1991. The mean ratio between lamb and milk in Denmark becomes: 11.8. This ratio was used for calculating the lamb levels in 1986-1989. The level in 1987 was the mean of 41 measured samples of lamb; for this year the lamb-milk ratio was 9.3, i.e. in good agreement with 11.8. Finland: The mean level in Finnish lamb was measured in 1986-1988. For the years 19891991, the levels were estimated by STUK (Rantavara 1993). Iceland: The mean concentration of 137Csin Icelandic lamb was determined in 1990 and 1991 The mean ratio between 137Csin lamb and milk in Iceland was 22f 1 in these two years. This ratio was used for the calculation of Icelandic lamb levels in 1986-1989. Norway: In a food-basket survey in 1986-1987 the countrywide mean level in mutton and lamb was 210 Bq IJ7Cskg-' in the first year after Chernobyl. In RAD-3 ( 3.6) the aggregated transfer factor from soil to lamb in Norway in 1990-1992was determined to 42 Bq 13'Cs kg-I pr kBq 137Csm-'. For a mean deposition of 7 kBq 137Csm-' in Norway, this corresponds to 290 Bq 137Cskg-I lamb in 1990-1992. The concentrations in 1987-1989 were estimated as the

mean of 210 and 290, i.e. 250 Bq 137Cskg-l. Sweden: The Swedish 137Cslevels in lamb were calculated as the mean of the concentrations found in Finland and Norway because the RAD-3 data (cf. 3.6) seem too high as a country mean. Faroe Islands: The Faroes had a complete set of data for 137Csin lamb for the years 19861991. The mean ratio between 137Csin lamb and milk was 15&5 (&lSD; n=6).

445

Fish The concentrations of L37Csin Danish fish for 1989-1990 were calculated from measurements of fish caught in the North Sea, the Kattegat and the Baltic Sea. The mean concentrations found in these areas were weighted by the relative catch of fish for human consumption i.e. 0.65, 0.10 and 0.25 respectively. For 1986-1988 the levels were the means of samples from the North S e a and the Kattegat only. The estimated levels in Norwegian marine fish were assumed identical with measured concentrations in Faroese fish. The Icelandic fish levels were unaffected by Chernobyl fallout (Kershaw et al., 1992) and the mean level for 1986-1991 was 0.3 Bq 137Cskg-I according to UK measurements. The 137Cslevels in Finnish fish were obtained from the STUK reports and from unpublished information from STUK (Rantavara, 1993). The levels are those found in fish from the Bothnian and Finnish Bays. The levels in Swedish marine fish were assumed equal to the Danish levels.

Non-farmed animals Reindeer, most of which are actually domestic livestock, are found in all Nordic countries except Denmark. The 137Cslevels were reported for the entire period (1986-1991) in Finland by STUK, (Rissanen 1993). For Norway a countrywide mean was available for 1986 and it was assumed that the concentrations in the remaining years were proportional to the Finnish levels. The levels in Sweden were calculated as the mean of the data from Norway and Finland. The Icelandic levels were estimated from measurements in 1990 and 1991. Elk are found in Finland, Norway and Sweden. Swedish observations (cf. 4.3) during 1986-

1991 indicated no significant reduction in the levels. The observed aggregated transfer-factor T,,

(Bq 137C~kg-1 per Bq 137Csm-2) for elk in Sweden in 1986-1991 varied between 0.018 and

0.024 (Johanson, & Bergstrom, 1993). In the present calculations for elk in Sweden (Tag,) 0.02 is used. In Norway 137Cslevels in elk are assumed to be equal to those in Finland and Sweden i.e. 200 Bq 137Cskg-'. T,,

for

deer (mean

of roe deer and red deer) was 0.04 (Howard ef al, 1993) and hare

showed a T,, of 0.004. From the consumption of non-farmed animals given in table 5.5.2., a weighted T,,

of 0.01 was obtained for Denmark. The concentrations for non-farmed

animals in Denmark thus represent this weighted mean level of deer, hare and other game.

446

Mushrooms (wild) As in the case of non-farmed animals use was made of the T,,’s proposed by the IAEA (Howard et al. 1993) for mushrooms in the calculation of 137Csmean concentrations. The median Tag,of the

nine species mentioned in the IAEA report is 0.06 Bq 137Cskg-’ per Bq 137Csm-2 (fresh-weight basis). We multiply this coefficient by the Chernobyl deposit in Bq 137Csm-2 for the various countries in order to obtain the levels in mushrooms in 1986. The Finnish measured value was 94% of the calculated value, which was 960 Bq kg-I, suggesting that the method of calculation is reasonable. The concentrations since 1986 were calculated assuming an effective halflife of 137Csin mushrooms of 20 years.

Wild berries According to Finnish information (Rantavaara 1993) wild berries contained about 100 Bq kg”7Cs kg-’ in the period 1986-1991. The same mean was used for Sweden and Norway.

Non-farmed freshwater fish were measured systematically in Finland throughout the years. The Finnish data were also used for Norway and Sweden. In Iceland, the Finnish pre-Chernobyl level in 1985, i.e. 20 Bq 137Cskg-’ fish, was used. In Denmark measurements were carried out in 1987 and 1992. Values for the missing years were estimated by interpolation. Freshwater fish from fishfarms were assumed to show the same concentrations as marine fish and thus not included under non-farmed freshwater fish.

Estimating the contribution to the diet of Chernobyl I3’Cs The relative contribution of Chernobyl I3’Cs to the total diet intake of 137Csdiffered in the Nordic countries. In Iceland nearly all the L37Csin the diet organized from global fallout, whereas in Norway, Finland and Sweden nearly all the ”’Cs was of Chernobyl origin; Denmark and the Faroes take a median position. The contribution of Chernobyl 137Csto the Danish diet was calculated as shown in Table 5.5.3B. The Chernobyl percentages of 137Csin the various Danish diet components were assumed

to be the same as for milk (shown at the top of table 5.5.3B.). From 1992 and in the future, it is assumed that the percentage of 137Csin the Danish diet will remain constant at 70% of the total 137Csin this diet. This is based on the expectation that the environmental availabilities of Chernobyl

and global fallout 137Cswill no longer differ and that the effective ecological halflife of the 13’Cs from the two sources becomes similar. Table 5.5.3. also shows the intake of 134Csin Denmark.

Table 5.5.3B. Radiocaesium intake from the Danish diet in Bq cap-’ from global fallout and from the Chernobyl accident 0% 1985

96% 1986

94% 1987

84% 1988

74% 1989

72% 1990

-70% 1991

86-91

-70% 1992-~0

C

ECO t%

137cs fallout 137CsCher ‘MCS

7 174 88

6 96 37

8 40 11

8 23 5

6 15 2

5 11 1

40 359 144

36 79 2

5 5 1.5

137cs fallout 137CsCher 1”cs

4 85 44

15 24 1 92

1 7 2

2 7 1

3 6 1

2 7 1

27 353 141

29 101 2

10

‘37csfallout 137CsCher ‘“CS

4 86 44

1 20 7

2 11 3

4 11 2

3 8 1

3 7 1

17 143 58

87 202 3

20 20 2

137cs fallout 137CsCher 1”CS

2 61 31

8

Meat

130 50

6 33 9

10 29 6

20 50 8

8 18 2

54 321 106

58 130 4

5 5 1.5

Fish

137cs fallout 137CsCher 1”CS

2 50 25

3 48 18

6 33 9

13 37 8

14 35 5

17 39 4

55 242 69

245 563 12

10 10 2

Total

137cs fallout 137CsCher ‘“CS

19 456 232

33 535 204

23 I24 34

37 107 22

46 114 17

35 82 9

193 1418 518

455 1075 23

Milk cheese

+

Cereals

Pot.Veg. Fruit

The percentages shown at the top of the table are the Chernobyl contributions of 137Csin Danish milk in the different years. ECO,,h is the effective ecological halflife in years.

10 1.7

P P 4

Table 5.5.4. Individual intakes of radiocaesium from the Nordic diet (based on tables 5.5.2. and 5.5.3.) ~~~~~

COUNTRY

PERIOD

Milk

products

DEN

FIN

ICE

NOR

SWE

0.04 0.1 0.5

0.5

86-87 88-9 1 92+w

10

86-87 88-91

0.8

5

1

0.2 0.02 0.1

0.1

86-87 88-91 92+ 00 86w

8 8 15 31

86-87 88-91 924 00 8600

0.3 0.5

0.2

92+ 00 8600

86-87 88-9 1 92-00

0.1 0.05

4 19

5

0.4 0.2

Potatoes vegetables fruit

1.5 0.6 3

2 12 15

8600

FAR

0.3

8600

86-

Cereals products

0.1

0.1

Beef, pork lamb

Fish

Wild produce

0.2 0.2 0.2 0.6

0.1 0.2 0.8 1.1

0.02 0.03 0.2 0.3

3 2

3 8

0.3

0.4 0.5

3 6 19 28

0.2 0.1 0.2 0.5

1 0.5 2 4

7 10 32 49

2 1 1 4

0.3

2

0.3

1.4

0.03

1.8 5

~

k Bq I3'Cs cap-'

0.3

86-87 88-9 1 92-00

P

%

0.1 0.2 0.6

0.1 0.4

0.05

1 0.6

3 5

1.2

1.5

4 6

3 2 3 7 4 3 4 11

1 2 5

8 0.02

0.03 0.1 0.2 0.02

0.03 0.06 0.1 0.2 0.3 1 2 0.1 0.1

0.3 0.5

10 17 60 87

0.05 0.1 0.3 0.5 4 7 22

33 7 12

64 83

TOTAL '37Csintake (dose in mSv) 1 0.6 1.7 3.4

L0.04) 26 27 75 128 (1.6) 5 8

33 45 20 26 12 118 (1.5) 13 16 72 101

8 6 10 24

10.3)

TOTAL W s intake (dose in mSv)

0.5 0.1 0.03 0.7 (0.01) 11 7

1 19

10.3) 0

0

0 0

m9 8 1 18 (0.3) 6 4 2 12 (0.2) 4 1 0.2 5

10.1)

Table 5.5.5.

Collective integrated intakes of radiocaesium from Nordic-produced diet since 1986 1T Bq 13'Cs + 12500 Sv ; 1 TBq IMCs+ 16700 Sv

(from tables 5.5.1. and 5.5.3. with corrections for fractions actually consumed)

TBq 137cs

'MCS

I37Cs

+w

DENMARK 1986-00

0.014

0.002

0.001

0.010

0.017

0

0.044

0.01

700

FINLAND

"

0.19

0.004

0.012

0.04

0.035

0.19

0.47

0.07

7000

ICELAND

"

0.009

0

0

0.007

0.003

0

0.02

0

0.06

0.001

0.010

0.13

0.003

0.09

0.29

0.05

4000

0.07

0.014

0.021

0.034

0.012

0.50

0.65

0.07

9000

0.34

0.02

0.04

0.22

0.07

0.78

1.47

0.20

22000

NORWAY SWEDEN

"

"

Nordic countries

s

300

P P \o

450 The observed 134Cs/'37Cs ratios in Danish diet samples were: 1986:0.49, 1987:0.36, 1988:0.23, 1989:O. 15, 199O:O. 11 and 1991:0.08. In Iceland there was no

intake with the diet. In Finland, Norway and Sweden the

relative 134Csintake compared with the intake of Chernobyl-derived 137Cs,was assumed to be the same as in Denmark (see table 5.5.3B)i.e. for the years 1986-1991 : 37% of the Chernobyl '37Cs intake (518/1418), and from 1992 and onwards 2% (23/1075). Production and consumption data used for estimating intake Two methods were applied in order to estimate the intake of 137Cswith the Nordic diet. Table 5.5.2. gives information on the individual intakes of the various foods and this can be multiplied by the respective concentrations given in table 5.5.3. The other method applies the production data given in table 5.5.1. Multiplication of these (corrected for the fraction consumed of the various products) by the concentrations in table 5.5.3. yields the collective intake of 137Csfrom foodstuff produced in the Nordic area. If we divide these collective intakes by the population numbers in the respective countries, we get the ratios shown in table 5.5.6. A ratio less than 1 will usually indicate that production is greater than consumption, i.e.

some of the production is exported, and a ratio greater than 1 would suggest import. However, there may be other reasons for deviation of the ratios from unity. A ratio may thus be larger than unity if the consumption figures in table 5.5.2. overestimate the actual intake of a product, because part of it is not eaten. This may in particular be the case for vegetable produce, meat, fish and wild produce. Another reason for a ratio larger than unity could be that production figures do not include private activities, such as angling, hunting and non-commercial collecting of berries and mushrooms. CRITICAL GROUPS Of the Nordic countries, Finland, Norway and Sweden received most of the Chernobyl fallout. The

mean deposition in these three countries was of the order of 10 kBq 137Csm-*. All have population groups that to a considerable extent obtain their food from natural and seminatural ecosystems, e.g. the Lapps. In order to estimate the dietary intake of '37Csby a group of people obtaining most of their food from a Nordic seminatural ecosystem, we will postulate a diet based mainly on products from such an ecosystem. We assume that all milk products come from goats instead of cows. The consumption of grain products and potatoes is identical to that of the general population, but instead of vegetables the group eats wild mushrooms and the only fruit consumed is wild berries.

45 1 Table 5.5.6. Comparison between individual mean intakes of I3’Cs based on consumption data (table 5.5.4) and production data (table 5.5.5.) Ratio: consumptiondatdproductiondata Milk products

Cereal products

Potatoes vegetables fruit

Beef pork lamb

Fish

Wild produce

DENMARK

0.2

1.3

2.6

0.3

0.3

FINLAND

0.5

1.0

2.1

1.0

1.1

ICELAND

0.4

1.0

0

NORWAY

2.2

2.1

1.7

1.6

0.1

1.5

SWEDEN

0.5

0.4

2.0

1.7

1.4

1.4

2.3

The meat eaten consists of equal portions of lamb, elk and reindeer and all fish consumed are wild freshwater fish. Table 5.5.7. shows the annual mean of a member of such a critical group and furthermore the infinite time integrals in Bq kg-I year. These values are the mean of the figures for Finland, Norway and Sweden calculated from tables 5.5.3. and 5.5.3A. The 137Csconcentration in goats’ milk was assumed to be 30 times that of cows’ milk (Russel, 1966).This hypothetical critical group is thus not supposed to live in one of the highly contaminated areas of Scandinavia. It is only the consumption of this diet that makes the group critical. It appears from table 5.5.7. that mushrooms contribute nearly half of the dose. Next in importance comes goat milk and then wild berries. The high consumption of these three food items is probably unrealistic, but should be considered as a supposition. The meat and fish group contributes 20% only. Lapps usually show a significantly higher intake of reindeer than this hypothetical group. A Norwegian study of Lapp reindeer herdsmen from southern and central Norway (Strand et al. 1992) thus assumes annual individual intakes of 1 kg lamb, 8 kg elk or deer, 47 kg reindeer, 2 kg freshwater fish and 7 kg wild berries. Using the time-integrated values in table 5.5.7., such a group would have an intake of 540000 Bq 137Cscorresponding to a dose of 7 mSv, i.e. a third of that of the above hypothetical group. It should however, be borne in mind that the time-integrated levels for the Norwegian Lapp groups were probably higher than those obtained from table 5.5.7., which were countrywide mean values for Finland, Norway and Sweden. Strand (1992) thus estimated the dietary intakes for 1987-1989 to 2.25.106Bq provided no dietary measures had been

452 implemented. This suggests that fallout in this region has been of the order of 100 kBq L37Csm-*, rather than the 10 kBq m-* used for the calculations in table 5.5.7. However, as countermeasures were taken the actual intake of L37Csby the Sami herdsmen was reduced by a factor of five.

Table 5.5.7. Diet composition and infinite time-integrated 13'Cs intake since 1986 of a hypothetical Scandinavia critical group Diet component

kg year-' cap"

Bq 137Cskg-'y

Intake Bq 137Cs

209

1700

350000

Goat cheese

12

2200

26000

Grain products

64

10

640

Potatoes

70

Wild mushrooms

49

19000

930000

Wild berries

75

3500

260000

Lamb

15

3300

50000

Elk

15

4100

6 1000

Reindeer

15

10000

150000

Wild freshwater fish

18

6500

120000

~~~

Goat milk

TOTAL

25

1700

1.95 * lo6

Corresponding dose 0.02 Sv.

In conclusion: critical groups in Scandinavia may have received individual mean doses that are 1-2 orders of magnitude higher than those of the general population, which means that individual doses in such critical groups may have reached values of 0.02-0.2 Sv, corresponding to 5-50 years of background radiation.

DISCUSSION Estimates have earlier been made of the doses to the Nordic population from dietary radiocaesium.

In Sweden (Moberg, 1991) it was thus estimated that the mean doses in mSv for the whole country were 0.023, 0.027 and 0.12 for the first year, the second year, and for 50 years respectively. These doses may be compared with those in table 5.5.4., which for 1986-1987, i.e. the first two years after Chernobyl, were 0.26 mSv, i.e. five times higher than the Swedish estimate, and the

453

infinite dose commitment in table 5.5.4. (1.5 mSv) was more than 10 times the Swedish 50-year dose. The Swedish 50-year dose estimate was based on the assumption that the rate of decrease of the 13’Cs in the Swedish diet was 2.5 years. This is significantly less than assumed in the present calculation (cf. table 5.5.3A.). Hence it was to be expected that the present estimate would be higher. If we look at the Finnish estimates (STUK A74, 1991) the dose in 1987 based on diet intake was 0.3 mSv, which is compatible with the dose derived from table 5.5.4. for 1986 -1987 (0.51 mSv). The Finnish report estimates the total external and internal commitment to be 2-3 mSv, which is also consistent with the estimate of internal dose in table 5.5.4. (1.8 mSv). For Norway, the first year dose from intake of radiocaesium with diet was estimated to 0.15k0.02 mSv (Strand et al., 1987), which is compatible with the 1986-1987 estimate in table 5.5.4., (0.40 mSv).

Finnish and Swedish wholebody measurements are reported in 5.6. The doses calculated for 1986-1990were 0.25 mSv to the Finns and 0.054 niSv to the Swedes. These doses are respectively a factor of 4 and 10 lower than those obtained from table 5.5.4. for 1986-1991. This suggests that the intakes of radiocaesium with diet as calculated in table 5.5.4. are higher than those received

by the groups used for the wholebody measurements in the two countries. If wild produce was excluded from the intakes in table 5.5.4., the Finnish intake of radiocaesium with diet would be reduced by a factor of two and the Swedish intake by a factor of three. However, the wholebody doses would still have been too low, viz. a factor of 2 and 3 lower than the doses estimated from the diet. This discrepancy should encourage dietary studies for the wholebody groups in the two countries. For instance how much wild produce do the group members consume and what are the lB7Cslevels in these products?

It was also observed earlier that wholebody measurements gave dose estimates lower than the doses based on diet measurements. The reason may be an overestimate of the amounts of food actually eaten on neglection of loss of caesium during foodpreparation, but may also be due to a bias connected with activity measurements because samples may preferentially be collected in the most heavily contaminated regions. Hence we cannot exclude such a bias in the dose estimates given in tables 5.5.4. and 5.5.5. It should, in this connection, also be recalled that the consumption statistics for foods from natural and seminatural ecosystems in particular are unreliable. As the concentrations in these products are 1-2 orders of magnitude higher than in agricultural products, the reliability of the dose estimates is influenced significantly by this uncertainty. Further the determination of the effective halflife of ‘37Csin various ecosystems needs verification in order to calculate the long-term doses in various diets. Again this is crucial for naturally and seminaturally produced foodstuffs in particular. Finally it should be kept in mind that the wholebody dose

454 estimates may be biased to the low side, if consumers of wild produce are not adequately represented in the group representing the average population. The internal doses from diet may be compared with those received from external radiation. In Sweden the effective dose commitment to man as a result of the Chernobyl accident was estimated to 0.6 mSv from external radiation (Moberg, 1981). In Norway the external dose commitment was estimated to be 1.0 to 1.2 mSv (Strand et al., 1987). In Finland the external individual dose commitment from external radiation was 1.7 mSv (STUK-A74, 1991). In Denmark the external dose commitment was estimated to 0.14 mSv (Risra National Laboratory, 1986-1992 and UNSCEAR, 1988). Hence the ratio between the above external doses and the internal doses calculated in table 5.5.4. was for Denmark: 0.14/0.05 = 2.7, for Finland: 1.7/1.8 = 0.9, for Norway: 1.U1.8 = 0.6 and for Sweden: 0.6/1.5 = 0.4. For nuclear weapons fallout the ratio between external and internal '37Csdose was 2.2 (UNSCEAR, 1982). This may suggest that the internal doses calculated in table 5.5.4. in general are too high. The relative contribution from wild produce to the total internal dose from 137Cswas 0.3/3.4 = 9% in Denmark; 87/128 = 68% in Finland; 0 3 4 5 = 1% in Iceland; 33/118 = 28% in

Norway, and 831101 = 82% in Sweden. Disregarding the doses from wild produce, the calculated ratios between external and internal dose would change to 3.0 for Denmark, 3.0 for Finland, 0.8 for Norway and 2.2 for Sweden. These ratios are in better agreement with the UNSCEAR estimate, suggesting that this does not include a significant intake from natural and seminatural ecosystems. The total collective dose from intake of radiocaesium with diet in the Nordic countries is calculated to 30000 man Sv from table 5.5.4. (consumption data) and to 22000 man Sv from table 5.5.5. (production data). This suggests that the individual intakes in table 5.5.4. are probably an

overestimate because it is less likely that the Nordic countries receive significant amounts of radiocaesium with import of foodstuffs. As appears from table 5.5.6., the estimates of wild produce consumption may be too high as well, and this will influence the individual mean doses significantly.

CONCLUSION Based on consumption data, the mean individual dose commitment from dietary radiocaesium in the Nordic countries after the 1986 Chernobyl accident is calculated to 1.3 mSv. The external population-weighted mean dose in the Nordic countries was 0.8 mSv. If the internal dose was based on production data and exports of food were neglected, the dose commitment would be 1.O mSv. It is believed that the dose based on consumption data is an overestimate because of the lack

455 of reliable information especially on wild produce, both with regard to amounts actually eaten and,

so far as radiocaesium concentrations are concerned, in particular for future levels, because the exact effective halflives are not known. Nordic critical groups with high consumptions of mushrooms, wild berries, reindeer, freshwater fish, elk, lamb and goat products may receive dose commitments from dietary intake that are 1-2 orders of magnitude higher than those of the general population. Such groups are found in Norway, Sweden and Finland, in particular among the Lapp population. It should, however, be kept in mind that remedial measures introduced in the Nordic countries after Chernobyl significantly reduced the exposure of these population groups.

REFERENCES Danmarks Statistik, Statistisk Arbog 1990-1993, Copenhagen 1990-1993. Danmarks Statistik, Landbrugsstatistik 1990-1992, Copenhagen 1990-1993. Finnish Game and Fisheries Research Institute, Hunters Central Organization. Personal Communication Helsinki 1992. Fiskifdag Islands. Egir Reykjavik 1992. Howard, B.S.; Johanson, K.J.; Linsley, G.S.; Hove, K.; Prohl, G. & Horyna, J.: Transfer of radionuclides by terrestrial food products from seminatural ecosystems to man. IAEAVAMP,Vienna, June 1993. (Draft report) Haggstrom, Ake, (personal communication) Fiskeriverket, Stockholm. Icelandic Museum of Natural History, personal communication, Reykjavik 1992. Institute of Freshwater Fisheries, personal communication, Reykjavik 1992. Johanson, K.J. & Bergstrom, R.: Radiocaesium transfer to man from moose and roedeer in Sweden. Sci. Total Environment. 1993. (in press) Jordbrugsverket. Reports 1992:2, 1992: 10, 1992: 14, Stockholm 1992. Kershaw, P.J.; Penetreath, R.J.; Woodhead, D.S. & Hunt, G.J.: A review of radioactivity in the Irish Sea. Aquatic Environment Monitoring Report No 32, 65pp, MAFF, Lowestoft 1992. Magnusson, S.M.; Palsson, S.E.; Indridason, K.; Gudmundsson, O.,Magnusson, B. and Thorsson, J.: Transfer of radiocaesium from soil and plants to lamb in Iceland. Det 6’te Nordiske Radio~kologiseminar.Torshavn 14-18 juni 1992. Manneldisrab Islands, Konnen A mataraeb Islendinga. Reykjavik 1992. Moberg, L. (editor) The Chernobyl fallout in Sweden; 633pp, Swedish Radiation Protection Institute, Stockholm 1991. Passon, S. Personal communication. Geislavarnir Rikisins, Reykjavik 1993. Rantavaara, A. Personal Communication, STUK Helsinki 1993. Rissanen, K. Personal communication, STUK, 1993. Risca National Laboratory. Environmental radioactivity in the North Atlantic Region Including the Faroe Islands and Greenland 1985-1991. Risra Reports Nos 541, 550, 564,571 and 622 (19861993). Risca National Laboratory. Environmental radioactivity in Denmark 1985-1991. R i s ~Reports Nos. 540, 549, 563, 570 and 621 (1986-1992). Russell, R.S.: Radioactivity and Human Diet. 552pp. Pergamon Press, Oxford 1966. Skogstyrelsen. Skogstatistisk Arsbok 1993, Sveriges oficiella statistik, Jonkoping 1993. Statens Ernaeringsrid, Oslo 1991. Statistics Finland. Household survey 1990.

a,

456 Statistics Finland. Production statistics 1986-1990. Statistics Finland. Yearbook of farm statistics, Helsinki 1990. Statistics Finland. Statistical Yearbook of Finland, Helsinki 1991. Statistiska centralbydn. Fiske 1991 & Fiske 1988 - en oversikt. SCB, Stockholm 1991 & 1988. Statistisk Sentralbyrh, Oslo 1992. Steingrimsd6ttir L., Thorgeirsd6ttir H. & Egisdottir S . A survey of dietary habits in Iceland 1990. Icelandic Nutrition Council. 104pp. Reykjavik, 1991. Strand. P.; Selnos, T.D.; B0e, E.; Harbitz, 0. & Andersson-Snrlie, A. Chernobyl fallout: internal doses to the Norwegian population and effect of dietary advice. Health Physics 63, 385-392 (1992). Strand, T.; Strand, P. & Baarli, J. Radioactivity in foodstuff and doses to the Norwegian population from the Chernobyl fallout. Radiation Protection Dosimetry 20, 21 1-220. (1987). STUK-A62, Rantavaara, A.; NygrCn, T.; NygrCn, K.; Hyvonen, T. Radioactivity of game meat in Finland after the Chernobyl accident in 1986. Supplement 7 to Annual Report STUK-ASS. Helsinki, 1987. STUK-A78 Rantavaara, A. Radioactivity of foodstuffs in Finland in 1987-88. Supplement 4 to Annual Reports STUK-74 and STUK-A89. Helsinki 1991. STUK-A74 Sumela M.; Blomquist, L., Rahola, T. and Rantavaara, A. Studies on environmental radioactivity in Finland in 1987. Annual Report Helsinki, 1991. STUK-A58 Rantavaara, A.; Haukka, S . Radioactivity of milk, meat, cereals and other agricultural products in Finland after the Chernobyl accident in 1986. Supplement 3 to Annual Report STUK-A55. Helsinki, 1987. STUK-A55 Studies on environmental radioactivity in Finland in 1986. Annual Report. Helsinki, 1987. STUK-A66 Ilus, E.; Sjoblom K-L.; SaxCn, R.; Aaltonen, H.; Taipale, T.K. Finnish studies on radioactivity in the Baltic S e a after the Chernobyl accident in 1986. Supplement 11 to Annual Reports STUK-ASS. Helsinki, 1987. STUK-A82 IEheimonen, T.K.; Ilus, E.; SaxCn, R. Finnish studies in radioactivity in the Baltic Sea in 1987. Supplement 8 to Annual Report STUK A-74. Helsinki, 1988. STUK-A90 SaxCn, R.; Ikiiheimonen, T.K.; Ilus, E. Monitoring of radionuclides in the Baltic Sea in 1988, Supplement 1 to Annual Report STUK-89. Helsinki 1990. STUK-A59 Rantavaara, A. A Radioactivity of vegetables and mushrooms in Finland after the Chernobyl accident in 1986. Supplement 4 to Annual Report STUK-ASS. Helsinki, 1987. STUK-A61 SaxCn, R.; Rantavaara, A. Radioactivity of fresh water fish in Finland after the Chernobyl accident in 1986. Supplement 6 to Annual Report STUK-ASS. Helsinki, 1987. STUK-A77 Saxtn, R.; Rantavaara, A. Radioactivity of surface water and fresh water fish in Finland in 1987. Supplement 3 to Annual Report STUK-A74. Helsinki, 1990. STUK-94A. SaxCn, R. & Koskelainen, U. Radioactivity of surface water and freshwater fish in Finland in 1988-1990. Suppl. 6 to Annual Report STUK A89, Helsinki, 1992. The Icelandic Agricultural information service, "Icelandic Agricultural Statistics", Reykjavik 1992.TradgArdsnaringens riksforbund. (1992) Tuunainen, Leinonen, Tuumainen. Sisavesien Kalatalous 60, 70-90, Helsinki 1992. UNSCEAR. Sources, effects and risks of ionizing radiation; United Nations, New York (1988). UNSCEAR, Ionizing radiation: sources and biological effects. 773pp. United Nations 1982.

457

5.6. INTERNAL RADIATION DOSES TO THE NORDIC POPULATION BASED ON WHOLEBODY COUNTING

MATTI SUOMELA and TUA RAHOLA Finnish Centre for Radiation and Nuclear Safety P.O. Box 14, FIN-00881 Helsinki, Finland SUMMARY The amounts of radionuclides incorporated in humans can be determined by whole-body counting, and the internal radiation doses calculated .on the basis of these results. Another method to estimate internal doses makes use of the statistics on food consumption and the data on activity concentrations in foodstuffs. Both methods were used in the Nordic countries but the calculation methods differed from one country to another. The aim of this internal dose estimation project was to re-estimate the internal doses using the same calculation method for all the countries and to compare the doses obtained by using dietary or whole-body counting data. After the Chernobyl accident whole-body counting measurements were made in Denmark, Finland, Norway and Sweden. This paper presents the mean internal effective doses arising from the '"Cs and 137Cs originating in the Chernobyl accident, calculated using the results of whole-body counting measurements performed in the Nordic countries after the accident. The mean effective internal dose from '"Cs and 137Csdelivered to the Danish population was 0.014 mSv in 1986 - 90, to the Finnish population 0.25 mSv, and to the Swedish population 0.054 mSv using whole-body counting results. The corresponding dose to the Norwegian population was estimated to lie between 0.095 and 0.23 mSv. The calculation of internal doses using food intake data gives doses higher by a factor varying from 1.2 for Denmark to 8 for Sweden compared with doses calculated from whole-body counting data. If the doses from food intake data are corrected by taking into account the fact that the biological half-life of radiocaesium in the Nordic countries is shorter than the value given by the ICRP, then the doses become smaller. Hence also the ratio between the doses calculated by the two methods becomes smaller. INTRODUCTION Internal radiation doses are caused by the radiation emitted by radionuclides in the body. The intake routes of radionuclides are inhalation and ingestion. In a fallout situation the first route of intake of radionuclides by man is inhalation directly from the cloud. The radioactive material in the cloud takes the form of a gas or an aerosol. This period of intake is usually relatively brief, as was the case after the Chernobyl accident when the radioactive cloud remained over the Nordic countries for a few days only. The absorption and behaviour of inhaled radioactive material in the body depend on its

458 physical and chemical properties, especially particle size and solubility. Larger particles are retained in the upper parts of the respiratory tract and transported by ciliac movement to the gastrointestinal tract or excreted from the nasopharyngeal tract. Only the smallest particles are transported into the pulmonary region of the lungs. The second route of intake is via ingestion with food and drink. The rate at which different foodstuffs will be contaminated by radionuclides varies depending on the deposition situation. In the first phase livestock is contaminated by inhalation of radionuclides directly from the cloud. Usually this route of contamination of foodstuffs are of minor importance. In addition to surface contamination of plants resulting from the radioactive cloud, plants are contaminated via root uptake and the radionuclide concentration increases slowly. The contamination rate of plants also determines the contamination rate of foodstuffs of animal origin. Thus some time elapses before the ingestion route becomes important. The factors affecting body burden are activity intake, physical and biological half-life, and the physical and chemical form of the radionuclide. The internal radiation dose depends on several factors: the energy and type of radiation, the activity concentrations in different organs and tissues and the time the radionuclide remains in these organs and in the body. Internal doses can be estimated indirectly by using statistics on food consumption and data

on activity concentrations in foodstuffs, or by measuring directly the activity of radionuclides in the body. Both methods have been used in the Nordic countries but the calculation methods differ from one country to another (UNSCEAR,1988, ICRP 30,1978, Kendall et al., 1987, NCRP, 1977, Leggett et al.,1984). When comparing radiation doses, special care should be taken not to compare committed effective doses with the effective doses delivered during the period considered. The aim of this internal dose estimation project was to re-estimate the internal doses using the same calculation methods for all the participating countries. The body content of gamma-emitting radionuclides at different times can be determined by whole-body counting. Results of periodically repeated measurements allow the calculation of activity time integrals for different radionuclides in the body, The internal doses can be calculated from these results. To calculate mean internal radiation doses for a population, a representative group of individuals should be chosen for whole-body counting using statistical methods.

An intercalibration of the whole-body counting systems in the Nordic countries was performed in I991 - 93, as reported by Rahola et al. in Chap. 5.3. This intercalibration showed a good agreement between the different systems. The present paper reports on the mean internal effective doses resulting from the '"Cs and I3'Cs originating from the Chernobyl accident, calculated using the results of whole-body counting

measurements carried out in the Nordic countries after the accident. Iceland was excluded because

459

no measurable amounts of Chernobyl fallout could be detected there. POPULATION GROUPS Whole-body counting measurements of different population groups were performed in Denmark, Finland, Norway and Sweden after the Chernobyl accident. The deposition of fallout radionuclides after the accident was very unevenly distributed in Finland, Norway and Sweden, as can be seen in the maps presented in Figs 5.6.1

- 5.6.3. This uneven distribution strongly influenced the

selection of population groups to be measured by whole-body counting. Interest was focused on the groups receiving the highest doses instead of on groups representing the whole population. Only in Finland and Sweden were groups representing the whole population studied. In Denmark and Norway only special groups were studied.

Denmark

In Denmark whole-body counter measurements of a reference group were re-started at Risa National Laboratory after the accident. Measurements of the reference group that first started in 1963 had been discontinued in 1978 when the levels of '37Cs resulting from the fallout of atmospheric nuclear weapons testing fell below the detection level. After the Chernobyl accident the reference group was measured from 3 to 11 times a year. The group consisted of 14 adults - eight men and six women - all employees at the Risra National Laboratory (S~agaard-Hansenand Lauridsen,l988, 1989, Aarkrog et al.,1991,1992). As stated by Aarkrog in the previous chapter 5.5, the deposition of '"Cs and 137Csin Denmark was very low (the 137Csdeposition 1.2 kBq/m2) and relatively evenly distributed. Therefore it could be assumed that there were no large variations in the radiocaesium body burdens in different parts

of the country. To get a rough estimate of internal radiation doses in Denmark, the results of the reference group measurementswere used to represent the whole population in estimating populati-

on doses.

Finland The Helsinki reference group, consisting of ten women and sixteen men employed at the Finnish Centre for Radiation and Nuclear Safety (STUK), was measured for the first time at the end of 1965 (Suomela,1984). Two to four measurements were then made annually.

In 1986 the Research Institute for Social Security, of the Social Insurance Institution, selected for whole-body counting measurements 380 people from different parts of Finland to represent the whole population. The selection method chosen was stratified random sampling; the strata were the number of people in the provinces and the sample size was self-weighting. In 1988

460

Fig. 5.6.1. The estimated 137Cssurface activity (kBq m-') resulting from the Chernobyl fallout in Finland on October 1, 1987. Finland is divided into five fallout regions (1 - 5) according to increasing surface activity (Arvela et al., 1990).

an additional group of 180 persons from the Helsinki region was selected in the same way. To keep up the numbers of people measured annually, an additional random sampling of 500 people was made in 1990. Because participation in the measurements was on a voluntary basis, the number of people measured varied from 161 in 1989 to 323 in 1990. The groups chosen consisted

of children aged 5 - 14 years and adults aged 15 - 65 years. Most of the measurements were made in the period from November to April each year (Rahola et al.,1987, 1989, 1991, 1993). The surface-weighted mean deposition of 137Cswas 10.7 kBq/m2 in 1986 (Arvela et a1.,1990).

461

Norway When selecting groups for whole-body counting, main emphasis was laid on special groups representing the areas of high radiocaesium deposition and thus high intake. Only two of the groups, Oslo and Sel, were randomly selected. The Oslo group represents a low deposition area and the Sel group a high deposition area. The 137Csdeposition in Oslo was 1.3 kBq/m2, in Sel 9 kBq/m* (Selnax, 1993) and the population-weighted mean, calculated by us, for the whole country was 5.1 kBq/m*. The Oslo group, consisting of 16 men and 22 women, was measured once only in May 1987. The Sel group consisted of 52 men and 53 women and was measured once annually from 1987 onwards. Those measured were over 14 years of age (Strand et

al. ,1989,1992). No group representing the whole population was whole-body counted in Norway.

Fig. 5.6.2. The estimated 137Cssurface activity (kBq m2) resulting from the Chernobyl fallout in Norway on June 30, 1986 (National Institute of Radiation Hygiene, Norway).

462

Sweden The Stockholm reference group consists of 36 people employed at the Swedish Radiation Protection Institute (SSI). Measurements of this group have been performed since 1959. By stratified random sampling 10oO individuals aged between 1 and 75 years were selected in December 1986. A stratified sampling method was applied to ensure an over-representation of the most highly contaminated areas (the counties of Uppsala, Vastmanland, Gavleborg, Vaster-

Fig. 5.6.3. The estimated I3’Cs surface activity (kBq m”) resulting from the Chernobyl fallout in Sweden based on the measurements made by.the Swedish Geological Company (SGAB) between May 1 and October 10, 1986. Sweden is divided into two fallout regions A and B according to the surface activity. Region A consists of the 5 counties most affected by fallout (Uppsala, Vastmanland, Gavleborg, Vasterbotten and Vasternorrland) and B the remaining 19 less affected counties.

463

botten and Vistemorrland) and also to ensure a correct age representation for the rural areas. After a randomly performed subsampling procedure, 250 people from this group were selected for whole-body counting. Of these, 218 took part in the measurements in March-April 1987, and 109 people from this group returned for a second measurement in April-May 1988 (Falk et al.,1991). Measurements were not repeated after that. The mean deposition of '"Cs in Sweden calculated from the data given by Edvards (1991) was 10 kBq/m2. INTERNAL DOSE CALCULATION METHOD

In order to make the dose estimations comparable, the method of dose calculation used in Finland was chosen. Committed effective doses for 137Cswere calculated using the dose factor 2.5 x 10" Sv per (Bq a kg-') given in UNSCEAR. To calculate the committed effective dose for '"Cs, use was made of the dose factor 3.6 x 10" Sv per (Bq a kg-I), obtained by multiplying the dose factor for '"Cs by the ratio of the ingestion dose fattors for '"Cs and '"Cs in ICRP 30. The same dose factors were used for both adults and children. Depending on the data available, the internal radiation doses were calculated using either individual '"Cs and 137Csvalues, or if individual values were not available using the mean values given for the groups. To avoid the effect of body size, values expressed as Bq/kg body weight were used. Assuming that intake rate is constant throughout the year, the effective internal dose delivered in a specified year is the fraction (0.61) of the committed effective dose to be received for 50 years following the intake. The doses given are the mean values of female and male groups. The activity time integrals were calculated separately for '"Cs and '"Cs. When calculating the activity time integrals for the reference groups, it was assumed that the body burdens of '"Cs and 137Cschanged linearly between the subsequent measurements. The activity time integrals for members of the population groups were calculated assuming that the individual body burdens of

'"Cs and 137Csvaried relatively in the same way as the mean body burden in the reference group each year. RESULTS

Denmark The individual results of the whole-body counter measurements of the Risa reference group are taken from the annual reports. In the reports individual body burdens of radiocaesium are given as the sum of '"Cs and "'Cs and expressed in Bq per kg potassium in the body, and the amounts of potassium expressed in grams per kilogram body weight. These values and the ratio of the activities of '"Cs and 137Cscorrected for the date of measurement are used to calculate body burdens of '"Cs and '"Cs in Bqlkg body weight. For the ratio of '"Cs and 137Cson 26 April

464 1986, 0.55, was used as the initial value. .The variation of mean L37Csbody burden of the

reference group is given in Fig. 5.6.4. The internal effective doses calculated from the results of the reference group measurements, and assumed to represent mean internal doses delivered each year to the whole population, are given in Table 5.6.1. For comparison, the annual doses together with the corresponding doses from the other Nordic countries are given in Fig. 5.6.7.

* Cs-137 body content (Bq/kg)

35 30 25 20

15 10

Fig. 5.6.4. The variation of 137Cs body burdens in the control groups at Helsinki (F), Rise (D), Sel (N) and Stockholm (S) in 1986 -1990. Table 5.6.1. Mean effective internal doses (mSv) from '"Cs and I3'Cs delivered to the Danish population in 1986 -1990.

YEAR

'"CS

137cs

TOTAL

1986 1987 1988 1989 1990

0.0008 0.0022 0.0012 0.0004 0.0002

0.0012 0.0042 0.0029 0.0013 0.0010

0.0020 0.0064 0.0041 0.0017 0.0012

465

Finland The changes of the 137Csbody burden in the Helsinki reference group are presented in Fig. 5.6.4. As can be seen, the maximum body burdens appeared in the summer of 1987. After the Chernobyl accident Finland was divided into five fallout regions, 1 - 5, according to increasing amounts of 137Csdeposition, as illustrated in Fig. 5.6.1. The results of whole-body counter measurements of the population group were divided into five subgroups, according to the

Fig. 5.6.5. Mean body burdens of 137Csin the Finnish population at the end of each year from 1986 to 1990 as mean values of the five fallout regions (1 -5) and the whole population (F). Table 5.6.2. Mean effective internal doses (mSv) from '"Cs and 137Csdelivered to the Finnish population in 1986 -1990. YEAR

'"CS

1986 1987 1988 1989 1990

0.020 0.028 0.016 0.009 0.005

'37Cs

0.031 0.050 .0.037 0.030 0.022

TOTAL 0.051 0.078 0.053 0.039 0.027

fallout region in which members resided. Variation of mean body contents of "%s

in the

subgroups and in the whole population group are presented in Fig. 5.6.5. For comparison, all the results were normalized to the end of each year with the aid of the results of the reference group.

466

The relative numbers of men, women and children were used as weighting factors when the mean values for each subgroup were calculated. The weighted mean value for the whole population was calculated using the number of people residing in the corresponding fallout region as the weighting factor. Effective internal doses delivered each year are given in Table 5.6.2 and the total doses in Fig. 5.6.7. Norway The results of whole-body counter measurements of the two randomly selected groups Sel and Oslo are shown in Fig. 5.6.4. There is only one measurement of the Oslo group, made in May 1987. The mean 137Csbody burden for women and men was then 11.9 Bq/kg, being about 40 per cent of the corresponding value for the Sel group. The mean value for the whole population was estimated to lie between these values. The internal effective dose for the Sel group from the '"Cs and 137Csdelivered in 1986 was 0.017 mSv, in 1987 0.065 mSv, in 1988 0.060 mSv, in 1989 0.051 mSv and in 1990 0.040 mSv. The doses are shown for comparison in Fig. 5.6.7. The total dose delivered during this period was 0.23 mSv. Assuming that the body burdens of radiocaesium during the period 1986

-

1990 behaved in Oslo relatively in the same way as in Sel, the

corresponding dose for people in Oslo would be 0.095 mSv. Figure 5.6.7 gives the annual doses from Sel and shows the upper Iimit of the dose for the whole population. Sweden The results of the I3'Cs body burdens from the whole-body counting of the Stockholm reference

Fig. 5.6.6. Weighted mean body burdens of '"Cs in the Swedish population in 1987 and 1988 based on the whole-body measurements. Region A contains the 5 counties most affected by fallout and region B the remaining 19 counties less affected. Region Sweden is the weighted mean for the whole population.

467

group after the Chernobyl accident are shown in Fig. 5.6.4. In Sweden the population group wasonly measured in 1987 and 1988. The weighted mean values of 137Csbody burdens for the population group were 7.7 Bq/kg in March 1987 and 5.6 Bq/kg in April 1988. The variation of the weighted mean body burdens for p p l e residing in the areas of low and high 137Csdeposition and for the whole population is presented in Fig. 5.6.6 for March 1987 and April 1988. Table 5.6.3. Mean effective internal doses (mSv) from '"Cs and 1377cs delivered to the Swedish population in 1986 -1990. YEAR

'"CS

'37Cs

TOTAL

1986 1987 1988 1989 1990

0.003 0.007 0.004 0.002 0.001

0.006 0.01 1 0.009 0.006 ,0.005

0.009 0.018 0.013 0.008 0.006

For dose calculation, the activity time integrals for the years 1986, 1989 and 1990 were obtained by extrapolation with the aid of the reference group values. The estimated effective internal doses from '"Cs and 137Csreceived by the Swedish population are given in Table 5.6.3. The total doses delivered each year are shown in Fig. 5.6.7.

Fig. 5.6.7. Effective internal doses from '"Cs and 137Csdelivered to the Danish, Finnish and Swedish populations and to the Sel group (Norway) in 1986 -1990.

468 DISCUSSION The whole-body counting results take into account the variations of radionuclide intake caused by changes in the activity concentrations of foodstuffs and changes in the composition of the diet. The changes can be made voluntarily or be the result of countermeasures of which the effect is very difficult to take into account when the radiation dose estimate is based on food consumption statistics. The results also take into account differences in the biological half-lives of the radionuclides affecting the accumulation and excretion of activity in the body. According to investigations in Finland (Suomela,1971, Hasiinen and Rahola,1971) and in Sweden (Falk et al.,1991), the mean biological half-life of radiocaesium is about 65 days for women and 85 days for men, or clearly shorter than the mean half-life of 110 days given by the ICRP. For estimation of the internal radiation dose to the whole population in a country, a group of people should be selected for whole-body counting by means of statistical methods in such a way that it represents the whole population living in areas with different radiocaesium deposition.

In practice such a large group cannot be measured often enough to follow temporal changes of body activity after an accident. Thus smaller reference groups that can be measured several times a year are needed to supplement the measurements of the population group. The temporal behaviour of the 134Csand '"Cs body burdens was studied by means of such reference group measurements. To be able to follow the changes of radiocaesium body burdens reliably, the group should be measured at least four times a year. The number of measurements made in a year is normally limited by practical and economic considerations. The results of the measurements in Denmark, Finland and Sweden show that the maximum radiocaesium body burdens were reached in 1987. The measurement frequency in Norway does not allow any conclusion regarding the time when the maximum level was reached or the precise temporal behaviour after that. Especially in 1987 this leads to a small difference between the delivered dose converted from the committed dose reported by Strand (1992) and by us. After 1987 the differences were negligible. In Norway, where several countermeasures were taken, the measurement of a group representing the whole population would have been especially useful. The mean committed effective doses given by Falk (1991) for Sweden were 0.023 and 0.027 mSv respectively for the first and second year after the accident. The corresponding doses calculated using the method described in this paper gave 0.021 and 0.031 mSv, respectively. The results do not differ from each other significantly. The doses calculated in this paper (Table 5.6.3) are the effective doses delivered to the population in the calendar year. The '%Csand 137Csintake values for the years 1986-91 are given by Aarkrog in Table 5.5.4 in the previous chapter. When these intake values are converted to delivered effective doses, a comparison can be made with the corresponding doses obtained by whole-body counting. This

469

shows that the mean internal dose estimate from food intake for the Danish population is about twenty per cent higher than that estimated from the whole-body counting results. In Finland the corresponding dose calculated from food intake is approximately twice as high as that calculated from the whole-body counting results. The ratio is in agreement with the data given by Suomela et al. (1991). Assuming that the delivered mean effective dose for the Norwegian population lies between that for Oslo (0.095 mSv) and that for Sel (0.23 mSv), the ratio varies from 3.4 to 1.4 and is in agreement with the ratio calculated by Strand et al. (1992).

For Sweden the

corresponding ratio is about eight using the intake estimate given by Aarkrog. In Sweden no countrywide surveys of foodstuffs were made and thus this estimate is of necessity more uncertain than the corresponding estimates for the other countries. Differences greater than a factor two or more have been reported by other investigators (Steger et al.,1989, MAlatovA et al.,1989). However, if the shorter biological half-lives (65 days for women and 85 days for men) reported in Finland and Sweden are used instead of the value (110 days) given by the ICW, the mean activity time integral based on food intake decreases about 20 per cent. This leads to a smaller difference between the doses calculated from food intake and whole-body counting data.

CONCLUSIONS With the data available it was not possible to reach better agreement between the internal dose estimates based on food intake and on whole-body counting data. The next step would be to study special groups more carefully regarding dietary habits and activity intake and to compare the body activity levels calculated from these data with the results of the whole-body counting measurements of the same group. It must be possible to measure this group frequently enough to make accurate calculations of activity time integrals for the activity in the body. The biological half-life of radiocaesium for this group should be measured and used in the calculation of activity time integrals and radiation doses.

REFERENCES Aarkrog A., L. Better-Jensen, C.Q. Jiang, H. Dahlgaard, H. Hansen, E. Holm, B. Lauridsen, S.P. Nielsen and J. Sragaard-Hansen (1991). Environmental radioactivity in Denmark in 1988 and 1989. Ris0-R-570. Risra National Laboratory, Denmark, 172-179. Aarkrog A., L. Better-Jensen, C.Q. Jiang, H. Dahlgaard, H. Hansen, E. Holm, B. Lauridsen,

S.P. Nielsen, M. Strandberg and J. Sragaard-Hansen (1992). Environmental radioactivity in

Denmark in 1990 and 1991. Risra-R-621(EN). Risra National Laboratory, Denmark, 118-120.

Arvela, H., M. Markkanen and H. Lemmela (1990). Mobile survey of environmental gamma radiation and fallout level in Finland after the Chernobyl accident. Radiation Protection Dosimetty 32, NO. 3,177-184.

470 Eidvardson, K. (1991). Fallout over Sweden from the Chernobyl accident. In: The Chernobyl fallout in Sweden. Ed. L. Moberg, The Swedish Radiation Protection Institute, Sweden, 47-65. Falk, R., G. Eklund, H. Gem and I. Ostergren (1991). Cesium in the Swedish population after Chernobyl: Internal radiation, whole-body counting. In: The Chernobyl fallout in Sweden. Ed. L. Moberg, The Swedish Radiation Protection Institute, Sweden, 547-577. Hashen, E. and T. Rahola (1971). The biological half-life of '"Cs and 24Nain man. Annals of Clinical Resarch 3,236-240. International Commission on Radiological Protection (1978). Limits for intakes of radionuclides by workers. Publication 30, Annuls of the ICRP 2 No. 3/4, 91-93. Kendall, G.M., B.V. Kennedy, J.R. Greenhalgh, N. Adams, and T.P. Fell (1987). Committed dose equivalent to selected organs and committed effective dose equivalent from intakes of radionuclides. Report No. NRPB-GS7. 114 pp. Leggett, R.W., K.F. Eckerman, D.E. Dunning, M. Christy, D.J. Crawford-Brown and L. Williams (1984). Dose rate to organs as a function of age following internal exposure to radionuclides. Report ORNL/TM-8265. 134 pp. MalAtovA, I., I. BuEina, I. CespirovA, D. DrAbovh and J. Thomas (1989). Committed effective dose equivalents from internal contamination of the Czechoslovak population after the Chernobyl accident. Radiation Protection Dosimetry 28, No. 3,291-301. National Council on Radiation Protection and Measurements (1977). Cesium 137 from the environment to man: Metabolism and dose. NCRP Report No. 52. Rahola, T., M. Suomela, E. Illukka, M. Puhakainen, and S . Pusa (1987). Radioactivity of people in Finland after the Chernobyl accident in 1986. Report STUK-A64. Finnish Centre for Radiation and Nuclear Safety, Finland. 35 pp. Rahola, T., M. Suomela, E. Illukka and S. Pusa (1989). Radioactivity of people in Finland in 1987. Report STUK-A81. Finnish Centre for Radiation and Nuclear Safety, Finland. 69 pp. Rahola, T., M. Suomela, E. Illukka and S. Pusa, (1991). Radioactivity of people in Finland in 1988. Report STUK-A91. Finnish Centre for Radiation and Nuclear Safety, Finland. 69 pp. Rahola, T., M. Suomela, E. Illukka, M. Puhakainen and S. Pusa (1993). Radioactivity of people in Finland in 1989 - 1990. Report STUK-A96. Finnish Centre for Radiation and Nuclear Safety, Finland. 104 pp. Segaard-Hansen, J. and B. Lauridsen (1988). Radiocesium in the human body. Rise-R-549. Risnr National Laboratory, Denmark, 202-205. Segaard-Hansen, J. and B. Lauridsen (1989). Radiocesium in the human body. Risnr-R-563. Risnr National Laboratory, Denmark, 113-118. Steger, F., K. Muck and K. Duftschmid (1990). Comparison of dose estimates derived from whole-body counting and intake calculations based on average food activity concentration. In: Environmental contamination following a major nuclear accident. IAEA-SM-306, Vol. 2 339-349.

47 1

Strand, P., E. B B ~ L. , Berteig, T. Berthelsen, T. Strand and 0. Harbitz (1991). Whole-body counting and dietary surveys in Norway during the first year after the Chernobyl accident. Radiation Protection Dosimetly 27, No 3, 163- 171. Strand, P., T.D. Selnies, E. Bw, 0. Harbitz and A. Andersson-Ssrlie (1992). Chernobyl fallout: Internal doses to the Norwegian population and the effect of dietary advice. Health Phys. 63(4) ,385-492. Selnies, T. (1993). Databank of the National Institute of Radiation Hygiene (not published). Norway. Suomela, M. (1971). Eliminationshastighet for 137Csi enskilda mhniskor och i kontrollgrupp. (In Swedish). In: The 3rd ordinary meeting of the Nordic Society for Radiation Protection, Copenhagen, Aug. 18-20 1971. 13 pp. Suomela, M. (1984). Whole-body counter studies in radiation protection and clinical research. STL-A45. Finnish Centre for Radiation and Nuclear Safety, Finland. 53 pp.

This Page Intentionally Left Blank

413

DEFINITIONS, TERMS AND UNITS

To reduce the risk of misinterpretation, the following list gives explanations of selected technical terms used in the present book. This is necessary because, for some terms, e.g. transfer factors and half-lives, there are significant differences in the perception of their definitions.

aggregated transfer coefficient: The relationship between the ground deposition of e.g. 137Csand the content of this in a product, for instance milk:

Tag (or Tags) = Bq 137Cskg-'

milk / Bq I3'Cs m-* deposited. allochthonous material: In aquatic ecology: material produced outside the lake system, e.g. the leaves of terrestrial plants carried by inflowing streams. autochthonous material: In aquatic ecology: material produced within the lake system, e.g. phytoplankton or aquatic macrophytes. Bioconcentration factor, see concentration factor. B,: bioavailability: being available for biological uptake and turnover. Generally, only part of the total amount of an element in an environment is bioavailable. Bq:

becquerel: the unit of activity for a radioactive element: s-', i.e. the number of decays per second.

coarse particulate organic matter (CPOM): Plant material 2 1 mm, terrestrial as well as aquatic. concentration factor: In aquatic ecology: CF = Bq kg-' sample (dry or fresh weight) / Bq kg-' water at assumed equilibrium. As no general convention exists, it is essential always to indicate dry or fresh weight basis. Because true equilibrium hardly ever occurs under natural conditions, the terms concentration ratio (CR)or bioconcentration factor (Bo are in practice used synonymously. critical group: That group of the population assumed to receive the highest radiation doses from a given source. fractionation: Separation by e.g. ultrafiltration of the total radioisotope content into various particle-size fractions.

F,:

See: transfer coefficient.

414

half-life (T,,J: Time (e.g. in years) during which the original amount or concentration of an element is reduced to half by an assumed exponential reduction. The physical half-life

(Tlh,Jis the characteristic time (e.g. in years) during which a radioactive element is reduced to half its original amount by physical decay (T,h,p= In 2 /

h, where h is

the decay constant (yr-I)). Provided the amount or concentration of the radionuclide in a component of the environment, e.g. a plant species, follows a single exponential reduction with time, the effective half-life (T,h,efl) (synonyms: observed half-life

(T,h,obs),effective ecological half-life (T,,*,-), effective environmental half-life), where T1h,e, = In 2 / heff, describes the decrease in the observed concentration of a radionuclide with time, including the effects of environmental and ecological processes

as well as physical decay. The biological half-life (T,h,b)in an organism is that part of the effective half-life attributed to biological excretion, i.e., it does not include the effect of physical decay. Likewise, the ecological half-life (T,,J

(synonym:

environmental half-life) is that part of the effective half-life attributed to ecological or environmental processes. In other words, TH,band Tth, are equivalent to the halflife of a stable element without any decrease due to physical decay. Biological and ecological half-lives may be calculated after making a correction for physical decay to one common time, or by applying the relations he, =

+ hb and heff= + he or

the equivalent relations (Tlh,eff)-'= (TIh,J-'+ (Tlh,b)-'and (TH,eff)-'= (T,h,p)-'+ (TH,e)-l. Unfortunately, the above definitions are not universally followed here.

K,:

A distribution coefficient defined as the quotient between element concentrations at assumed equilibrium in solid materials (e.g. sediment) and water (Bq kg-' dry material / Bq kg-' water).

low molecular weight fraction (LMF): Components in the water phase with a nominal molecular diameter of

< 1.2 nm.

macromolecular fraction (HMF): Particles in the water phase with a nominal molecular diameter of

> 100 nm.

pseudocolloidal fraction (CF): Components in the water phase with a nominal molecular diameter in the range 1.2 - 100 nm.

radiation dose: Health physicists use an abundant nomenclature to define various aspects of the absorbtion of ionizing radiation by the human body. Only the chief terms will be discussed here. The effective dose eqivalent (Sv) is corrected for various differences

in radiation quality and in the different sensitivity of various organs to radiation damage to give the same magnitude of effect per unit. The effective dose commitment or the committed effective dose describes the assumed accumulated effective dose

475

integrated over 50 years, whereas the dose rate (Sv/yr) gives e.g. the annual dose at a given time. Internal radiation doses result from radiation emitted by radionuclides inside the body; external doses are caused by penetrating radiation from radionuclides outside the body, e.g. on the soil. The doses can be expressed as individual doses (Sv) to single individuals in a population or as collective doses (man-Sv), i.e. the sum of all individual doses in a given population. radioecological sensitivity (Bq kg" yr / Bq m-') (synonym: transfer coefficient from fallout to sample): is defined as the ratio of the infinite integral of concentrations in the sample to the integrated deposition. This is equivalent to UNSCEAR's transfer coefficient P,, = IC /

IF,, where IC is the integrated concentration of the radionuclide in the diet and

IF, the integrated deposition of that radionuclide.

statistical terms:

standard deviution: S.D.

=

s

=

coefficient of variation: CV: relative standard deviation.

sv:

sievert: the unit of dose equivalent = 1 J kg''.

Tag = Tag,: See: aggregated transfer coefficient. T,:

See: transfer factor from ground deposition.

TF,: TF,:

See: transfer factor from deposition to fish.

See: transfer factor from ground deposition

TF,:

See: transfer factor from deposition to surface water.

transfer factor from deposition to surface water: TF, = C, / D, where C, is the annual average concentration of 137Csin water (Bq m-3) and D is the average deposition of I3'Cs (Bq m-*) corrected annually for the physical decay. transfer factor from deposition to fmh: TF, = Cf / D, where C, is the annual average concentration of 137Csin fish (Bq kg-' fresh weight) and D is the average deposition of 137Cs(Bq m-2) corrected annually for the physical decay. transfer coefficient from fallout to sample: See: radioecological sensitivity (Bq kg-' yr / Bq m-'). transfer factor from ground deposition (TF,) or synonym transfer factor from soil to plants (Tf) is the ratio of the concentration in a product, e.g. in pIant material, and the

476 deposition on the ground surface; Bq kg-' I Bq

= m2 kg-I. Weight basis, dry- or

fresh-weight, must be specified. transfer coefficient to milk (Fm): The milk transfer coefficient (F, = Bq L-'milk / Bq ingested d-') represents the fraction of the daily intake of a nuclide that is transferred to a litre of milk at equilibrium.

UNSCEAR: United Nations Committee on the Effects of Atomic Radiation. whole-body counting: Colloquial term for the measurement of X and y radiation emitted from radionuclides in the human body.

Prefures: E

exa

1018

P T

Peh tera

1015

G M

gigs

109

mega

106

k

kilo

103

m

milli

10-3

Lc

micro

10"

n

nano

10-9

P

pic0

10-12

f

femto

1045

a

atto

10-18

10'2

417

INDEX

Common words used throughout the book, such as 13’Cs, radiocaesium, potassium, Chernobyl, etc. are not included in the index. Words appearing several places in each paper refer to the first page of the paper or the first or main appeareance. Refer also to the species index, page 481. Page numbers in bold refer to the section on definitions, terms and units, page 473.

7Be

%r

385

biological effects

324

157, 239

biological uptake

96 94

95~r

173

BIOMOVS

1311

158

bioturbation

144Ce

173

BlocksLtjkn (lake)

128

45, 129, 166

2IOPo

119, 385

Blomhojden

169

222Rn

119

body burden

45 8

absorbtion of 13’Cs

214

bogs

actinides

168

boreal forest

335

Bothnian Sea

106

aggregated transfer factor 187, 197, 211, 287, 320, 445, 473

63, 288, 335

Bottentjiirn (lake)

56

agricultural ecosystem

143

catchment

air samplers

385

cellulose filter

387

Alstadhaug

174

centrifugal pumps

385

analytical quality assurance

383

chemical fractionation

annual crops

185

coarse particulate organic

Baltic sea

42, 68, 105, 119

Belorus

181

bioaccumulation factor

96

biological half-life 85, 214, 233, 265, 303, 458, 474 biological availability 165, 239, 271, 335, 473

29, 63, 80, 129, 335

material(CP0M) coastal waters collective doses colloids complexation

45 29, 473 105

287, 433, 457, 475 29, 166 166

concentration factor

211, 473

concentration ratio

113, 473

478 condensed particles

168

Gulf of Finland

106

countermeasures

239, 298, 321

Gulf of Bothnia

106

critical group

287, 450, 473

Hamstasjon (lake)

deposition velosity

150, 232

deposition of 137Cs 9, 220, 239, 266, 338 diet (human)

185, 433

diffusion

45

direct contamination

151, 271

distribution coefficient (Kd)

82

HELCOM / MORS

107

high-volume air samplers

385

Hillesjon (lake)

29, 91

hot particles

165

indicator organism

105

indirect contamination

149

individual doses

87, 96, 129, 474 dose assessment 19, 119, 372, 433, 474 drainage basins

63, 93

dynamic model

229

ecological half-life

433, 451, 475

inert fraction

165

intercalibration

383, 407, 425

interception

151, 232, 341

intercomparison

385, 425

internal dose

433, 457, 474

ions 240, 265, 295, 474

ecological tracer

127

effective ecological half-life 63, 185, 197, 211, 287, 303, 442, 474 efficiency calibration external dose

408 433, 457, 475

29, 165

Is0 Valkjiirvi (lake)

97

Jokionen

169

labile fraction

165

lake sediments

29, 45, 87, 93, 127

lateral transport

268, 345

lichen-reindeer-man food-chain

303

external radiation

239

Lierne

169

fertilization

239

liming

239

field loss

151

long-term perspective

46

Flatsjon (lake)

Loppesjon (lake)

387

Liivsjon (lake)

food production

433

low molecular weight

food web

127

fraction (LMF)

FOA glass fibre filter

265, 287

forest ecosystem

281, 354 56, 128 82 34, 165, 474

macromolecular fraction (HMF) 34, 165, 474

fractionation

29, 45, 166, 473

fuel particles

165

marine environment

gamma spectrometry

425

mobile fraction

165

grain size (lake sediments)

54

mobility factor

170

Grundvattnet (lake)

82

model

105, 119

93, 229, 364

479 molecular weight fractions

29, 165

retention

monitoring organism

105

Ribe

natural ecosystems

433

root uptake

observed ratio Ortrasket (lake) 0vre Heimdalsvatn (lake) particle retention

169 185, 233, 239, 271

230, 373

runoff

179, 352

54

Russia

181

29, 97

Saarisjarvi (lake)

29

129, 165

seasonal variation

287, 316

particle tracers

127

seasonality

particle settling flux

129

sediment traps

particle focusing

131

sedimentation rate

particulate fractions

151

46

149 45 45, 87, 97, 127

Sellafield

107

Petrianov filter

387

seminatural ecosystems

263, 433

phantoms

407

sequential extraction

175, 270

45

Siggeforasjon (lake)

46

physical fractionation physico-chemical forms

29, 165, 339, 458

Simpnh

111

soil characteristics

217

soil-herbage-lamb food-chain

ploughing

239

polypropylene filter

387

population groups

459

sorption

positive displacement pumps

385

speciation

pseudocolloidal fraction

34, 474

211, 229 166 29, 165

SpellemannsAkern

174

165

quality assurance

383

stable isotopes

RAD-1

383

Storsjouten (lake)

RAD-2

23

82

surface adsorption on algae

115

RADJ

143

surface water

RAD-4

263

Tj~tta

169

tracer

127

radioecological sensitivity

18, 475

rainsplash

149

transfer factor

63

63, 79, 165, 197,

211, 230, 248, 265, 319, 373, 475

RAksjon (lake)

53

redistribution

266

translocation

refractory fission products

168

transport mechanisms (lakes)

remobilisation

130

trophic level

residence time

79

resuspension

29, 45, 131, 149, 185, 232, 303, 350

155, 343 45

79, 96, 122, 127

trophic dynamics

127

Ukraine

181

ultrafiltration

34

480 uptake from soil

185 94

VAMP

vertical distribution weathering

269, 339 151, 166, 232

Whatman GFIA glass fibre filter whole-body counting

387

407, 457, 476

wild produce

433

zeolite

254

48 1

SPECIES INDEX

Agrostis

224

Alces alces

287, 435

cabbage (Brassica oleracea)

185

Calluna vulgaris

288

alder (Alnus incana)

336

Canada goose (Branta canadensis)

291

Alectoria ochroleuca

31 1

capelin (Mallotus villosus)

120

algae

105

Capreolus capreolus

287

Ahus incana

336

Carex rostrata

272, 310

Amanita

275

Carex nigra

272, 310

Anas platyrhynchos

291

Carex sp.

arctic charr (Salvelinus alpinus)

79

caribou

303, 435

carrot (Daucus carota)

186

Ceramium gobii

115

272, 310

Cervus elaphus

290

122

Cetraria nivalis

307

94

Cladina stellaris

307

Cladina mitis

311

Cladophora glomerata

115

Clupea sprattus

122

aspen (Populus tremula) 272, 288, 310

Astragalus frigidus

224

baltic herring benthos barley (Hordeum vulgare)

156, 185, 244

berries

435

Betula

288, 336

Clupea harengus

120

Betula nana

272, 310

cod (Gadus morruha)

120

Betula pubescens

272, 310

Coelocaulon divergens

311

Columbia palumbus

29 1

Coregonus lavaretus

120

bilberry (Vaccinium myrtillus) 288, 343

birch (Betula sp.)

Cortinarius

275

black grouse (Lyrusus tetrix)

29 1

cows’ milk

158, 197, 240, 435

bladder-wrack (Fucus vesiculosus)

105

cucumber

248

Branta canadensis

29 1

Dama dama

290

Brassica oleracea

185

Daucus carota

186

brown algae

105

deer

435

brown hare (Lepus europeus)

290

Deschampsia

224

267, 288, 336

482 Deschampsia ceaspitosa Deschampsia flexuosa

272, 310 272, 288, 310

diatoms

105

elk (Alces alces)

287, 435

Lactarius

275

Lagopus lagopus

291

lamb

211, 229, 435

Leccinum

275

Enteromorpha intestinalis

115

Lepus timidus

290

epilithic diatom communities

105

Lepus europeus

290

Epilobium angustifolium

288

lettuce

248

37, 367

lichens

37, 303

Equisetum Eriophorum angustifolium

lingonberry (Vaccinium vitis-idaea) 288

272, 310, 367 Esox lucius

72, 79, 120

Lyrusus tetrix

29 1

mallard (Anas platyrhynchos)

291

fallow deer (Dama dama)

290

Mallotus villosus

120

Festuca

224

man (Homo sapiens)

Festuca ovina

288

287, 303, 372, 401, 433, 457

fireweed (Epilobium angustifolium) 288

Melampyrum ssp.

fish, freshwater

Menyanthes trifoliata 23, 63, 79, 93, 435

fish, marine

119, 435

flounder (Platichthys flesus)

122

Fucus vesiculosus

105

fungi

265, 287, 306, 435

19, 119, 185, 272, 310 272, 310, 367

moose (Alces alces)

287, 435

moss

37, 306

mountain ash (Sorbus aucuparia) 288, 336 mountain hare (Lepus timidus)

290 120

Gadus morruha

120

mussel (Mytilus edulis)

game animals

287

Myriophyllum

37

272, 310

Mytilus edulis

120

Geranium silvaticum grain

159, 240, 435

grass

21 1, 229, 240, 271

Norwegian spruce (Picea abies) 288, 336

green algae

115

Nuphar

heather (Calluna vulgaris)

288

oats

herbage

211, 229

Herring (Clupea harengus) Homo sapiens

12’0 19, 119, 185,

287, 303, 372, 407, 433, 457

36 159, 244

parsley

248

pasture, natural

271

pea (Pisum sativum) Perca fluviatilis

186, 244 72, 79, 98, 120

Hordeum vulgare

156, 185

perch (Perca fluviatilis) 72, 79, 98, 120

Juncus trifidus

272, 310

Phragm ites

37

272

Picea abies

288, 336

Juniperus communis

483 pike (Esox lucius)

72, 79, 120

Pilayella littoralis pine

Solidago virgaurea

272, 310

115

Sorbus aucuparia

288, 336

267

sprat (Clupea sprattus)

122

Pinus sylvestris

288, 336

Stereocaulon paschale

31 1

Pisum sativum

186, 244

Sui1I us

275

plaice (Pleuronectus platessa)

120

tomato

248

plants

265

Tricholoma

275

Platichthys flesus

122

Trifolium

224

Pleuronectus platessa

120

trout (Salmo trutta)

79, 98

272, 288, 310

Vaccinium myrtillus

272, 288, 343

Populus tremula

potato (Solanum tuberosum)

185, 435

Vaccinium vitis-idaea

288

ptarmigan (Ptarmigan mutus)

29 1

vegetables

Ptarmigan mutus

291

water lilly (Nuphar)

radish

248

wheat

159

Rangifer tarandus

303, 435

160, 185, 248, 435 36

white mustard

244

raspberry (Rubus idaeus)

288

whitefish (Coregonus lavaretus)

120

red algae

115

willow grouse (Lagopus lagopus)

291

red deer (Cervus elaphus)

290

willow (Salix spp.)

reindeer (Rangifer tarandus)

303, 435

223, 272, 288, 310

roe deer (Capreolus capreolus)

287

wood pigeon (Columbia palumbus) 291

Rozites

275

woodcock (Scolopax rusticola)

291

Rubus idaeus

288

Xerocomus

275

223, 272, 310

zooplankton

94

Rumex acetosa Russula

275

rye

159

ryegrass

242

Salix spp.

223, 272, 288, 310

Salmo trutta

79, 98

Salvelinus alpinus

79

Saussurea alpina

272,310

Scolopax rusticola

29 1

scots pine (Pinus sylvestris) sedge (Carex)

288, 336

224, 272, 310

sheep

211, 229

Solanum tuberosum

185. 435

This Page Intentionally Left Blank

Studies in Environmental Science Other volumes in thisseries 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33

Atmospheric Pollution 1978 edited by M.M. Benarie Air Pollution Reference Measurement Methods and Systems edited by T. Schneider, H.W. de Koning and L.J. Brasser Biogeochemical Cycling of Mineral-Forming Elements edited by P.A. Trudinger and D.J. Swaine Potential Industrial Carcinogens and Mutagens by L. Fishbein Industrial Waste Management by S.E. Jprrgensen Trade and Environment: ATheoretical Enquiry by H. Siebert, J. Eichberger, R. Gronych and R. Pethig Field Worker Exposure during Pesticide Application edited by W.F. Tordoir and E.A.H. van Heemstra-Lequin Atmospheric Pollution 1980 edited by M.M. Benarie Energetics and Technology of Biological Elimination of Wastes edited by G. Milazzo Bioengineering, Thermal Physiology and Comfort edited by K. Cena and J.A. Clark Atmospheric Chemistry. Fundamental Aspects by E. Meszaros Water Supply and Health edited by H. van Lelyveld and B.C.J. Zoeteman Man under Vibration. Suffering and Protection edited by G. Bianchi, K.V. Frolov and A. Oledzki Principles of EnvironmentalScience and Technology by S.E.J~ rgensen and I. Johnsen Disposal of Radioactive Wastes by 2.Dlouhjl Mankind and Energy edited by A. Blanc-Lapierre Quality of Groundwater edited by W. van Duijvenbooden, P. Glasbergen and H. van Lelyveld Education and Safe Handling in Pesticide Application edited by E.A.H. van Heemstra-Lequin and W.F. Tordoir Physicochemical Methods for Water and Wastewater Treatment edited by L. Pawlowski Atmospheric Pollution 1982 edited by M.M. Benarie Air Pollution by Nitrogen Oxides edited by T. Schneider and L. Grant Environmental Radioanalysis by H.A. Das, A. Faanhof and H.A. van der Sloot Chemistryfor Protection of the Environment edited by L. Pawlowski, A.J. Verdier and W.J. Lacy Determination and Assessment of Pesticide Exposure edited by M. Siewierski The Biosphere: Problems and Solutions edited by T.N. Veziroglu Chemical Events in the Atmosphere and their Impact on the Environment edited by G.B. Marini-Bettolo Fluoride Research 1985 edited by H. Tsunoda and Ming-Ho Yu Algal Biofouling edited by L.V. Evans and K.D. Hoagland Chemistry for Protection of the Environment 1985 edited by L. Pawlowski, G. Alaerts and W.J. Lacy Acidification and its Policy Implications edited by T. Schneider Teratogens: Chemicals which Cause Birth Defects edited by V. Kolb Meyers Pesticide Chemistry by G. Matolcsy, M. Nadasy and Y. Andriska Principles of EnvironmentalScience and Technology (second revised edition) by S.E. Jorgensen and I.Johnsen

34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55

56 57 58 59 60 61

Chemistry for Protection of the Environment 1987 edited by L. Pawlowski, E. Mentasti, W.J. Lacy and C. Sarzanini Atmospheric Ozone Research and its Policy Implications edited by T. Schneider, S.D. Lee, G.J.R. Wolters and L.D. Grant Valuation Methods and Policy Making in Environmental Economics edited by H. Folmer and E. van lerland Asbestos in Natural Environment by H. Schreier How t o Conquer Air Pollution. A Japanese Experience edited by H. Nishimura Aquatic BioenvironmentalStudies: The Hanford Experience, 1944-1984 by C.D. Becker Radon in the Environment by M. Wilkening Evaluation of Environmental Data for Regulatory and Impact Assessment by S.Ramamoorthy and E. Baddaloo Environmental Biotechnology edited by A. Blazej and V. Privarova Applied Isotope Hydrogeology by F.J. Pearson Jr., W. Balderer, H.H. Loosli, B.E. Lehmann, A. Matter,Tj. Peters, H. Schmassmann and A. Gautschi Highway Pollution edited by R.S. Hamilton and R.M. Harrison Freight Transport and theEnvironment edited by M. Kroon, R. Smit and J. van Ham Acidification Researchin The Netherlands edited by G.J. Heij and T. Schneider Handbook of Radioactive Contamination and Decontamination by J. Severa and J. Bar Waste Materials in Construction edited by J.J.J.M. Goumans, H.A. van der Sloot and Th.G. Aalbers Statistical Methods in Water Resources by D.R. Helsel and R.M. Hirsch Acidification Research: Evaluation and Policy Applications edited by T. Schneider Biotechniques for Air Pollution Abatement and Odour Control Policies edited by A.J. Dragt and J. van Ham Environmental Science Theory. Concepts and Methods i n a One-World, Problem-Oriented Paradigm by W.T. de Groot Chemistry and Biology of Water, Air and Soil. Environmental Aspects edited by J. Tolgyessy The Removal of Nitrogen Compounds from Wastewater by B. Halling-Sorensen and S.E.Jpcrgensen Environmental Contamination edited by J.-P.Vernet The Reclamation of Former Coal Mines and Steelworks by I.G. Richards, J.P. Palmer and P.A. Barratt Natural Analogue Studies in the Geological Disposal of Radioactive Wastes by W. Miller, R. Alexander, N. Chapman, I. McKinley and J. Smellie Water and Peace i n the Middle East edited by J. Isaac and H. Shuval EnvironmentalOriented Electrochemistry edited by C.A.C. Sequeira Environmental Aspects of Construction with Waste Materials edited by J.J.J.M. Goumans, H.A. van der Sloot and Th. G. Aalbers Characterization and Control of Odours and VOC i n the Process Industries edited by S. Vigneron, J. Hermia and J. Chaouki

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