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 36 63 13.21 8.32 Lake RBksion (sample 9 m depth, max 10 m) 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 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 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.
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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
This Page Intentionally Left Blank
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 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 (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