California State University, Northridge

CALIFORNIA STATE UNIVERSITY, NORTHRIDGE

Secret Nuclear Meltdown? Measuring Cesium-137 from Environmental Samples to Determine

Radiation Exposure from the Santa Susana Field Laboratory, Simi Valley, California

A thesis submitted in partial fulfillment of the requirements

For the degree of Master of Science in Geology

Justin Raul Pardo

August, 2020 The thesis of Justin Raul Pardo is approved:

Dr. Trevis Matheus Date

Dr. Priya Ganguli Date

Dr. Jennifer Cotton, Chair Date

California State University, Northridge

Acknowledgments

Before I joined the geology department, I was learning about stable isotopes and their applications from Dr. Jennifer Cotton’s geochemistry class. I knew I caught the “isotope bug” of wanting to do research in isotope geochemistry and would like to graciously thank my advisor,

Dr. Cotton for taking me as her first graduate student. Joining Dr. Cotton’s lab allowed me to participate in research projects, learned from top researchers in my field at the Isotopes in the

Spatial Ecology and Biogeochemistry Course at the University of Utah, and collaborate with new researchers.

I want to thank next my co-advisor Dr. Trevis Matheus, who has shown and taught me the dendrochronological world. This project would not have been completed without Dr.

Matheus in the field and lab mentoring me in how we collect and cross-date tree-cores and wood samples.

I would also like to thank Dr. Priya Ganguli for being on my thesis committee and directing me to take a class with my current advisor that began my academic journey. Dr.

Ganguli spoke highly about Dr. Cotton’s geochemistry class, which has since positively changed my life and led me to enroll in a doctorate program at the University of Ottawa

This project could not have existed without Dr. Joshua Schwartz informing Dr. Cotton about this idea and the financial support by Marilyn Hanna, CSUN Graduate Studies’ Programs

(i.e., Graduate Equity Fellowship, Graduate Student Travel Funding, and Thesis Support),

California State University Sally Casanova Pre-Doctoral Program, and Associated Students’

Scholarship in Recognition of Tom Piernik and Student Travel & Academic Research program.

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Table of Contents

Signature Page…………………………………………………………………………………...ii

Acknowledgments……………………………………………………………………………...iii

List of Tables……………………………………………………………………………………vi

List of Figures………………….………………………………………………………..………vii

Abstract…………………………………………………………………………………………viii

Section 1. Introduction………………………….………………………………………………..1

1. 1 Santa Susana Field Laboratory (SSFL) Secrecy ……………………………………1

1.2 Nuclear Reactors and Meltdowns…………………………………………………....2

Section 2 Background……………………………………………………………………………5

2.1 Site Selection History for SSFL……………………………………………………..5

2.2 Secret Nuclear Meltdown……………………………………………………………6

2.2.1 Formation of the Sana Susana Field Advisory Panel……………………...7

2.3 What is Cesium-137…………………………………………………………………8

2.4 Importance of Cesium-147 to Human Health …………….………………………...10

2.5 Radiocesium in the Environment from Atmospheric Testing……………………….11

2.6 Previous SSFL Studies………………………………………………………………13

2.7 Cesium-137 in the Ecosystem……………………………………………………….15

2.7.1 Cesium-137 in Wood………………………………………………………15

2.7.2 Cesium-137 in Wildlife……………………………………………………17

2.8 Dendrochronology and Dendrochemistry....…………………………………...……18

2.9 Potential Impact of this Project……………………………………………………...19

Section 3. Methods………………………………………………………………………………21

3.1 Field Site……………………………………………………………………………..21

3.2 Tree-Ring Process……………………………………………………………………22

3.2.1 Collecting Tree-Cores and Wood………………………………………..23

3.2.2 Mounting and Sanding Tree-Cores………………………………………25

3.2.3 Ring-Counting……………………………………………………………25

3.2.4 Cross-dating……………...………………………………………….…..26

3.3 Small Mammal Collections……………………………………………………….28

3.4 Background Concentration of Cesium-137……………………………………….31

3.5 Measuring and Reconstructing Cesium-137 Levels………………………...…….32

Section 4. Results……………………………………………………………………………..36

4.1 Dendrochronology………………………………………………………………...36

4.2 HpGe Radiation Detector Analyses………………………………………………38

4.2.1 Tree-Cores and Wood…………………………………………………..38

4.2.2 Small Mammals………………………………………………………...40

4.3 NaI Radiation Detector Analyses…………………………………………………42

Section 5. Discussion…………………………………………………………………………46

5.1 SSFL Analysis…………………….………………………………………...……46

5.2 Research Assumptions and Limitations…………………………………………..46

5.2.1. Assumptions in the Study………………………………………………46

5.2.2 Potential Analyses of Strontium-90…………………………………….47

5.2.3 Lack of Samples Closer to SSFL……………………………………….47

5.3 Small Mammals Surrounding SSFL vs. Chernobyl and Fukushima…………...... 48

5.4 Wood Near SSFL vs. Chernobyl and Fukushima.………………………...……...49

5.5 The Possibility of an Extensive Nuclear Release at SSFL………………………..51

5..5.1 Potential Explanation for a Spike in Cancer Rates……………………..54

Section 6. Conclusion…………………………………………………………………………55

References……………………………………………………………………………………..56

List of Tables

Table 1: Wood and tree-core sets location …………………………………………………..23

Table 2: Small mammals sets that were collected for gamma-ray analysis………………….30

Table 3: Tree-cores used in cross-dating and their mean ages……………………….……....36

Table 4: COFECHA results…………………………………………………………………..37

Table 5: 2 sets of tree-cores and 3 sets of wood samples assayed by HpGe detector………..39

Table 6: Cesium-137 results from the HpGe detectors for 5 small mammal sets……………40

Table 7: 2020 and 1959 estimated maximum radiocesium levels……………………………41

Table 8: NaI detector results………………………………………………………………….42

List of Figures

Figure 1: Santa Susana Field Laboratory location………………………………………..6

Figure 2: The decay scheme of Cesium-137……………………………………………..9

Figure 3: Map of properties surrounding SSFL…………………………………….……13

Figure 4: Tree-core and wood sample locations………………………………………....22

Figure 5: Small mammal sample locations…………………………………………..….28

Figure 6: NaI Energy Spectrum of Trinitite and Wood …………………...……………42

Figure 7: NaI Energy Spectrum of Backround…………………………..…………..….43

Figure 8: NaI Energy Spectrum of Small Mammals……………………………………44

Figure 9: NaI Cesium-137 Peaks of Trinitite, Background, and Samples………………44

Figure 10: Comparison of radiocesium in small mammals……………………………..48

Figure 11: Comparison of radiocesium in Wood……….………………………………50

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Abstract

Secret Nuclear Meltdown? Measuring Cesium-137 from Environmental Samples to Determine

Radiation Exposure from the Santa Susana Field Laboratory, Simi Valley, California

Justin Raul Pardo

Master of Science in Geology

The Santa Susana Field Laboratory (SSFL) was a nuclear reactor development, rocket, and missile testing facility built ~48 km (30 miles) northwest of downtown Los Angeles in Simi

Valley, CA. For two weeks in July of 1959, highly radioactive material was potentially emitted into the atmosphere following a secret nuclear incident. The partial meltdown was withheld from the public for three decades and what occurred there is surrounded in secrecy. Since then, independent studies hired by Boeing and the SSFL advisory panel, made up of researchers from the UCLA school of public health have produced conflicting reports of an estimated radionuclide

viii release. Previous soil studies examining radiocesium a half-life later, or 50% original material remaining, could not identify any soil samples nearby to the SSFL facilities or local communities that had concentrations statistically different from the mean of 0.087 pCi g-1 or 3.219 Bq kg-1 for non-contaminated background samples (McLaren and Hart, 1993; Hamilton, 1997).

However, radiocesium is highly soluble in water and bound to soil particles, so 30 years is an ample amount of time to remove it from the surface via rain, wind, or wildfires.

In this study, we focused on environmental samples present during the 1959 incident, and shown to have great promise in retaining radiocesium by collecting wood of three species of the genus Quercus (Oak) from sites 2 – 55 km away from SSFL, as well as preserved small mammals at local museums from the families of Cricetidae and Soricidae, that were collected from 1959-1964 about 10 – 60 km around SSFL. The samples were analyzed by both NaI and

HpGe gamma-ray detectors, to estimate the concentration of the radioactive isotope Cs-137. All samples, measured in the years 2019 and 2020, were below detection limit for Cs-137.

Depending on the mass of these samples and the sensitivity of the detector, these samples all contained Cs-137 levels < 4.16 Bq kg-1 and indistiguishable from uncontaminated background levels. All these samples support the assertion that if any released nuclear material from SSFL entered the ecosystem was negligible amounts in comparison to “natural” radioactivity from the

Atmospheric Bomb Testing in Nevada from the 1940s.

Section 1: Introduction

1.1 Santa Susana Field Laboratory (SSFL) Secrecy

In 1989, an internal DOE study finding widespread radioactive and chemical contamination at the Santa Susana Field Laboratory (SSFL) was leaked to the public (Wing et al., 2006; Rucker, 2009). The Boeing Company, formerly Rockwell International Rocketdyne

Division, operated SSFL. Additionally, the Department of Energy (DOE), Atomic Energy

Commission (AEC), and National Aeronautics and Space Administrations (NASA) all participated in various operations and research at SSFL. Contamination at SSFL is believed to have been released from either a secret partial nuclear meltdown at the experimental nuclear power reactor, a facility that fabricated fuel from plutonium and uranium, a “hot lab” that researched the irradiated nuclear fuel shipped from other nuclear facilities owned by AEC or

DOE, or an open-air pit that burned items enriched and coated with sodium, trash, and combustible liquid waste (Wing et al., 2006; Vitkus, 2008; Rucker, 2009; Weitzberg, 2014). In one instance, a series of fires involving radioactive materials occurred in the “hot lab”, potentially releasing them into the environment. The facility had inadequate cleanup plans for dealing with any resulting radioactive contamination (Wing et al., 2006; Vitkus, 2008; Rucker,

2009). In a separate instance, it is speculated that the AEC was aware of high radiation readings during a potential partial meltdown in the nuclear reactor room and that SSFL employees were intentionally venting radioactive materials to the environment for weeks (Wing et al., 2006;

Rucker, 2009; Weitzberg, 2014). SSFL’s accident history is poorly documented because all public technical reports released by AEC lacked detailed information (e.g., frequency, duration,

1 severity, wind-direction, and amount of released radionuclide etc.) and any other reports are withheld by SSFL’s current owner, Boeing (Wing et al., 2006; Rucker, 2009; Weitzberg, 2014).

1.2 Nuclear Reactors and Meltdowns

It is important to differentiate nuclear fusion and fission, where nuclear fusion follows a similar process in our sun, that requires high temperatures (> 4,000,000 C ) and two light atoms’ protons (hydrogen) to fuse into one heavy nucleus (helium) and two positrons, that produces a substantial amount of energy (Moses, 1964; Lavrukhina et al., 1967; Gilmore, 2008).

Technology has not yet advanced enough to produce fusion reactors, which remain in the experimental stages, but are of great interest because of the reactor’s potential for clean energy

(zero emission of CO2 and air pollutants) and minimal waste (Bowen and Gibbons, 1963;

Quittner, 1972; Gilmore, 2008). Modern nuclear reactors focus instead on fission, which releases thermal energy by splitting atoms (i.e., opposite of fusion) of uranium-235 or plutonium-239 unevenly, which releases energy to be harnessed for electricity production. The most common by-product of this nuclear fission is Cs-137 (Moses, 1964; Quittner, 1972; Gilmore, 2008). The fission process involves a large unstable isotope in a fuel rod colliding with high-speed particles, usually neutrons, that cause the isotope to break into smaller atoms (Bowen and Gibbons, 1963;

Gilmore, 2008). In a nuclear power plant, the fission reaction is important because the splitting of atoms releases heat, which boils the water the fuel rod is submerged in, to produce steam that turns the turbines to produce electricity (Bowen and Gibbons, 1963; Gilmore, 2008). When the fuel rod is reaches the end of its liftspan, it can release dangerous heat level from the splitting of

U-235 or Pl-239, and it is typically removed and stored as nuclear waste for several years, to minimize and prevent any possible release of radiation (Gilmore, 2008).

Nuclear incidents are classified as either a partial or total meltdown. Both types are of great concern to the general public, because of its impact on human health. In a partial meltdown, the reactor burns out of control and begins to melts the fuel rods, but the coolant is able to stop the fuel rods from completely melting and restore the reactor to containment operative levels (Gilmore, 2008). The 1979 Three Mile Island incident is a partial nuclear meltdown example, where a mechanical issue prevented water from cooling the radioactive fuel, allowing it to build heat and pressure. The relief valve in the reactor was supposed to close once pressures stabilized, but was temporarily stuck open, letting nuclear fuel melt through its container and trace amounts of radioactive noble gasses escape from the nuclear plant (U.S.

Nuclear Regulatory Commission, 2018). This open valve is important because it allowed fission products to be released into the reactor coolant, then travel to the auxiliary building beyond the containment boundary, and escape to the local environment (U.S. Nuclear Regulatory

Commission, 2018). A total nuclear meltdown is when the nuclear core and containment is damaged beyond repair as the result of a neglected partial meltdown, which causes a massive release of various radionuclides into the environment. Chernobyl is the classic example of a total nuclear meltdown, where the reactor’s core melt breached beyond the casing unit’s threshold, ventured to other parts of the facility, and was not quickly or at all stopped by a coolant

(International Advisory Committee, 1991; Bouville et al., 2002; Aleksakhin et al., 2006; Gupta and Walther, 2017).

The severity of the meltdown will determine the amount of radioactive material released and is dependent on the amount of energy that was produced during nuclear fission. Nuclear

3 meltdowns release both radioactive gases and a particulate matter cloud shaped like a

“mushroom”, until it reaches a stable height in the upper stratosphere that varies by location and meteorology (Bouville et al., 2002; Simon et al., 2006). The radioactive material will then follow a downwind path and scatter in all directions (Bouville et al., 2002; Simon et al., 2006; Kerr et al., 2013). The largest particles will settle closest to the source (local fallout) that varies depending on meteorological conditions, so areas with higher precipitation will receive higher radioactive fallout (Bouville et al., 2002; Simon et al., 2006; Malá et al., 2013). The smaller particles and gas can reside in the stratosphere for months or years before settling, which could travel for thousands of kilometers via wind (global fallout) (Bouville et al., 2002; Simon et al.,

2006; Malá et al., 2013).

Section 2. Background

2.1 Site Selection History for SSFL

In 1949, the Atomic Energy Commission (AEC), a division of North American Aviation evaluated several sites across the United States to conduct nuclear research, and ranked the San

Fernando Valley, California 5th out of 6 potential sites because of the many meteorological, hydrological, and growing population safety concerns (Wing et al., 2006). Despite several safety factors, it is speculated that the AEC (a division of North American Aviation), the predecessor of

Rocketdyne, picked their second-lowest choice for shorter commute times between SSFL and other local universities participating in research (Wing et al., 2006; Rucker, 2009). Thus, the

2268-acre SSFL was built about 30 miles (48 km) northwest of downtown Los Angeles atop the

Simi Hills, about 1.4 miles (2.3 km) from the nearest city, Simi Valley (Figure 1) (Wing et al.,

2006; Vitkus, 2008; Rucker, 2009; Weitzberg, 2014). The SSFL facility is divided into four areas

(Area I – IV), with the nuclear research being conducted at Area IV. Within this section, the

AEC and Rocketdyne operated ten nuclear reactors that powered a local city (Moorpark), experimented with radioisotopes, and prepared the disposal of radioactive materials, while Areas

I-III conducted rocket and missile development and testing (Wing et al., 2006; Vitkus, 2008;

Rucker, 2009; Weitzberg, 2014).

Figure 1: Santa Susana Field Laboratory location (yellow star) on the border of Los Angeles and Ventura counties, Southern California.

2.2 Secret Nuclear Meltdown

AEC’s Sodium Reactor Experiment (SRE) was a pioneering nuclear power plant at

SSFL; a 20-Megawatt test reactor that powered the city of Moorpark (Rucker, 2009). Area IV’s

SRE may have been the most dangerous operation that occurred at SSFL, because the details of an accident on July 12, 1959, remain in a “fog” and are publicly unknown (Wing et al., 2006).

For two weeks, the SRE underwent a partial nuclear meltdown. During the incident, operators of the reactor were unsure how the SRE lost control (power rose uncontrollably), why the coolant channels were unresponsive, and the operators were unaware that the temperature within the reactor was rising to dangerous levels for fuel melt and radionuclide release (Wing et al., 2006;

Weitzberg, 2014). The SRE operators were unaware of what caused the energy surge, so they

6 continued operations (Weitzberg, 2014). During these two weeks, the SRE was repeatedly shut off for various reasons, such as a buildup of radioactivity in the building above the reactor, which required the development of new equipment to examine the reason(s) why the reactor was malfunctioning (Rucker, 2009). Public records provided by the AEC and DOE claim 13 of the 43 fuel rods were partially melted, and some rods had “swelled” from the excessive heat, but no radiation was released beyond the building above the reactor and that the incident posed no radiological threat to the employees or environment (Oldenkamp and Mills, 1991). Conflicting reports from independent studies claim operators were intentionally venting radioactive gasses into the atmosphere two weeks after the initial incident, to release pressure and heat in the reactor (Boice et al., 2006; Wing et al., 2006; Rucker, 2009; Weitzberg, 2014).

2.2.1 Formation of the Santa Susana Field Advisory Panel

Three decades after the 1959 nuclear incident, residents became aware of an internal

DOE study finding widespread radioactive contamination at SSFL (Wing et al., 2006;

Weitzberg, 2014). The local community demanded that legislators form an independent health study, which led to the formation of the SSFL Advisory Panel, composed of experts (e.g., nuclear physicists, hydrologists, nuclear engineers, epidemiologists, geologists, and atmospheric modelers) from various institutions around the country. The UCLA School of Public Health conducted an intensive study (Wing et al., 2006) in 1997, and found unusually high

“radionuclide sensitive” cancer (blood and lymph systems) rates in both SSFL employees and local residents, whose rates were greater than the general Los Angeles county population (Wing et al., 2006; Weitzberg, 2014). Based on one model, 30% of the reactor core was damaged, and

7 the SSFL Advisory Panel estimated that up to 2600 curries of radiocesium may have been released from the reactor during the 1959 event (Lochbaum, 2006; Wing et al., 2006; Weitzberg,

2014). Additionally, the SSFL Advisory Panel report included details on the 1999 UCLA health study that found that SSFL employees’ radiation risks or rates of dying from cancer were up to 8 times higher than atomic-bomb survivors, because of their exposure to chemicals used at rocket test stands (Wing et al., 2006; Weitzberg, 2014). The most common radioisotopes that could have potentially been emitted into the environment are, but not limited to, thorium (228 and

232), uranium (234, 235, 238), plutonium (238, 239, 240, 241), promethium-147, beryllium-10, cadmium-113, americium-241, curium-244, strontium-90, and lastly the most common byproduct of fission, cesium-137 (Bowen and Gibbons, 1963; Fischer, 2006; Wing et al., 2006;

Rucker, 2009). Additionally, the SSFL Advisory Panel lead to the public demands for an ongoing clean-up of the contamination (Rucker, 2009).

2.3 What is Cesium-137?

Unstable (134 and 137) radiocesium isotopes are produced primarily as nuclear fission products from a variety of human activities such as: atmospheric nuclear weapons testing, rocket testing, nuclear reactors, or from the production of radionuclides for research (Quittner, 1972;

Bouville et al., 2002; Fischer, 2006; Simon et al., 2006). Cesium-137 (Cs-137) is one of the most common radionuclides released to the environment by nuclear activities (Quittner, 1972;

Markert, 1993; Fischer, 2006). The decay scheme of Cs-137 (Figure 2) is a simple single beta- emitting radionuclide with 2 components of parent and daughter (Gilmore, 2008). The beta decays in an uneven split, where 7% of the material goes directly to the ground stable state of

Barium-137 and the other 93% goes to an excited nuclear state of Barium-137m (Gilmore,

2008). At the excited state, this isotope releases gamma-ray radiation at 661.7 kiloelectron volts

(keV), then de-excites and drops to the ground state (Gilmore, 2008). Cs-137 has a very high yield during nuclear fission (most common by-product), and a relatively long half-life compared to other radionuclides derived from nuclear fission (~30.17 years). Its high-energy decay pathway makes this radionuclide able to cause severe environmental and health issues (Moses,

1964; Quittner, 1972; Fischer, 2006). Specifically, radiation from Cs-137 mimics potassium as it enters mammals and humans and mutates DNA and vital cells, which can become cancerous

(Simon et al., 2006; Wing et al., 2006; Weitzberg, 2014).

Figure 2: The decay scheme of Cesium-137 follows a simple single beta-emitting radionuclide decay with 2 components of parent and daughter. The beta decays unevenly, where 7% goes directly to the ground stable state of Ba-137 and the other 93% goes to an excited nuclear state of Barium-137 (Stony Brook University – Radioactive Decays Lab).

2.4 Importance of Cesium-137 to Human Health

Radiation from Cesium-137 (Cs-137 or radiocesium) and other radionuclides are typically expressed in two ways: 1. how much radioactive material is present in a sample as

Becquerels or Bq (1 Bq = 1 decay per second), or 2. the absorbed dose (millisieverts or mSv)

(Chesser et al., 2001; Bouville et al., 2002; Fischer, 2006; Simon et al., 2006; Gilmore, 2008).

The largest concern with any nuclear meltdown is the increased cancer risk associated with long- term radiation exposure (Bouville et al., 2002; Boice et al., 2006; Fischer, 2006; Simon et al.,

2006; Kerr et al., 2013; Weitzberg, 2014). Cs-137 is a great concern for human health due to its prevalence as the largest byproduct of nuclear fission and its ability to mimic potassium that travels to humans’ soft tissues (e.g., lymph system, blood, muscles, lungs, and kidneys) causing various cancers (Fischer, 2006; Kerr et al., 2013; Weitzberg, 2014). The most common way people become contaminated and harmed by Cs-137 is through internal irradiation from inhalation, absorption through open wounds, and diets of both plants and animals that were contaminated (Bouville et al., 2002; Simon et al., 2006; Kerr et al., 2013). There is an apparent increase in the risk of cancer (thyroid, lymph, and leukemia) associated with exposure to radionuclides from a nuclear accident, in which women and young children are more susceptible

(Simon et al., 2006; Kerr et al., 2013). For example, there were a recorded 4837 cases of thyroid cancer in the most contaminated regions near the Chernobyl incident (e.g, Belarus, Ukraine, and

4 areas of Russia) in the < 17 years old population (Cardis et al., 2006). After two decades of studies on various age groups exposed to radiation, it is speculated thyroid cancers associated with Chernobyl will continue to increase in all age groups, but the long-term risk or magnitude is still unknown and unquantifiable (Moysich et al., 2002; Cardis et al., 2006).

Atomic bomb survivors and Chernobyl power plant workers and firefighters received high radiation doses of 700 to 13,400 mSv from various radionuclides (International Advisory

Committee, 1991; Bouville et al., 2002; Kerr et al., 2013). To put this into perspective, a single dose of 5000 mSv of radiation is fatal without treatment, 1000 mSv can cause nausea, even doses of 50 – 100 mSv is associated with increased risk of thyroid tumors and cancer in children, and the average background radiation as of 2006 (~2 half-lives later) from atmospheric testing before

SSFL in the United States is about 3 mSv per year (Bouville et al., 2002; Simon et al., 2006;

Sinnott et al., 2010).

2.5 Radiocesium in the Environment from Atmospheric Testing

Nuclear weapons testing in the atmosphere between 1945 and 1963 performed by the

United States, United Kingdom, and Russia was the first largest (global) radionuclide exposure for the environment (Bouville et al., 2002; Simon et al., 2006). In the United States, one hundred reported nuclear tests occurred at the Nevada Test Site (NTS). Nuclear weapons tests at NTS accounted for >90% of all radionuclides released in the United States (Bouville et al., 2002;

Simon et al., 2006; Kerr et al., 2013). Almost all radionuclides that were produced by NTS or other nations were released into the stratosphere, then descended to the surface years later. These radionuclides were predominantly deposited by precipitation and rarely by sedimentation or impaction processes (Bouville et al., 2002; Simon et al., 2006; Kerr et al., 2013).

Models of Cs-137 deposition pattern from all tests conducted at NTS exhibit a decrease in concentration with distance, specifically from NTS to the east following the Westerlies

(Bouville et al., 2002; Simon et al., 2006; Kerr et al., 2013). Additionally, both global and NTS

11 radiocesium fallout and deposition models consistently show the western states, specifically in the more arid regions (i.e., Southern California and Arizona) having the lowest Cs-137 concentrations in soils. The estimated deposition of radiocesium in 1990 (one half-life later) is between 0 – 200 becquerels (Bq) per square meter (m-2), which compared to the Midwestern and

Northeastern United States of ~1300 Bq m-2 (Farber and Hodgdon, 1991; Bouville et al., 2002;

Simon et al., 2006). Any deposition in the western states is the result of dry processes because

NTS strategically planned their testing to avoid precipitation events. All radioactive clouds followed the prevailing winds, which on average move north-eastward from the NTS. Almost all the radionuclides were deposited throughout the central and eastern U.S, varying by local precipitation because Cs-137 has a high water solubility and higher precipitation will lead to higher deposition (Bouville et al., 2002; Simon et al., 2006; Kerr et al., 2013). Thus, modeling studies in 1991 estimate Los Angeles and the greater Southern California region to have the least amount of radiocesium derived from atmospheric nuclear testing (0 -30 Bq m-2 ) in soils compared to the rest of the country (Farber and Hodgdon, 1991; Bouville et al., 2002; Simon et al., 2006). At the time of writing in 2020, the values from the 1991 estimate would be approximately another half-life later or 0 to ~15 Bq m-2.

Figure 3: Map of properties surrounding SSFL’s Areas I-IV - Hambrick et al., 2007

2.6 Previous SSFL Studies

Sage Ranch is located ~1.5 miles or 2.4 kilometers northeast of SSFL, and surface water drainages from the northern portions of Area I (Figure 3) flow through Sage Ranch (Hambrick et al., 2007). A 1992 and 1994 study (approximately 1 half-life after the incident) detected radiocesium levels in four of 118 soil samples between .23 - .34 pCi g-1 or 8.51 – 12.58 Bq kg-1 near Sage Ranch Park in the Brandeis-Bardin watersheds along the SSFL property lines, this was above the 95th percentile of the measured background concentrations of ~.21 pCi g-1 or 7.8 Bq kg-1 (McLaren and Hart, 1993; Hambrick et al., 2007). This 1992 study, unfortunately, did not determine if the Cs-137 levels found in the soil are from the result of any offsite migration of Cs-

137 from SSFL, or just background levels from NTS (Hambrick et al., 2007). Additionally, the

EPA concluded that the samples’ that had radiocesium concentrations higher than the local background levels, but were still below typical levels found throughout the United States pose no risk to human health (Hamilton, 1997; Hambrick et al., 2007; Boeing, 2019). A 1994 soil follow- up study revisited SSFL, Sage Ranch Park, and its watershed, and could not identify any samples nearby to SSFL facilities that had radiocesium concentrations different from the mean of 0.087 pCi g-1 or 3.219 Bg/kq for background soil samples (Hambrick et al., 2007).

Soils are a common environmental sample used to identify radionuclide releases, but they are not suited for use in identifying a nuclear incident many years after it has occurred, such as at

SSFL with the SRE partial meltdown occurring 60 years ago. Radiocesium is highly soluble in water and binds to soil particles, so years of heavy precipitation could cause any radiocesium to percolate to the deeper soil horizons below into the Chatsworth bedrock aquifer 23 -61 meters

(or 75 – 200 feet) below the ground surface, or runoff into the ocean (Hamilton, 1997; Pay, 2006;

Boeing, 2019). Thirty years (for the previous soil studies) is an ample amount of time for natural events (e.g., wind, rain, and fires) to remove radiocesium from the surface soils. Additionally, radiocesium is relatively volatile (>380ºC) and the area is prone to wildfires triggered by local

Santa Ana winds, which may have re-released Cs-137 to the atmosphere and be transported and deposited in the Pacific Ocean via these strong wind events (Koh et al., 1999). Surviving and recently dead trees and small mammals are better environmental recorders of radiocesium over longer timescales than soils because cesium should persist in these organisms longer than in soils.

2.7 Cesium-137 in the Ecosystem

Radiocesium is highly mobile in ecosystems because of its solubility and ease of absorption to both wildlife and vegetation. The mechanism responsible for Cs-137 entering vegetation is through the xylem and phloem vessels (Markert, 1993; Speer, 2010). A tree’s growth is dependent on the xylem vessels that deliver water and nutrients, but can also transport metals or radionuclides through the roots to the stem (Brownridge, 1984; Markert, 1993; Fischer,

2006; Speer, 2010). The phloem system that is also responsible for providing water and nutrients can translocate radionuclides during photosynthesis in the stomata throughout the tree

(Brownridge, 1984; Markert, 1993; Kagawa et al., 2002; Fischer, 2006; Balouet et al., 2007;

Speer, 2010). Additionally, Cs-137 is believed to also be easily absorbed (mimics potassium) in wildlife and possibly “bioaccumulates” or increases in concentration with increases in trophic level (Fischer, 2006; Rucker, 2009; Weitzberg, 2014; Murakami et al., 2014). Thus, it is important to investigate both plants and animals, because they are biological monitors of a nuclear release (Markert, 1993; Antonopoulos-Domis et al., 1996; Chesser et al., 2001;

Aleksakhin et al., 2006; Fischer, 2006).

2.7.1 Cs-137 in Wood

Trees are an ideal proxy for monitoring anthropogenic contamination fluxes, because their immobility provides a reproducible radiocesium signal (Markert, 1993; Barci-Funel et al.,

1995; Kagawa et al., 2002; Speer, 2010). Cycling of radionuclides in trees is not fully understood and is still being studied, because it is difficult to track highly mobile Cs-137 in tree-rings

(Antonopoulos-Domis et al., 1996; Clouvas et al., 2007; Zhiyanski et al., 2010; Rulík et al.,

2014). Radiocesium concentrations in trees are highest in the bark and outer rings, then steadily decline as it approaches the pith (center of the tree) (Brownridge, 1984; Chigira et al., 1988;

Antonopoulos-Domis et al., 1996; Zhiyanski et al., 2010). Studies suggest one pathway for Cs-

137 uptake in trees is from the transportation by the xylem vessels, that transports radiocesium from the roots in the contaminated soils to the stem and leaves (Markert, 1993; Barci-Funel et al., 1995; Papastefanou et al., 1999; Kagawa et al., 2002; Fischer, 2006). Additionally, studies also find that the stomata possibly absorb Cs-137 metabolically during photosynthesis, and then the phloem vessels transport radiocesium throughout the tree (Markert, 1993; Papastefanou et al.,

1999; Kagawa et al., 2002; Fischer, 2006).

Radiation from Cs-137 released into the atmosphere and environment during the

Chernobyl and Fukushima incidents was about 8 x 1016 Bq and 1.5 x 1016 Bq, respectively, and environmental concentrations varied depending on the distance from the source and type of deposition (wet or dry) (International Advisory Committee, 1991; Aleksakhin et al., 2006; Gupta and Walther, 2017). Cs-137 concentrations vary in different tree species, which is believed to be due to differences in tree anatomy, age, and geological conditions (e.g., slope, type of soil, leeward vs. windward side, proximity to water source etc.) (Brownridge, 1984; Soukhova et al.,

2003; Zhiyanski et al., 2010; Ohashi et al., 2014).

Chernobyl studies were able to detect Cs-137 in various gymnosperms and angiosperms and found the most contamination in the trees between the bark at breast height (~1.3 m) and their roots (Antonopoulos-Domis et al., 1996; Chesser et al., 2001; Clouvas et al., 2007;

Zhiyanski et al., 2010). One study from Bulgaria’s Mediterranean region, the Maleshevska

Mountain (1500 km southwest of Chernobyl) mirrors two important site characteristics of SSFL

(Mediterranean climate and oak ecosystems). In this Chernobyl study, the researchers detected

16 radiocesium throughout various parts of the trees one half-life later, the highest concentration of

Cs-137 (18-23 Bq kg-1) were located in the bark and the lowest concentrations (~1 ± 0.2 Bq kg-1

) were located in old branches (Zhiyanski et al., 2010). Additionally, various trees species were measured for Cs-137 shortly after the Fukushima incident in 2011, which found similar results to

Chernobyl studies of the highest concentration in the bark (Kuroda et al., 2013; Ohashi et al.,

2014). Oak trees ~50 km away from the Fukushima epicenter detected high concentrations of radiocesium in their bark of 1076 ± 703 to 1 ± 0 Bq kg-1 (Kuroda et al., 2013; Ohashi et al.,

2014).

2.7.2 Cs-137 in Wildlife

Small mammals have the potential to record radionuclide releases in the ecosystem, because of their small migration range and ability to accumulate Cs-137 (Markert, 1993;

Antonopoulos-Domis et al., 1996; Chesser et al., 2001; Aleksakhin et al., 2006). Accumulation of radiocesium in small mammals’ tissue is from a combination of three different pathways

(Antonopoulos-Domis et al., 1996; Chesser et al., 2001). Small mammals are exposed to radioactive material from (1) their contaminated habitat (e.g., nests) constructed in the soil and leaf litter, (2) inhalation of radionuclide gasses, and (3) their diet of water, vegetation, and insects that are enriched in radiocesium (Markert, 1993; Antonopoulos-Domis et al., 1996;

Chesser et al., 2001; Aleksakhin et al., 2006, 2006; Fischer, 2006; Murakami et al., 2014).

Previous studies from the 2011 Fukushima incident found radiocesium concentrations in leaf litter between 30,000 – 300,000 Bq kg-1, with decreasing concentrations with increasing distance from the nuclear meltdown. Detritivores (e.g., earthworms, ants, flies etc.) and their

17 predators have a high radiocesium uptake (Ryabokon et al., 2005; Murakami et al., 2014). The detrital food chain records higher radiocesium concentrations (~7,000 – 40,000 Bq kg) than grazing food chains (~500 - 8,000 Bq kg), because of their interactions with contaminated plants and soils (Ryabokon et al., 2005; Aleksakhin et al., 2006; Murakami et al., 2014). In this study, preserved specimens, specifically in the families of Cricetidae and Soricidae, that were collected from 1959-1964 and within a 150 km radius of SSFL were analyzed for Cs-137 concentrations.

The purpose of using small mammals was to look for a Cs-137 concentration gradient across the

LA region. The goal of using small mammals in addition to wood samples is to measure two different independent types of environmental samples to aid in determining the spatial extent and potentially affected regions of a nuclear material release from SSFL.

2.8 Dendrochronology and Dendrochemistry

In this study, we collected wood samples from Valley Oak (Quercus lobota), Black Oak

(Quercus kelloggii) and Coast Live Oak (Quercus agrifolia) to measure for Cs-137 concentrations resulting from the potential partial nuclear meltdown at SSFL in 1959.

Dendrochronology and dendrochemistry are important tools in extracting information recorded by trees on tree age and environmental changes (Markert, 1993; Barci-Funel et al., 1995;

Kagawa et al., 2002; Speer, 2010). The main principle of dendrochronology is that both tree-ring structures and wood from temperate forests are considered a high-resolution temporal record, because their annual-ring growth can record important environmental changes (e.g., contamination fluxes or climate change) (Brownridge, 1984; Balouet and Oudijk, 2006; Speer,

2010). Tree rings are formed annually. The width is dependent on the tree’s interaction with the environment (e.g., climate, proximity to water source etc.) and trees’ species (Speer, 2010).

After cross-dating (statistically verifying tree age), one can perform dendrochemical analyses to track and monitor a contaminant from nearby sources (Markert, 1993; Papastefanou et al., 1999; Kagawa et al., 2002; Balouet et al., 2009). Chemicals (anthropogenic and natural) are present in wood samples, because of the trees’ uptake and assimilation (Papastefanou et al.,

1999; Kagawa et al., 2002; Fischer, 2006; Balouet et al., 2009). Trees ability to store radionuclides in their wood makes them a vital component in measuring Cs-137 in the environment decades after a nuclear release (Chigira et al., 1988; Momoshima and Bondietti,

1994; Antonopoulos-Domis et al., 1996; Kagawa et al., 2002; Soukhova et al., 2003). For example, dendrochemical studies in oaks from the Fukushima, Chernobyl, and Hiroshima all had detectable radiocesium amounts, but concentration declined towards the pith (Chigira et al.,

1988; Momoshima and Bondietti, 1994; Barci-Funel et al., 1995; Kagawa et al., 2002; Clouvas et al., 2007; Kuroda et al., 2013; Ohashi et al., 2014).

2.9 Potential Impact of this Project

This project is the first to study the concentration of radionuclides 60 years after a nuclear reactor failure. In addition, this research will fill a fundamental knowledge gap locally and globally. Locally, many nearby residents still have inconclusive results on what occurred in

1959, as the incident was simply “swept under the rug” (Simmon, 2018). Some Los Angeles residents have no knowledge of this nuclear facility, or even worse, that a nuclear incident took place in their backyard, and that clean up from contamination associated to rocket enegine

19 testing (e.g., trichloroethlyene and perchlorate) is ongoing (Rucker, 2009; Simmon and 2018,

2018). Globally, researchers interested in other areas like Chernobyl and Fukushima might be interested in our results, because they could consider it a glimpse into the future for those regions.

Section 3. Methods

3.1 Field Site

SSFL is located in the Santa Susana Mountains, a Transverse Range (runs east-west) in

Southern California, approximately 30 miles (48 km) northwest of downtown Los Angeles

(Smith et al., 1978; Squires et al., 1983; Yerkes and Campbell, 1995). The range is dominated by high, narrow ridges, and deep canyons that are mostly covered in Chaparral shrubland, grasslands, and oak savanna/woodlands (Smith et al., 1978; Squires et al., 1983; Yerkes and

Campbell, 1995). The Santa Susana Mountains are comprised mostly of sandstone from the late

Cretaceous Chatsworth Formation, 70 million years of age, a deep-sea turbidite deposit that consist of primarily sandstone interbedded with shale, siltstone, and conglomerates (Smith et al.,

1978; Squires et al., 1983; Yerkes and Campbell, 1995).

Groundwater is present at SSFL in the alluvium and weathered bedrock a few feet below the ground surface (depending on precipitation received that year) (Hambrick et al., 2007). The aquifer’s extends across most of the SSFL and its depth range typically between 75-200 feet (23

– 61 meters) (Hambrick et al., 2007). Some depths of this aquifer are as shallow as 20 feet or 6 meters (Hambrick et al., 2007).

Climate at the SSFL and the surrounding area falls within the Mediterranean climate, which allows for a range of shrubs and riparian species to thrive, specifically coast live oak (Q. agrifolia), valley oak (Q. lobata), and coastal scrub oak (Q. dumosa) (Mealey and Brodie, 2007).

Diverse vegetation, seasonal streams produced from wet winters, groundwater-fed springs, and

21 deep canyons allow for various species of birds, amphibians, and mammals to permanently reside in the area (Mealey and Brodie, 2007).

3.2 Tree-Ring Processing

Figure 4: Tree-core and wood sample locations and their proximity to SSFL. Sage Ranch Park and Upper Las Virgenes Canyon were used to create the chronology.

3.2.1 Collecting Tree-Cores and Wood

Q. lobata, Q. kelloggii, and Q. agrifolia trees were chosen because of their age longevity, abundance, proximity to SSFL, and potential of wood to store radionuclides (Brownridge, 1984;

Momoshima and Bondietti, 1994; Zhiyanski et al., 2010). Permission for tree-core extracting and collection of wood samples was granted by the Santa Monica Mountains National Recreation

Area for three different sites that surround SSFL (Figure 4). Due to restrictions on destroying living trees, this study collected wood samples from trees that were uprooted or had breakage in the trunks, possibly from weak and shallow roots that were affected by various factors (e.g., wind, rot, depth, and fires etc.), Additionally, we collected wood from oak trees up to 55 kilometers (34 miles) away at Liebre Mountain, to compare wood samples from a locality that should not have been affected by potential Cs-137 emissions from SSFL (Table 1)

Table 1: Wood and tree-core sets location, distance, and direction from SSFL. Q. agrifolia, Q. lobata, and Q. kelloggii were collected from Sage Ranch Park, Upper Las Virgenes Canyon, and Liebre Mountain, respectively.

For each tree, 2-4 tree-cores were collected for cross-dating. A 5 mm (.2 inch) wide increment borer was used at breast height (Speer, 2010). All coring followed a straight path to

23 the center of the tree and perpendicular to the trunk. Once the borer was in the tree, the borer turned in a clockwise direction until the tip reached the center of the tree or felt a sudden ease in resistance, indicating a pocket of rot (Stokes, 1968; Fritts, 1976; Speer, 2010). The core and any fragments extracted from the borer were preserved in a paper straw to protect the core during transportation to the lab. For each tree, we recorded GPS coordinates in a field notebook, and labeled with site characteristics (e.g., SPQL01A or Sage Park = SP, Quercus Lobata = QL, 01 = first tree, A = first core from this tree).

We followed previous studies’ methods of using a chainsaw on or near the stump (which contains the highest radiocesium concentrations) to collect wood (Kagawa et al., 2002; Kuroda et al., 2013; Ohashi et al., 2014). The chainsaw method to gather wood was only permitted on naturally fallen trees, so a limited number of samples per site of sufficient age were attainable and ~1 kg of material was gathered from Sage Ranch Park, Upper Las Virgenes Canyon, and

Liebre Mountain. We collected tree-cores and wood <20 feet (6 m) from each other from trees of the same oak species and of relatively similar ages, in order to have a reproducible radiocesium signal of the area (i.e., increase the confidence of detecting Cs-137 concentrations). Different oak species in the same area may differ in the uptake of Cs-137, because of biological reasons that are outside the scope of this project. The goal of this research is to detect a radiocesium signal from an area that should otherwise have little to no radiation from the Nevada bomb testing.

3.2.2 Mounting and Sanding Cores

Following Speer (2010) procedures, all tree-cores were removed from their paper straws and mounted on a premanufactured wood mount. All mounts were labeled to match with the collected tree-cores’ code, to prevent misidentification of samples. The mounts were filled in with a thin line of Elmer’s water-soluble white glue, then large binder clips were used to hold the cores in place and have both tree-cores and glue naturally dry together (Speer, 2010). The tree- cores were mounted in a cross-sectional view facing up. Cross-sectional mounting allows for an easier examination of the tree-rings after being sanded (Speer, 2010).

Tree-cores needed to be sanded in progressive finer sandpaper to polish the wood in order to count the annual rings (Stokes, 1968; Fritts, 1976; Speer, 2010). The first sandpaper grit (120) was on a belt sander to flatten the tree-core (Speer, 2010). The next finer sandpaper grit (240) on the belt sander begins was to remove any striations caused by the increment bore or previous sanding belt (Speer, 2010). The final sandpaper with the finest grit of 400 (15.3 μm) allowed the surface to be polished enough to view all cellulose cells and rings under a microscope.

3.2.3 Ring-Counting

A technique used by many dendrochronologists is to mark all decades on the wood before cross-dating. Assuming each year corresponded to a tree-ring, we count “backwards in time” from the outermost ring next to the bark to the pith, center of the tree (Yamaguchi, 1991; Balouet et al., 2009; Speer, 2010). To count rings, we utilized a visual cross-dating system, where one dot symbolizes a decade, two dots for mid-century rings, three dots for each century, and 4 dots to

25 indicate a millennium ring (Speer, 2010). After all rings were counted and dotted for a rough estimate of the age of the tree, that value was written on the side of mount.

The tree-core was placed on a velmex stand micrometer, a linear sliding stand that measures in micrometers. The annual ring width measurements were recorded by the

MEASURE J2X program, a multi-platform program that captures and edits annual ring measurements. All ring-width measurements were measured to the nearest .001 mm. This process aided in later determining the level of confidence in the assignment of years to the annual rings.

3.2.4 Cross-dating

We cross-dated the wood from each site to determine that the wood collected from each tree was minimally 60 years old (i.e., confirming tree was alive during the 1959 incident). A total of 19 tree cores were utilized to produce a chronology for cross-dating in this study. Tree-ring width data is critical for cross-dating, a fundamental dendrochronology technique, that matches ring-width and other structural characteristics among the local trees, to allow precise assignment of the year of formation to each tree-ring (Fritts, 1976; Speer, 2010). This process is important in dendrochronology to statistically increase confidence of a tree’s age (Fritts, 1976; Speer, 2010).

Cross-dating is the most important technique in dendrochronology that ensures each individual tree ring is assigned its exact year of formation. In this study, we use Richard Holmes’

1982 statistical evaluation computer program, COFECHA. Dendrochronologists use COFECHA

26 as a segmented time-series correlation technique, to assess the quality or accuracy of the annual- rings’ measurements (Holmes, 1983; Speer, 2010).

One of the benefits of COFECHA is the ability to detect outliers, which could be the result of an error during measuring of the tree-rings or climatic events such as droughts or abnormally high precipitation years (Holmes, 1983; Speer, 2010). Additionally, COFECHA was used because of its ability to increase confidence levels of our cross-dating (Holmes, 1983;

Balouet et al., 2009; Speer, 2010). COFECHA statistically creates a master chronology from all the cores’ ring-width measurements that were entered into the program, which by default, “fits a

32-year cubic smoothing spline to the cores for standardization”, by averaging all the index series (Holmes, 1983; Speer, 2010). All cores’ ring-width measurements are divided into 50-year segments that each have 25 years of overlap, and automatically statistically correlates each segment against the master chronology (Holmes, 1983; Speer, 2010). A benefit of COFECHA is the ability to highlight the time segment(s) that have a poor correlation to the master chronology and any outliers (Holmes, 1983; Speer, 2010).

All errors or outliers identified by COFECHA were remeasured, to re-record and re- confirm the correct tree-ring width value. After all errors were fixed and a master correlation was produced, which encompassed a COFECHA score (i.e., intercorrelation value) from the correlations to the master chronology (Holmes, 1983; Speer, 2010). It is important to note that we maximized the intercorrelation value for each species to be comparable to other oak studies in nearby areas from the International Tree Ring Data Bank (ITRDB, 2020).

With the average tree age being far greater than 60 years, the necessity to determine the true age of all trees was not as important as confirming the tree was present approximately 10 years older

27 than the potential 1959 radionuclide release. Thus, this study focused on the intercorrelation value between 1920 and ~2018, to increase confidence that the tree was alive in 1959. This strategy was selected because radiocesium is highly mobile, travels throughout the trunk and this study is measuring whole wood concentrations rather than individual annual rings (Kagawa et al., 2002; Kuroda et al., 2013; Ohashi et al., 2014). After a chronology was created from the tree- cores, then we were able to anchor or assign years to chainsawed cross-sections with unknown death dates.

3.3 Small Mammal Collections

Figure 5: Small mammals analyzed by an HpGe detector and its proximity to SSFL. Five sets of preserved small mammals (N. crawfordi, P. boyllii, and N. lepida) were collected both < 5 years after 1959 and between ~11 – 60 km from SSFL.

Natural history museum collections contain millions of preserved specimens, and curators began collecting in the early 1900s from various localities. This study focused on the families of Cricetidae and Soricidae (Figure 5), because of their potential to record the highest amount of Cs-137 through their diet of detritovores and vegetation on forests floors (which should contain the largest radiocesium reservoirs), their ability to bioaccumulate radiocesium

(Murakami et al., 2014). Additionally, these were the most abundance specimens collected to be housed in natural history museums shortly after the incident (August 1959 – December 1965)

(Chesser et al., 2001; Ryabokon et al., 2005; Fischer, 2006). Loans of the preserved small mammals were granted by the Los Angeles County Natural History Museum, Santa Barbara

Natural History Museum, and California State University, Long Beach. All samples were accompanied by a GPS location and a detailed biological survey (Table 2). During the preservation process, these mammals’ organs and skeletons were removed and stuffed with cotton, which we expect to not have any radiocesium. We weighed the mammals with their

29 stuffing intact, to maintain our agreement with the local museums of non-destructive analyses and estimated that ~ 75% of the mass measured is associated with cotton.

Table 2: Small mammals sets that were collected for gamma ray analysis by HpGe or NaI detectors. Gamma-ray spectroscopy differ between Sets 1 - 5 and 6, which were analyzed by an HpGe and NaI detector, respectively.

3.4 Background Concentration of Cs-137

As reported in previous studies, Cs-137 background from atmospheric nuclear weapons tests that were detected in soils in Los Angeles one half-life later in 1991 range between 0 – 3.8

Bq kg-1 (Farber and Hodgdon, 1991; McLaren and Hart, 1993; Wallo III et al., 1994; Bouville et al., 2002; Simon et al., 2006). Radiocesium deposition from the fallout of the Chernobyl incident is detectable, but limited to the East coast of the U.S. and not significant in comparison to the deposition from weapons testing fallout (Wallo III et al., 1994; Bouville et al., 2002; Simon et al., 2006; Gupta and Walther, 2017).

After the total nuclear meltdown at Fukushima, measurable radiocesium concentrations were detected in trees and small mammals within a 60 km radius of the powerplant (Kuroda et al., 2013; Ohashi et al., 2014; Murakami et al., 2014). We use this distance as a guideline for how far radiation may have been deposited during the nuclear incident at SSFL. For this study, any gamma-ray radiation from Cesium-137 detected in samples <60 km from SSFL could contain remnants of radiocesium by dry deposition from NTS (Farber and Hodgdon, 1991;

Bouville et al., 2002; Simon et al., 2006) in addition to radiation from the nuclear incident at

SSFL. Samples >60 km from SSFL that are downwind of the site should not have been affected by any incident at SSFL, and so any detectable radiation would be attributed to the NTS (Farber and Hodgdon, 1991; Bouville et al., 2002; Simon et al., 2006), though we expect the radiation from NTS in southern California to be low or possibly even non-detectable. To measure for any background radiocesium levels from the NTS, we measured Cs-137 in both preserved small mammals from Ojai and wood from Liebre Mountains that were ~ 62 km and 55 km away from

SSFL, respectively.

3.5 Measuring and Reconstructing Cesium-137 levels

Sample preparations for small mammals and wood in both a high-purity Germanium or sodium iodide gamma-ray detector are minimal and non-destructive. Small mammals were removed from their original case and placed in a labeled plastic container, to prevent any lead contamination from the detector. Additionally, all wood samples were collected from previously fallen trees and stumps and after being cross-dated were sliced into smaller pieces (12 in. x 12 in. or 30.5 cm x 30.5 cm) to fit in the detector.

Cesium-137 has several pathways in which it can interact with the type of matter in a gamma-ray detector. In one of these pathways, a gamma ray photon in a low energy state hits an outer electron, transfers all of its energy to that electron, and forces it to leave its orbit

(photoelectric effect) that causes a scintillation or a detection (Gilmore, 2008). At high energy, the photon has an elastic collision after ricocheting off an electron, which makes an excited electron (i.e., Compton scattering) (Gilmore, 2008). When gamma-rays hit a radiation detector, they are usually detected, though the detection rate is dependent on the thickness of the detector

(Quittner, 1972; Knoll, 1989; Gilmore, 2008). In this study, we attempted to detect radiocesium concentrations in wood and small mammals in two commonly used gamma-ray spectrometers, using detectors made of High-purity Germanium (HpGe) and Sodium iodide (NaI). The

Canberra Ultra Low-Background HpGe system 60% Stainless Steel Cap with Compton

Suppression was housed in the Nuclear Reactor Facility in the Department of Chemistry at the

University of California, Irvine. The Harshaw 7.6 cm x 7.6 cm Sodium Iodide (NaI) Integral

Line detector was housed in the Department of Physics at California State University,

Northridge. Additionally, The NaI detector was able to display and analyze data by ADMCA, the

Analog and Digital Acquisitions Software (MC8000A/DP4/X123/GAMMA-

RAD/PX4/X123SDD).

Both detectors are commonly used methods in measuring gamma radiation from Cesium-

137. The greatest benefit of a HpGe detector is germanium’s high atomic number and few centimeter thickness allows for a large linear attenuation coefficient, or how easily Ge can absorb a large energy range of gamma-rays (Gilmore, 2008). An HpGe detector can have up to a

30x higher resolution than NaI, because it can distinguish or separate gamma-rays with close energies and allow to identify radionuclides or different decays in the energy spectrum (Knoll,

1989). NaI detectors are also highly popular because iodine’s high atomic number assures the most important process, photoelectric absorption, occurs; which is responsible for its high detection efficiency (Knoll, 1989; Gilmore, 2008). Additionally, NaI detectors have a higher efficiency (3%) than a HpGe (1 - 2%), which is important to convert gamma-rays into signals given by the ratio of detected events per emitted gamma-rays. All detectors are enclosed by pure lead bricks (most widely accepted materials for detector shields), to reduce the number beta particles hitting the detector (i.e., to make a cleaner energy spectrum), and remove or reduce any radiation background in the analysis (Moses, 1964; Knoll, 1989; Alfassi, 1994; Gilmore, 2008).

In this study, two HpGe systems were used that differ in their detection limit between the tree cores and both the small mammals and larger wood samples of 1 pCi/0.037 Bq and 1.5 pCi/0.0555 Bq, respectively.

An HpGe and NaI gamma-ray detector were used for about 15 hours and 17 hours, respectively, to increase resolution/efficiency of recording the entire gamma-ray energy spectrum emitting from the samples. HpGe’s low detection limit allowed to easily convert disintegrations per seconds (DPS) to becquerels, where 1 DPS = 1 Bq. For the NaI detector, we

33 analyzed the gamma-ray energy spectrum by converting the net area or the counts of radiocesium’s gamma ray energy peak (661.7 keV) to Becquerels. We were able to calculate current Bq levels by using equation 1, where 푁푟푎푡푒 is the net area of all gamma-ray emission counts under radiocesium’s peak that excludes any radiation from the background of the room where the detector is located in and is then divided by live time or amount of time on the detector, ⍵ is the solid angle of an object to the detector’s axis,  is the absolute efficiency percentage of the NaI detector, and Tr is the ratio of the photoelectric to the total detection efficiency (Knoll, 1989). To account for errors, we followed equation 2 of multiplying Nrate by

NaI 3% detection error and converted to Bq using equation 1.

1 1 1 퐸푞푢푎푡푖표푛 1 ∶ 푁푟푎푡푒 ∗ ∗ ∗ = 퐵푞 ⍵ ℇ 푇푟

퐸푞푢푎푡푖표푛 2 ∶ 퐸푟푟표푟 = 푁푟푎푡푒 ∗ 3%

Comparing our results to environmental samples from Fukushima and Chernobyl first required using the re-arranged decay rate equation 3, to reconstruct the possible radiocesium amount release in 1959 of two half-lives ago (Knoll, 1989; Gilmore, 2008). In Eq. 3, N0 is the initial radiocesium amount, N is the amount detected from the sample or detection limit for samples with radiocesium below the detection limit, t1/2 is the value of the radiocesium half-life of 30.17 years, and λ is the decay constant of 0.02297 years-1.

푁 퐸푞푢푎푡푖표푛 3 ∶ 푁0 = 푒−휆푡1/2

Studies in nuclear disaster areas report their radiocesium amounts in the ecosystem as Bq kg-1 , which we followed by a simple division of the reconstructed Bq amount from the samples’ mass using equation 4.

퐸푞 3: 푅푒푐표푛푠푡푟푢푐푡푒푑 퐵푞 퐿푒푣푒푙푠 퐸푞푢푎푡푖표푛 4 ∶ = 퐵푞 푘푔−1 푆푎푚푝푙푒푠′ 푀푎푠푠 (푘푖푙표푔푟푎푚푠 표푟 푘푔)

Section 4. Results

4.1 Dendrochronolgy

A total of 20 tree cores from Sage Ranch Park and Upper Las Virgenes Canyon Open

Space Preserves were used in cross-dating (Table 3). Due to radiocesium’s mobility throughout the tree, we prioritized determining whether the tree was alive in 1959 rather than determining the actual age. We first found at Sage Ranch Park that the Q. agrifolia rings were a mean age of

93.7 years and median age of 102 years.We focused on a shorter chronology for cross-dating in

COFECHA, because the priority goal of visual cross-dating was to confirm our trees contained at least 60 individual tree-rings (i.e., alive in 1959), rather than determine an exact age.

Table 3: Tree-cores used in cross-dating and their earliest ring date. Sage Park Quercus agrifolia have a mean age of 93.7 years and Upper Virgenes Canyon Quercus lobata have mean age of 145.7 years. Tree-cores highlighted in gray were utilized for gamma-ray spectroscopy.

At Sage Ranch Park, we focused on a chronology between 1950 and 2018, which produced a relatively high intercorrelation value of 0.647 (p <0.05) from 6 tree cores. Cross- dating for Upper Las Virgenes Canyon was also shortened, so the Q. lobata mean and median ages were higher, at 145.2 years and 117 years, respectively. At Upper Las Virgenes Canyon, given the higher mean age of the Quercus lobata, we expanded the chronology series to between

1920 and ~2018, which produced an intercorrelation value of .518 from 14 tree-cores. One large wood trunk sample was collected from a stump of a fallen tree at Sage Ranch Park (sample number: SPQA09). Determining the age of sample SPQA09 required cross-dating the trunk sample to the master chronology of all tree-cores’ ring-width from both Sage Ranch and Upper

Las Virgenes, which produced a series intercorrelation value of 0.465 (p < 0.05). Table 4 highlights the intercorrelation values between each site and the master chronology associated with assigning the years to annual rings for sample SPQA09.

Table 4: COFECHA results that display the series intercorrelation values of each site and the Master Chronology used to assign the years to the tree-rings for SPQA09. Average mean sensitivity is the measure of the year-to-year variability in the master chronology. Mean length of series is the average age of the trees in the chronology.

4.2 HpGe Radiation Detector Analyses

2 sets of wood samples and 5 sets of mammal specimens were assayed for 54,000 seconds (15 hours) on a HpGe radiation detector, to measure any Cs-137 activity above the background level, respectively. A Marinelli standard source was used for calibration with an efficiency of 1.90%. The lowest detection limit (95%) computation for the calibration source calculated to about 0.0555 Bq (1.5 picocuries) and 0.037 Bq (1 picocurie) for the small mammals and wood samples, respectively.

4.2.1 Tree-Cores and Wood

Preliminary tree-cores were selected by having >70 distinct annual rings to confirm age was > 60 years old and were sliced into wood fragments that contained 10 rings before and after

1959, to account for radiocesium’s mobility (Chigira et al., 1988; Soukhova et al., 2003). Table 5 presents both the two sets of wood fragments (~1 gram each) from tree-cores and 3 wood sets that were analyzed by the HpGe detector. Other studies from Fukishima and Chernobyl analyzed kilograms of wood (cross-sections) than a gram of individual rings, so we revisited the study areas to collect larger masses of wood from fallen trees and measured for radiocesium in both a sodium iodide (NaI) spectrometer and a second assay with a HpGe detector (Zhiyanski et al.,

2010; Ohashi et al., 2014). The cross-sections larger mass, produced a higher gamma-ray energy resolution than tree-cores in the HpGe detector. Set 1 tree-cores (SPQA 03A, SPQA04B, and

SPQA07A) were assayed together and was below the detection limit, which 2020 estimate maximum levels are < 23.09 Bq kg-1 or 0.62 pCi g-1. Recalculating initial estimate values of Set

1 tree-cores using EQ 3, produced a maximum 1959 value of 92.68 Bq kg-1 or 2.50 pCi g-1. Set

1’s cross section (SPQA09A) was divided into three pieces, to fit in the detector and were all below the detection limit, so 2020 estimated maximum radiocesium is < 0.32 Bq kg-1 or 0.0086 pCi g-1. The 1959 estimated maximum levels from Set 1’s cross-section are < 1.3 Bq kg-1 or

0.035 pCi g-1. Set 2 tree-cores (VCQL01A, VCQL03D, and VCQL04B) were assayed together and was below detection limit, which 2020 estimated maximum radiocesium level is < 29 Bq kg-

1 or 0.79 pCi g-1. Tree-cores from Set 2 calculate to < 118 Bq kg-1 or 3 pCi g-1 for their 1959 estimate maximum values. Set 2’s cross-section (VCQL13) was below detection limit for 2020 levels that estimated to be < 0.040 Bq kg-1 or 0.0011 pCi g-1, which calculate the 1959 estimated maximum levels of < 0.16 Bq kg-1 or 0.0044 pCi g-1. Controlled site or Set 3 (LMGQ01) was below detection limit for 2020 levels of < 0.049 Bq kg-1 or 0.0013 pCi g-1, which calculates the

1959 estimated maxmimum level of < 0.20 Bq kg-1 or 0.0052 pCi g-1. Set 3 results were similar to both Sage Ranch Park (Set 1) and Upper Las Virgenes Canyon (Set 2).

Table 5: 2 sets of tree-cores and 3 sets of wood samples were assayed by an HpGe detector and all samples contained Cs-137 concentrations that were below the detection limit.

4.2.2 Small Mammals

Table 6: Cs-137 results from the HpGe detector for 5 small mammals sets. Brandeis Ranch, with largest mass, produced the lowest potential value < 4.16 Bq kg-1.

All specimens of each locality were analyzed together, to increase both mass and precision of detecting radiocesium. A total of 5 sets of samples were analyzed (Table 6) using the lowest detection limit for the instrument (1.5 pCi/0.0555 Bq): Brandeis Ranch in Simi Valley

(13.1 km/8.1 miles northeast of SSFL), Brea Canyon in Simi Valley (11.3 km/7 miles northwest of SSFL), Ventura River in Ojai (61.6 km/38.3 miles northwest of SSFL), Highway 126 in

Fillmore (23.3 km/14.5 miles northwest of SSFL), and Little Sycamore Canyon in Malibu (24.5 km /15.2 miles southwest of SSFL).

All sets were non-detectable for Cs-137, so Table 7 compares the possible values at the time of this study (two half-lives later) and estimated 1959 levels using EQ 3. Even though

40 current Cs-137 levels in samples from Set 1 Brandeis Ranch are non-detectable the 2020 estimated maximum levels is < 4.16 Bq kg-1 or 0.11 pCi g-1. Set 2 Brea Canyon’s small mass reduced the signal or ability to measure radiocesium, so 2020 estimated maximum levels is <

330.26 Bq kg-1 or 8.93 pCi g-1. Controlled site or Set 3 Ventura River’s 2020 estimated maximum radiocesium level is <11.95 Bq kg-1 or 0.323 pCi g-1. Set 4 Filmore/Hwy 126’s 2020 estimated maximum radioceisum level is 5.37 Bq kg-1 or 0.145 pCi g-1. Set 5 Sycamore

Canyon’s 2020 estimated maximum radiocesium level is 8.02 Bq kg-1 or 0.217 pCi g-1.

Table 7: 2020 and 1959 estimated maximum radiocesium levels in both the wood and small mammals analyzed by the HpGe detector. All samples were below detection limits, so 2020 estimates were derived by using detector limits in equation 4. We then used the 2020 estimates in equation 3 to calculate the range of 1959 levels.

4. 3 NaI Radiation Detector Analyses

Table 8: NaI detector results of Sage Ranch Park wood sample that was divided into 3 parts, to fit in the detector, and Camarillo State Hospital’s small mammals. All samples were indistingushable the background and less than trinitite.

Figure 6: Energy spectrum of trinitite (red circle) and wood (black circle) for ~ 16 hours. The purple highlights radiocesium's peak in trinitite and wood

A NaI gamma-ray detector was used for the additional wood and small mammal samples

(Table 8), because of this instrument’s ability to have a higher efficiency (3%) than a HpGe (1-

2%). Trinitite (atomsite), or Alamogordo glass, a residue left on the desert floor at NTS after the plutonium-based Trinity Nuclear bomb test on July 16, 1945 was used for calibration, to aid in comparing radiocesium gamma-ray energy 66.17 MeV (661.7 keV) of the wood sample (Figure

6). The background analysis measured Cs-137 levels in the room of Live Oak Hall at Caliornia

State University, Northridge using the NaI detector with no samples loaded, which detects the radioactivity naturally from the Earth, cosmogenic, and NTS (Figure 7) (Knoll, 1989).

Additional small mammal specimens from Camarillo State Hospital were assayed for radiocesium (Figure 8). The background Cs-137 radioactivity level was indistinguishable from both the wood and small mammal samples and less than trinitite, the NTS source.

Figure 7: Energy spectrum of background radiation or the gamma-ray energy in the room where the detector is placed without any samples for 1 hour. The purple highlights radiocesium's peak in the background

Figure 8: Energy spectrum of small mammal from Camarillo State Hospital that was assayed for ~3 hours. The purple highlights radiocesium's peak in the small mammals.

Figure 9: Comparing Cesium-137 peaks in Trinitite, background, and samples, which displays that the wood and small mammals were indistinguishable from background and lower than trinitite.

ADMCA software allowed the isolation radiocesium’s MeV from the energy spectrum, which both the wood and small mammals showed the concentration of Cs-137 to be below that of trinitite. Additionally, figure 9 highlights the peak of counts of radiocesium decay in the wood and small mammals compared to the background and the standard trinitite. Both wood and small

44 mammal samples are indistinguishable from the background and less than trinitite. Radiocesium in our Trinitite standard was calculated to be 2.08 Bq or 56.08 pCi and background is estimated at about 0.25 Bq or 6.8 pCi. Then, using the ADMCA software we were able to estimate 2020 radiocesium concentrations in the wood to be < 0.46 Bq kg-1 or 0.016 pCi g-1 and calculating

1959 estimates to 1.9 Bq kg-1 or 0.062 pCi g-1.

Section 5. Discussion

5.1 SSFL Analysis

The hypothesis of this project was that if radiation was released at SSFL in 1959, then there should be a measurable concentrations of Cs-137 in the local ecosystem from small mammals and wood that were present during the incident. The activity of Cs-137 in either spectrometer reports no values above background levels of 0.087 pCi g-1 or 3.22 Bq kg-1 noted from the previous soil analyses, which does not immediately rule out that no radionuclides were released into the environment from SSFL (Hambrick et al., 2007). The most probable reason why we are not able to detect radiocesium specifically in the wood or small mammals is that if cesium-137 was released during the nuclear incident in 1959, it may have been contained, which has been claimed by Boeing.

5. 2 Research Assumptions and Limitations

5.2.1 Assumptions in this Study

Several assumptions were made throughout this project, which were taken into consideration. First, radiocesium in wood samples from Fuksuhima and Chernobyl studies were different oak species or confier trees, which in this study we assumed our oak species would behave similarly in its ability to store radionuclides (Kagawa et al., 2002; Zhiyanski et al., 2010;

Ohashi et al., 2014). Our small mammals in this study, we assumed that our species behaved similar to other small mammals from Fukushima and Chernobyl studies and that they should be able to retain radiocesium in their skin, skeleton, and hair too.

5.2.2 Potential Analyses of Strontium-90

Strontium-90 is another product of nuclear fission of great concern to human health with a relatively long half-life of 28.8 years that mimics calcium and can be stored in bones, causing bone cancers and leukemia. This isotope can also potentially determine the precise year of a radionuclide release in the annual-rings of trees because of their immobility (Chigira et al., 1988;

Momoshima and Bondietti, 1994; Kagawa et al., 2002). Sr-90 is not as likely to be released from a nuclear reactor accident than radiocesium because it is less volatile and has a smaller yield from nuclear fuel (< 2 – 7 %) (Chesser et al., 2001; Kagawa et al., 2002; Simon et al., 2006).

Detecting Sr-90 in this study would be more challenging than radiocesium, because of the destructive process for beta scintillators of needing to grind-up irreplaceable small mammal bones. Thus, in this study, we wanted to first focus on detecting radiocesium from its higher yield, non-destructive gamma-ray detector technique, and importance to human health, then re- examine samples for Sr-90 or other radionuclides if a signal was found. Future studies could revisit the wood samples, to correlate a Sr-90 spike in the annual-rings to year(s) of fallout.

However, given the very low concentrations of Cs-137, it is unlikely that there will be detectable levels of Sr-90 in the samples. This analysis however could provide further support that NTS is the main source of “natural” radiation background in the local ecosystem.

5.2.3 Lack of Samples Closer to SSFL

The small mammals collected shortly after the incident were gathered from public property, but SSFL at that time was considered private property. Thus, the only available small mammals at the local museums were as close as 11 km or 7 miles from SSFL. Furthermore, 60

47 years after the event it is now impossible to collect any specimens from the site that would have been present during the incident. Wood samples were also limited to trees that already fell or began deteriorating by natural causes. There are gaps in our sample coverage (e.g. west of Area

IV) that can be revisited to seek other trees that may be old enough and able to store radiocesium if a radionuclide release did occur.

5.3 Small Mammals Surrounding SSFL vs. Chernobyl and Fukushima

Figure 10. Comparison of radiocesium in small mammals between the Murakami et al. (2014) Fukushima study of local herbivores and detritivores ~2 years after and 53 km or 33 miles from the disaster, and the Beresford et al., (2008) Chernobyl study of C. glareolus (Bank vole) 19 years after and ~ 8.5 km or 5.3 miles from the disaster, and 6 sites surrounding the Santa Susana Field Lab that were collected <5 years after 1959 at distances of ~11 - 62 km or 7 – 39 miles. All small mammals were below the detection limit. The calculated 2020 maximum estimate levels are based on mass and the 1.5pCi detection limit (gray). The predicted 1959 levels (light blue) are substantially lower than levels observed in animals surrounding Chernobyl or Fukushima shotly after the incident (dark blue) and their estimate values 60 years later (black).

The small mammals analyzed by an HpGe or NaI detector (Figure 10) were below the detection limit or similar to the background, respectively. We utilized the detector limit for

Equation 4, to determine the maximum possible concentration of cesium-137 for each sample.

To put this into perspective, Chernobyl released ~8.0 x 1016 Bq of radiocesium into the

48 environment, where small mammals in the Cricetidae family, such as Clethrionomys glareolus

(Bank vole) from the incident contained up to 22.6 x 105 Bq kg-1 19 years after the incidence, or

~ 35 x 105 Bq kg-1 for a 1986 original amount estimate (International Advisory Committee,

1991; FASSET, 2001; Beresford et al., 2008). Fukushima is estimated to only have released ~1.5 x 1016 Bq of radiocesium, and Murakami et al. (2014) were able to detect up to 700 and 50,000

Bq kq-1 in herbivores and detritivores < 2 years after the incidence, respectively (Malá et al.,

2013; Murakami et al., 2014). The maximum 2020 Cs-137 estimate, as well as the reconstructed maximum 1959 Cs-137 level, is far lower than the animals analyzed near Chernobyl and

Fukushima after their incidents (Figure 8).

5.4. Wood Near SSFL vs. Chernobyl and Fukushima

Wood samples had a larger mass, making it possible to have a higher resolution of Cs-

137 than the small mammals or tree-cores. For the wood samples, radiocesium was still below the detection limit and indistingushiable from the background. We utilized the detector limit to estimate maximum value for concentrations of Cs-137 in the wood in 2020, as well as at the time of the potential event in 1959 (Figure 11). Comparably, Fukushima’s oak wood < 20 km away contained Cs-137 concentrations 1.5 years after the incident to be between 1 and 99 Bq kg-1 and we estimate that after 60 years between 0.25 and 25 Bq kg-1 should be detected in the local ecosystem. A Chernobyl study 1200 km away in South Bulgaria shortly after the incident had radiocesium concentrations ~2 Bq kg-1, but this an approximation from the initial detection of 1.2

Bq kg-1 23 years after the event. We can estimate that 60 years after their incident there would be

~ 0.5 Bq kg-1 in the wood (Zhiyanski et al., 2010). A closer Chernobyl study 200 km away in a

Scots pine woodland close to Vetka, Belarus 16 years after the event had radiocesium concentration up to 5385 Bq kg-1 and estimate that 60 years later would retain as much as 1357

Bq kg-1 (Goor and Thiry, 2004).The Sage Ranch Park wood sample was located 2 km away and our cross-dating results support that the tree was alive throughout SSFL entire operation history.

All wood samples (excluding tree-cores because of their small mass) 2020 estimates are lower than both Chernobyl and Fukushima concentrations at the time of their disaster and the 60-year later estimate. If there was a radionuclide release, then there should have been detectable amounts above our 2020 estimates. The HpGe detector is able to reduce the amount of background radiation on the detector, but all samples were still below the detection limit. The radiocesium concentrations in 2020 are indistinguishable from the cosmic and NTS background, which coud support the conclusion that no radionuclide release beyond the facility may have occurred.

Figure 11: Comparison of radiocesium in wood between the Ohasi et al. (2014) Fukushima study 1.5 years after and <20 km or 12.4 miles from the disaster, the Zhiyanski et al., (2010) Chernobyl study ~23 years after and 1200 km or 750 miles from the disaster, Goor and Thiny (2004) Chernobyl study 16 years after and 200 km or 124 miles away, and this study of Sage Ranch Park, Liebre Mountain, and Upper Las Virgenes Canyon of ~60 years after and <55 km or 34 miles from the Santa Susana Field Lab. All wood samples were below the detection limit, and 2020

50 maximum estimate Cs-137 levels (light blue), as well as 1959 maximum Cs-137 levels, are smaller than levels from Chernobyl or Fukushima disasters (dark blue).

5.5 The Possibility of an Extensive Nuclear Release at SSFL

Commercial nuclear reactors’ thermal power for electricity generation typically have levels of near 3,000 MW (Bowen and Gibbons, 1963; International Advisory Committee, 1991;

Rucker, 2009). Three Mile Island, Chernobyl, and Fukushima produced 1725 MW, 3,200 MW and 4,700 MW of thermal power, respectively (International Advisory Committee, 1991;

Yasunari et al., 2011; U.S. Nuclear Regulatory Commission, 2018). The reactors operated at

SSFL, specifically at Area IV in comparison had lower power levels: six reactors produced <0.1

MW (100 kW), three between .6 – 1 MW (600 kW – 1000 kW, and the highest from the SRE of

20 MW (Vitkus, 2008; Rucker, 2009; Weitzberg, 2014). A severe or extensive radionuclide release from a relatively small reactor at SSFL in comparison to a commerical nuclear reactor seems unlikely (Oldenkamp and Mills, 1991).

If radiocesium concentrations were detectable in any of our samples and significantly higher than the estimated Cs-137 in southern California soils (2020 levels would be ~ 0.044 pCi g-1 or 1.62 Bq kg-1), then it would certainly have been possible for Cesium-137 to have been released from the SRE’s partial meltdown and not limited to within the reactor coolant system

(Hamilton, 1997; Boeing, 2006; Cochran, 2009). Previous research measuring radiocesium in soils surrounding Area IV’s facilities (the Ground Prototype Test Facility Building 059 , the “hot lab”, and the soil surrounding the sodium disposal facility that handled radionuclides and waste) one half-life later, or ~50% original material remaining, calculated a low mean concentration of

0.02 pCi g-1 (.74 Bq kg-1 ) (Oldenkamp and Mills, 1991; McLaren and Hart, 1993; Rucker,

2009). Boeing claims the remaining (60 curies or 2.2 x 103 GBq of various unspecified radionuclides) radioactivity at SSFL are confined within Area IV. They estimate that at one-half life only one millicurie or 0.037 GBq of Cs-137 may be present in the sodium disposal facility

(Oldenkamp and Mills, 1991). A soil study < 7 km East of SSFL at the former Boeing Inc.,

Employees’ Recreational and Fitness Center in Canoga Park at one-half life after the event detected on average 0.087 pCi g-1 or 3.219 Bq kg-1, and concluded that this radiaiton is background associated with NTS (McLaren and Hart, 1993; Hamilton, 1997). Rocketdyne’s claim that they contained most of their fission products (5,000 – 10,000 curries of various radionuclides) in their primary sodium cooling system after the partial meltdown could be accurate (Vitkus, 2008; Rucker, 2009). What about the estimates from the SSFL Advisory Panel stating that up to 30% of fission products were potentially released into the environment?

According to independent studies the SSFL Advisory Panel failed to acknowledge that the SRE fuel was immersed in a 2 meter or 6 foot closed pool of a sodium coolant capable of trapping volatiles (iodine-131, strontium-90, and cesium-137) and estimated the amount of released radionuclides by the SSFL Advisory Panel through an assumption that the fission products in some way migrated up the fuel tank and was not stopped by the coolant (Boeing, 2006;

Hambrick et al., 2007). This coolant may explain why there could have been a partial nuclear meltdown without a large release of radioactive material and that the levels of radiocesium are too low to detect beyond natural background from NTS or pose a risk to the health of the public

(Oldenkamp and Mills, 1991; Hamilton, 1997; Hambrick et al., 2007; Cochran, 2009). If the amount of Cs-137 released was low, then this may also explain why radiocesium in our wood and small mammals were below the detection limit.

More recently, a $42 million survey of >38,000 soil samples in 2007, and decades of radiological monitoring, the U.S. Environmental Protection Agency (EPA) and California

Department of Toxic Substances Control (DTSC) did not find any evidence of offsite radiation

(e.g., Chatsworth Reservoir, Brandeis-Bardin Institute, North of the Santa Monica Conservancy,

Rocketdyne Recreation Center in West Hills, and various private homes surrounding SSFL) exceeding their acceptable risk levels of .2 pCi g-1 or 7.4 Bq kg-1 (McLaren and Hart, 1993;

U.S. Environmental Protection Agency, 2002; Liu et al., 2005; Boeing, 2019). Our results of wood and small mammal samples that, unlike soils, should not have been affected by the removal of water-soluble radiocesium and were still below the detection limit from an Ultra-Low background HpGe system. Our samples being below detection limit may provide support that radionuclide contamination in the local ecosystem now, and also decades ago seems unlikely

(Oldenkamp and Mills, 1991; McLaren and Hart, 1993; Hamilton, 1997).

Cesium-137 could potentially be re-released from the wood to the environment from rain washing it out of bark or a wildfire but this does not explain why the small mammals also had

Cs-137 concentrations that were non-detectable. Small mammals sorrounding SSFL had similar distances of < 60 km from the site of radionuclide release, as well as similar diets to other omnivores in the Chernobyl and Fukushima studies. Additionally, the preserved specimens from the local museums were shielded away from any factors that could have reduced or removed radiocesium from their hair, skin, and skeleton. Thus, these samples support the assertion that if any released nuclear material from SSFL entered the ecosystem was negligible amounts in comparison to background radiation from NTS.

5.5.1. Potential Explanation for a Spike in Cancer Rates

Thyroid, bladder, and lymph cancer spikes in the surrounding communities (e.g., Simi

Valley, Chatsworth, or Canoga Park etc.) of SSFL may not be directly linked to Cs-137, but rather other contaminants (Trichloroethylene or TCE, perchlorate etc.) that were recorded to have been released into the environment. For example, perchlorate, a component of rocket fuel with significant concentrations in the soil and groundwater at SSFL may play a role in the spike of thyroid cancer. SSFL has received dozens of violations for perchlorate (the only known user in the area), that contaminated numerous pathways for surface runoff to enter the surrounding communities (Wing et al., 2006; Boeing, 2019).

Section 6: Conclusion

We hypothesized that if radiation was released at SSFL in 1959, then there should be measurable concentrations of Cs-137 in the local ecosystem from small mammals and wood present during the incident that would be detectable 60 years after the event. Our results point to the conclusion that no radiocesium was released beyond the SSFL site. Other studies from the

Chernobyl and Fukushima areas that used small mammals and wood as biomarkers for a radionuclide release provided useful insight, such as radiocesium virtually contained in the Santa

Susana Field Laboratory facility and no evidence supporting radiation in the local ecosystem.

There is no refuting that a partial nuclear meltdown occurred in 1959, but all our small mammal and wood samples have Cs-137 concentrations below the detection limit, and using an Ultra-Low

Background HpGe system gives insight on the magnitude of this incident. The null results imply that no radionuclides were released beyond Area IV’s facilities and were essentially contained within the coolant system as claimed by Boeing. Cancer spikes in the local residential communities surrounding SSFL may have been attributed to other contaminants associated with chemicals for engine testing (e.g., TCE and perchlorate) that Rocketdyne was cited for and

Boeing has actively been cleaning up. Future studies could analyze Sr-90 in the local annual-tree rings, to determine an exact year of a radionuclide release and if it correlates to 1959 or NTS.

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