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Technical Report
TR-13-26
Chlorine cycling and fates of 36Cl in
terrestrial environments
David Bastviken, Teresia Svensson, Per Sandén, Henrik Kylin
Department of Thematic Studies – Water and Environmental
Studies, Linköping University
December 2013
Svensk Kärnbränslehantering AB
Swedish Nuclear Fuel
and Waste Management Co
Box 250, SE-101 24 Stockholm Phone +46 8 459 84 00
ISSN 1404-0344
Tänd ett lager:
SKB TR-13-26
P, R eller TR.
ID 1414778
Chlorine cycling and fates of 36Cl in
terrestrial environments
David Bastviken, Teresia Svensson, Per Sandén, Henrik Kylin
Department of Thematic Studies – Water and Environmental
Studies, Linköping University
December 2013
This report concerns a study which was conducted for SKB. The conclusions
and viewpoints presented in the report are those of the authors. SKB may draw
modified conclusions, based on additional literature sources and/or expert opinions.
A pdf version of this document can be downloaded from www.skb.se.
Abstract
Chlorine-36 (36Cl), a radioisotope of chlorine (Cl) with a half-life of 301,000 years, is present in some
types of nuclear waste and is disposed in repositories for radioactive waste. As the release of 36Cl
from such repositories to the near surface environment has to be taken into account it is of interest to
predict possible fates of 36Cl under various conditions as a part of the safety assessments of repositories
for radioactive waste. This report aims to summarise the state of the art knowledge on Cl cycling in
terrestrial environments. The view on Cl cycling in terrestrial environments is changing due to recent
research and it is clear that the chloride ion (Cl–) is more reactive than previously believed. We group
the major findings in three categories below according to the amount of data in support of the findings.
From the result presented in this report it is evident that:
• There is an ubiquitous and extensive natural chlorination of organic matter in terrestrial ecosystems.
• The abundance of naturally formed chlorinated organic compounds (Clorg) frequently exceeds the
abundance of Cl–, particularly in soils. Thereby Clorg in many cases dominates the total Cl pool.
• This has important implications for Cl transport. When reaching surface soils Cl– will not be
a suitable tracer of water and will instead enter other Cl pools (Clorg and biomass) that prolong
residence times in the system.
•Cl– dominates import and export from terrestrial ecosystems while Clorg and biomass Cl can
dominate the standing stock Cl within terrestrial ecosystems.
• Both Cl and Clorg pools have to be considered separately in future monitoring programs addressing
Cl cycling.
Further, there are also indications (in need of confirmation by additional studies) that:
• There is a rapid and large uptake of Cl– by organisms and an accumulation in green plant parts.
A surprisingly large proportion of total catchment Cl (up to 60%) can be found in the terrestrial
biomass.
• Emissions of total volatile organohalogens could be a significant export pathway of Cl from
the systems.
• Some of the Clorg may be very persistent and resist degradation better than average organic matter.
This may lead to selective preservation of some Clorg (with associated low bioavailability).
• There is a production of Clorg in tissues of e.g. plants and animals and Cl can accumulate as
chlorinated fatty acids in organisms.
Most other nevertheless important aspects are largely unknown due to lack of data. Key
unknowns include:
• The development over time of major Cl pools and fluxes. As long as such data is lacking
we cannot assess net changes over time.
• How the precesses behind chlorination, dechlorination and transport patterns in terrestrial
systems are regulated and affected by environmental factors.
• The ecological roles of the chlorine cycling in general.
• The ecological role of the microbial chlorination in particular.
• The chlorine cycling in aquatic environments – including Cl– and Clorg pools in sediment and
water, are largely missing.
Given the limited present information available, and particularly the lack of data with a temporal
dimension and the lack of process understanding, predictive models are challenging.
We also summarize currently available methods to study Cl in the environment.
SKB TR-13-263
Sammanfattning
Klor-36 (36Cl), en radioisotop med en halveringstid på 301 000 år, förekommer i vissa typer av radioaktivt avfall. För att kunna förutse vad som händer om 36Cl når markytan är det viktigt att veta hur
klor kan omvandlas och transporteras i olika ekosystem. Denna rapport syftar till att sammanfatta
kunskapsläget om klor i naturmiljöer med fokus på landmiljöer.
Synen på klor i naturen är under omfattande förändring till följd av de senaste decenniernas forskning.
Det står nu helt klart att klorid (Cl–) som tidigare betraktats som icke-reaktiv och totalt dominerande,
istället är i hög grad reaktiv och inte alltid utgör den dominerande klorformen.
Utifrån de studier som presenteras i rapporten är det tydligt att:
• Det sker en omfattande naturlig klorering av organiskt material i många miljöer och inte minst i
ytliga marklager.
• Mängden organiskt bunden klor (Clorg) är i många miljöer betydligt högre än mängden Cl–.
Därmed dominerar Clorg ofta det totala klorförrådet i exempelvis mark.
• Detta har stor inverkan på transporten av klor eftersom Clorg till stor del finns i partikulärt organiskt
material medan Cl– är mycket vattenlösligt. Cl– som når ytliga marklager är t ex inte lämpligt som
spårämne för markvattenflöden såsom tidigare antagits. Cl– kommer till stor del att bindas in till
Clorg -förrådet och därmed förlänga uppehållstiden i ekosystemen.
•Cl– dominerar både importen och exporten från terrestra ekosystem medan Clorg kan dominera
stationära klorförråd i systemen.
• Framtida mätningar med syfte att klargöra kloromsättning och klorflöden behöver beakta Cl– och
Clorg separat.
Därtill finns ett antal troligen viktiga indikationer som skulle behöva bekräftas av ytterligare studier.
Dessa inkluderar att:
• Det sker ett snabbt och omfattande upptag av Cl– av organismer och klor tycks ackumuleras i
grön växtbiomassa. En stor andel av den totala klormängden i avrinningsområden (upp till 60% i
en studie) har påträffats i den terrester biomassa.
• Avgång av flyktiga klorerade kolväten kan vara en stor okänd exportväg för klor från ekosystem.
• En del Clorg verkar vara betydligt mer motståndskraftigt mot nedbrytning än det genomsnittliga
organiska materialet. Detta kan leda till att Clorg bevaras selektivt i mark och därmed också
mindre tillgängligt för mikroorganismer.
• Det sker en klorering av organiskt material i levande biomassa och klor kan ansamlas som
klorerade fettsyror i organismer.
Övriga aspekter på klor i naturen är till stora delar okända. Centrala okända aspekter inkluderar:
• Hur klor-förråden utvecklas över tid. Detta är centralt för att förstå förändringar över tid och
reglering i förråden.
• Reglering av klorerings-, deklorerings- och transportprocesser, samt hur dessa påverkas av olika
miljövariabler och miljöförhållanden.
• Den ekologiska förklaringen till varför så många organismer utför klorering av organiskt material.
• Omsättning av klor i akvatiska system. Här saknas separata data gällande Cl– och Clorg både i
sediment och i vattenfasen.
Rapporten fokuserar framför allt på terrestra aspekter av klorcykeln och innehåller också information
om vanliga metoder för mätning av olika klorföreningar.
4
SKB TR-13-26
Contents
1Introduction7
2
2.1
Fundamental chemical aspects of chlorine9
9
Cl isotopes and sources of 36Cl 3
Major Cl reservoirs and large scale cycling11
4
4.1
4.4
4.5
Chlorine in terrestrial ecosystems13
Input of Cl to terrestrial ecosystems
13
4.1.1Deposition
13
4.1.2Weathering
14
4.1.3 Input from irrigation, fertilization and road de-icing
14
Gaseous efflux from terrestrial systems
15
Terrestrial reservoirs of chlorine
15
4.3.1Soil
15
4.3.2Sediment
16
4.3.3Water
16
4.3.4Biomass
16
4.3.5Litter
18
Translocation within systems and hydrological export
18
Ecosystem Cl budgets
19
5
Chlorine transformation processes 23
6
6.1
Chlorine in organisms25
General uptake by plants and microorganisms
25
4.2
4.3
7
Techniques for studying Cl in the environment29
7.1 Ion chromatography
29
7.2 AOX 29
7.3 TOX and TX 29
7.4 VOCl 30
7.5 XSD and AED to identify Cl compounds in complex matrices
30
7.6ICP
31
7.7 Stable chlorine isotopes
31
7.8 Sample preparation and preservation
32
7.9 Process rate studies
32
8
8.1
8.2
8.3
8.4
Future challenges35
What is the spatio-temporal variability of Cl– and Clorg distribution in
landscapes?35
35
Which conditions and processes control Cl– and Clorg levels and transport?
Cl cycling in inland waters
36
How to model Cl cycling in terrestrial environments?
36
References37
SKB TR-13-265
Frequently used abbreviations
Clchlorine
chloride ion
Cl–
Clorg total organochlorine
VOCl volatile organochlorine
36
Cl chlorine-36; a radioisotope of Cl emitting primarily beta radiation
36
Cl– chloride-36 ion
36
Clorg organically bound chlorine-36
35
Cl chlorine-35; a stable isotope of Cl
37
Cl chlorine-37; a stable isotope of Cl
TX
total halogens; an operational definition based on an analysis method which strictly defined
detect chlorine, bromine and iodine
TOX total organic halogens; an operational definition based on an analysis method which strictly
defined detect chlorine, bromine and iodine
AOX adsorbable organic halogens; an operational definition based on adsorption on activated
carbon prior to analysis.
EOX Extractable organic halogens
d.m. dry mass
HCl hydrochloric acid
6
SKB TR-13-26
1Introduction
Chlorine (Cl) is one of the 20 most abundant elements on earth. It is essential for life for various
reasons. Chloride (Cl–), the only stable ionic form of Cl, is the major anion in blood and is present at
concentrations of approximately 100 mmol L–1 in plasma and interstitial fluid (Yunos et al. 2010). Cl–
participates in osmoregulation of cells (White and Broadley 2001), and is as an important electrolyte
for regulation of muscle function and synaptic transmission in the neural system. The adult human
dietary intake of Cl– in the USA is 6–12 g d–1 (Yunos et al. 2010). Cl– also functions as an essential
co-factor in enzymes involved in photosynthesis related to the oxidation of water by the PSII photosystem (Winterton 2000). Thereby, Cl is a critical nutrient and a suggested minimum requirement of
Cl for crops is 1 g kg–1 dry mass (d.m.) (White and Broadley 2001).
Many of the most debated organic pollutants, including the “dirty dozen” highly toxic and now
internationally banned persistent organic pollutants, are chlorinated (Godduhn and Duffy 2003).
Although natural halogenated organic compounds have been know since the late 19th century
(Gribble 2003), this was forgotten in the environmental debate, and the dominating view was that
chlorinated organic compounds (organochlorines; Clorg) in the environment were primarily of anthropogenic origin and often toxic. It is now evident that there is also a large natural production of Clorg.
Nearly 5,000 naturally produced chlorinated organic compounds have been identified and chemically
characterized, and their production has been associated with fungi, lichen, plants, marine organisms
of all types, insects, and higher animals including humans (Gribble 2003, 2010, Öberg 2002). Some
of these have well known physiological functions, including several important antibiotics (e.g. vancomycin). Others have important effects in the environments. For example volatile organochlorines
(VOCl) enhance atmospheric ozone destruction (Winterton 2000). However, the ecological functions
of most Clorg in nature, and the reasons for its production, are largely unknown.
Another research area where Cl has been central is hydrology. Cl– is the dominating chlorine pool
globally, is highly soluble in water, and has a high enrichment factor when comparing oceanic and
riverine concentrations (i.e. sea water concentrations are in the order of 2,500 times larger than
freshwater concentrations; Winterton 2000). At the first glance this indicates that Cl– is unreactive
in the environment and this has been a prevailing view for a long time (e.g. White and Broadley
2001). Accordingly, Cl– has been seen as an inexpensive and suitable tracer of soil and ground water
movements (Herczeg and Leaney 2011, Hruška et al. 2012) and studies using Cl– as a water tracer has
been a foundation for contaminant transport models (e.g. Kirshner et al. 2000). However, as discussed
below, there is now clear evidence that Cl– is highly reactive in some environments.
Recently 36Cl, a radioactive isotope with a half-life of 301,000 years, has attracted interest because of
its presence in radioactive waste. The long half-life in combination with high mobility in the geosphere
and the potential for substantial biological uptake creates a need for long-term risk assessments related
to handling and storage of radioactive waste (Limer et al. 2009). Previous assumptions that 36Cl in soils
primarily occurs as 36Cl– and is highly soluble and unreactive has been questioned along with growing
awareness of a more complex cycling of Cl in terrestrial environments.
Several aspects of Cl, including the physiological role (e.g. Yunos et al. 2010, White and Broadley
2001), the persistent organic pollutant perspective (e.g. Winterton 2000) and the global Cl cycle
(Graedel and Keene 1996) have been summarized previously, and will not be the primary focus here.
This text rather provides an update and supplement to previous reviews focusing on the terrestrial Cl
cycling (e.g. Öberg 2002, Clarke et al. 2009). A primary motivation for this is the recent interest in
36
Cl behavior in soils (e.g. Sheppard et al. 1996, Limer et al. 2009, Van den Hoof and Thiry 2012).
SKB TR-13-267
2
Fundamental chemical aspects of chlorine
Cl is the 20th most abundant element on Earth (Winterton 2000). It has atomic number 17 and belongs
to the halogen group in the periodic table. Cl has a high electron affinity and electronegativity and
thereby molecular Cl is a strong oxidant. Consequently, molecular Cl is rare in nature and instead
the inorganic ion from Cl– which is highly soluble in water typically dominates in the hydrosphere
and in minerals (but frequently not in soil as discussed below).
2.1
Cl isotopes and sources of 36Cl
Cl occurs in nature as primarily two stable isotopes, 35Cl (ca 76%) and 37Cl (ca 24%). Besides those
isotopes seven radioactive isotopes exist of which 36Cl has a very long half-life, 3.01·105 years. The
half-lifes of the other six radioactive Cl isotopes are less than one hour; these isotopes are, therefore,
not of interest in the context of Cl cycling in the environment. 36Cl decays with a maximum energy
of 709.6 keV either by emitting a beta particle (98.1%) or by electron capture (1.9%) resulting in
the end products argon-36 (36Ar) and sulphur-36 (36S), respectively (Rodríguez et al. 2006, Peterson
et al. 2007).
In the environment, 36Cl is produced by natural nuclear reactions; in the atmosphere by the spallation
of argon with cosmic ray protons, and in soil and rock by neutron activation of potassium, calcium and
chlorine (White and Broadley 2001). The resulting radiological dose to individuals is calculated by the
ratio of 36Cl to stable chlorine (36Cl/Cl) in the surface environment but it varies between geographical
locations. The natural 36Cl/Cl ratio is between 10–15 and 10–12 (Campbell et al. 2003). The dose can
thereby differ by several orders of magnitude between coastal and inland areas due to the difference
in concentration of stable Cl. 36Cl/Cl ratios exceeding 10–12 (up to 2·10–11) have been found in a 100 km2
area in the Tokai-mura region, Japan, where four nuclear power reactors and one nuclear fuel
reprocessing plant had been operated (Seki et al. 2007).
Cl was produced in large amounts by neutron activation of seawater upon nuclear weapon tests
between 1952 and 1958 (Peterson et al. 2007). These peaks in 36Cl have been used for dating ground
water (Campbell et al. 2003, White and Broadley 2001). 36Cl is also produced during nuclear power
reactor operation due to neutron capture of stable 35Cl that may be present at trace levels in core
materials, graphite, coolant water, and construction materials such as steel and concrete (Fréchou
and Degros 2005, Hou et al. 2007). In addition, 36Cl can be produced in considerable amounts via
spallation reactions of other concrete components, such as Potassium (K) and Calcium (Ca), primarily
in fast reactors where high-energy particles such as fast neutrons are present (Aze et al. 2007).
Although 36Cl levels are typically low, the active uptake of organisms and high concentration ratios
in plants relative to soils (Kashparov et al. 2007a, b, White and Broadley 2001) makes information
about Cl cycling in soils and sediment layers including bioavailability and residence end exposure
times relevant for risk assessments (Limer et al. 2009).
36
SKB TR-13-269
3
Major Cl reservoirs and large scale cycling
The largest Cl reservoirs on the earth’s surface are the crust and the ocean (Graedel and Keene 1996;
Table 3-1). Inorganic Cl by far dominates these reservoirs. Estimates for the other reservoirs are also
largely based on Cl– concentration measurements. This assumption of a general dominance of Cl– is
problematic for the pedosphere because Clorg levels have been shown to range from 11 to near 100%
of the total Cl pool in a large range of soil types (Gustavsson et al. 2012, Johansson et al. 2003a,
Redon et al. 2011, 2013; see also Section 4 below). This means that the pedosphere Cl pool may
be at least twice as large as proposed by Greadel and Keene (1996).
Graedel and Keene (1996) also propose fluxes between the reservoirs based on available data.
These values are poorly constrained. For example, to balance the overall budget a yearly loss of
30 Tg Cl yr–1 from the pedosphere (equivalent to 1.25‰ of the pedosphere reservoir) had to be
assumed. This would lead to a rapid depletion of the pedosphere stock which is clearly unrealistic,
and therefore this budget illustrates substantial lack of knowledge regarding large scale fluxes in
combination with bias from ignoring Clorg formation in terrestrial environments.
In the large scale inorganic Cl cycle mineral weathering contributes with Cl– to freshwaters and
later the ocean. The largest contribution of Cl– to the atmosphere is sea salt aerosols while minor
contributions include HCl from volcanic activity and biomass burning, mineral aerosols, and volatile
organochlorines (VOCls) of natural or anthropogenic origin. Cl– is transported to oceans and soils by
wet and dry deposition (further described in Section 4 below).
Table 3-1. Major Cl reservoirs on earth, how they were estimated, and theoretical residence times
based on data from Graedel and Keene (1996; corrected values for the cryosphere). Note that
major pools of organic Cl are not considered and therefore the pedosphere reservoir is highly
uncertain (see text for details).
Reservoir
Cl content (g) Reservoir was estimated from:
Residence time
(years)
Mantle
22 × 1024
Meteorite Cl:Si ratio; mantle mass.
1.1 × 1013
Crust
60 × 10
21
Meteorite Cl:Si ratio; crustal mass.
3.4 × 108
Oceans
26 × 10
21
Cl concentration; water volume.
4.3 × 106
Pedosphere
24 × 10
15
Average soil Cl 100 mg kg d.m. mean soil depth and density 2 m
and 1.0 g cm–3.a
5.3 × 102
Freshwater
320 × 1015
Average Cl– concentration in rivers and ground water; water volume.
1.5 × 103
Cryosphere
0.5 × 10
15
Cl concentration in rain or snow; ice volume.
8.3 × 101
Troposphere 5.3 × 10
12
Concentrations of HCl, CH3Cl and Cl aerosols; troposhere volume.
8.8 × 10–4
Stratosphere 0.4 × 10
12
Cl concentration; stratosphere volume.
1.3 × 101
a
–
–1
–
–
Assumptions by Graedel and Keene (1996); no published references in support of these values provided.
SKB TR-13-2611
4
Chlorine in terrestrial ecosystems
During the past decades there has been a rapid development towards improved qualitative understanding of the terrestrial Cl cycle. We are, however, far from being able to present a detailed quantitative
picture due to two main reasons. The first one is due lack of data – in most cases the reported fluxes
are extrapolated to whole ecosystems or regions based on a few measurements at specific points in
time and space. Secondly, there is still lack of knowledge regarding some qualitative aspects of the
biogeochemical Cl cycle, and the processes behind the Cl cycling (e.g. formation and degradation
of Clorg) as well as their regulation are still uncertain.
The terrestrial environment includes biomass, litter and surface soil layers (often characterized
by higher content of organic/humic matter than below layers), mineral soil layers, and soil water.
Figure 4-1 show these reservoirs and also fluxes of Cl to, from, and within the system.
4.1
Input of Cl to terrestrial ecosystems
4.1.1Deposition
Prior to the mid-1950s, the chemical composition of surface waters was considered to be a result
of physical land-use history in combination with the geochemical, hydrological, and features of
the surrounding area (Öberg and Bäckstrand 1996). In the mid-1950s, it was suggested that the
chemical composition of rivers mirrors the chemical composition of precipitation (Eriksson 1955).
The arguments were based on extensive Cl– and sulphate data and implied that these compounds
originated from oceans, as the oceans produce sea salt aerosols when the waves break the ocean’s
surface (Eriksson 1960). The aerosols are carried away with the winds to the atmosphere and are
either transported back to the sea or deposited on land by precipitation that washes out Cl– from the
atmosphere. There is a clear pattern of decreasing wet deposition of Cl– with increasing distance to
the sea and with consideration to prevailing wind directions (Clarke et al. 2009). Gases and particles
can also contain Cl–; they can either be deposited directly on the ground or stick to the crown of
trees or washed with the precipitation to the soil. The input of Cl– by gases and particles is called
“dry deposition” compared to “wet deposition”, which is deposition of Cl– by e.g. rain and snow.
The deposition to soil is generally higher in forested areas than over open land because atmospheric
particles are intercepted by vegetation and there is also possible leaching from the vegetation.
Gas emission
Aboveground
biomass
Belowground
biomass
Throughfall
Lierfall
Wet deposon and
dry deposion
Surface
organic/humus
layer
Soil leaching
Mineral soil
Stream
output
Figure 4-1. A schematic diagram of pools and fluxes of chlorine in terrestrial ecosystems.
SKB TR-13-2613
The quantification of wet deposition of Cl– can be done with high precision and is relatively well constrained but highly variable depending on distance to the sea (higher wet deposition close to the sea).
In Europe wet Cl deposition varies from 0.5 to 220 kg ha–1 yr–1 (Clarke et al. 2009). Dry deposition
includes inputs Cl– via gases, aerosols, and particles. Because dry deposition includes several forms
of Cl and is affected by interception by surfaces such as tree canopies, the estimate of dry deposition
is difficult to measure and considerable more uncertain but has been estimated to 15–73% (average
43%) of total deposition based on data from North America and Europe (Svensson et al. 2012).
It is well known that precipitation, in addition to Cl–, also contains Clorg (Enell and Wennberg 1991,
Grimvall et al. 1991, Laniewski et al. 1995). Measurements of individual halogens in organic matter
derived from precipitation have revealed that most of the organically bound halogens detected as
AOX are chlorinated compounds (Laniewski et al. 1999) (see Section 7.1 for description of AOX
analysis). Brominated compounds are widespread but less prevalent, and organically bound iodine
has only been detected at sites close to the sea (Laniewski et al. 1999).
Characterization of the Clorg present in rain and snow has shown that the major part of Clorg is found
in fractions of relatively polar and non-volatile to semi-volatile compounds, particularly organic
bases and acids (Laniewski et al. 1999). Chloroacetic acids can occasionally explain up to 6% of
the Clorg in precipitation (Laniewski et al. 1995, von Sydow et al. 1999, Svensson et al. 2007a),
while the relative contribution from volatile organochlorines (VOCls) usually is smaller, with
concentrations often at ppt (ng L–1) levels (Schleyer 1996, Svensson et al. 2007b).
Very little is known about the origin of the Clorg in precipitation. Known industrial pollutants, such as
flame retardants (e.g. chlorinated alkyl phosphates) and pesticides (e.g. lindane) are typically present
at ppt levels (Stringer and Johnston 2001), i.e. in concentrations about three orders of magnitude less
than observed total AOX concentrations. In addition, throughfall contains higher concentrations of
Clorg than precipitation only; a study conducted at Klosterhede in northwest Denmark suggests that
Clorg in throughfall mainly originates from internal sources rather than from dry deposition (and thus
external) sources (Öberg et al. 1998).
4.1.2Weathering
As mentioned previously, Cl– has long been believed to participate in geochemical processes only,
i.e. transported from oceans via soil back to the oceans again, being only negligibly affected by biological processes or interactions with organic matter. Riverine Cl– has likewise often been considered
to originate primarily from the atmosphere only, despite possible weathering processes during the
pathway through the soil (Eriksson 1960, Schlesinger 1997). There are limited analyses of Cl– in
rocks, but felsic bedrocks such as granite contains low amounts of Cl–, and the highest amounts are
found in mafic bedrocks (Melkerud et al. 1992) and obviously in halide rich evaporite minerals.
Acidic minerals can be considered to have a lower chemical weathering rate than alkaline minerals.
The weathering rate has been estimated for a small stream at Hubbard Brook, New Hampshire, USA,
with bedrock consisting of mainly granite. Approximately 2% of the Cl– stream output originated
from weathering, which can be considered as small compared to the atmospheric contribution
(Lovett et al. 2005).
Land rise in previously glaciated regions can result in soils that were originating as marine sediments
and therefore are rich in Cl–. Release of Cl– from such marine deposits constitutes a special case
with significant subsurface contribution of Cl– to soils, water and organisms. In the Forsmark area,
in central East Sweden, investigated thoroughly by SKB, leaching from marine deposits could have
contributed up to 20% of the Cl– exported from the area (Tröjbom and Grolander 2010).
4.1.3 Input from irrigation, fertilization and road de-icing
Anthropogenic Cl– input from irrigation and fertilization can represent substantial inputs to terrestrial
environments. Irrigation with low salinity water can contribute in the order of 1,000 kg ha–1 yr–1 and
thereby anthropogenic contributions can be the major Cl input in some areas (White and Broadley
2001). Irrigation of crops can also lead to accumulation of salt in soil (Rengasamy 2010). Cl is
an essential element for plants (Broyer et al. 1954) and is known to be an important anion in crop
production (Engel et al. 2001).
14
SKB TR-13-26
Since the start of de-icing of roads in mid-twentieth century, studies have shown increased Cl–
concentrations in both surface water and groundwater in the vicinity of roads. In the LaxemarSimpevarp area in South East Sweden, 35–56% of the total Cl– input was estimated to come from
road salt (Tröjbom et al. 2008). Road salt effects can be chemical (e.g. induce ion exchange affecting
acidification and metal and nutrient leaching; Löfgren 2001, Bäckström et al. 2004) or biological
(e.g. effects on mussels and pond food webs; Todd and Kaltenecker 2012, Van Meter et al. 2011).
4.2
Gaseous efflux from terrestrial systems
Volatile organochlorines (VOCls) are produced in a wide variety of ecosystems such as wetlands,
salt marshes and forests (Dimmer et al. 2001, Rhew et al. 2002, Albers et al. 2011), and in different
climatic regions including boreal, temperate and tropical areas (Yokouchi et al. 2002, Rhew et al.
2008, Albers et al. 2011). Despite the growing knowledge in the field, data on VOCl emission rates
are scattered and inconsistent. Budget and transport estimates on various scales are highly uncertain,
partly because low concentrations of each specific VOCl make sampling and analyses challenging
(Pickering et al. 2013; see Section 7 below). Furthermore, VOCl sources and sinks are not well
understood, and continuous observations over time are scarce.
VOCls have been found in several terrestrial biomes such as tropical forest, grasslands, deciduous
forest, taiga, tundra, and rice fields (Khalil et al. 1998, Varner et al. 1999, Laturnus et al. 2000, Redeker
et al. 2000, Haselmann et al. 2002, Dimmer et al. 2001, Hoekstra et al. 2001, Wang et al. 2007, Rhew
et al. 2008). Most of the available information has been gathered in the northern hemisphere. Previous
studies in terrestrial ecosystems have primarily considered seven different VOCls. The most commonly
studied compounds are chloromethane (CH3Cl) and chloroform (CHCl3). Other VOCl compounds
reported include CCl4 (tetrachlorometane), C2H3Cl (chloroethylene), CH2Cl2 (dichloromethane),
CH3CCl3 (methyl chloroform), and C2H3Cl3 (trichloroethane) (Haselmann et al. 2002, Hoekstra et al.
2001, Wang et al. 2007, Mead et al. 2008, Rhew et al. 2008). In addition, other halogenated compounds
such as bromomethane, iodomethane trichlorofluoromethane (freon-11) and dichlorodifluoromethane
(Freon-12) have also been reported to be released from terrestrial sources (Khalil and Rasmussen 2000,
Rhew et al. 2000, Keppler et al. 2003, Varner et al. 2003).
Emissions are considered to be small compared to wet and dry deposition of Cl. However, a 36Cl radio­
tracer study indirectly indicated substantial release of VOCl in soils corresponding to 0.18 g Cl m–2 yr–1
or 44% of the annual wet deposition (Bastviken et al. 2009). This is a high number that needs validation,
but interestingly it includes all possible VOCls in contrast to other estimates that measure specific
VOCl compounds only. Previous studies showed average emission of chloroform and chloromethane
corresponding to 0.13 and 0.04 g Cl m–2 yr–1, respectively, from a coniferous forest soil (Dimmer
et al. 2001), or < 0.01 g Cl m–2 yr–1 for chloroform from a Scots pine (Pinus silvestris) forest soil
(Hellén et al. 2006). Other ecosystems also show VOCl emissions. For instance coastal salt marshes
are releasing chloromethane indicating fluxes of 0.2–1.2 g m–2 yr–1 (Rhew et al. 2000). Given this, the
formation of VOCl would not only represent a substantial proportion of the emission to the atmosphere
but also a significant part of the chlorine cycle. In addition, such high fluxes might explain some of the
observed Cl– imbalances found in catchments in Europe and North America (Svensson et al. 2012).
4.3
Terrestrial reservoirs of chlorine
4.3.1Soil
Total Cl typically range from 20 to > 1,000 mg Cl kg–1 d.m. in non-saline soils (Table 4-1). The concentrations of Clorg in surface soil layers are in most cases higher than Cl– levels (Table 4-1). The dry
mass fraction of Clorg in surface soils (0.01–0.5%), is as large as for phosphorous (0.03–0.2%) and
only slightly lower than nitrogen (1–5%) and sulphur (0.1–1.5%) (Öberg 2002). Measurements of
bulk density, horizon thicknesses etc are often difficult to obtain for many studied soils and often
affect total storage calculations more than concentration differences. Total storage is usually largest
in the mineral soil layer because of its greater thickness compared to the organic surface layer, despite
the fact that Clorg concentrations are typically 2–5 times higher in the organic surface soil layer (Redon
et al. 2011). The presently available data only separates organic and mineral soil layers and more
SKB TR-13-2615
detailed discussions about specific soil profiles is therefore not possible at this point. The data in
Table 4-1 indicate that total Cl levels are higher in soils with more organic matter while the percentage
Clorg is frequently higher and above 80% in mineral soils. Based on this information higher soil Cl
levels may be connected with soil organic matter content, probably because soil chlorination processes
stabilize Cl as Clorg as proposed by Gustavsson et al. (2012). However, the available information is too
scarce to be conclusive on a general basis.
The soil organic matter is mainly composed of high molecular weight substances, usually larger than
1,000 Dalton (Hjelm and Asplund 1995). Very few studies have tried to determine Clorg content in
different types of soil organic matter. Lee et al. (2001) concluded that the Clorg was associated with
organic matter with a molecular weight of < 10,000 Dalton, while most organic matter in the studied
soil had higher molecular weight (> 10,000 Dalton). Bastviken et al. (2007, 2009) found that 1–10%
of the Clorg in coniferous soil was associated with water leachable fractions of the organic matter.
Interestingly Redon et al. (2011) found that the Cl:C ratio in Clorg increases with soil depth, ranging
from 0.08 to 2.7 mg Clorg g–1 C in the humus layer and from 0.6 to 6.1 mg Clorg g–1 C in mineral
soil (0–40 cm). Johansson et al. (2001) also report increasing Cl:C ratios with soil depth. Hence,
some of the chlorinated organic molecules can be resistant to degradation and selectively preserved
compared to average soil organic matter.
4.3.2Sediment
Analysis of sediment Clorg have focused on contamination from industrial activities. There is
a large body of literature on specific chlorinated pollutants (e.g. PCBs and DDTs). Among the bulk
Clorg measures, adsorbable organic halogens (AOX; see Section 8 about description of analytical
techniques) have been used to study sediment pore waters, but such efforts in non-contaminated
sediments are relatively rare. There has also been a suggestion to avoid AOX analyses for sediment
pore waters when it was discovered that it cannot discriminate between natural and anthropogenic
Clorg (Müller 2003). Extractable organic halogens (EOX; extraction of sediments with cyclohexane–
isopropanol under sonication) yielded concentrations of 5–70 mg kg–1 sediment (probably dry mass
sediment but this was unclear) in the upper 2 cm of Bothnia bay sediments in a pulp and paper
mill contamination area (Pöykiö et al. 2008). Another study reported Clorg concentrations of < 10 to
843 µg Cl g–1 organic matter in seven non-polluted inland water sediments (Suominen et al. 1997).
The analysis methods differed (AOX and EOX after various extractions) which makes comparisons
uncertain. Analyses using similar methodology as for soils (TOX; see Section 8 for description)
appear to be largely missing and therefore Cl and Clorg levels in sediments are presently unclear.
4.3.3Water
In contrast to soils, Cl– concentrations generally exceed Clorg concentrations in water. For example,
the Cl– concentration in various waters is measured in mg L–1, while Clorg is typically measured in
µg L–1 and VOCls are in the range of ng L–1 (Table 4-2) (Eriksson 1960, Asplund and Grimvall 1991,
Enell and Wennberg 1991, McCulloch 2003). This means that the atmospheric deposition of Clorg
is in the order of 1000-fold lower than deposition of Cl– and thereby often assumed to be negligible
from a bulk Cl perspective. While ground water has the highest Cl– concentrations in comparison
with rain water and surface waters, Clorg and VOCl concentrations can be highest in surface waters
(Table 4-2). The environmental quality criteria with regard to Cl– levels in groundwater published
by Swedish food agency use a Cl– threshold level of 100 mg L–1.
4.3.4Biomass
The Cl– content of plant biomass varies among plant species. For plant growth a general Cl–
requirement of 1 mg g–1 d.m. has been suggested, but deficiency symptoms have been observed at
0.1–5.7 mg g–1 d.m., while toxic levels between 4–50 mg g–1 d.m. have been reported (White and
Broadley 2001). This means that extrapolations across species and locations are highly uncertain.
Plant Clorg content has been estimated to 0.01–0.1 mg g–1 d.m. (Öberg et al. 2005), but this is based
on scattered measurements from beech leaves, spruce needles, sphagnum moss and bulk samples
of grass and the variability between species and plant parts are unknown at present.
16
SKB TR-13-26
Table 4-1. Examples of total Cl and the fraction Clorg in various soils. Soil depth is denoted by soil
layer (e.g. humus and mineral layers) or by distance from soil surface.
Ecosystem, country
Total Cl
(mg kg–1)
Clorg
(%)
Soil layer or depth
Reference
Coniferous forest, Sweden
99–274
67–73
humus
Gustavsson et al. 2012
Conif. forest, Sweden
154
86
humus
Bastviken et al. 2009
Conif. forest, Sweden
331
95
humus
Bastviken et al. 2007
Conif. forest, Sweden
127
69
humus + mineral
Svensson et al. 2007b
humus
Öberg and Sandén 2005
Conif. forest, Sweden
Conif. forest, Sweden
369–458
81–85
humus
Johansson et al. 2003a,
Johansson et al. 2003b
Conif. forest, Sweden
310
68
humus
Johansson et al. 2001
Conif. forest, Denmark
206–772
67–85
humus
Albers et al. 2011
Conif. forest, China
45
38
15 cm
Johansson et al. 2004
Conif. and decideous forests, Francea
45–1,041
40–100
humus
Redon et al. 2011
Mixed deciduous forest, Sweden
224
85
humus
Johansson et al. 2003a,
Johansson et al. 2003b
Mixed forestsa
34–340
89
mineral, 0–30 cm
Redon et al. 2013
Conif. and decideous forests, Francea
25–210
29–100
mineral 0–10 cm
Redon et al. 2011
Pasture, Sweden
46–65
85–90
5–15 cm
Gustavsson et al. 2012
Grassland, Francea
13–1,248
83
mineral, 0–30 cm
Redon et al. 2013
Agricultural soil, Francea
19–100
87
mineral, 0–30 cm
Redon et al. 2013
Agricultural soil, Sweden
45–49
84–89
5–15 cm
Gustavsson et al. 2012
Paddy soil, China
38
34
15 cm
Johansson et al. 2004
Peat bog, Canada
30–1,177
43–84
Surface – 6 m
Silk et al. 1997
Peat bog, Chile
366–1,084
82–93
Surface – 2 m
Biester et al. 2004
a
Includes study sites at different distances to the sea.
Table 4-2. Chloride (Cl–), organochlorines (Clorg) and chloroform concentrations in various waters,
primarily in Sweden. Chloroform is one of the most frequently detected volatile chlorinated
organic compounds (VOCl) in surface water.
Cl–
(mg L–1)
Clorg
(µg L–1)
Chloroform
(ng L–1)
Rain water
0.2–3.5a
1–15d
11–97g
Groundwater
10–300
5–24
Surface water (lakes and rivers)
0.74–11
b
c
5–1,600h
e
5–200
f
4–3,800i
(a) Minimum and maximum concentrations obtained from 6 precipitation stations in different regions of Sweden
1983–1998 (Kindbohm et al. 2001).
(b) Minimum and maximum concentrations from 20,100 wells (dug wells and drill wells) in Sweden sampled during
1984–1986 (Bertills 1995).
(c) Concentrations (10th and 90th percentiles) obtained from analyses of Swedish lakes during 1983–1994
(Wilander 1997).
(d) Minimum and maximum concentrations in rain and snow at 7 sites in Sweden (Laniewski et al. 1998, 1999),
combined with typical range given in Öberg et al. (1998).
(e) Minimum and maximum concentrations in groundwater from 14 wells in Denmark (Grøn 1995).
(f) Minimum and maximum concentrations in 135 lakes (Asplund and Grimvall 1991) and rivers in Sweden (Enell
and Wennberg 1991).
(g) Minimum and maximum concentrations of chloroform obtained from precipitation measurements in Germany
1988–1989 (Schleyer et al. 1991, Schleyer 1996).
(h) Minimum and maximum concentrations obtained from groundwater measurements at one site in Denmark
(Laturnus et al. 2000).
(i) Minimum and maximum concentrations compiled from rivers and lakes in Belgium, Canada, France, Germany,
The Netherlands, Switzerland, UK, USA (McCulloch 2003).
SKB TR-13-2617
In another study, standing stock Cl in trees in a pine forest in Belgium, based on measurements of different plant parts, were estimated to 4.7 and 5.5 kg Cl ha–1 for wood plus leaves, and roots, respectively
(Van den Hoof and Thiry 2012). This study also found that fresh leaves had the highest Cl concentration (0.59 mg g–1 d.m.) corresponding to 35% of the Cl in the trees and that Clorg accounted for less than
10% of the Cl in the leaves and the bark but constituted 20% of the total biomass Cl of the whole tree.
Interestingly, monitoring of total Cl in various landscape compartments, including soil, sediment,
water, and biomass, in the Forsmark area indicated that the terrestrial biomass Cl pool dominated
over the other pools and accounted for in the order of 60% of the total catchment Cl (Tröjbom and
Grolander 2010). Cl was substantially enriched in biomass compared to other comparable elements
(e.g. bromine and iodine) and nutrients (nitrogen, phosphorus, potassium, calcium). It is known that
Cl is an essential element, but this level of enrichment indicates that the roles of Cl for organisms
may not be fully understood, and that a large part of potential contaminant 36Cl reaching terrestrial
parts of the landscape will be taken up by biota. The pool of Cl in the limnic biota was negligible
compared to Cl pools in sediments and water.
Another interesting finding is reported by Tröjbom and Nordén (2010). In a comparison between
two areas in central East Sweden and South East Sweden, Forsmark and Laxemar-Simpevarp, respectively, total Cl:C ratios in the green parts of terrestrial plants were 10-fold higher in the Forsmark
area for unknown reasons. In the Forsmark area the green parts of rooted terrestrial plants also had
a 10-fold higher total Cl:Br ratio compared to heterotrophs, most dead biomass, sea water and soils.
Altogether, this indicates active uptake and accumulation of Cl in green parts of terrestrial plants.
4.3.5Litter
Litter is represented by detached dead or dying plant biomass. Simultaneous leaching of Cl– and
formation of Clorg has been shown in litter (Myneni 2002). A recent study of senescent leaves from
white oak (Quercus alba) showed Cl– and Clorg contents of 335 and 165 mg kg–1, respectively (Leri
and Myneni 2010). In this study Cl fractions were quantified with X-ray absorption near-edge structure
(XANES) spectroscopy. This enabled the discoveries that (1) total Clorg content in the leaves increased
during the senescence and gradual degradation, (2) aliphatic Clorg was present a stable levels over time
and seems contributed by plant processes and stable to degradation, and (3) that water soluble aromatic
Clorg was first leached from the leaves followed by later accumulation of non-soluble aromatic Clorg
during senescence.
Spruce needle litter in Denmark contained 51–196 (median 101) µg Clorg g–1 d.m. (Öberg and Grøn
1998). In a study of 51 different forest sites in France with both coniferous and broad leaf tree species
total Cl content in the litter was 46–528 (median 147) mg kg–1 d.m. and the percentage Clorg was
11–100% (median 40%; Redon et al. 2011). Again, available data suggest substantial variability
within and between species and locations.
4.4
Translocation within systems and hydrological export
There are scattered indications of extensive internal cycling of Cl in terrestrial ecosystems. For
example, the annual root uptake of Cl by Scots pine (Pinus silvestris) was found to be 9-fold larger
than Cl demand by the tree (Van den Hoof and Thiry 2012). The excess Cl was returned primarily as
Cl– in throughfall and to some extent by litterfall. Similarly, a study integrating data from 27 forests of
different types in France show that throughfall was highly variable between different forests, but on
an average twice as high as the total atmospheric Cl deposition (41 and 20 kg ha–1 yr–1, respectively;
Redon et al. 2011). For comparison the average Cl in litterfall in these systems ranged from 0.1 to
2.5 kg ha–1 yr–1. Again this indicates that tree Cl uptake can be much greater than the net internal
demand to supply growth of new biomass. This cycling, if being a general pattern, will prolong Cl
residence times in the systems. The reasons for excess uptake of Cl relative to needs is unknown but
could be related with evapotranspiration if Cl– enters the plant with the water without discrimination.
There is also an extensive cycling of Cl in soils. In the upper soil layers microbial activity results in
formation of Clorg from Cl– (i.e. chlorination; e.g. Bastviken et al. 2009, Öberg et al. 2005). Measured
rates in incubations from 14 sites varied from 1.4–90 ng Cl g–1 d.m. d–1 or, expressed as fraction of
the Cl– pool being transformed to Clorg per day, from 0.0002 to 0.003 d–1) (Gustavsson et al. 2012).
18
SKB TR-13-26
Table 4-3. Examples of estimated soil organic matter chlorination rates.
Source
Type of study and experiment time Specific chlorination Mass-based rate
Area-based rate
(ng Clorg g–1 Corg d–1) (kg Cl ha–1yr–1)
(d–1)
Lee et al. 2001
Lab: arable soil; 11 weeks
0.00199
–
–
Silk et al. 1997
Lab: peat; 8 weeks
0.00066
–
–
Bastviken et al. 2007
Lab: forest soil; 78 days
0.00029
20
–
Bastviken et al. 2009
Lab: forest soil; 6 months
0.0007–0.0034
78–311
–
Gustavsson et al. 2012 Lab: conif. forest soil. 4 months
0.00094–0.0014
37–90
–
Gustavsson et al. 2012 Lab: pasture soil. 4 months
0.00021–0.00074
3.5–4.9
–
Gustavsson et al. 2012 Lab: agricultural soil. 4 months
0.00032–0.00055
2.6–5.0
–
Öberg et al. 1996
Field experiment: spruce litter
–
1,002
0.5
–
1,469–7,517
0.35
Öberg and Grøn 1998
Field study: spruce litter
Öberg 2002
Lysimeter mass balance
2.7
Öberg et al. 2005
Mass balance of catchment
2.0
Field based estimates are rare. Öberg et al. (2005) estimated the chlorination to be 2 kg ha–1 yr–1 from
mass balance calculations while laboratory measurements in the same area yielded chlorination rates
of 2–13 kg ha–1 yr–1 (corresponding to 50–300% of the wet deposition at this catchment; Bastviken
et al. 2007, 2009).
Table 4-3 gives an overview of some published soil organic matter chlorination rates of relevance
for field conditions. Available data indicates higher chlorination rates on a weight basis in litter
compared to in deeper soil layers. Extensive chlorination has been shown upon litter degradation
(Myneni 2002) and seasonal patterns have been suggested (Leri and Myneni 2010).
Dechlorination processes (transformation from Clorg to Cl– by either organic matter decomposition or
by selective removal of Cl atoms from organic molecules) have been extensively studied in relation
to organochlorine pollution and bioremediation (e.g. van Pée and Unversucht 2003). However, while
well known for specific Clorg compounds there are yet no known studies reporting directly measured
dechlorination rates for bulk Clorg in terrestrial environments. In spite of this, dechlorination is believed
to be important and in mass balance calculations steady state conditions with similar chlorination
and dechlorination rates are often assumed (Öberg et al. 2005). Based on indirect evidence a recent
study suggested that the balance between chlorination and dechlorination is more important for soils
Cl– levels than Cl– deposition (Gustavsson et al. 2012).
Migration across different depth zones can be important for internal Cl cycling in soils. Intensive
chlorination has been observed in surface soil and litter layers, while Clorg levels decrease with soil
depth and the form of Cl– dominating in the hydrological export from catchments is Cl– (Figure 4-2).
This has led to the suggestion that Clorg is leached from surface soils and either absorbed (and preserved) or transformed to Cl– in deeper soil layers (Figure 4-3) (Öberg and Sandén 2005, Rodstedth
et al. 2003, Svensson et al. 2007b).
4.5
Ecosystem Cl budgets
There have been a few attempts to make ecosystem or catchment scale Cl budgets. The ecosystem
budgets are typically based on concentration measurements in combination with information about
carbon and water cycling that can support estimates of Cl transport and transformation. Figure 4-4
illustrate published budgets. While they all converge regarding the order of magnitudes, many
potentially important reservoirs and fluxes remain unknown or uncertain because of lack of data.
Net ecosystem budgets of Cl–, i.e. comparison of atmospheric deposition and stream export are common
because Cl– is often covered in monitoring efforts. A recent summary of such data reveal that there is
an imbalance in many catchments (Table 4-4) (Svensson et al. 2012). This is not surprising given the
new knowledge of several processes that can retain Cl– (plant uptake, formation of Clorg) or release Cl–
(decay of biomass, dechlorination), and imbalances were most striking in areas with a wet Cl– deposition
below 6 kg ha–1 yr–1.
SKB TR-13-2619
Clorg
Cl–
0.007
10
0.4
Clorg
Cl–
0.007
5
0.5
Figure 4-2. Major paths of Cl transport to and from the Stubbetorp catchment, near Norrköping, Sweden
(Svensson et al. 2007b). Fluxes to and from the catchment are dominated by chloride (Cl–) while the standing Cl stock in soil is dominated by Clorg. Reservoirs are in g m–2 and fluxes represent g m–2 yr–1.
Figure 4-3. Estimated organic chlorine transport in soil. Concentration data and flux estimations for
top-soil are based on data from Rodstedth et al. (2003) and Svensson et al. (2007b). Clorg is leached from
the top soil and gradually lost from the soil water by precipitation or adsorption to the solid phase or by
organic matter degradation while the water moves downward through the soil column. Cl– shows an opposite pattern with lower relative concentrations in surface soils and increasing concentrations downward
partly due to the release of Cl– from Clorg. This model can explain why the water released from soils has
higher concentrations of Cl– than Clorg, while Clorg dominates in surface soils.
20
SKB TR-13-26
Figure 4-4. Examples of terrestrial Cl budgets. Units are kg ha–1 (reservoirs) and kg ha–1 yr–1 (fluxes). Note
that the budget by Öberg et al. 2005 is not independent from the budget by Öberg and Grøn (1998). Cl– and
Clorg denotes chloride organochlorines, respectively, and VOCl is volatile organic chlorines e.g. chloroform.
A couple of laboratory lysimeter studies specifically addressed Cl– balances by irrigating soil cores
with artificial rain water having well known Cl– concentrations and monitoring efflux from the cores
(Rodstedth et al. 2003, Bastviken et al. 2006). Both these studies noted substantial imbalances indicating substantial Cl transformation in the soil, but the patterns were not clear (sometimes a net loss and
sometimes a net accumulation) and could not be easily explained.
A field scale 36Cl tracer study was performed by Nyberg et al. (1999) which injected 36Cl– in a small
catchment. When injections were made in surface soils only 47% could be recovered over 30 days
while 78% of simultaneously injected radioactive water (3H2O) was recovered. Upon injection in
deeper soils, the 36Cl– recovery was greater (83%), but still lower than for 3H2O (98%). Clearly, Cl–
is preferentially retained in surface soil.
The Cl– retention affects residence times of Cl in catchments and implies longer residence times than
for water which previously were assumed to be reflected by Cl– transport through soils. Few studies
have addressed this issue yet, but estimates from one study of forest Cl cycling conclude that overall
Cl residence times considering both Cl– and Clorg pools and fluxes were 5-fold higher than residence
times considering only Cl– and neglecting Clorg formation (Redon et al. 2011).
As noticeable above, in all parts of Chapter 4, data have primarily been collected in forests and to
some extent also pastures, agricultural land, and peat bogs. Unfortunately, wetlands, sediments and
discharge areas where potential 36Cl contaminants will leave underground aquifers are poorly studied;
to our knowledge, no studies on Cl cycling including Clorg, Cl uptake by organisms, food web transfer
and loss by emission of VOCl from wetlands, streams, reservoirs or lakes have been published.
SKB TR-13-2621
Table 4-4. Chloride (Cl–) balances of catchments. Negative numbers indicate loss of Cl– from the
catchment while positive numbers indicate accumulation in the catchment. Note that balances
can be biased by underestimated dry deposition in some areas depending on the sampling
methods.
Site and location
Forest type
Sampling
years
Cl balance
(input-output)
(kg ha–1 y–1)
Cl balance
% of input
Ref.
Aneboda, Sweden
Coniferous
9
–5.7
41
1
Bear Brook, ME, USA
Deciduous
3
–4.2
27
2
Birkenes, Norway
Coniferous
19
–19.2
57
3
Coweeta-14, NC, USA
Deciduous
4
0.6
9
4
Coweeta-18, NC, USA
Deciduous
12
–0.6
11
4
Coweeta-2, NC, USA
Deciduous
12
–1.1
22
4
Coweeta-27, NC, USA
Deciduous
12
0.2
2
4
Coweeta-32, NC, USA
Deciduous
4
–0.1
1
4
Coweeta-34, NC, USA
Deciduous
4
0.8
11
4
Coweeta-36, NC, USA
Deciduous
12
–3.7
60
4
Fernow forest, w-4, WV, USA
Deciduous
15
–1.5
71
5
Forellenbach, Germany
Coniferous
16
–3.8
83
3
Gammtratten, Sweden
Coniferous
6
–0.9
45
1
Gårdsjön, Sweden
Coniferous
10
2.7
3
6
Hietajärvi, Finland
Coniferous
15
–0.3
30
3
Hubbard Brook, w-3, USA
Deciduous
29
–1.1
35
7
Hubbard Brook, w-6, USA
Deciduous
33
–0.8
23
7
Kindlahöjden, Sweden
Coniferous
9
–0.3
3
1
Kosetice Observatory, Czech Republic
Coniferous
18
2.5
46
8
Lehstenbach, Germany
Coniferous
6
0.4
4
9
Llyn Brianne catchment C16, Great Britain
Heath/grass
2
–2.2
3
10
Lysina, Czech Republic
Coniferous
1
4.5
45
11
Maryland, HCWS, USA
Deciduous
1
–4.4
102
12
Panola Mountain, GA, USA
Deciduous
12
–3.1
111
13
Pluhuv Bor, Czech Republic
Coniferous
1
0.8
13
11
Plynlimon catchment Afon Cyff, Great Britain
Heath/grass
2
–5.3
4
10
Plynlimon catchment, Upper Hafren,
Great Britain
Moorland
18
–5.8
4
16
Plynlimon catchment, Nant Tanllwyth,
Great Britain
Coniferous
5
9.9
6
17
Saarejärve, Estonia
Coniferous
13
1.8
33
14
Steinkreuz, Germany
Deciduous
6
–0.1
1
9
Strengbach, France
Mixed coniferous/
deciduous
3
–0.2
1
15
Valkea-Kotinen, Finland
Coniferous
15
–1.7
121
3
References:
1. Integrated monitoring in Sweden, Swedish Environmental Research Institute, 2007-October.
2. (Rustad et al. 1994; Colin Neal, Centre for Ecology and Hydrology, UK, personal communication, September 2010).
3. ICP Integrated Monitoring Programme Centre, Finnish Environment Institute 2009-September.
4. (Swank and Crossley 1988).
5. USDA Forest service (Mary Beth Adams, USDA Forest Service, personal communication, October 5, 2006).
6. (Hultberg and Grennfelt 1992).
7. Cary Institute of Ecosystem Science, 2007-October.
8. Czech Hydrometeorologicla Institute, 2009-September.
9. (Matzner 2004).
10. (Reynolds et al. 1997).
11. (Krám et al. 1997).
12. (Castro and Morgan 2000).
13. (Peters et al. 2006).
14. The Estonian Environment Information Centre 2009-October.
15. (Probst et al. 1992).
16. (Neal et al. 2010; Colin Neal, Centre for Ecology and Hydrology, UK, personal communication, Spetember 2010).
17. (Neal et al. 2004; Colin Neal, Centre for Ecology and Hydrology, UK, personal communication, Spetember 2010).
22
SKB TR-13-26
5
Chlorine transformation processes
Recent results show that most of the soil chlorination is driven by enzymes and thereby organisms, but
abiotic chlorination at significant rates seems also to have occurred (Bastviken et al. 2009). Other studies indicate that besides the dominating biotically driven Cl– retention there is support for additional
abiotic processes related to iron cycling in soils (Keppler et al. 2000, Fahimi et al. 2003). However,
since the redox cycling of iron is usually a consequence of microbial activity, the proposed abiotic
processes may be indirectly linked to biological processes. It is also clear that chlorination capacity is
widespread among various groups of organisms including bacteria, fungi, algae, insects, mosses, and
vascular plants (e.g. Clutterbuck et al. 1940, Hunter et al. 1987, de Jong and Field 1997, Öberg 2002).
The chlorination of organic matter can occur both inside and outside cells. The intracellular chlorination seems strictly regulated by enzymatic processes. Enzymes known to mediate the intracellular
chlorination include enzyme groups named FADH2-dependent halogenases and perhydrolases (FADH2
is a cofactor necessary for the enzyme function). The underlying process for the extracellular chlorination seems to be a formation of reactive chlorine (e.g. hypochlorous acid, HOCl), from reactions
between hydrogen peroxide and Cl–. The reactive Cl is a strong oxidant and reacts with surrounding
organic matter which renders an unspecific chlorination of various organic compounds in the large and
complex pool of soil organic matter (Hoekstra 1999, van Pée and Unversucht 2003). Even though
the extracellular chlorination thereby is less rigorously controlled by enzymes, it still depends on
enzymes such as heme and vanadium containing haloperoxidases for the production of reactive
chlorine (van Pée 2001).
Given the rapid chlorination rates (i.e. the rapid Cl– retention) in soil, the high abundance of organochlorines, and the widespread capacity among organisms to chlorinate organic material there should
be a fundamental ecological explanation for the organic matter chlorination. Yet, it is still unknown
why such processes occur in soil. Intracellular chlorination processes have been explained as ways of
detoxification or are believed to represent production of compounds serving as chemical defense (e.g.
antibiotics), hormones, or pheromones (Hoekstra 1999). However, direct verification of these hypotheses is limited. Extracellular chlorination represents a different process, although it is well documented
that reactive chlorine species such as hypochlorous acid are potent bactericides used by phagocytes to
kill invading microorganisms (e.g. Apel and Hirt 2004) and that many microorganisms and plants produce allomones, i.e. substances that deter or kill competing or pathogenic organisms. Hence, the ability
to use reactive chlorine in the chemical warfare between competing microorganisms could provide a
substantial advantage and become a general strategy. In support of this, one screening of genetic databases found that many of the identified haloperoxidases from terrestrial environments are originating
from organisms that are associated with living plants or decomposing plant material (Bengtson et al.
2009). Hence, the ability to produce reactive chlorine could be especially common in environments that
are known for antibiotic-mediated competition for resources (interference competition). Yet, the ability
to produce haloperoxidases is also recorded, for example, for plant endosymbionts and parasites, and
there is little or no empirical evidence that suggests that these organisms are antagonistic.
Another hypothesis relate to microbial processing of organic material representing their substrates.
There is a general perception that chlorinated organic matter is less bioavailable than non-chlorinated
organic compounds. However, chloroperoxidases, like many other oxidases, catalyse production of
small reactive molecules (hypochlorous acid in the case of chloroperoxidase) that can break C-C
bonds in complex and refractory organic compounds (Hoekstra 1999, van Pée and Unversucht 2003).
Thereby, smaller and more bioavailable parts of the refractory compounds may be formed. In support
of this it has been shown that exposure of lignin to reactive chlorine enhance its biodegradability
(Johansson et al. 2000), and that fungal chloroperoxidase activity resulted in depolymerization and
breakdown of synthetic lignin (Ortiz-Bermúdez et al. 2003). Similarly, biodegradation of chlorinated
bleachery effluent lignin were greater than the degradation of corresponding chlorine-free lignin
(Bergbauer and Eggert 1994). After dechlorination these compounds should be highly preferred as
substrates by microorganisms. Hence, promoting formation of Clorg could be a way of increasing
the organic substrate supply for microorganisms.
SKB TR-13-2623
A third potential reason for the chlorination could be connected to defense against oxygen radicals.
Formation of reactive chlorine is related to consumption and thereby detoxification of reactive oxygen
species including hydrogen peroxide and oxygen radicals. Therefore extracellular chlorperoxidase
mediated formation of e.g. hypochlorous acid that can be prevented from entering the cell may form
one line of defence against reactive oxygen species. Interestingly, and in support of this hypothesis,
repeated oxidative stress exposure have been found to induce the expression of chloroperoxidase genes
and increase the production of reactive chlorine in some algae and bacteria (Bengtson et al. 2013).
As explained above the first rate estimates of chlorination in soils are now available, and there are
several hypotheses, all with some support, regarding the reasons for the chlorination. However,
it is still not known how the chlorination is regulated and how environmental variables influence
chlorination rates. Tests with different nitrogen levels have yielded ambiguous results (Rodstedth
et al. 2003, Bastviken et al. 2006), and local variability seems large. A few studies have found that
rates are slower under anoxic conditions (Bastviken et al. 2009), which is reasonable given that
chlorination or organic matter is an oxidative process. This points at an indirect regulation of soil
moisture, but apart from that the regulation of natural chlorination is still unclear. Chlorination
rates in sediments and other environments (e.g. in plants) is yet unknown as well.
Recent findings indicate that chlorinated compounds can be used as terminal electron acceptors
in microbial metabolism. Interestingly, the Gibbs free energy yield of this process is similar to the
energy yield with nitrate as the electron acceptor, and thereby only slightly lower than the energy
yield of oxic respiration (Smidt and de Vos 2004). Hence, chlorinated organic compounds can be
very potent as electron acceptors. Dechlorination could therefore be the result of either degradation
of the chlorinated organic matter or the microbial use of chlorinated organic molecules as electron
acceptors (i.e. halorespiration) and there is a wide literature regarding dehalogenation processes
in terms of biochemistry and related to specific compounds (e.g. Pries et al. 1994, Fetzner 1998,
Dolfing 2000, Olivas et al. 2002, van Pée and Unversucht 2003, Smidt and de Vos 2004).
However, the rates and regulation of dechlorination of bulk Clorg in nature is still unknown.
24
SKB TR-13-26
6
Chlorine in organisms
6.1
General uptake by plants and microorganisms
Growing plants rapidly take up large amounts of Cl–. The ratio of Cl– concentrations in fresh plant
tissue to concentrations per dry mass in the top 20 cm of soil ranged from 1.5 to 305 for common
agricultural plants (Kashparov et al. 2007b). Similar ranges were found by Kashparov et al. (2007a;
Table 6-1). This is in line with the high proportion of total catchment Cl found in biomass from the
Forsmark area (60%, Tröjbom and Grolander 2010) and a Danish forest (10%, Öberg and Grøn 1998);
both sites dominated by coniferous forest. Further, a Cl– uptake grossly exceeding cellular needs by
pine trees followed by rapid leaching, and thereby an extensive internal cyclingthrough the biomass
has been proposed (Van den Hoof and Thiry 2012).
Recent evidence also indicates that soil microorganisms rapidly take up Cl– during growth phases.
In the only known experiment studying this so far 20% of the 36Cl added to experiment soil was
incorporated into microbial biomass within 5 days (Bastviken et al. 2007). It was suggested that
rapid microbial growth following as system disturbance (e.g. a rain event, leaf fall in autumn etc)
could lead to rapid microbial uptake of Cl– based on physiological need. It is unclear if this can
affect Clorg formation rates.
6.2
Cl compounds with potentially long residence time in biota
The purpose of this section is to identify possible mechanisms for 36Cl to become “resident“ in biota.
Today a large number of natural products containing organically bound halogens are known (Gribble
2003, 2010). In terrestrial vascular plants alone, a couple of hundred chlorinated compounds have
been identified (Gribble 2010), but many of these are relatively short lived with a specific function,
such as the chlorinated auxins, “death hormones“, that trigger senescence (Engvild 1986). For the
purpose of this review we exclude such more or less ephemeral compounds as they do not undergo
transfer through the food chain and likely have little effects on anything but their target organism. Of
interest for this review are compounds that have characteristics that allow them to accumulate and have
a long half-life in biota, and which may be transferred between organisms in the food chain. The reason
for this interest is to identify possible compounds that might contribute to a prolonged exposure of
organisms to radiation from 36Cl. Thus, focus will be on pools of chlorine that may have a long half-life
in organisms rather than chloride ions that in comparison undergo a fairly rapid turnover.
Halogenated natural products (Drechsel 1896) and enzymatic oxidation of halide ions (Chodat and
Bach 1902) have been known for over a century. However, essentially due to analytical limitations
the understanding of the role of halogenated natural products was limited until the middle of the
20th century. Thus, when environmental chemistry began as a discipline in its own right in the early
1960s following the publication of “Silent Spring“ (Carson 1962), many environmentalists were
unaware of the occurrence of halogenated natural products (Mu et al. 1997), and there were much
discussion about what possible unknown anthropogenic pollutants proliferated in the environment
as only 5–10% of the extractable organically bound chlorine (EOCl) could be accounted for by
known compounds (Bernes and Naylor 1998).
Table 6-1. Average concentration ratio (CR; concentrations in fresh plants divided with
concentrations in dried soil in the upper 20 cm layer) of Cl– in food compared to in surface
soils on the area supporting the food production. Adopted from Kashparov et al. (2007a).
Food
CR
Leafy vegetables
Fruit vegetables
Root vegetables
Beans
Straw
Fruits
Milk and beef
28–71
17–18
8–98
16–20
149–237
0.6–4
48–57
SKB TR-13-2625
The past two decades has seen the identification of an increased number of halogenated compounds
that have structural similarities with the well-known persistent organic pollutants (POPs). POPs
include the chlorinated pesticides and polychlorinated biphenyls and were the first compounds to
become of interest to environmental chemists. Some of the compounds that have come under interest
recently, specifically methoxylated polybrominated diphenyl ethers (MeO-PBDEs), occur at very high
concentrations at high trophic levels (Mwevura et al. 2010, Alonso et al. 2012), and by analyzing the
carbon isotopes it has been shown that they are natural products. A number of compounds that contain
both chlorine and bromine have also been identified, and although less seems to be known about
these than the MeO-PDBEs it is likely that they are natural products. A number of mixed brominated
chlorinated bipyrroles and terpenoids have also been identified as likely persistent and bioaccumulating
natural products (Tittlemier et al. 1999, Rosenfelder and Vetter 2012). It is generally believed that these
brominated and mixed brominated chlorinated compounds are produced at low trophic levels in algae
and sponges, and are used as allomones, e.g. to deter grazing animals (Gribble 2010). As is clearly
shown by the very high concentrations of MeO-PBDEs that may occur in cetaceans (Mwevura et al.
2010, Alonso et al. 2012) these compounds undergo biomagnification and are persistent with a long
half-life in biota. Although less is known about, e.g. mixed brominated chlorinated dibenzo-p-dioxins
(Unger et al. 2009), bipyrroles, and terpenoids, these have such physical-chemical properties that they
generally should undergo biomagnification. If such compounds are formed in the presence of 36Cl–,
the turnover of 36Cl in the organism will to some degree depend on the half-life of these compounds.
Although these compounds may accumulate to high concentrations, especially in homeothermic
organisms at high trophic levels, the major halogen is bromine with limited contribution of chlorine.
Thus, these compounds may be of relatively little importance as pools of 36Cl that will be persistent in
biota. However, ongoing work on persistent halogenated natural products should be followed as new
compounds in which chlorine plays a more central role may be found in unexpected environments.
Several types of halogenated carboxylic acids1 are known as natural products (Dembitsky and
Srebnik 2002). Most of these, however, are structurally complex, often containing reactive functional
groups that shorten their half-lives. There are, however, some polyhalogenated homosequiterpene
acids that potentially are persistent, but we have found no studies that investigate food chain
transfer and accumulation actually takes place with these complex acids.
The one group of halogenated carboxylic acids that have been shown to have a long half-life and
undergo food-chain transfer is the chlorinated fatty acids (ClFAs). These compounds become especially
interesting as they do exert toxicity, but they are not generally recognized by biota as xenobiotics
(Ewald 1998). ClFAs were identified as constituting the bulk of EOCl (up to 70% or even more) in
environmental samples in the 1990s (Håkansson et al. 1991, Mu et al. 1997). To date the origin of
ClFAs is not fully understood. There clearly are anthropogenic influences as EOCl concentrations are
particularly high close to point sources. An important type of point source was identified as production
of chlorine bleached pulp and paper (Håkansson et al. 1991). During chlorine bleaching direct chlorination of unsaturated FAs takes place, and ClFAs were identified in the effluent from bleaching facilities.
Some reports also indicate that the ClFA-profile depends on the type of anthropogenic pollutants
present (Vereskuns 1999). In spite of these caveats on the concentrations and profile, ClFAs are, as
far as known today, ubiquitous and present also in areas where the concentrations of anthropogenic
POPs are low. It is also noteworthy that away from point source of chlorinated organic material EOCl
is relatively evenly distributed in the aquatic environment (Bernes and Naylor 1998). In view of this
general distribution of EOCl, a de novo synthesis of ClFAs in organisms, also those that are not subject
to any anthropogenic source of organically bound chlorine, is likely. However, although formation
of CLFAs via chloroperoxidase action have been shown in some marine organisms (Mu et al. 1997),
a general cause for the presence of CLFAs in organisms away from point sources appears not yet
to have been identified. Food chain transfer and biomagnification of ClFAs has been shown in wild
populations of fish (Mu et al. 2004), and in laboratory experiments from chironomids (midges) to
fish (Björn 1999).
Some authors, e.g. Dembitsky and Srebnik (2002), call even very complex molecules “fatty acids” as long as
they contain a carboxylic functional group. We prefer to reserve the term “fatty acids” to such carboxylic acids
that actually may occur in lipids.
1 26
SKB TR-13-26
In contrast to traditional POPs, ClFAs do not primarily accumulate in fatty tissues, but are incorporated
in membrane lipids. Thus the concentrations of ClFAs relative other fatty acids (FAs) are higher in
muscle tissues than in depot fat, at least in mammals (Björn 1999, Åkesson Nilsson 2004). To date
there is no information of any specific role of ClFAs in membrane lipids, but the three-dimensional
structure of ClFAs is, at least in some cases, very similar to unsaturated FAs; e.g. threo-9,10-dichlorooctadecanoic acid (threo-9,10-dichlorostearic acid) has a three-dimensional structure similar to cis9-octadecenoic acid (oleic acid), which has a double bond between carbons 9 and 10. Thus, ClFAs may
fill similar functions as unsaturated FAs in membrane lipids. However, there are also reports suggesting
that a high proportion of ClFAs in the membranes causes membrane leakage (Gunilla Åkesson Nilsson,
personal communication).
ClFAs are metabolized via β-oxidation, but the metabolic process seems to stop at chlorinated
myristic acid (tetradecanoic acid) (Mu et al. 1997). Indeed, dichloromyristic acid accumulates
in some systems after addition of longer chained ClFAs (Mu et al. 1997). This accumulation of
dichloromyristic acid is not fully understood, but seemingly the enzymes responsible for β-oxidation
are sterically hindered by the presence of the chlorine atoms. A further implication of this is that
ClFAs may be more recalcitrant in biota than other FAs. Thus, incorporation of 36Cl into FAs may
lead to longer exposure time to radiation than would otherwise be expected, but the present lack of
data prevents quantification of this effect.
Of special interest for a discussion on human exposure to radiation from 36Cl is that dichloromyristic
acid accumulates also in human systems (Gustafsson-Svärd et al. 2001) and there is a possibility of
de-novo synthesis of ClFAs in inflammatory processes (Åkesson Nilsson 2004). The results as yet
are preliminary, but the wider implication is that if an individual that for some reason has an ongoing
inflammatory process is exposed to 36Cl, this radioisotope could be incorporated into compounds that
are recalcitrant in the organism. Further, preliminary experiments have indicated that ClFAs may be
particularly prone to accumulate in the myelin sheath of nervous tissue (Kylin, unpublished), why
targeted damage to the nervous system cannot be excluded.
SKB TR-13-2627
7
Techniques for studying Cl in the environment
This section will first describe various analysis techniques that can be used to study Cl in the
environments. This is followed by a discussion regarding sampling and preservation. Thereafter
examples of approaches to study Cl related processes are provided. Finally key aspects of Cl
monitoring in ecosystems will be addressed.
7.1
Ion chromatography
Cl in water samples can be determined by e.g. ion chromatography or ion-selective electrodes. Ion
chromatograpy with chemical suppression (such as MIC-2, Metrohm), according to standard procedure
for determination of Cl– of water with low contamination (ISO 10304-1:1992) is frequently used.
In short, water samples are filtered (0.15 µm filter, Metrohm) and separated on an anion column and
detected with a conductivity meter.
–
7.2AOX
Water samples are typically analyzed for Clorg as AOX (adsorbable organic halogen; ISO 9562:2004).
In short, the organic compounds in the water sample is adsorbed on activated carbon, and washed
with an acidified nitrate solution to remove remaining chloride. Thereafter the sample is combusted
at 1,000°C in an oxygen atmosphere and the formed halides is titrated with silver ions with microcoulometric titration with an AOX instrument (e.g. ECS3000, Euroglas). This AOX method measures the
sum of chlorine, bromide, and iodine but does not distinguish between the different halogens. Since Cl
is by far the most abundant of these halogens in the soil environment (Brady and Weil 2002), the mass
estimates are based on the assumption that Cl dominates in the samples. Therefore, Clorg is used as the
sum of organically bound chlorine and if other halides are present in considerable amounts, the AOX
method will overestimate the Clorg in the samples. There were previous concerns that chloride might
interfere with the AOX measurements and thereby overestimate the Clorg, but such interference is not
significant unless there are high Cl– concentrations such as in brackish waters (Asplund et al. 1994).
Furher tests of how to perform AOX analyses in high Cl– samples may be necessary.
The AOX method has been adapted for other types of bulk Clorg analyses including extractable
organic halogens (EOX) and volatile organic halogens (VOX). Apart from the pretreatments
the combustion and coulometric detection is similar. Two other adaptions of the AOX method is
described in the below section.
7.3
TOX and TX
The concentration of Clorg (TOX) in solid samples is now frequently analyzed according to Asplund
et al. (1994). This method is similar to the method of analyzing sludge or contaminated soils and
sediments. In short, a dry sieved and milled soil sample is washed with an acidic nitrate solution to
remove remaining chloride. The residual soil is combusted and TOX is detected according to the
procedure for AOX analysis described above. A recent suggestion, based on results that significant
amounts of Cl– can be trapped inside microbial cells in soils (Bastviken et al. 2007), is to also have
a sonication step before the final washing, but if the milling is done thoroughly this may disrupt most
cells using the original approach. The total amount of chlorine (TX) can be determined by adding
sieved and milled soil to a small crucible followed by direct combustion in the AOX instrument. Cl–
can then be calculated as TX minus TOX and this represents an alternative method of determining
chloride in soil rather than measuring Cl– in water soil leachate (Johansson et al. 2001). There is a
risk of overestimation of chloride by using “TX minus TOX” because TX analyzes not only Clorg and
porewater chloride but possibly also mineral chloride and this has to be considered when deciding
methods. Additional methods that has been used for studies of Clorg, although requiring much more
SKB TR-13-2629
expensive equipment, are X-ray absorption spectroscopy (XAS: e.g. Reina et al. 2004) and neutron
activation (Xu et al. 2004). Redon et al. (2013) cross checked Clorg analyses of mineral soil samples
by neutron activation and by the TOX method and found no significant differences (unpublished
data), but additional cross comparisons of several techniques for different types of samples would
be beneficial.
7.4VOCl
There are several difficulties in estimating VOCls including low ambient air concentrations, and
sampling and storage without contamination. The concentrations of VOCls are in the lower ppt
range, compared to carbon dioxide (CO2), which is found in hundreds of ppm levels in the atmosphere. Despite the low concentrations, there are now instruments capable of analyzing VOCls in
ppq-levels (Dobrzyńska et al. 2010). Therefore, the analytical challenges for VOCl detection are now
more related to sample collection (enabling a reliable and large enough sample amount for detection)
and storage of samples (without loss or contamination of the compounds of interest) than to the
analytical detection in the laboratory. For measuring VOCl soil-atmosphere exchange there will also
be additional difficulties such as related to sampling methodology in the field (capturing the VOCl
soil-atmosphere exchange) and estimating the VOCl fluxes (calculating the flux from the observed
soil-atmosphere exchange).
There are some recent papers presenting analytical methods and sample preparation for determining
VOCls in air samples (e.g. Dobrzyńska et al. 2010, Ramírez et al. 2010, Dewulf and Van Langenhove
2006, Demeestere et al. 2007). As with all trace gases, it is essential to take precautions during
sampling, storage, and analysis to avoid loss of analytes. Despite high-precision instruments, the
concentrations of VOCls in the air are so low that sample pre-concentration is typically needed to
be able to detect the compounds of interest. Methods commonly used for the determination of VOCl
concentrations in air are based on either adsorption onto a sorbent column followed by thermal desorption with subsequent analysis using gas chromatography (GC) (e.g. Valtanen et al. 2009, Hoekstra
et al. 2001) or on cryogenic concentration of VOCls directly from a sampled volume of air followed
by GC (e.g. Rhew et al. 2008). The disadvantages of the sorption tubes are that sorption and desorption
efficiencies may not be 100%, which may limit the detection limit for samples at low concentrations. In addition, storage of sorbent tubes before desorption and analysis can result in losses of e.g.
chloroform. Determination of VOCls using methods based on cryogenic trapping can be influenced
by moisture condensation in the trap. The other way of sampling VOCls in air is to use canisters. This
has been successfully done by many researchers (e.g. Redeker et al. 2003). Evacuated canisters used
to collect air samples in the field for transport to the laboratory must be carefully cleaned to avoid
contamination of the sample. Since canister sampling is not a pre-concentration technique, it requires
a subsequent sorbent and/or cryogenic concentration stage in the laboratory (Wang and Austin 2006).
The more recent technique, solid phase microextraction (SPME) has been shown to be useful for the
determination of volatiles in air (Dobrzyńska et al. 2010). This technique is based upon the absorption
of chloroform into a polymer coated on a silica fiber. Following equilibration of the fiber with the
atmosphere, the volatile is released via thermal desorption in the injection port of a gas chromatograph.
Sample preparation is very easy and fast with this technique although sample collection conditions
and thermal desorption conditions must be carefully controlled for the best precision. The canisters are
difficult to use if a large amount of samples is needed, due to canisters’ large volume and limited space
available for storage and transportation. In the reviewed papers there seems to be almost equal numbers
using sorbents or canister techniques. There is only one study that has used SPME (Yassaa et al. 2009).
7.5
XSD and AED to identify Cl compounds in complex matrices
An interesting technique to identify chlorinated organic compounds in a complex matrix is the
halogen selective detector (XSD) for GC (Åkesson Nilsson et al. 2001, Åkesson Nilsson 2004).
The selectivity of this detector is much higher than other GC detectors why unknown compounds
containing halogens can be identified in a complex matrix. The XSD can be used either to select
chromatographic peaks that should be identified with high-resolution spectrometric techniques, or
30
SKB TR-13-26
to monitor for known compounds in complex matrices that would otherwise require more expensive
and laborious methods, e.g. GC-high resolution mass spectrometry (MS). Thus, the sample throughput can be increased.
GC coupled to an atomic emission detector (AED), in which a microwave induced plasma is used
to excite the elements, has also been used to identify halogenated compounds in the environment
(Flodin et al. 1997a, Hjelm et al. 1996, Laniewski et al. 1998, 1999), including to select peaks for
characterization with MS (Flodin et al. 1997b, Johansson et al. 1994). The AED will give information about which type of halogen atom the individual peak contains, and it is also possible to identify
other non-carbon elements in the compounds. In contrast, the XSD will only give information that
the peak contains an unspecified halogen. The AED is, however, more expensive and complex
to operate, with a lower sample throughput as potential consequence. With some very complex
mixtures it may also be less selective than the XSD.
7.6ICP
Several techniques for the simultaneous determination of multiple elements have been built around
the inductively coupled plasma (ICP). Traditionally, the ICP has been used to excite the elements for
atomic emission spectroscopy (AES), also called optical emission spectroscopy (OES). However,
until recently ICP-AES instruments have not been well suited for the determination of halogens; the
low emission wavelengths of the halogens have not been compatible with the instrument configuration and the limits of detection have not compared favourably with other techniques (Strellis et al.
1996, Teledyne Leeman Labs 2013). Recently, developments of instrument configurations and detectors have made it possible to efficiently include the halogens in multi-element determinations by
ICP-AES (Teledyne Leeman Labs 2013, Naozuka et al. 2003, Oliveira et al. 2012), and the number
of scientific publications using this technique is growing.
Coupling the ICP with MS (ICP-MS) provides a powerful technique to not only determine the
elements, but individual stable isotopes (Barnes and Straub 2010, Pereira Barbosa et al. 2013).
During the 2000s applications in which ICP-MS has been coupled to chromatography that allows
characterization of the elemental and isotopic composition of individual compounds (González-Gago
et al. 2007, Kretschy et al. 2012) have become increasingly common. It is also possible to couple
ICP-MS with, e.g. laser ablation to determine the elemental and isotopic composition of very small
targets areas (Fietzke et al. 2008, Seo et al. 2011).
7.7
Stable chlorine isotopes
Isotopes of an element participate in the same chemical reactions but react at slightly different rates
due to their difference in mass. This difference in reaction rates between the isotopes are referred to as
isotopic fractionation and usually result in discrimination against the heavier isotope in the reactions.
This fractionation can differ widely between different reactions e.g. depending on what enzymes
catalyze the reaction. Therefore comparisons of stable isotope ratios before and after reactions can
be very useful for identifying roles and contributions of different processes in elemental cycles. This
approach is extensively used for light elements where the relative mass difference between stable
isotopes is large, such as H, C and N. The mass difference between stable Cl isotopes is smaller but
the Cl stable isotope ratio, e.g. 37Cl proportion of the total Cl expressed as
δ37Cl (‰) = (Rsample/Rreference – 1) · 1000
where R is the abundance ratio of 37Cl/35Cl in the sample and reference material, respectively, have
recently shown promising for separating natural and anthropogenic sources of some chlorinated
compounds (Holmstrand et al. 2010). In addition there are indications that chlorination catalyzed by
chlorperoxidases fractionate strongly with reaction products having δ37Cl values of –12‰ (Aeppli
et al. 2013, Reddy et al. 2002), while Clorg in fish and polar bear fat have δ37Cl values closer to reference values (–1 to –3‰: Wassenaar and Koehler 2004).
SKB TR-13-2631
This field is in its early phase and analyses are demanding in terms of sample preparation (Cincinelli
et al. 2012). Most efforts have targeted individual compounds (Aeppli et al. 2010). One approach
described for bulk Cl is based on the following steps (Wassenaar and Koehler 2004):
1. Separation of Cl– and Clorg (by e.g. extraction or adsorption techniques).
2.Clorg samples are converted to Cl– by combustion.
3. The Cl– is precipitated to AgCl.
4. AgCl is converted to methyl chloride (CH3Cl) by reaction with iodomethane.
5. The CH3Cl in injected in a multicollector continuous-flow isotope ratio mass spectrometry
system (CF-IRMS).
Other approaches outlined in Cincinelli et al. (2012) include various GC-MS, GC-IRMS (gas
chromatography isotope ratio mass spectrometry) and TIMS (thermal ionization mass spectrometry)
techniques ensuring that sample sizes and mass resolution are large enough. ICP-based techniques
have also been proposed (Fietzke et al. 2008).
7.8
Sample preparation and preservation
The procedures for sample preparation and possible preservation differ widely depending on the specific analysis of interest. While e.g. sampling and sample handling for measuring VOCl or specific
Clorg compounds is highly demanding due to e.g. low concentrations, the sample procedures for bulk
Cl samples are easier and straight-forward. To completely cover all these aspects is beyond the scope
of this work, and we will here just highlight a few aspects related to sampling and preservation.
Regarding bulk Cl analyses, there is increasing awareness that organic and inorganic Cl pools are
both large and behave very differently in the environment, and should therefore preferably be separated in the analyses. Many of the analyses are specific in their detection, e.g. ion chromatography
detects only Cl– and not Clorg while AOX only detect adsorbable Clorg. However, when making
pretreatments for bulk analysis such as TOX, TX or ICP, details may be important. For example if
fresh soil is extracted with water (at low pH and with nitrate to facilitate ion exchange to liberate
Cl–), it is likely that the soil will not release the Cl– present in microbial cells (Bastviken et al. 2007).
To fully extract most of the Cl– it is therefore better to dry, freeze, and/or sonicate the soil to break up
microbial cells before the extraction.
In terms of preservation, this again strictly depends on the analyses. For the high concentration Cl
pools (Cl– and bulk Clorg) it has become common practice to freeze liquid samples and dry solid
phase samples which are often milled prior to analysis. However, effects of drying or freeze-thaw
cycles on analysis results have seemingly not been systematically evaluated.
7.9
Process rate studies
Studying process rates are important for mass balance studies to determine turnover rates and
residence times. Further, to allow predictive modeling knowledge about how critical processes and
their rates are regulated or influenced by environmental variables is often critical. A number of
fundamental approaches for process studies have been used for Cl as outlined below.
A classical way to study processes such as chlorination, dechlorination, Cl uptake or release, has been
to monitor changes in concentrations over time in a confined sample or system. This approach applied
to leaf litter bags was behind breakthrough findings such as simultaneous loss and gain of Cl– and
Clorg, respectively in decaying litter (Myneni 2002), and for studies of the fates of Cl in soil cores
(Rodstedth et al. 2003, Bastviken et al. 2006). This approach depends on concentration changes that
are large enough to be detected during the time frames in focus and if net changes are of interest.
32
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For processes near steady state, or that occur slowly, other approaches have to be used. This applies
for example to soil organic matter chlorination and dechlorination, which can occur simultaneously
possibly resulting in a small or slow net change in spite of highly active and relatively rapid
processes. In such cases tracer studies become an important tool. In the case of Cl, radiotracer
experiments with 36Cl provide a highly sensitive way to measure gross transformation rates. This
approach was used to confirm soil organic matter chlorination and determine rates by adding 36Cl–
to soils and measure the formation of 36Clorg (Bastviken et al. 2007, Gustavsson et al. 2012) as well
as for studying the temperature sensitivity of soil organic chlorination (Bastviken et al. 2009).
At the ecosystem scale, mass balance modelling by combining concentration measurements with
material fluxes in the ecosystem is the most common way to estimate Cl cycling. It is powerful in
its logical simplicity and provides god overviews of fluxes and stocks. Ideally, to support predictive
modelling and study influence of environmental variables such modelling should also have a time
dimension based on regular repeated measurements, and preferable possibilities to make sub-scale
experimental treatments, although this is rarely possible because of resource constraints. To date,
there seems to be no such mass balance studies with repeated sampling of both Cl– and Clorg pools,
which severely limits the possibilities of predictive modelling.
SKB TR-13-2633
8
Future challenges
The research about natural Cl cycling is still in its infancy, especially in the light of recent findings
of the importance of the Clorg pool in the terrestrial environment. Therefore several fundamental
questions remain. Some of these are discussed below based on the interest to understand transport,
uptake and/or exposure, and residence times of 36Cl in terrestrial and aquatic ecosystems. For all
these questions it is clear that Cl– and Clorg pools can be expected to behave very different, e.g. in terms
of solubility, bioavailability, residence times. Therefore inorganic and organic Cl pools have to be
considered separately in order to understand and model the dynamics of Cl in landscapes over time.
8.1
What is the spatio-temporal variability of Cl– and Clorg
distribution in landscapes?
This is a very fundamental question and includes inputs, standing stocks and outputs from varius
types of landscape units. Previous studies, generally non-repeated snapshot measurements, have
revealed differences between soil types, as well as local variations within sample plots (Johansson
et al. 2003a, b). Currently, there are no systematic assessments of the temporal variability at different
spatial scales. Such measurements are a key asset for all further studies including; i) the design of
sampling programs, ii) validate data comparability from different locations or collected at different
times, and iii) accurately assessing data uncertainty in models. It is important that seasonal patterns
are considered to test for data comparability across seasons. Although seemingly simple and basic,
such studies are fundamental for building appropriate environmental models.
8.2
Which conditions and processes control Cl– and Clorg levels
and transport?
The regulation of Cl– and Clorg levels are key topics for long term modelling with ambitions to
include varying environmental conditions. Fundamental environmental variables such as water
content, organic matter content, primary productivity, dominating vegetation, nutrient levels, temperature, etc could have large impacts on the relative distribution of Cl– and Clorg, in turn affecting
bioavailability, transport pathways and residence times in ecosystems. In addition, a better understanding of why so many organisms perform chlorination of organic matter would greatly facilitate
modelling of chlorination rates, i.e. what is the adaptational value of biological chlorination?
Dechlorination rates, i.e. release of organically bound Cl from the organic matter, of bulk soil Clorg
is at present unknown. This means that we do not know if the Clorg we observe in soils represent primarily a dynamic pool that is turned over rapidly or if the measured soil Clorg pool largely consists of
very stable compounds that have slowly accumulated. Therefore, rates and regulation of dechlorination are important to understand and predict the fate of Clorg. Persistence versus lability of the Clorg
is a related question that determines bioavailability and residence times. From a modelling point of
view an interesting question is if Clorg can be modelled as average organic matter. The indications of
increasing Cl:C ratio with material age indicates that this is not the case, but data is very limited at
this point.
In terms of transport, circumstantial evidence presented above highlight that the emissions of VOCl
from various landscape compartments needs to be better quantified. Several measurements of individual compounds exist, but new techniques now make studies on the total VOCl emissions possible.
Circumstantial evidence as outlined above indicates that total VOCl emissions could be a substantial
part of the export of chlorine from terrestrial ecosystems, but this remains to be confirmed with
additional measurements.
SKB TR-13-2635
Most of the questions regarding the regulation of Cl– and Clorg levels and transport are hitherto not
addressed and will require experimental assessments. However, other approaches can also provide
valuable information. For example better understanding of chemical composition and structure of
Clorg would greatly facilitate interpretations regarding how chlorination and dechlorination can occur
from a chemical perspective, what enzymes are likely to be involved and thereby under what conditions various processes may occur.
8.3
Cl cycling in inland waters
Soil and ground water discharge areas, e.g. wetlands, streams and lakes, are likely recipients of subsurface sources of 36Cl in humid climate zones. At present, while Cl– measurements are common in
aquatic monitoring, there is, to our knowledge, no data regarding Cl cycling of the separate Cl– and
Clorg pools.
8.4
How to model Cl cycling in terrestrial environments?
As explained above, Cl cyling models need information about both Cl– and Clorg. The presently
available data for both Cl– and Clorg is typically based on single measurement campaign “snap-shot”
information from a limited number of environments. These data support static mass balance models
only, models that are restricted by a number of assumptions, including the steady-state assumption,
or that Clorg behaves as organic matter. Further, model uncertainty is difficult to assess given the poor
knowledge concerning spatial and temporal variability. Also, the unexpected indications that large
parts of the total catchment Cl can be found in biomass show that many of our previous beliefs
guiding our model assumptions need to be re-evaluated.
Thus, without data covering the temporal dimension for both Cl– and Clorg separately along with
relevant environmental variables in sediment, water, soil and biomass pools, and without information
of how major processes (e.g. chlorination and dechlorination) is regulated, long-term dynamic models
have to be based on educated guesswork.
36
SKB TR-13-26
References
SKB’s (Svensk Kärnbränslehantering AB) publications can be found at www.skb.se/publications.
Aeppli C, Holmstrand H, Andersson P, Gustafsson Ö, 2010. Direct compound-specific stable
chlorine isotope analysis of organic compounds with quadrupole GC/MS using standard isotope
bracketing. Analytical Chemistry 82, 420–426.
Aeppli C, Bastviken D, Andersson P, Gustafsson Ö, 2013. Chlorine isotope effects and composition
of naturally produced organochlorines from chloroperoxidases, flavin-dependent halogenases, and in
forest soil. Environmental Science & Technology 47, 6864−6871.
Apel K, Hirt H, 2004. Reactive oxygen species: metabolism, oxidative stress, and signal transduction.
Annual Reviews of Plant Biology 55, 373–399.
Albers C N, Jacobsen O S, Flores É M M, Pereira J S F, Laier T, 2011. Spatial variation in natural
formation of chloroform in the soils of four coniferous forests. Biogeochemistry 103, 317–334.
Alonso M B, Eljarrat E, Gorga M, Secchi E R, Bassoi M, Barbosa L, Bertozzi C P, Marigo J,
Cremer M, Domit C, Azevedo A F, Dorneles P R, Torres J P M, Lailson-Brito J, Malm O,
Barceló D, 2012. Natural and anthropogenically-produced brominated compounds in endemic
dolphins from Western South Atlantic: another risk to a vulnerable species. Environmental
Pollution 170, 152–160.
Asplund G, Grimvall A, 1991. Organohalogens in nature, more widespread than prevoiusly
assumed. Environmental Science & Technology 25, 1347–1350.
Asplund G, Grimvall A, Jonsson S, 1994. Determination of the total and leachable amounts of
organohalogens in soil. Chemosphere 28, 1467–1475.
Aze T, Fujimura M, Matsumura H, Masumoto K, Nakao N, Matsuzaki H, Nagai H, Kawai M,
2007. Measurement of the production rates of 36Cl from Cl, K, and Ca in concrete at the 500 MeV
neutron irradiation facility at KENS. Journal of Radioanalytical and Nuclear Chemistry 3, 491–494.
Barnes J D, Stroub S M, 2010. Chlorine stable isotope variations in Izu Bonin tephra: implications
for serpentinite subduction. Chemical Geology 272, 62–74.
Bastviken D, Sandén P, Svensson T, Ståhlberg C, Magounakis M, Öberg G, 2006. Chloride
retention and release in a boreal forest soil: effects of soil water residence time and nitrogen and
chloride loads. Environmental Science & Technology 40, 2977–2982.
Bastviken D, Thomsen F, Svensson T, Karlsson S, Sandén P, Shaw G, Matucha M, Öberg G,
2007. Chloride retention in forest soil by microbial uptake and by natural chlorination of organic
matter. Geochimica et Cosmochimica Acta 71, 3182–3192.
Bastviken D, Svensson T, Karlsson S, Sandén P, Öberg G, 2009. Temperature sensitivity indicates
that chlorination of organic matter in forest soil is primarily biotic. Environmental Science &
Technology 43, 3569–3573.
Bengtson P, Bastviken D, de Boer W, Öberg G, 2009. Possible role of reactive chlorine in
microbial antagonism and organic matter chlorination in terrestrial environments. Environmental
Microbiology 11, 1330–1339.
Bengtson P, Bastviken D, Öberg G, 2013. Possible roles of reactive chlorine II: assessing biotic
chlorination as a way for organisms to handle oxygen stress. Environmental Microbiology 15,
991–1000.
Bergbauer M, Eggert C, 1994. Degradability of chlorine-free bleachery effluent lignins by two
fungi: effects on subunit type and on polymer molecular weight. Canadian Journal of Microbiology
40, 192–197.
Bernes C, Naylor M, 1998. Persistent organic pollutants: a Swedish view of an international
problem. Stockholm: Swedish Environmental Protection Agency.
Bertills U, 1995. Grundvattnets kemi i Sverige. Solna: Naturvårdsverket. (In Swedish.)
SKB TR-13-2637
Biester H, Keppler F, Putschew A, Martinez-Cortizas A, Petri M, 2004. Halogen retention,
organohalogens and the role of organic matter decomposition on halogen enrichment in two Chilean
peat bogs. Environmental Science & Technology 38, 1984–1991.
Björn H, 1999. Uptake, turnover and distribution of chlorinated fatty acids in aquatic biota. PhD
thesis. Lund University, Sweden.
Brady N, Weil R, 2002. The nature and properties of soils. 13th ed. Upper Saddle River, NJ:
Prentice Hall.
Broyer T C, Carlton A B, Johnson C M, Stout P R, 1954. Chlorine – a micronutrient element for
higher plants. Plant Physiology 29, 526–532.
Bäckström M, Karlsson S, Bäckman L, Folkeson L, Lind B, 2004. Mobilisation of heavy metals
by deicing salts in a roadside environment. Water Research 38, 720–32.
Campbell K, Wolfsberg A, Fabryka-Martin J, Sweetkind D, 2003. Chlorine-36 data at Yucca
Mountain: Statistical tests of conceptual models for unsaturated-zone flow. Journal of Contaminant
Hydrology 62–63, 43–61.
Carson R, 1962. Silent spring. Boston: Houghton Mifflin.
Castro M S, Morgan R P, 2000. Input-output budgets of major ions for a forested watershed in
western Maryland. Water, Air, and Soil Pollution 119, 121–137.
Chodat R, Baach A, 1902. Untersuchungen über die Rolle der Peroxyde in der Chemie der
lebenden Zelle. Erste Mittheilung: Über das Verhalten der lebenden Zelle gegen Hydroperoxyd.
Berichte der deutschen chemischen Gesellschaft 35, 1275–1279.
Cincinelli A, Pieri F, Zhang Y, Seed M, Jones K C, 2012. Compound Specific Isotope Analysis
(CSIA) for chlorine and bromine: a review of techniques and applications to elucidate environmental
sources and processes. Environmental Pollution 169, 112–127.
Clarke N, Fuksová K, Gryndler M, Lachmanová Z, Liste H H, Rohlenová J, Schroll R,
Schröder P, Matucha M, 2009. The formation and fate of chlorinated organic substances in
temperate and boreal forest soils. Environmental Science and Pollution Research 16, 127–143.
Clutterbuck P W, Mukhopadhyay S L, Oxford A E, Raistrick H, 1940. Studies in the bio­
chemistry of micro-organisms. The Biochemical Journal 34, 664–677.
de Jong E, Field J A, 1997. Sulfur tuft and turkey tail: biosynthesis and biodegradation of organo­
halogenes by basidiomycetes. Annals Reviews of Microbiololgy 51, 375–414.
Dembitsky V M, Srebnik M, 2002. Natural halogenated fatty acids: their analogues and derivatives.
Progress in Lipid Research 41, 315–367.
Demeestere K, Dewulf J, De Witte B, Van Langenhove H, 2007. Sample preparation for the
analysis of volatile organic compounds in air and water matrices. Journal of Chromatography A
1153, 130–144.
Dewulf J, Van Langenhove H, 2006. Developments in the analysis of volatile halogenated
compounds. TrAC Trends in Analytical Chemistry 25, 300–309.
Dimmer C, Simmonds P, Nickless G, Bassford M, 2001. Biogenic fluxes of halomethanes from
Irish peatland ecosystems. Atmospheric Environment 35, 321–330.
Dobrzyńska E, Pośniak M, Szewczyńska M, Buszewski B, 2010. Chlorinated volatile organic
compounds – old, however, actual analytical and toxicological problem. Critical Reviews in
Analytical Chemistry 40, 41–57.
Dolfing J, 2000. Energetics of anaerobic degradation pathways of chlorinated aliphatic compounds.
Microbial Ecology 40, 2–7.
Drechsel E, 1896. Beiträge zur Chemie einiger Seetiere. Zeitschrift für Biologie 33, 85–107.
Enell M, Wennberg L, 1991. Distribution of halogenated organic-compounds (AOX) – Swedish
transport to surrounding sea areas and mass balance studies in 5 drainage systems. Water Science &
Technology 24, 385–395.
38
SKB TR-13-26
Engel R E, Bruebaker L, Emborg T J, 2001. Chloride deficient leaf spot of durum wheat. Soil
Science Society of America Journal 65, 1448–1454.
Engvild K C, 1986. Chlorine-containing natural compounds in higher plants. Phytochemistry 25,
781–791.
Eriksson E, 1955. Air borne salts and the chemical composition of river waters. Tellus 7, 243–250.
Eriksson E, 1960. The yearly circulation of chloride and sulfur in nature; metoerological, geo­
chemical and pedological implications. Part II. Tellus 12, 63–109.
Ewald G, 1998. Chlorinated fatty acids – environmental pollutants with intriguing properties.
Chemosphere 37, 2833–2837.
Fahimi I J, Keppler F, Schöler H F, 2003. Formation of chloroacetic acids from soil, humic acid
and phenolic moieties. Chemosphere 52, 513–520.
Fetzner S, 1998. Bacterial dehalogenation. Applied Microbiology and Biotechnology 50, 633–657.
Fietzke J, Frische M, Hansteen T H, Eisenhauer A, 2008. A simplified procedure for the determination of stable chlorine isotope ratios (δ37Cl) using LS-MC-ICP-MS. Journal of Analytical Atomic
Spectrometry 23, 769–772.
Flodin C, Johansson E, Borén H, Grimvall A, Dahlman O, Mörck R, 1997a. Chlorinated
structures in high molecular weight organic matter isolated from fresh and decaying plant material
and soil. Environmental Science & Technology 31, 2464–2468.
Flodin C, Ekelund M, Borén H, Grimvall A, 1997b. Pyrolysis-GC/AED and pyrolysis-GC/
MS analysis of chlorinated structures in aquatic fulvic acids and chlorolignins. Chemosphere 34,
2319–2328.
Fréchou C, Degros J-P, 2005. Measurement of 36Cl in nuclear wastes and effluents: validation of
a radiochemical protocol with an in-house reference sample. Journal of Radioanalytical and Nuclear
Chemistry 263, 333–339.
Godduhn A, Duffy L K, 2003. Multi-generation health risks of persistent organic pollution in the
far north: use of the precautionary approach in the Stockholm Convention. Environmental Science &
Policy 6, 341–353.
González-Gago A, Manuel Marchante-Gayón J, García Alonzo J I, 2007. Determination of
trihalomethanes in drinking water with GC-ICP-MS using compound independent calibration with
internal standard. Journal of Analytical Atomic Spectrometry 22, 1138–1144.
Graedel T E, Keene W C, 1996. The budget and cycle of Earth’s natural chlorine. Pure and Applied
Chemistry 68, 1689–1697.
Gribble G W, 2003. The diversity of naturally produced organohalogens. Chemosphere 52,
289–297.
Gribble G W, 2010. Naturally occurring organohalogen compounds – a comprehensive update.
Wien: Springer.
Grimvall A, Borén H, Jonsson S, Karlsson S, Sävenhed R, 1991. Organohalogens of natural
and industrial origin in large recipients of bleach-plant effluents. Water Science & Technology 24,
373–383.
Grøn C, 1995. AOX in groundwater. In Grimvall A E, de Leer E W B (eds). Naturally-produced
organohalogens. Dordrecht: Kluwer Academic Publishers, 49–64.
Gustavsson G, Karlsson S, Öberg G, Sandén P, Svensson T, Valinia S, Thiry Y, Bastviken D,
2012. Organic matter chlorination rates in different boreal soils: the role of soil organic matter
content. Environmental Science & Technology 46, 1504–1510.
Gustafsson-Svärd C, Åkesson-Nilsson G, Mattsson M, Sundin P, Wesén C, 2001. Removal of
xenobiotic dichlorostearic acid from phospholipids and neutral lipids in cultured human cell lines
by β-oxidation and secretion of dichloromyristic acid. Pharmacology and Toxicology 89, 56–64.
Haselmann K F, Laturnus F, Grøn C, 2002. Formation of chloroform in soil. A year-round study
at a Danish spruce forest site. Water, Air, and Soil Pollution 139, 35–41.
SKB TR-13-2639
Hellén H, Hakola H, Pystynen K-H, Rinne J, Haapanala S, 2006. C2-C10 hydrocarbon emissions
from a boreal wetland and forest floor. Biogeosciences 3, 167–174.
Herczeg A L, Leaney F W, 2011. Review: Environmental tracers in arid-zone hydrology.
Hydrogeology Journal 19, 17–29.
Hjelm O, Asplund G, 1995. Chemical characterization of organohalogens in a coniferous forest
soil. In Grimvall A E, de Leer E W B (eds). Naturally-produced organohalogens. Dordrecht: Kluwer
Academic Publishers, 105–111.
Hjelm O, Borén H, Öberg G, 1996. Analysis of halogenated organic compounds in coniferous
forest soil from a Lepista nuda (wood blewitt) fairy ring. Chemosphere 32, 1719–1728.
Hoekstra E J, 1999. On the natural formation of chlorinated organic compounds in soil: chloroform,
trichloroacetic acid and chlorinated phenols, dibenzo-p-dioxins, dibenzofurans. PhD thesis. Dept. of
Analytical Chemistry, TNO Institute of Environmental Sciences, Free University of Amsterdam.
Hoekstra E J, Duyzer J H, de Leer E W B, Brinkman U A T, 2001. Chloroform – concentration
gradients in soil air and atmospheric air, and emission fluxes from soil. Atmospheric Environment
35, 61–70.
Holmstrand H, Zencak Z, Mandalakis M, Andersson P, Gustafsson Ö, 2010. Chlorine isotope
evidence for the anthropogenic origin of tris-(4-chlorophenyl)methane. Applied Geochemistry 25,
1301–1306.
Hou X, Frøsig L, Nielsen S P, 2007. Determination of 36Cl in nuclear waste from reactor decommissioning. Analytical Chemistry 79, 3126–3134.
Hruška J, Oulehle F, Šamonil P, Šebesta J, Tahovská K, Hleb R, Houška J, Šikl J, 2012. Longterm forest soil acidification, nutrient leaching and vegetation development: linking modelling and
surveys of a primeval spruce forest in the Ukrainian Transcarpathian Mts. Ecological modeling 244,
28–37.
Hultberg H. Grennfelt P, 1992. Sulfur and seasalt deposition as reflected by throughfall and runoff
chemistry in forested catchments. Environmental Pollution 75, 215–222.
Hunter J, Belt A, Sotos L, Fonda M, 1987. Fungal chloroperoxidase method. United States
Patent 4,707,447.
Håkansson H, Sundin P, Andersson T, Brunström B, Dencker L, Engwall M, Ewald G, Gilek M,
Holm G, Honkasalo S, Idestam-Almquist J, Jonsson P, Kautsky N, Lundberg G, LundKvemheim A, Martinsen K, Norrgren L, Personen M, Rundgren M, Stålberg M, Tarkpea M,
Wesén C, 1991. In vivo and in vitro toxicity of fractionated fish lipids, with particular regard to
their content of chlorinated organic compounds. Pharmacology and Toxicology 69, 459–471.
ISO 9562:2004. Water quality – Determination of adsorbable organically bound halogens (AOX).
Geneva: International Organization for Standardization.
ISO 10304-1:1992. Water quality – Determination of dissolved fluoride, chloride, nitrite, ortophosphate, bromide, nitrate and sulfate ions, using liquid chromatography of ions – Part 1: Method for
water with low contamination. Geneva: International Organization for Standardization.
Johansson C, Pavasars I, Borén H, Grimvall A, Dahlman O, Mörck R, Reimann A, 1994.
A degradation procedure for determination of halogenated structural elements in organic matter
from marine sediments. Environment International 20, 103–111.
Johansson E, Krantz-Rülcker C, Zhang B X, Öberg G, 2000. Chlorination and biodegradation of
lignin. Soil biology and biochemistry 32, 1029–1032.
Johansson E, Ebenå G, Sandén P, Svensson T, Öberg G, 2001. Organic and inorganic chlorine in
Swedish spruce forest soil: influence of nitogen. Geoderma 101, 1–13.
Johansson E, Sandén P, Öberg G, 2003a. Spatial patterns of organic chlorine and chloride in
Swedish forest soil. Chemosphere 52, 391–397.
Johansson E, Sandén P, Öberg G, 2003b. Organic chlorine in deciduous and coniferous forest soils
in southern Sweden. Soil Science 168, 347–355.
40
SKB TR-13-26
Johansson E, Xin Z B, Zhengyi H, Sandén P, Öberg G, 2004. Organic chlorine and chloride in
submerged paddy soil: a case study in Anhui province, southeast China. Soil Use and Management
20, 144–149.
Kashparov V, Colle C, Levchuk S, Yoschenko V, Svydynuk N, 2007a. Transfer of chlorine from
the environment to agricultural foodstuffs. Journal of Environmental Radioactivity 94, 1–15.
Kashparov V, Colle C, Levchuk S, Yoschenko V, Zvarich S, 2007b. Radiochlorine concentration
ratios for agricultural plants in various soil conditions. Journal of Environmental Radioactivity 95,
10–22.
Keppler F, Eiden R, Niedan V, Pracht J, Schöler H F, 2000. Halocarbons produced by natural
oxidation processes during degradation of organic matter. Nature 403, 298–301.
Keppler F, Borchers R, Elsner P, Fahimi I, Pracht J, Schöler H F, 2003. Formation of volatile
iodinated alkanes in soil: results from laboratory studies. Chemosphere 52, 477–483.
Khalil M A K, Rasmussen R A, 2000. Soil-atmosphere exchange of radiatively and chemically
active gases. Environmental Science and Pollution Research 7, 79–82.
Khalil M A K, Rasmussen R A, Shearer M J, Chen Z-L, Yao H, Yang J, 1998. Emissions of
methane, nitroux oxide, and other trace gases from rice fields in China. Journal of Geophysical
Research 103, 25241–25250.
Kindbohm K, Svensson A, Sjöberg K, Pihl Karlsson G, 2001. Trends in air concentration and
deposition at background monitoring sites in Sweden – major inorganic compounds, heavy metals
and ozone. IVL report B 1429, IVL Swedish Environmental Research Institute.
Kirchner J W, Feng X, Neal C, 2000. Fractal stream chemistry and its implications for contaminant
transport in catchments. Nature 403, 524–527.
Krám P, Hruška J, Wenner B S, Driscoll C T, Johnson C E, 1997. The biogeochemistry of basic
cations in two forest catchments with contrasting lithology in the Czech Republic. Biogeochemistry
37, 173–202.
Kretschy D, Koellensperger G, Hann S, 2012. Elemental labelling combined with liquid
chromatography inductively coupled plasma mass spectrometry for quantification of biomolecules:
a review. Analytica Chimica Acta 750, 98–110.
Laniewski K, Borén H, Grimvall A, Jonsson S, Sydow L, 1995. Chemical characterization
of adsorbable organic halogens (AOX) in precipitation. In Grimvall A E, de Leer E W B (eds).
Naturally-produced organohalogens. Dordrecht: Kluwer Academic Publishers, 113–130.
Laniewski K, Borén H, Grimvall A, 1998. Identification of volatile and extractable chloroorganics
in rain and snow. Environmental Science & Technology 32, 3935–3940.
Laniewski K, Borén H, Grimvall A, 1999. Fractionation of halogenated organic matter present in
rain and snow. Chemosphere 38, 393–409.
Laturnus F, Lauritsen F R, Grøn C, 2000. Chloroform in a pristine aquifer system: toward
an evidence of biogenic origin. Water Resources Research 36, 2999–3009.
Lee R T, Shaw G, Wadey P, Wang X, 2001. Specific association of 36Cl with low molecular weight
humic substances in soils. Chemosphere 43, 1063–1070.
Leri A C, Myneni S C B, 2010. Organochlorine turnover in forest ecosystems: the missing link in the
terrestrial chlorine cycle. Global Biogeochemical Cycles 24, GB 4021. doi:10.1029/2010GB003882.
Limer L, Albrecht A, Bytwerk D, Marang L, Smith G, Thorne M, 2009. Cl-36 Phase 2:
dose assessment uncertainties and variability. Version 2.0 (Final). Report D.RP.CSTR.09.0026,
Andra, France. Available at:
http://www.bioprota.org/wp-content/uploads/2012/06/Chlorine-phase-2-dose-assessment.pdf
Löfgren S, 2001. The chemical effects of deicing salt on soil and stream water of five catchments
in southeast Sweden. Water, Air, and Soil Pollution 130, 863–868.
Lovett G M, Likens G E, Buso D C, Driscoll C T, Bailey S W, 2005. The biogeochemistry of
chlorine at Hubbard Brook, New Hampshire, USA. Biogeochemistry 72, 191–232.
SKB TR-13-2641
Matzner E (ed), 2004. Biogeochemistry of forested catchments in a changing environment:
a German case study. Berlin: Springer. (Ecological studies 172).
McCulloch A, 2003. Chloroform in the environment: occurrence, sources, sinks and effects.
Chemosphere 50, 1291–1308.
Mead M I, Khan M A H, Nickless G, Greally B R, Tainton D, Pitman T, Shallcross D E, 2008.
Leaf cutter ants: a possible missing source of biogenic halocarbons. Environmental Chemistry 5,
5–10.
Melkerud P-A, Olsson M T, Rosén K, 1992. Geochemical atlas of Swedish forest soils. Uppsala:
SLU. (Rapporter i skogsekologi och skoglig marklära 65).
Mu H, Wesén C, Sundin P, 1997. Halogenated fatty acids: I. Formation and occurrence in lipids.
Trends in Analytical Chemistry 16, 266–274.
Mu H, Ewald G, Nilsson E, Sundin P, Wesén C, 2004. Fate of chlorinated fatty acids in migrating
sockeye salmon and their transfer to arctic grayling. Environmental Science & Technology 38,
5548–5554.
Mwevura H, Amir O A, Kishimba M, Berggren P, Kylin H, 2010. Organohalogen contaminants
in blubber of Indo-Pacific bottlenose dolphin (Tursiops aduncus) and spinner dolphin (Stenella longirostris) from the coastal waters of Zanzibar, Tanzania. Environmental Pollution 158, 2200–2207.
Müller G, 2003. Sense or no-sense of the sum parameter for water soluble “adsorbable organic halogens” (AOX) and “absorbed organic halogens” (AOX-S18) for the assessment of organohalogens in
sludges and sediments. Chemosphere 52, 371–379.
Myneni S C, 2002. Formation of stable chlorinated hydrocarbons in weathering plant material.
Science 295, 1039–1041.
Naozuka J, Mesquita Silva da Veiga M A, Vitoriano Oliveira P, de Oliveira E, 2003.
Determination of chlorine, bromine and iodine in milk samples by ICP-OES. Journal of Analytical
Atomic Spetrometry 18, 917–921.
Neal C, Reynolds B, Neal M, Wickham H, Hill L, Williams B, 2004. The water quality of streams
draining a plantation forest on gley soils: the Nant Tanllwyth, Plynlimon mid-Wales. Hydrology and
Earth System Sciences 8, 485–502.
Neal C, Robinson D, Reynolds B, Neal M, Rowland P, Grant S, Norris D, Williams B, Sleep D,
Lawlor A, 2010. Hydrology and water quality of the headwaters of the River Severn: stream acidity
recovery and interactions with plantation forestry under an improving pollution climate. Science of
the Total Environment 408, 5035–5051.
Nyberg L, Rodhe A, Bishop K, 1999. Water transit times and flow paths from two line injections
of 3H and 36Cl in a microcatchment at Gårdsjön, Sweden. Hydrological processes 13, 1557–1575.
Olivas Y, Dolfing J, Smith G B, 2002. The influence of redox potential on the degradation of
halogenated methanes. Environmental Toxicology and Chemistry 21, 493–499.
Oliveira A A, Nóbrega J A, Pereira-Filho E R, Trevizan L C, 2012. Evaluation of ICP OES with
axial of radial views for determination of iodine in table salt. Quimica Nova 35, 1299–1302.
Ortiz-Bermúdez P, Srebotnik E, Hammel K E, 2003. Chlorination and cleavage of lignin structures
by fungal chloroperoxidases. Applied Environmental Microbiology 69, 5015–5018.
Pereira Barbosa J T, Moreira Santos C M, dos Santos Bispo L, Henrique Lyra F, David J M,
das Graças Andrade Korn M, Moraes Flores E M, 2013. Bromine, chlorine, and iodine determination in soybean and its products by ICP-MS after digestion using microwave-induced combustion.
Food Analytical Methods 6, 1065–1070.
Peters N E, Shanley J B, Aulenbach B T, Webb R M, Campbell D H, Hunt R, Larsen M C,
Stallard R F, Troester J, Walker J F, 2006. Water and solute mass balance of five small, relatively
undisturbed watersheds in the U.S. Science of the Total Environment 358, 221–243.
Peterson J, MacDonell M, Haroun L, Monette F, 2007. Radiological and chemical fact sheets
to support health risk analyses for contaminated areas. Available at:
http://www.remm.nlm.gov/ANL_ContaminantFactSheets_All_070418.pdf
42
SKB TR-13-26
Pickering L, Black T A, Gilbert C, Jeronimo M, Nesic Z, Pilz J, Svensson T, Öberg G, 2013.
Portable chamber system for measuring chloroform fluxes from terrestrial environments – methodological challenges. Enviromental Science & Technology 47, 14298–14305.
Pries F, van der Ploeg J R, Dolfing J, Janssen D B, 1994. Degradation of halogenated aliphatic
compounds: the role of adaptation. FEMS Microbiology Reviews 15, 279–295.
Probst A, Viville D, Fritz B, Ambroise B, Dambrine E, 1992. Hydrochemical budgets of a small
forested granitic catchment exposed to acid deposition: the Strengbach catchment case study (Vosges
massif, France). Water, Air, and Soil Pollution 62, 337–347.
Pöykiö R, Nurmesniemi H, Kivilinna V A, 2008. EOX concentrations in sediment in the part of
the Bothnian Bay affected by effluents from the pulp and paper mills at Kemi, Northern Finland.
Environmental Monitoring and Assessment 139, 183–194.
Ramírez N, Cuadras A, Rovira E, Borrull F, Marcé R M, 2010. Comparative study of solvent
extraction and thermal desorption methods for determining a wide range of volatile organic
compounds in ambient air. Talanta 82, 719–727.
Reddy C M, Xu L, Drenzek N J, Sturchio N C, Heraty L J, Kimblin C, Butler A, 2002. A chlorine
isotope effect for enzyme-catalyzed chlorination. Journal of the American Chemical Society 124,
14526–14527.
Redeker K R, Wang N, Low J C, McMillan A, Tyler S C, Cicerone R J, 2000. Emissions of
methyl halides and methane from rice paddies. Science 290, 966–969.
Redeker K R, Meinardi S, Blake D, Sass R, 2003. Gaseous emissions from flooded rice paddy
agriculture. Journal of Geophysical Research 108. doi:10.1029/2002JD002814.
Redon P-O, Abdelouas A, Bastviken D, Cecchini S, Nicolas M, Thiry Y, 2011. Chloride and
organic chlorine in forest soils: storage, residence times, and influence of ecological conditions.
Environmental Science & Technology 45, 7202–7208.
Redon P-O, Jolivet C, Saby N P A, Abdelouas A, Thiry Y, 2013. Occurrence of natural organic
chlorine in soils for different land uses. Biogeochemistry 114, 413–419.
Reina R G, Leri A C, Myneni S C B, 2004. Cl K-edge x-ray spectroscopic investigation of
enzymatic formation of organochlorines in weathering plant material. Environmental Science &
Technology 38, 783–789.
Rengasamy P, 2010. Soil processes affecting crop production in salt-affected soils. Functional Plant
Biology 37, 613–620.
Reynolds B, Fowler D, Smith R I, Hall J R, 1997. Atmospheric inputs and catchment solute fluxes
for major ions in five Welsh upland catchments. Journal of Hydrology 194, 305–329.
Rhew R C, Miller B R, Weiss R F, 2000. Natural methyl bromide and methyl chloride emissions
from coastal salt marshes. Nature 403, 292–295.
Rhew R C, Miller B R, Bill M, Goldstein A H, Weiss R F, 2002. Environmental and biological
controls on methyl halide emissions from southern California coastal salt marshes. Biogeochemistry
60, 141–161.
Rhew R C, Teh Y A, Abel T, Atwood A, Mazéas O, 2008. Chloroform emissions from the Alaskan
Arctic tundra. Geophysical Research Letters 35. doi:10.1029/2008GL035762.
Rodríguez M, Piña G, Lara E. 2006. Radiochemical analysis of chlorine-36. Czechoslovak Journal
of Physics 56, D211–D217.
Rodstedth M, Ståhlberg C, Sandén P, Öberg G, 2003. Chloride imbalances in soil lysimeters.
Chemosphere 52, 381–389.
Rosenfelder N, Vetter W, 2012. Stable carbon isotope composition (δ13C values) of the halogenated
monoterpene MHC-1 as found in fish and seaweed from different marine regions. Journal of
Environmental Monitoring 14, 845–851.
Rustad L E, Kahl J S, Norton S A, Fernandez I J, 1994. Underestimation of dry deposition by
throughfall in mixed northern hardwood forests. Journal of hydrology 162, 319–336.
SKB TR-13-2643
Schlesinger W, 1997. Biogeochemistry: an analysis of global change. 2nd ed. San Diego, CA:
Academic Press.
Schleyer R, 1996. Beeinflussing der Grundwasserqualität durch Deposition anthropogener
organischer Stoffe aus der Atmosphäre. Langen: Institut für Wasser-, Boden- und Lufthygiene.
Schleyer R, Renner I, Muehlhausen D, 1991. Beeinflussung der Grundwasserqualität durch
luftgetragene organische Schadstoffe. Ergebnisse eines vom Umweltbundesamt geförderten
Forschungsvorhabens. Langen: Institut für Wasser-, Boden- und Lufthygiene.
Seki R, Matsuhiro T, Nagashima Y, Takahashi T, Sasa K, Sueki K, Tosaki Y, Bessho K,
Matsumura H, Miura T, 2007. Isotopic ratios of 36Cl/Cl in Japanese surface soil. Nuclear
Instruments and Methods in Physics Research Section B 259, 486–490.
Seo H J, Guillong M, Aerts M, Zajacz S, Heinrich C A, 2011. Microanalysis of S, Cl, and Br
in fluid inclusions by LA-ICP-MS. Chemical Geology 284, 35–44.
Sheppard S C, Johnson L H, Goodwin B W, Tait J C, Wuschke D M, Davison C C, 1996.
Chlorine-36 in nuclear waste disposal – 1. Assessment results for used fuel with comparison to 129I
and 14C. Waste Management 16, 607–614.
Silk P J, Lonergan G C, Arsenault T L, Boyle C D, 1997. Evidence of natural organochlorine
formation in peat bogs. Chemosphere 35, 2865–2880.
Smidt H, de Vos W M, 2004. Anaerobic microbial dehalogenation. Annual Review of Microbiology
58, 43–73.
Strellis D A, Hwang H H, Anderson T F, Landsberger S, 1996. A comparative study of IC,
ICP-AES, and NAA measurements of chlorine, bromine, and sodium in natural waters. Journal of
Radioanalytical and Nuclear Chemistry 211, 473–484.
Stringer R, Johnston P, 2001. Chlorine and the environment: an overview of the chlorine industry.
Dordrecht: Kluwer Academic Publishers.
Suominen K P, Jaakkola T, Elomaa E, Hakulinen R, Salkinoja-Salonen M S, 1997. Sediment
accumulation of organic halogens in pristine forest lakes. Environmental Science and Pollution
Research 4, 21–30.
Svensson T, Laturnus F, Sandén P, Öberg G, 2007a. Chloroform in runoff water – a two-year
study in a small catchment in Southeast Sweden. Biogeochemistry 82, 139–151.
Svensson T, Sandén P, Bastviken B, Öberg G, 2007b. Chlorine transport in a small catchment in
southeast Sweden during two years. Biogeochemistry 82, 181–199.
Svensson T, Lovett G M, Likens G E, 2012. Is chloride a conservative ion in forest ecosystems?
Biogeochemistry 107, 125–134.
Swank W T, Crossley D A (eds), 1988. Forest hydrology and ecology at Coweeta. New York:
Springer. (Ecological studies 66).
Teledyne Leeman Labs, 2013. Determination of the halogen elements in the deep UV region of
the spectrum by ICP-OES. Technical Note. Hudson, NH: Teledyne Leeman Labs.
Tittlemier S A, Simon M, Jarman W M, Elliott J E, Norstrom R J, 1999. Identification of a novel
C10H6N2Br4Cl2 heterocyclic compound in seabird eggs. A bioaccumulating marine natural product?
Environmental Science & Technology 33, 26–33.
Todd A K, Kaltenecker M G, 2012. Warm season chloride concentrations in stream habitats of
freshwater mussel species at risk. Environmental Pollution 171, 199–206.
Tröjbom M, Grolander S, 2010. Chemical conditions in present and future ecosystems in Forsmark
– implications for selected radionuclides in the safety assessment SR-Site. SKB R-10-27, Svensk
Kärnbränslehantering AB.
Tröjbom M, Nordén S, 2010. Chemistry data from surface ecosystems in Forsmark and LaxemarSimpevarp. Site specific data used for estimation of CR and Kd values in SR-Site. SKB R-10-28,
Svensk Kärnbränslehantering AB.
44
SKB TR-13-26
Tröjbom M, Söderbäck B, Kalinowski B, 2008. Hydrochemistry of surface water and shallow
groundwater. Site descriptive modeling, SDM-Site Laxemar. SKB R-08-46, Svensk Kärnbränsle­
hantering AB.
Unger M, Asplund L, Haglund P, Malmvärn A, Arnoldsson K, Gustafsson Ö, 2009.
Polybrominated and mixed brominated/chlorinated dibenzo-p-dioxins in sponge (Ephydatia
fluviatilis) from the Baltic Sea. Environmental Science & Technology 43, 8245–8250.
Valtanen, A, Solloch S, Hartikainen H, Michaelis W, 2009. Emissions of volatile halogenated
compounds from a meadow in a coastal area of the Baltic Sea. Boreal Environment Research 14,
915–931.
Van den Hoof C, Thiry Y, 2012. Modelling of the natural chlorine cycling in a coniferous
stand: implications for chlorine-36 behaviour in a contaminated forest environment. Journal of
Environmental Radioactivity 107, 56–67.
Van Meter R J, Swan C M, Leips J, Snodgrass J W, 2011. Road salt stress induces novel food
web structure and interactions. Wetlands 31, 843–851.
van Pée K H, 2001. Microbial synthesis of halometabolites. Archieves of Microbiology 175,
250–258.
van Pée K-H, Unversucht S, 2003. Biological dehalogenation and halogenation reactions.
Chemosphere 52, 299–312.
Varner R K, Crill P M, Talbot R W, 1999. Wetlands: a potentially significant source of atmospheric methyl bromide and methyl chloride. Geophysical Research Letters 26, 2433–2435.
Varner R K, White M L, Mosedale C H, Crill P M, 2003. Production of methyl bromide in
a temperate forest soil. Geophysical Research Letters 30, 1521. doi: 10.1029/2002GL016592.
Vereskuns G, 1999. Chlorinated fatty acids in freshwater fish and some biological effects of
dichlorostearic acid. PhD thesis. Swedish University of Agricultural Sciences, Uppsala.
von Sydow L, Borén H, Grimvall A, 1999. Chloroacetates in snow, firn and glacier ice.
Chemosphere 39, 2479–2488.
Wang D K W, Austin C C, 2006. Determination of complex mixtures of volatile organic compounds
in ambient air: canister methodology. Analytical and Bioanalytical Chemistry 386, 1099–1120.
Wang J, Qin P, Sun S, 2007. The flux of chloroform and tetrachloromethane along an elevational
gradient of a coastal salt marsh, East China. Environmental Pollution 148, 10–20.
Wassenaar L I, Koehler G, 2004. On-line technique for the determination of the δ37Cl of inorganic
and total organic Cl in environmental samples. Analytical Chemistry 76, 6384–6388.
White P J, Broadley M R, 2001. Chloride in soils and its uptake and movement within the plant:
a review. Annals of Botany 88, 967–988.
Wilander A, 1997. Referenssjöarnas vattenkemi under 12 år: tillstånd och trender. Stockholm,
Naturvårdsverket. (In Swedish.)
Winterton N, 2000. Chlorine: the only green element – towards a wider acceptance of its role in
natural cycles. Green Chemistry 2, 173–225.
Xu S, Leri A C, Myneni S C B, Jaffe P R, 2004. Uptake of bromide by two wetland plants (Typha
latifolia L. and Phragmites australis (Cav.) Trin. ex Steud). Environmental Science & Technology
38, 5642–5648.
Yassaa N, Wishkerman A, Keppler F, Williams J, 2009. Fast determination of methyl chloride and
methyl bromide emissions from dried plant matter and soil samples using HS-SPME and GC-MS:
method and first results. Environmental Chemistry 6, 311–318.
Yokouchi Y, Ikeda M, Inuzuka Y, Yukawa T, 2002. Strong emission of methyl chloride from
tropical plants. Nature 416, 163–165.
Yunos N M, Bellomo R, Story D, Kellum J, 2010. Bench-to-bedside review: chloride in critical
illness. Critical Care 14. doi:10.1186/cc9052.
SKB TR-13-2645
Åkesson Nilsson G, 2004. Determination of chlorinated fatty acids using SPE, XSD and GC/MS
with particular regard to cultured human cells. PhD Thesis. Swedish University of Agricultural
Sciences, Uppsala, Sweden.
Åkesson Nilsson G, Nilsson O, Odenbrand I, Wesén C, 2001. New halogen-specific detector
applied to the analysis of chlorinated fatty acids. Journal of Chromatography A 912, 99–106.
Öberg G, 2002. The natural chlorine cycle – fitting the scattered pieces. Applied Microbiology and
Biotechnology 58, 565–581.
Öberg G, Bäckstrand K, 1996. Conceptualization of the acidification theory in Swedish environmental research. Environmental Reviews 4, 123–132.
Öberg G, Grøn C, 1998. Sources of organic halogens in a Danish spruce forest soil. Environmental
Science & Technology 32, 1573–1579.
Öberg G, Sandén P, 2005. Retention of chloride in soil and cycling of organic matter-bound
chlorine. Hydrological Processes 19, 2123–2136.
Öberg G, Johansen C, Grøn C, 1998. Organic halogens in spruce forest throughfall. Chemosphere
36, 1689–1701.
Öberg G, Holm M, Sandén P, Svensson T, Parikka M, 2005. The role of organic-matter-bound
chlorine in the chlorine cycle: a case study of the Stubbetorp catchment, Sweden. Biogeochemistry
75, 241–269.
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