REPORT
The environmental fate of
linear alkylbenzene
sulphonates (LAS) and other
anionic surfactants in
sediments in the city of
Stockholm and its inner
archipelago.
Using manganese speciation as a tool for assessing the role of surface
sediment redox conditions for degradation of anionic surfactants.
Organization
IVL Swedish Environmental Research Institute Ltd.
Report Summary
Project title
Address
P.O. Box 21060
SE-100 31 Stockholm
Project sponsor
Telephone
+46 (0)8-598 563 00
Author
Title and subtitle of the report
.
Summary
.
Keyword
Bibliographic data
IVL Report
The report can be ordered via
Homepage: www.ivl.se, e-mail: publicationservice@ivl.se, fax+46 (0)8-598 563 90, or via IVL, P.O. Box 21060,
SE-100 31 Stockholm Sweden
.
IVL report
Contents
. ......................................................................................................Error! Bookmark not defined.
Introduction ........................................................................................................................................2
LAS ..................................................................................................................................................2
Other anionic surfactants .............................................................................................................3
Manganese as an indicator of past redox conditions in surface sediments ...........................4
Biogeochemistry of manganese in sediments ........................................................................4
Aim ..................................................................................................................................................8
Materials and methods ......................................................................................................................8
Sampling ..........................................................................................................................................8
Chemical analysis ............................................................................................................................ 10
Sediment dating ........................................................................................................................... 10
Sediment redox conditions ........................................................................................................ 10
Samples selected for the analysis of LAS and other anionic surfactants ............................ 11
Results............................................................................................................................................... 13
Sediment characteristics/dating ................................................................................................ 13
Sediment redox conditions ........................................................................................................ 13
Anionic surfactant concentrations ........................................................................................... 16
LAS ........................................................................................................................................... 16
Discussion ........................................................................................................................................ 23
Sediment characteristics and redox conditions ...................................................................... 23
Lake Mälaren ........................................................................................................................... 23
Saltsjön ..................................................................................................................................... 24
Torsbyfjärden .......................................................................................................................... 25
Trends between stations ........................................................................................................ 26
Spatial distribution of linear alkyl sulfonates .......................................................................... 27
Vertical Profiles and aerobic versus anaerobic degradation ................................................. 29
Ecotoxicity ....................................................................................................................................... 31
Conclusions...................................................................................................................................... 33
References ........................................................................................................................................ 34
Appendix .......................................................................................................................................... 39
1
.
IVL report
Introduction
Surfactants are vital components of detergent formulations. Given the globally extensive
use of washing detergents, it is not surprising that surfactants are among the most widely
used chemicals in the world with a worldwide market of 24.33 billion US$ in 2009 (Acmite,
2010). The two dominating types of surfactants are nonionic and anionic, together making
up of around two thirds of the global market value for surfactants (Acmite, 2010). The
market value of nonionic surfactants has during the last 10 years become larger than that of
anionic surfactants, and is now the largest of all surfactants whereas nonionic share of the
world market was around 30% in 2009(Acmite, 2010). Global demand for anionic
surfactants was 6.5 million tons in 2010 tons (Ceresana Research, 2012). The main use of
surfactants is domestic use, around 70% as household cleaners and detergents and 9.5% as
body care products and cosmetics in 2010 (Ceresana Research, 2012). Thus for these
compounds wastewater is the main source to aquatic environment. Irrespective if it is in
the marine, estuarine or lacustrine environment into which the surfactants are discharged, a
relative rapid degradation in the water column follows. Because of the surfactants´ high
affinity for organic carbon, the remaining load is eventually incorporated into the
sediments. Accordingly one key property when assessing environmental risks of surfactants
is biodegradation in water and sediment. Particularly, the ability of the surfactant to
undergo biodegradation under all prevailing sediment redox conditions is of great interest.
Since degradation in oxygenated sediment pore water is as rapid as in the water column, the
focus has been on biodegradation under anoxic and anaerobic conditions. Anoxic
conditions will prevail immediately below the oxic and suboxic layer, which exist in the top
few millimeters if no bioturbation of the sediments occur and if transport of oxygen only
occurs through diffusion from the bottom water. With the presence bioturbating meio and
macro fauna, the pore water of the top centimeters of the sediments will be oxygenated.
Eventually, with increasing sediment depth, strict anaerobic conditions exist. The question
to what extent one of the main surfactants, linear alkylbenzene sulphonates (LAS) is
biodegradable under anaerobic conditions, has attracted some attention from the scientific
community, policymakers and environmental organizations, particularly in Sweden.
Another key property is the toxicity of the surfactant itself and that of the metabolites
produced during its degradation. However, several other aspects have to be regarded if one
wants to assess the combined environmental impact of the whole life cycle of the
surfactant. Some of these aspects include: washing efficiency, production of raw materials
for the surfactant, production itself and properties important to the operations of
wastewater treatment plants. Considering the vast volumes of surfactants used and their
widespread usage, the selection of cleaning agent has to be selected carefully to ensure
minimum impact on the total environment.
LAS
Linear alkylbenzene sulphonates (LAS) is globally the primary cleaning agent used in many
washing detergents and cleaners. Detergents based on LAS were developed during the
1960’s to replace the branched tetrapropylenbenzensulfonates (TPBS), that due to high
persitance caused considerable aggregations of foam in the water recipients. The LAS
molecule contains an aromatic ring sulphonated at the para position and attached to a
2
.
IVL report
linear alkyl chain that typically has 10 to 14 carbon atoms. Concerning LAS, an extensive
database of studies demonstrates rapid and complete (ultimate) biodegradation of LAS in
many of the available aerobic biodegradation tests, including soil and the aqueous
environment. In several tests, LAS has been shown to be readily biodegradable, and has
passed the 10-day biodegradation window in mineralization tests for most ready tests. LAS
is removed in biological wastewater treatment at percentages ranging from 77- 82% for
trickling filters up to 99% for activated sludge (OECD 2005).
During the 80’s 10000-15000 tonnes of LAS were used per year in detergents in Sweden
(KEMI, 2012a). However, ecolabeling organizations, such as Good Environmental Choise
(Bra Miljöval), Nordic Ecolabel (Svanen) and EU-Ecolabel have not apporved LAS in
laundry detergents based on a too low anaerob degradation rate. The critera set are, for
instance, that ”60% of the cleaning agent must be anaerobially degraded according to ISO
11734 or another equivalent test”. (Naturskyddsföreningen, 2006) or ”All surfactants must
be anarobically and aerobially biodegradable” (Nordic Ecolabelling, 2011). Sweden is the
only country where ecolabeling of detergnts has considered the anaerobic degradation as
the determining criteria for acceptance. As a consequence, in Sweden the use of LAS
decreased with 95% between the years 1989 and 1999, compared to a decrease of 15% in
other European countries. The Swedish Society for Nature Conservation estimated a
decrease in usage of LAS in household chemicals from 6300 tonnes for the year 1988 to
260 tonnes in 1996 (SSNC, 1999). The import of LAS in all chemical products has
increased during recent years, from 407 tonnes in 2001 to 1048 tonnes in 2007 (KEMI,
2012a). LAS is also imported as raw material (248-564 tonnes per year) and produced in
Sweden. Information on the total amount of raw material produced and used in Sweden is
however not publicly available.
In other European contries as well as USA, South America and Asia, LAS is the primary
cleaning agent due to its advantages compared to other anionic surfactants. LAS enables
the possility to lower the washing temperature from 60 ºC to 30 ºC for most type of
materials, that consequently saves energy for the households. With the use of LAS a smaller
dose of washing detergent per volume of laundry can be used. A lower consumption of
detergent will in turn reduce the transports which is a gain for the climate. However, LAS
is manifactured through sulphonication of linear alkylbenzene (LAB). LAB is derived from
benzene and linear paraffins both of which are petroleum derivatives. Thus LAS is not
from a renewable source but from a petrochemical feedstock.
Other anionic surfactants
For the industry there is no rationallity in developing specific detergents that does not
contain linear alkylbenznene sulphonates LAS for the Swedish, and possibly the Finish,
market. Furthermore, the impact on third coutries needs to be considered. The anionic
surfactants used in detergents on the Swedish market are primarly the alkyl ether sulphates
(AES) which are sodiums salts of alkyl polyethylene ether sulphates e.g. sodium lauryl ether
sulphate(SLES) and alkyl ether sulphates, e.g. sodium dodecyl sulphates (SDS)/sodium
lauryl sulphates (SLS)and soft soap. These surfactants can also be derived from different
3
.
IVL report
types of oleochemical feedstock, generally palm oil and palm oil derivatives, pine needle oil
or from animal residues and can, thus, come from renewable sources. However, the
production of palm oil and palm oil derivatives have resulted in large global environmental
consequences. To give space for palm oil plantages large areas of tropical rain forests in
Indonesia (Borneo), Philipines and Malaysia have been devastated. The destruction of the
rain forests have adverse impact on both the biodiveristy of species and the climate. At
present only a few percent of the production of palm oil producion in the world is certified
according to Roundtable on Sustainable Palm Oil (RSPO). In addition, an increased
demand for palm oil will compete with the local food production and may make Indonesia,
Malaysia and the Philipines more dependet on imported food. The properties for the
anionic surfactants used are not as known as for LAS, e.g. concerning toxicity and
bioaccumulation (Woldegiorgis, 2011). These surfactants, e.g. AES, are easily degraded in
anaerobic standard tests but still detected in both freshwater and marine sediments (LaraMartin et al., 2006).
For sodium dodecyl sulphate and sodium dodecyl ether sulphate usage in household
products was estimated to 1600 tonnes in 1988 and 5600 tonnes in 1996 (SSNC, 1999).
The import of sodium dodecyl sulphate in chemical products was 2191 tonnes in 2007
(KEMI, 2012b), of sodium dodecyl ether sulphate, 430 tonnes in 2009 (KEMI, 2012c), and
of coco amidopropyl betaine, 289 tonnes in 2005 (KEMI, 2012d).
Manganese as an indicator of past and present redox
conditions in surface sediments
By studying the changes in speciation (different chemical forms) of Mn with sediment
depth, information on past bottom water and surface sediment redox conditions will be
gained (Calvert et al., 2001 Morford et al., 2001). In order to understand how Mn
speciation can be used, a summary of the biogeochemistry of manganese in sediments
follows.
Biogeochemistry of manganese in sediments
A schematic figure of Mn-cycle in marine sediments is shown in Fig 1 below.
4
.
IVL report
Fig. 1. Schematic figure of Mn –fluxes in marine sediments (Wang and Van Cappellen, 1996). The sizes of the
arrows represent the magnitude of the fluxes. The relative sizes of these fluxes depend on a number
of factors, e.g. whether there is bioturbation or not and salinity (marine, brackish, lacustrine), and
can therefore vary from location to location.
Dissolved Mn is scavenged out of the oxygenated bottom water and deposited to the
sediment surface by the adsorption to FeOOH and other particles (Turner and Millward,
2000) and by the adsorption to and precipitation of oxyhydroxides, MnOOH) and oxides
MnO2 (Reed et al., 2011). Manganese oxyhydroxides and oxides are however mainly formed
in the oxic upper part of the sediment from upward diffusing dissolved manganese, Mn2+
(Berner, 1980; Mouret et al., 2009). The dissolved manganese is also adsorbed to the
manganese oxyhydroxides and oxides (Canfield et al., 1993). The dissolved manganese
originates where Mn (III,IV) is reduced to Mn(II). The reduction of manganese occurs as
part of a sequence of well-established diagenetic reactions. This sequence has a vertical
distribution determined by preferential use of electron that yields the highest amount of
free energy for the bacterially mediated oxidation of organic matter. At the sediment-water
interface, oxygen is reduced, followed by the reduction of nitrate and manganese oxide,
5
.
IVL report
then reactive iron oxyhydroxides, FeOOH and finally the reduction of sulphate (Berner,
1980; Tromp et al., 1995; Hyacinthe et al., 2001).
In addition to the redox reactions involving the microbial degradation of organic matter
there as some important secondary redox reactions. Nitrate can oxidise upward diffusing
Mn2+ and MnO2 is formed (Aller et al., 1998). Mn-oxides and Mn-oxyhydroxides can be
reduced by upward diffusing Fe2+(Hyacinthe et al., 2001). They can also be reduced by
ammonia forming either nitrate under anaerobic conditions or dinitrogen under aerobic
conditions (Aller et al., 1998).
Depth sequence of bacterially-mediated oxidation of organic matter (Hyacinthe et al., 2001)
O.M. = C106H263O110N16P
1. Oxygen consumption by oxic respiration and nitrate production
138O2+ O.M. + 18HCO3-→124CO2 + 16NO3- + HPO42- + 140H2O
2. Nitrate consumption by denitrification
94.4 NO3- + O.M.→13.6 CO2 +92.4 HCO3- + 55.2N2 + 84.8H2O + HPO42
3. Reduction of Mn oxides by anaerobic respiration
236MnO2+O.M.+364CO2+104H2O→470HCO3-+8N2+236Mn2++HPO424. Reduction of Fe-oxides and production of ammonia
424Fe(OH)3+O.M.+740CO2→846HCO3-+424Fe2++16NH3+320H2O+HPO425. Production of sulphide and ammonia by sulfatoreduction
53SO42-+O.M.→39CO2+67HCO3-+16NH4++53HS-+39H2O+HPO42Secondary redox reactions coupled to iron and nitrogen (Hyacinthe et al., 2001):
6. Production of nitrate by nitrification:
NH4++2O2→NO3-+2H++H2O
7. Oxidation of Mn2+ with oxygen
2Mn2++O2+2H2O→2MnO2+4H+
4Mn2++O2+6H2O→4MnOOH+8H+
8. Oxidation of Mn2+ with nitrate
5Mn2++2NO3-+4H2O→5MnO2+N2 +8H+
9. Oxidation of Fe2+ with nitrate
5Fe2++NO3-+12H2O→5Fe(OH)3 +1/2N2+9H+
10. Oxidation of Fe2+ with Mn-oxides
Fe2++MnOOH + H2O→Fe(OH)3 + Mn2+
2Fe2++MnO2+4H2O→2Fe(OH)3 + Mn2+ + 2H+
11. Reduction of Mn-oxide by ammonia to give dinitrogen
2MnOOH+NH4++3H+→2Mn2++1/2N2+4H2O
3/2MnO2+NH4++2H+→3/2Mn2++1/2N2 + 3H2O
12. Reduction of Mn-oxide by ammonia, production of nitrate
8MnOOH+NH4++14H+→8Mn2+ + NO3-+13H2O
4MnO2+NH4+ + 6H+→ 4Mn2++NO3- + 5H2O
Secondary redox reactions of carbon and sulphur (Wang and Van Cappellen, 1995)
6
.
IVL report
13.
14.
15.
16.
17.
18.
19.
20.
≡S-Mn+ + ½ O2 + OH- → ≡S-H0 + MnO2
≡S-Fe+ + ¼O2 + 3/2H2O + OH-→ ≡S-H0 + Fe(OH)3
H2S + MnO2 → Mn2+ + S0
H2S + 2O2 → SO42- + 2H+
H2S + 2Fe(OH)3 → 2Fe2+ + So + 4OH- + 2H2O
FeS + 2O2 → Fe2+ + SO42CH4 + 2O2 → CO2 + 2H2O
CH4 + CO2 + SO42- → H2S + 2HCO3-
There are also adsorption reactions, which influences the concentrations of the dissolved
species of nitrogen, iron and manganese (Wang and Van Cappellen, 1996):
21. NH4+↔ NH4+ (ads)
22. ≡S-H0 + Mn2+ + OH- ↔ ≡S-Mn+ + H2O
23. ≡S-H0 + Fe2+ + OH- ↔ ≡S-Fe+ + H2O
Below the zone of sulfatoreduction in marine sediments, where sulphate is exhausted, the
decomposition of organic matter continues with the production of methane (Berner, 1980,
Tromp et al., 1995). This diagenetic reaction yields the least amount of free energy (Tromp
et al., 1995) and occurs only if sulphate is not present. In the case of lake sediments and
brackish water where the supply of sulphate is less, methane can be produced just below
the sediment (Berner, 1980; Whiticar, 1999; Thiessen et al., 2006).
24. Production of methane (Wang and Van Cappellen, 1996)
O.M. + 14H2O → 53CH4 + 35CO2 + 14HCO3- + 16NH4+ + HPO42In the anoxic part of the sediment the reduced manganese can precipitate as authigenic
minerals, predominantly rhodocrosite, MnCO3 (Berner, 1980; Wang and Van Cappellen,
1996). This reaction and that of the iron analogue (siderite) and iron sulphide are three
important precipitation and dissolution reactions influencing the concentrations of
dissolved iron and manganese.
Precipitation and dissolution reactions (Wang and Van Cappellen, 1996)
25. Mn2+ + 2HCO3- ↔ MnCO3 + CO2 + H2O
26. Fe2+ + 2HCO3- ↔ FeCO3 + CO2 + H2O
27. Fe2+ + 2HCO3- + H2S ↔ FeS + 2CO2 + 2H2O
7
.
IVL report
The precipitation of a mixed Ca-Mn solid carbonate seem to control the concentration of
dissolved Mn2+ in the pore water (Magen et al., 2011). However, in coastal sediment with
high concentrations of organic matter and low carbonate contents, complexation with
dissolved organic matter can lead to porewater concentrations well above supersaturation
with respect to MnCO3 (Mucci and Edelborn, 1992). The dissolved organic matter (DOM),
produced during the microbial degradation of sediment organic matter, interferes with the
nucleation and growth of MnCO3 (Berner, 1980). The rate at which manganese is reduced,
and hence the preservation of Mn(IV,III) solid phases such as pyrolusite and manganite , in
anoxic environments seems to be a complex function where the supply of fresh organic
matter can be limiting (Fischer et al., 2008; Magen et al., 2011). Also, in low sulphate
environments such as in lakes and brackish environments, mixed Mn(II/III)-Fe(II/III)-P
solid phases formed in the oxic part of these sediment can remain stable as they become
buried in the deeper, anoxic part of the sediments (Hyacinthe and Van Cappellen, 2004). If
the concentrations of sulphates are higher, as in the marine environment, the ferric iron
form iron sulphides instead.
Aim
The aim of this project is to get better knowledge concerning the occurrence and
environmental fate of linear alkylbenzne sulphonates (LAS) for risk assessments in aquatic
envrionmnents. The specific need for an improved basis for sediment risk assessment in
Sweden has been identified by Woldegiorigis (2011). The result may contribute to a better
contribution for regulation and ecolabelling of washing detergents and other products
where surfactants are important components. Specifically, to assess the role of surface
sediment redox conditions on the degradation of LAS is a major objective.
Materials and methods
Sampling
The locations for sediment sampling were chosen in order to reflect the influence of two
factors. Firstly, the impact from the city of Stockholm on the total loads of linear alkyl
sulphonates (LAS) and other anionic surfactants and, secondly, the influence of sediment
redox conditions on the degradation of LAS at the time of deposition. The total load of
LAS and other anionic surfactants were estimated by taking sediment samples along a
transect: Lake Mälaren, upstream of assumed major sources, in Saltsjön immediately
downstream of city of Stockholm’s two WWTPs, (Bromma and Henriksdal) and finally in
Torsbyfjärden, approximately 20 km further downstream. The influence of sediment redox
conditions on the degradation of LAS was estimated by choosing two adjacent sampling
8
.
IVL report
points at each location with oxic and anoxic/suboxic condition, respectively, at the
sediment surface. Below the mixing (bioturbation) depth all sediments are anoxic. The
coordinates for the sampling locations are shown in Table 1 below
Table 1. Locations, water depths and redox conditions at the surface of the sediment samples.
Area
Lake Mälaren
M1
M2
Saltsjön
S1
S2
Torsbyfjärden T1
T2
Oxic
Anoxic
Anoxic
Oxic
Anoxic
Oxic
Depth
Coordinates
19
25
32
20
49
28
59 19.24; 18 00.03
59 19.18; 18 00.703
59 19.01; 18 06.78
59 19.23; 18 08.45
59 21.65; 18 26.86
59 20.53; 18 27.40
Location in relation
to the WWTPs
upstream
upstream
At WWTPs
At WWTPs
20km downstream
20km downstream
Fig. 2.Map over Stockholm area with Lake Mälare to the West and Stockholm Archipelag in the Baltic Sea to
the East. Sediment sampling locations are marked with red dots.
At each location two, in total 12, sediment cores were collected using a Kajak type gravity
corer of plexi glass; length 50 cm and diameter 80 mm. Upon retrieval both ends of the
sampling cylinder were sealed with air tight rubber caps and transported to the laboratory
the same day as the sampling took place. In laboratory the length of sediment cores were
measured and described. Two different clean up procedures followed depending on the
coming analyses, either speciation of Mn to assess sediment redox conditions, or counting
of 137Cs-activity for sediment dating.
9
.
IVL report
Chemical analysis
Sediment dating
One sediment core from each sampling point was analyzed for 137Cs-activity for sediment
dating, estimated based on the peak in deposition of 137Cs in 1986 after the Chernobyl
accident. The overlying bottom water was siphoned off and the top 10 cm of the sediment
core was sliced in 2 cm sections. The rest of the sediment core was sliced in 1cm sections.
The wet sediment samples were weighed and frozen. The frozen sediment samples were
freeze-dried and weighed. The water content and porosity of the sediment core were
calculated from the difference in the wet and dry weights. Each sample was grinded
manually with a mortar and pestle. The organic material and the grain size distribution of
the sample were classified by ocular inspection and containers were filled with clay, silt and
sand material. The activities of the samples were counted in a gamma counter,
Intertechnique CG 4000 with a 3´´ NaI detector. The measurement time was 100 min.
Sediment redox conditions
Two different methods were employed to discriminate between oxic and anoxic conditions
in the upper 0.1m of the sediment core. The first method to discriminate between the
different redox conditions of the surface sediments is ocular inspection of the core upon
retrieval. Through the transparent plastic cylinder of the sediment core, the colour of the
sediment and the presence of bioturbating animals and their burrows can been seen. In the
case of an anoxic environment the sediment has an overall black colour with lighter and
darker bands (lamina) representing winter and summer periods, respectively. The black
colour comes from the presence of reduced organic material and sulphides. There can also
be a smell of hydrogen sulphide (H2S), a gas formed during the decomposition of organic
matter, which contains sulphur, under reducing conditions. Sometimes one can see a
metallic lustre, coming from sulphur (formal oxidation state -2) in the form of sulphides.
The sulphides constitute a solid and very stable mineral phase formed by sulphur under
reducing conditions with a doubly charged metal cation such as Fe2+, Zn2+ and Cu2+. The
winter bands are lighter in colour due to the higher proportions of clay minerals to reduced
organic matter and sulphides. In contrast, a sediment sample from an oxic environment is
grey to olive green – brown, depending on the relative amount of inorganic material, clay
and hydrous iron oxide minerals, and organic material. Normally these sediment samples
are not laminated because of the presence of bioturbating macro fauna.
10
.
IVL report
Below the level of 0.1m in the sediments in the area of study, the sediments are always
anoxic because of the lack of oxygen diffusion even with the presence of bioturbating
macro fauna. Macro fauna such as priapulid worm Halicryptus spinulosus, Baltic clam Macoma
baltica and the amphipod crustacean Monoporeia affinis have their bioturbation depths
confined to the upper 10cm of the sediments in hypoxic sediments (Bradshaw et al., 2006).
The recent past and present redox conditions of the sediment samples were also analysed
by studying the speciation of manganese (Mn). Mn can occur in different solid phases in
the sediments, mainly as carbonates, oxides and oxyhydrooxides. The distribution of the
two groups of Mn-species (II and III/IV) is operationally defined as the amount of Mn
being extracted by 1 N HCl, Mn-HCl (both groups) and by ascorbate, Mn-Asc (only the
most easily reduced group, III/IV).
From each station one sediment core was analysed for speciation of Mn. The bottom seal
of the sediment core was removed and the top seal was briefly opened to directly transfer
and distribute the bottom end of the sediment core to shorter sampling pipes with the
same diameter, but varying lengths; 2, 3, 5 and 8 cm, respectively. The subsamples had air
tight rubber seals in the bottom and, after filled with sediment, capped with plastic film.
The subsamples were directly transferred to an anaerobic (90% N2, 5% H2 and 5% CO2)
chamber. In the anaerobic chamber, an equal volume of representative sediment samples
for pore water analysis were taken from each subsample, transferred to 15 ml
centrifugation tubes, capped and immediately centrifuged at 1500 rpm for 10 minutes. The
pore water samples were transferred to glass tubes after the centrifugation, acidified with
HCl to pH 2, and sealed with air-tight caps. The remaining sediment samples were put in
pre-weighed plastic containers, weighed, frozen and thereafter freeze-dried. The freezedried sediment samples were weighed and the samples’ porosities and water contents were
calculated as the difference between wet and dry weights.
Two subsamples of 100 mg each was retrieved from every freeze-dry sediment sample and
put in a glass tube with air-tight seals. One of the subsamples was extracted with 10 ml 1 N
HCl and the other with 10 ml ascorbate. One litre of ascorbate solutions was prepared by
dissolving 50 g NAHCO3, 50 g Na-citrate and 20 g ascorbic acid, buffered at pH 8. Both
fractions (Mn-HCl and Mn-Asc) were shaken continuously for 24 h at room temperature.
For the HCl extraction the supernatant was thereafter diluted with water and the ascorbate
extraction with 0.2 M HCl.
Mn in pore water and extraction samples was measured with flame atomic absorption
spectroscopy, using an external aqueous standard for calibration. Mn-HCl represents the
whole fraction of Mn-oxides and Mn associated with carbonates. Mn extracted with
ascorbate is the most reducible part of Mn (III, IV) oxides and oxyhydroxides.
Samples selected for the analysis of LAS and other
anionic surfactants
11
.
IVL report
Based on the results from the sediment dating, sediment sections representing recent years
(surface sediment), the mid 1980s (peak in Cs137) and sections between these periods, were
chosen for anionic surfactant analyses. At the sampling stations where the sedimentation
rates were high (S1) or low (T1 and T2), sections deposited before the peak in Cs137 were
also sampled. In the former case the high sedimentation rate allowed for expanding the
time period studied, while keeping the same number of analyses. In the latter case, the
sectioning of the sediment cores were not fine enough to resolve the period in between
sedimentation during recent years and the peak in Cs137. In the sediment cores sampled at
M1 and S2, bioturbation had occurred constantly and therefore impossible or difficult to
identify a peak in Cs137activity. Thus, for these sediment cores, sections were chosen based
on sedimentation rates inferred from nearby sites in this study and others (Östlund et al.,
1998 and Jönsson, 2011). Due to that larger sections were used for the analysis of sediment
redox conditions, the results from sediment dating from the replicate cores could not
always be matched.
The first analysis of the sediment cores revealed concentrations of individual linear alkyl
sulphonates (LAS) homologs above µg/g dw for the sediment cores sampled in Saltsjön,
which is above the concentration linear range of the analytical instrument (UFLC-MS/MS)
used to quantify the individual anionic surfactants. Therefore a second analysis focusing on
the content of LAS in the Saltsjön sediment cores were performed, in which either sample
extracts from the first set of analysis of Saltsjön were diluted or samples representing new
depths in the Saltsjön sediment cores were prepared with a lower level of preconcentration.
Also, samples representing resembling or new depths in the Torsbyfjärden sediment cores
were analysed to verify differences in content of LAS between sediment cores sampled at
the same location.
Sample extraction and clean-up
Three gram of freeze-dried sediment was spiked with 4-octylbenzenesulfonic acid (C8LAS) as surrogate standard and extracted twice with methanol: water (1:1). The
supernatants were combined and loaded on a SPE-cartridge containing graphitized carbon
as adsorptive phase. Prior to elution, the SPE-cartridge was washed with methanol. The
analytes were eluted from the cartridge utilizing 10 mM tetramethylammonium hydroxide
in dichloromethane: methanol (1:1). The elute was evaporated to dryness under a gentle
stream of nitrogen at 40 °C and reconstituted in 0.5 ml methanol: water (1:1).
Instrumentals
A binary liquid chromatography (UFLC) system with auto injection (Shimadzu, Japan) was
utilized for injection of samples and pumping of the mobile phase. The chromatographic
separation was performed on a C8- reversed phased column (dimensions 50 x 3 mm, 5 µm
particle size) (thermo scientific) at a temperature of 35°C. The mobile phase consisted of
10 mM ammonium acetate in water (A) and methanol (B). The system was programmed to
deliver a linear gradient with an initial composition of 30% B, which was kept for 2 min.
After 2 min the composition of B was increased from 30% to 100% in 6 min. This
composition was kept for 3 minutes and thereafter returned to 30% B. The total run-time
12
.
IVL report
was 15 min and the flow rate was 0.4 ml/min. The UFLC-system was coupled to an API
4000 tipple quadrupole (MS/MS) (Applied Biosystems) with an electrospray ionization
interface (ESI) operated in negative ion mode.
Results
Sediment characteristics/dating
The Cs137-activity data in the sediment cores from Mälaren, M1 oxic, (see appendix) had a
weak increase down to a depth of six cm followed by a weak decrease down to eight cm.
The diluted Cs137-activity depth profile indicates that the sediment has been bioturbated.
Thus the deposition rate cannot be determined. At the other site, M2 anoxic, the Cs137activity data indicate no or possibly very little bioturbation. A rather well defined peak in
Cs137-activity at 15 cm depth corresponds to an estimated annual average dry matter
deposition rate of 0.26 g/cm2.
In Saltsjön, the average annual dry matter deposition rate at S1 anoxic was estimated to be
0.21 g/cm2, assuming that the 15 cm sediment depth correspond to the year 1986. For the
other sediment core, S2 oxic, a maximum in Cs137-activity cannot be defined due to a
continuous bioturbation.
The annual average dry matter deposition rates in Torsbyfjärden at T1 anoxic was
estimated to be 0.18 g/cm2. Since the year 1986 ten cm of sediment was estimate to have
been deposited, as indicated by the Cs137-activity data. The corresponding depth at the
other site, T2 oxic, was six cm, which gives an estimated annual dry matter deposition of
0.11 g/cm2.
Sediment redox conditions
The manganse data from Mälaren, M1 oxic, (Fig. 3) indicated a redox-boundary between 5
and 10 cm. Above 10 cm sediment depth there was an increase in Mn-Asc up to the
sediment –water interface. Below 10 cm sediment depth the Mn-Asc inventory remained
the same whereas Mn-HCl increased somewhat.
The porewater Mn2+ concentrations at M2 anoxic (Fig. 3) did not vary with sediment depth
and was approximately 20 moles. The Mn-HCl concentrations vary only to a relative
small extent and no clear trend is discernible. At M2 the distinct positive increase in
concentrations of Mn-Asc close to the sediment surface could indicat a redox-boundary
somewhere in this interval (0-8 cm). However this is probably not the case due to the
negative difference between the two fractions Mn-HCl and Mn-Asc and the decoupling of
their variations with sediment depth.
13
.
IVL report
The sediment profiles of Mn-Asc and Mn-HCl in Saltsjön, S1 anoxic (Fig. 4) displayed a
distinct positive difference between the Mn-HCl and Mn-Asc concentrations. However,
below 10 cm sediment depth, this had changed to a negative difference as in the samples
from M2 in Lake Mälaren. The Mn-Asc data from S2 oxic (Fig. 4) displayed a peak in the
Mn-Asc inventory at a sediment depth of 3-4 cm and then stabilised at approximately 10
moles/g. The corresponding Mn-HCl inventory showed an increase from the sediment
surface down to 10 cm. The Mn-HCl inventory then remains steady at approximately 19
moles/g. Once again, there is a difference of Mn-HCl and Mn-Asc in the top 4 cm. The
sediment depth profile of porewater Mn2+ concentrations showed a peak at 3-4 cm,
followed first by a decrease and then an increase.
The Mn-data from station T1 anoxic in Torsbyfjärden (Fig. 5) displayed a maximum in the
Mn-Asc inventory in the top 4 cm. Below the Mn-Asc inventory decreased down to 14 cm
and thereafter increased down to 19 cm, where it stabilised at approximately 30 moles/g.
There was a simultaneous increase in the porewater Mn2+ concentration with depth below
10 cm.
At T2 oxic, the sediment Mn-data (Fig. 5) indicated a weak increase in both Mn-Asc and
Mn-HCl with sediment depth. The inventories of Mn-Asc and Mn-HCl in mole/g are
smaller compared to T1.
Fig. 3. Sediment profiles of Mn-Asc, Mn-HCl (mole/g) and Mn2+ (mole) in Lake Mälaren.
14
.
IVL report
0
Mn-Asc, Mn-HCl (uM/g); Mn2+ (uM)
10
20
30
40
50
0
Sediment depth (cm)
5
S2: Mn-Asc
S2: Mn-HCl
10
S1: Mn-Asc
15
S1: Mn-HCl
S2: Mn2+
20
25
30
Fig.4. Concentrations of Mn-Asc, Mn-HCl (mole/g) and Mn2+ (mole) in sediment samples from Saltsjön.
0
Mn-Asc, Mn-HCl (uM/g); Mn2+ (uM)
20
40
60
80
0
Sediment depth (cm)
5
T2: Mn-Asc
10
T2: Mn-HCl
T1: Mn-Asc
15
T1: Mn-HCl
20
T1: Mn2+
25
30
35
Fig. 5. Sediment profiles of Mn-Asc, Mn-HCl (mole/g) and Mn2+ (mole) in Torsbyfjärden.
15
.
IVL report
Anionic surfactant concentrations
LAS
Linear alkylbenzene sulphonates (LAS) was found in samples from all sampling sites and at
all sediment depths analysed (Figs. 6-11).
In both anoxic and oxic sediment cores from the eastern part of Lake Mälaren, the
concentrations of LAS were in the range 82-440 ng/g dw, (Fig. 6 and Fig. 7). Slightly
higher concentrations were found in the deepest sediment sections but no clear vertical
distribution pattern could be seen.
0
Concentration (ng/g dw)
200
400
0
Concentration (ng/g dw)
100
200
300
400
Sediment depth (cm)
C10-LAS
0_7
0_4
C11-LAS
C12-LAS
8_12
7_12
12_17
12_17
C13-LAS
Fig.6. Las concentrations in sediments at M2 (anoxic)
Sediment depth (cm)
0
Concentration (ng/g dw)
50
100
150
200
0
Concentration (ng/g dw)
200
400
600
0_2
2_4
0_6
4_6
Fig. 7. LAS concentrations in sediments at M1 (oxic)
16
C10-LAS
C11-LAS
C12-LAS
C13-LAS
.
IVL report
In sediment cores sampled in Saltsjön, LAS concentrations were higher than in Mälaren,
with the highest concentrations in the anoxic cores, (Fig. 8 and 9). In surface sections from
the oxic cores (S2), concentrations of 3900 and 7900 ng/g dm were found, whereas surface
sections from the anoxic cores (S1) contained 7000 and 12000 ng/g dm. Highest
concentrations were found in the deeper sediment sections with concentrations of 8300
and 11000 ng/g dm in the oxic cores and even higher concentrations in the anoxic cores,
32000 and 44000 ng/g dm.
As could be expected, the difference in concentrations between surface sediments and
deeper sections were more profound in the anoxic cores. In the oxic cores, concentrations
in deeper sediments were 1.4 and 2.1 times higher compared to surface sediment. In the
anoxic cores, the deeper sediments contained 2.6 and 6.2 times higher concentrations.
Sediment depth (cm)
0
Concentration (ng/g dw)
20000
40000
0
0_2
0_5
6_8
5_10
12_15
10_15
Concentration (ng/g dw)
20000
40000
60000
15_18
C10-LAS
C11-LAS
18_21
C12-LAS
C13-LAS
Fig. 8. LAS concentrations in sediments at S1 (anoxic).
Sediment depth (cm)
0
Concentration (ng/g dw)
5000
10000
15000
0
0_2
0_2
2_6
2_6
6_10
6_11
10_12
14_16
Fig.9. LAS concentrations in sediments at S2 (oxic).
17
Concentration (ng/g dw)
5000
10000
C10-LAS
C11-LAS
C12-LAS
C13-LAS
.
IVL report
Sediment depth (cm)
0
Concentration (ng/g dw)
10
20
30
40
0
0_2
Concentration (ng/g dw)
100
200
300
0_4
2_14
4_14
C10-LAS
C11-LAS
C12-LAS
C13-LAS
Fig. 10. LAS concentrations in sediments at T1 (anoxic).
Sediment depth (cm)
0
Concentration (ng/g dw)
500
1000
0
1500
0_2
0_5
2_6
5_10
Concentration (ng/g dw)
50
C10-LAS
C12-LAS
100
C11-LAS
C13-LAS
Fig. 11. LAS concentrations in sediments at T2 (oxic).
In the sediment cores from Torsbyfjärden, the concentrations of LAS were significantly
lower on a sum-basis as well as for the homologs C12 and C13 (Fig. 10 and 11) compared
to Saltsjön (Fig. 8 and 9). For surface sediments from both the oxic and anoxic site, the
concentrations of C10 and C11 homologs were about the same magnitude as in Saltsjön.
Overall, the distribution of LAS concentrations with sediment depth and between oxic and
anoxic sites displays a relative large heterogeneity or variance.
To test if there were any significant differences in mean concentrations in surface
sediments of linear alkylbenzene sulphonates (LAS) depending on redox environment
(oxic/anoxic) or distance from WWTPs (Saltsjön/Torsbyfjärden) a series of pairwise t-tests
were performed using both sum- and C10-C14-LAS concentration data. The results from
these tests are shown in tables 2 and 3 below.
18
.
IVL report
Table 2. Pairwise t-test to test differences between mean concentrations of sum, C10, C11, C12 and C13
LAS in oxic and anoxic surface (0 – 7 cm) sediments.
Anox/Ox
C10
C11
C12
C13
Sum
C10
0.56
C11
C12
C13
Sum
0.52
0.83
0.91
0.83
Table 3. Pairwise t-test to test differences between mean concentrations of sum, C10, C11, C12 and C13
LAS in Saltsjön (anoxic+oxic) and Torsbyfjärden (anoxic + oxic) surface (0 – 7 cm)
sediments.
Salt/Tor
C10
C11
C12
C13
Sum
C10
0.15
C11
C12
C13
Sum
0.02
3e-4***
0.00085***
0.00042***
*** 99.9 % significance
19
.
IVL report
Alkyl sulphate, alkyl ether sulphates and cocoamidopropyl
betaine
The concentrations of the anionic surfactants alkyl sulphate (AS) and alkyl ether
sulphate(AES) were below their respective detection limit (LOD) in the majority of the
sediment samples investigated (Appendix A6-A7). The concentrations of cocoamidopropyl
betaine (CAPB) were below the LOD in all of the sediment samples (Appendix A6). The
concentrations of the anionic surfactants AS, AES and CAPB are presented in the figures
below (Figs 12 – 17). Values below LOD are represented by half their respective LOD.
Detectable concentrations of AS were found in surface sediments at all sites except T2,
which was an oxic site. The concentrations were clearly highest at S1 and T1, both anoxic,
although the variance between cores and samples of the same core with depth was once
again clearly higher at T1. Only at S1 could AS be detected at sediment depths greater than
10cm or approximately the zone of bioturbation. All four ethersulphates (1,2,3,4-AES)
could only be detected in the top 10cm at S1 (anoxic) while 1,3,4-AES could be detected in
the sample from 10 – 15cm of sediment depth at S1. Thus the parameter sum-AES could
was only above its’ LOD in the top 15cm at S1.There was however an inter-core variance
at S1, since in the other core, only 1,3-AES could be detected in the surface sample (0-2cm)
and 1,4-AES in a deeper sample (6-8cm). In the sample from 12-15cm only 4-AES could
be detected in this core. At the other anoxic sites, 3,4-AES could be detected in samples T1
(0-4; 4-9cm) and M2 (0-7cm). At oxic sites only 4-AES could be detected at M1 (0-2cm)
and T2 (2-4cm).
Fig. 12. The bars represent concentrations of Alkyl sulphates (AS), alkyl ether sulphate (AES) and
cocoamidopropyl betaine (CAPB) in sediments at M1 (oxic). The concentrations represent half their
respective limit of detection (LOD) with the exception of AS in the 0 – 2cm sample from core 1
used for dating by Cs137 (left panel).
20
.
IVL report
Fig. 13. The bars represent concentrations of alkyl sulphates (AS), alkyl ether sulphate(AES) and
cocoamidopropyl betaine (CAPB) in sediments at M2 (Anoxic). The concentrations represent half
their respective limit of detection (LOD) with the exception of AS in the 8 – 12cm sample from
core 2 used for dating by Cs137 (left panel).
Fig. 14. The bars represent concentrations of Alkyl sulphate (AS), alkyl ether sulphate(AES) and
cocoamidopropyl betaine (CAPB) in sediments at S1 (Anoxic). All concentrations of CAPB, of AS
in sample 12-15cm from core 2 and all of AES from core 2 used for dating by Cs 137(left panel), are
represented by half their respective limits of detection (LOD)
Fig. 15. The bars represent concentrations of alkyl sulphate (AS), alkyl ether sulphate (AES) and
cocoamidopropyl betaine (CAPB) in sediments at S2 (oxic). The concentrations represent half their
respective limit of detection (LOD) with the exception of AS in the 0 – 2cm sample from core 2
used for dating by Cs137 (left panel).
21
.
IVL report
Fig. 16. The bars represent concentrations of alkyl sulphate (AS), alkyl ether sulphate (AES) and
cocoamidopropyl betaine (CAPB) in sediments at T1 (anoxic). The concentrations represent half
their respective limit of detection (LOD) with the exception of AS in the 0 – 2cm and 2-4 cm
samples from core 2 used for dating by Cs137 (left panel).
Fig. 17. The bars represent concentrations of Alkyl sulphate (AS), alkyl ether sulphate(AES) and
cocoamidopropyl betaine (CAPB) in sediments at T2 (oxic). The concentrations represent half their
respective limit of detection (LOD).
22
.
IVL report
Discussion
Sediment characteristics and redox conditions
Lake Mälaren
The Cs137-data from core M1 indicate that the sediments always have been bioturbated. The
interpretation of the surface sediments as being bioturbated and oxygenated, indicates a
low to moderate deposition of dry matter. The degradation of organic matter requires
oxygen, thus high deposition of organic matter usually leads to anoxic conditions in the
surface sediment (Kiirikki et al., 2006; Reed et al. 2011). The Mn-HCl and Mn-Asc
inventories (Fig. 3) are practically identical above the depth of bioturbation (10cm) and
reflecting the Mn content of deposited particles such as, refractory Mn-oxides and Iron
hydroxyoxides and the oxidation of Mn2+ to MnOOH and MnO2. Below the oxygenated
and bioturbated zone, reduction of bioturbated Mn-oxides from the oxic zone occurs
reduced Mn species mainly MnCO3, are formed in the anoxic zone. However, as described
abover it could also be Mn(IV,III) solid and mixed Mn(II/III)-Fe(II/III)-P solid phases
which are preserved in the anoxic environment due to limited supply of either sulphate
(Hyacinthe and Van Capellen, 2004) or fresh organic matter (Magen et al., 2011).
Considering the high amount of refractory terrestrial organic material that can deposited to
the sediments in freshwater lakes either explanation or both in conjunction are possible.
The refractory terrestrial organic material is not available for further degradation by the
microbes Hence, the difference in the inventories of Mn-Asc and Mn-HCl below the
bioturbated zone can be attributed to either preserved Mn oxides or hydrous oxides or
precipitated MnCO3.
The probably higher annual deposition rates of dry matter at M2 (0.26g/cm2) is consistent
with the interpretation of this site being anoxic and not bioturbated since (at least) 1986.
This deposition is comparable to what has previously been reported from the surrounding
area: 0.14 and 0.20 g /cm2 (Östlund et al., 1998), and 0.23 and 0.29 g /cm2 (Jönsson, 2011).
The location of the peak in Mn-Asc concentrations (Fig. 3) is poorly constrained, due to
the large interval of that section (0-8cm). The sediment core was anoxic from an ocular
inspection and devoid of bioturbation. The distribution of dissolved manganese and the
two extracted solid fractions Mn-Asc and Mn-HCl with sediment depth can lead to two
different interpretations. The first is that the actual location of the peak in Mn-Asc is
probably just below a very thin (some mm) oxidised layer. The almost constant porewater
Mn2+ concentrations (Fig. 3) with depth indicates pseudo-equilibrium between dissolved
Mn2+ and an authigenic Mn-phase, probably MnCO3 (Berner, 1980). The authigenic Mnphase is formed from the reduction of the Mn-oxides, which is indicated by the decrease of
Mn-Asc with sediment depth. However, as espoused above, high concentrations of
dissolved organic matter (DOM), produced during the microbial degradation of sediment
organic matter, interferes with the nucleation and growth of MnCO3 (Mucci and Edelborn,
1992). Although rhodocrosite super saturation is reached the precipitation is so slow due to
inhibition of DOM that pseudo-equilibrium is attained (Berner, 1980).
23
.
IVL report
However, this interpretation cannot explain the negative difference between Mn-HCl and
Mn-Asc. If manganese oxides are reduced to rhodocrosite at depth, then Mn-HCL would
be larger than Mn-Asc. Instead an alternative interpretation of the Mn-Asc and Mn-HCl
data is that the manganese oxides, hydrous oxides and Fe,Mn-P solid phases are preserved
from reduction either due to low amounts of sulphate or organic matter that can easily be
degraded (Hyacinthe and Van Cappellen, 2004; Magen et al., 2011). This pool is a part of
the HCl-fracation together with rhodocrosite. There is also a high concentration of
dissolved Mn2+ which at pH8 is extracted as together with a solid phase as adsorbed or
easily exchangeable. This solid phase could be birnessite (-MnO2) which is formed under
oxidising conditions, or a another solid Mn-phase which has survived in the anoxic part of
the sediment. Birnessite has an increasing adsorption of Mn2+ with increasing pH. From
pH2 to pH8 the adsorption of Mn2+ increases from 2 to 7 moles/m2 (*106) or 0.1 to 1.2
moles/kg MnO2 (Appelo and Postma, 1999). For soils, increased solubility of Mn at pH8
and above under reducing conditions have been demonstrated (Kabata-Pendias, 2004).
Thus in addition to the increased sorptive capacity of the solid Mn-phase, there could also
be an addition of dissolved Mn2+ which could then be adsorbed to the solid Mn-phase,
resulting a negative difference between the Mn-HCL and Mn-Asc inventories.
Saltsjön
The Cs137-data (appendix) do not show a distinct peak at around the year of maximum
deposition, 1986, when the Chernobyl accident occurred. Instead there is a weak decrease
upwards from a sediment depth of 12-15cm, which is assumed to correspond to
1986/1987. This vertical distribution can reflect either bioturbation meiofauna or
resuspension of sediments due to currents (Bradshaw et al., 2006). Considering the
lamination and grain size distribution of the core, bioturbation by meiofauna is the likely
candidate. The Baltic Sea Meiofauna has the same capacity of vertical mass transport of
sediment as macrofauna and are present also in anoxic surface sediments as long as the
bottom water is aerated (Bradshaw et al., 2006).
Similarly to the anoxic core from Lake Mälaren (M2), below 10 cm the negative difference
between Mn-HCl and Mn-Asc at S1 (Fig. 4) could be explained in this way. The organic
matter in the sediments in Saltsjön is to a large extent, 35% of Lake Mälaren origin
(Jönsson et al., 2005). Thus a there is a significant part of the organic matter in Saltsjön
which is of refractory terrestrial origin. The difference between the Mn-Asc and Mn-HCl
inventories in the upper could represent the inventory of MnCO3, which precipitates as
MnO2 is reduced as soon as it is buried beneath a thin oxidised zone. There is an increase
of Mn-Asc towards the surface, which could indicate a redox-boundary somewhere in the
top 5 cm of the sediments. Since the sediment core from S1 is considered to recently have
been bioturbated (see below) the oxic zone could be a few cm thick at most. One way to
explain the vertical overlap of reduced solid species in the Mn-HCl fraction and a redox
boundary is to consider the three dimensional distribution of the redox potential. The
distribution of the redox potential is determined by the burrows of the bioturbating
organisms which leads to both vertical and lateral gradients. Below the redox boundary the
sediment at S1 is anoxic. One explanation for the high concentrations of Mn(III, IV)
oxides and hydroxides (Mn-Asc) in the anoxic part is that there is an on-going reductive
24
.
IVL report
dissolution of Mn-oxides in the anoxic part of the sediment but the sediment accumulation
rate is high enough and the mixing depth is large enough to make the observed residence
time in the anoxic part short enough to preserve the Mn-oxides. On a multi-annual time
scale, the upward flux of Mn2+ and the downward flux of Mn-oxides are at steady state. If
there would have existed Mn2+ data for S1 we should then have seen an increase in Mn2+
concentrations with sediment depth. The flux of Mn2+ would then exceed the upward and
downward diffusive Mn(II,III) fluxes generated by Mn(II,III) re-oxidation and authigenic
carbonate formation respectively (Mouret et al., 2009).
According to Cs137-activity data (see appendix), its distribution with sediment depth could
indicate a recent re-distribution upward from a buried maximum (1986), which could have
been caused by bioturbation of macrofauna or meiofauna. By combining the Cs137-data
and the Mn-data, this model would imply that there is an on-going bioturbation in the top
centimetres at S1 which is consistent with the findings in other previously anoxic areas in
Saltsjön (Karlsson et al., 2011). The annual deposition rate of dry matter at S1 (0.21g/cm2)
is somewhat lower than the average value for Saltsjön, 0.29 (s=0.11) g/cm2 (Karlsson et al.,
2011). Other studies report similar values for the annual deposition rate of dry matter in
Saltsjön; 0.20g/cm2 (Östlund et al., 1998) and 0.52;0.40;0.31 g/cm2 (Jönsson, 2011).
The combined Mn-data and Cs137-activity data from S2 can be explained by the following
model (Mouret et al., 2009). There is an intensive bioturbation at S2, which transport
Mn(III, IV) –oxides and hydroxides down to the redox boundary. Throughout the oxic top
4 cm of the sediments, these Mn-oxides are stable. As they reach the anoxic part of the
sediment, they become reduced and a peak in porewater Mn2+ is produced. Thus the
explanation for the presence Mn-Asc in the anoxic part of the sediment is the same here as
for S1. In the case of S2, this explanation is supported by the Mn2+ data, which show a
local maximum in the top 4 cm of the sediment core, which is oxic and bioturbated. The
increase in porewater Mn2+ indicate that there is no equilibrium between aqueous Mn2+and
an authigenic Mn-phase. The negetative difference between Mn-HCl and Mn-Asc in the
upper part of the sediment core, could also here be explained by increased solubility of
manganese at pH8 under the reducing conditions followed by adsorption solid phase with
increasing sorptive capacity at pH 8 compared to pH 2.
Torsbyfjärden
In the sediment core from station T1, the maximum in the Mn-Asc inventory is in the top
4 cm (Fig. 5), which indicate a redox-boundary somewhere in that interval. There is an
increase in the porewater Mn2+ concentration with depth below 10 cm. This could indicate
that Mn(III,IV) oxides and hydroxides are being dissolved in the anoxic part of the
sediment, which produces an increase in dissolved Mn(II) at depth. The increase in
porewater Mn(II) indicates that there is no equilibrium between aqueous Mn(II) and an
authigenic Mn-phase. The increase of Mn2+ in the bioturbated oxic part, top 4cm, of the
sediment can be explained by this reductive dissolution of Mn-oxides in the anoxic part of
the sediment which exceeds the upward and downward diffusive Mn(II,III) fluxes
generated by Mn(II,III) re-oxidation and authigenic carbonate formation respectively
(Mouret et al., 2009), as can be the case for S1 and S2.
25
.
IVL report
According to the 137Cs-data (see appendix), the mixing-depth is 10 cm and the average
annual dry matter deposition rate is 0.18g/cm2, which is in the range (mean 0.26, s.d. 0.11)
of what is found in this and other parts of the Stockholm archipelago (Karlsson et al.,
2011). One model which encompass these findings is that bioturbation has recently started
in these sediments and thereby transported Mn-oxides down into the anoxic, formerly
laminated (i.e. non-bioturbated) sediments. This model is in line with the hypothesis of
Karlsson et al. (2011) on the recent bioturbation by Marenzelleria of previously anoxic
sediments in the Stockholm Archipelago.
At the other station in Torsbyfjärden, T2, the sediment Mn-data (Fig. 5) and 137Cs-activity
data from T2 also indicate a bioturbation which buries Mn-oxides in the anoxic part of the
sediment. However the inventories of Mn-Asc and Mn-HCl in mole/g are smaller
compared to T1. Unfortunately, there are no Mn2+ porewater data to indicate whether there
is an equilibrium between aqueous Mn2+ and an authigenic Mn-phase or not. A porewater
Mn(II) profile would also facilitate the interpretation of the depth of the redox-boundary.
Once again the presence of Mn-Asc in the anoxic part of the sediment can be explained by
the same mechanism as for S1, S2 and T2
The 137Cs-activity data (see appendix) from T2 indicate that sediments have always been
bioturbated and that the annual deposition rate of dry matter is rather low, 0.11g/cm2
compared to other areas in the Stockholm archipelago, see above.
Trends between stations
Over all trends for all three pair of stations: Mn-inventories are always higher at anoxic
than oxic stations. This relationship can be explained by higher particle deposition rates at
the anoxic stations, since all the main factors which control sediment redox conditions, e.g.
the deposition flux and reactivity of organic matter, the intensities of bioturbation, and the
manganese and iron deposition fluxes, all tend to correlate with sedimentation rate (Tromp
et al., 1995). However, it can also be explained by the adverse relationship between particle
mixing (caused by bioturbation) and manganese cycling. Particle mixing enhances the
transport of particle bound Mn2+ to the sediment-water interface where it is oxidised and
returned to the water column ((Wang and van Cappellen, 1996). Without bioturbation,
Mn2+ is produced from the reduction of Mn(III,IV) below the rather thin oxic sediment
layer caused by the downward diffusion of O2. This thin oxic zone and nitrate in a suboxic
zone below it effectively stops the Mn2+ from leaving the sediment since they are oxidised
into stable Mn(III,IV) phases, MnOOH and MnO2 (Wang and van Cappellen, 1996; Aller
et al., 1998; Mouret et al., 2009). Therefore there exists a negative relationship between
maximum concentration of MnO2 in the sediment and particle mixing coefficient (Wang
and van Cappellen, 1996). The stations in Lake Mälaren, M1 and M2 have the highest
inventories of manganese. This could be explained by both the favourable conditions for
precitpitaton of rhodocrosite, due to carbonate content of pore water and preservations of
solid Mn(IV,III) phases due to lower concentrations of either sulphate or fresh organic
matter, or both.
26
.
IVL report
Spatial distributions of linear alkylbenzene
sulphonates, alkyl sulphates, alkyl ether sulphates
and cocoamidopropyl betaine
Linear alkylbenzene sulphonates (LAS) has previously been measured in sediments from
Lake Mälaren by Kaj et al. (2008) and in sediments from the area Torsbyfjärden by Folke et
al. (2003). Kaj et al. (2008) found concentrations in the range 360-1600 ng/g dry matter
(dm) in surface sediment sampled in central Stockholm. In surface sediment from
Torsbyfjärden, Folke et al. (2003) determined a concentration of 400 ng/kg dm.
Concentrations in sediments from Lake Mälaren and from Torsbyfjärden determined in
previous studies were thus in the same range as found in the present study. No previous
data exist for the area Saltsjön. The spatial distribution in sediments in central Stockholm
of nonylphenols (NP), the metabolites of another type of synthetic surfactants, the
nonylphenol ethoxylates (NPEOs), show a strong concentration to the discharge points in
Saltsjön of the WWTPs in Henriksdal and Bromma (Jönsson, unpublished results). The
concentrations in Lake Mälaren and eastern part of Saltsjön, a few km upstream and
downstream respectively are significantly lower. An interesting aspect of this spatial
distribution of LAS in sediments is that it has been constant even when the consumption
of LAS in Sweden was much higher and therefore the loads from the WWTPs, as evident
from the higher concentrations of LAS further down in the sediment record (Figs. 6 – 11).
This result corroborates the hypothesis of Woldegiorgis (2011) that a return to previous
levels of LAS consumption in Sweden would not increase the areas where the benthic
fauna would be at risk due to concentrations of LAS in the sediments. The areas at risk
would be the same as currently: sediments close to the discharge points of WWTPs where
organic matter accumulates. Pronounced horizontal gradients from the discharge of
wastewater effluents were also found for LAS in an estuarine environment in Spain
(Gonzales-Mazo et al., 1998; Lara-Martin et al., 2006) and in Portugal (Hampel et al., 2009)
and in a marine (open sea) environment in Spain (Petrovic et al., 2002). The explanation is
the hydrophobicity of LAS which links its environmental fate with that of organic carbon,
that is LAS sediments where the particulate organic matter sediments (Leon et al., 2001;
Lara-Martin et al., 2006).
The high variability of LAS concentrations in Torsbyfjärden, between both oxic and anoxic
stations can possibly be explained by some stations receiving effluents from the nearby
located WWTP in Käppala on the island Lidingö, whereas the other station reflect the
distant load from the city of Stockholm, 20km upstream.
27
.
IVL report
Table 4. Previous measurements of linear alkylbenzne sulphonates LAS in sediments.
Area
Stora Essingen, Årstaviken and Riddarfjärden (eastern part of
Lake Mälaren, central Stockholm)
Fjords of Little Belt, Denmark
Concentration
(mg/ kg dw)
Reference
1.6, 0.36 and
0.53
Kaj et al., 2008
≤0.1-19
Folke et al., 2003
Torsbyfjärden (Stockholm) and Baltic Proper (open sea north of
Gotland)
0.4 and 0.8
German Baltic and North Sea coasts
<0.039-0.106
Bester et al.,
2001
Danish harbours, Copenhagen area
0.2-28
Nerpin et al.,
2005
0.3-8.4
Danish harbours
Jensen &
Gustavson, 2001
Tagus estuary (Lissabon area, Portugal), 40 stations sampled
2004
0.03-17.76 (July)
Hampel et al.,
0.09-9.57
2009
(December)
Cadiz Bay area (Spain), 5 stations sampled in 2002. Highest
conc. close to an untreated waste water discharge.
1.2-67.6
Cadiz Bay area, 11 stations sampled in 2002. Highest conc. same
station as Lara-Martín et al., 2006, but half year after WWTP
start.
1-10
Lara-Martín et al.,
2006
Lara-Martín et al.,
2005
Spanish mediterrian and atlantic coasts. Harbours and close to
industrial and waste water outflows. 39 samples analysed 19992000.
0.1-238
Petrovic et al.,
2002
Bay of Cadiz, Spain,21 samples along transect from municipal
discharge point
1.5 - 50
Gonzales-Mazo et
al., 1998
The statistical tests (tables 2 and 3) show that the major factor for explaining variations in
concentrations is distance from discharge points of WWTPs and not redox conditions of
surface sediments. There are only significant differences between the surface
concentrations of Saltsjön and Torsbyfjärden for the C11, C12, C13 and sum of homologs
C10-C13. This may indicate that C10 is transported longer distances and not deposited to
the same extent at the discharge points of the WWTPs due to weaker sorption to sediment
particles and having a lower degradation rate in the sediments compared to the longer
chain homologs. The capacity of C10 homolog to be transported longer and to be less
28
.
IVL report
degraded is also supported by the findings in other studies (Leon et al., 2001; Lara-Martin
et al., 2006).
For the alkyl ether sulphates (AES) as indicated by the parameter sum-AES and the
homolog 1-ethoxylate (EO) there is possibly a weak gradient from the discharge points of
the WWTPs at S1 compared to both upstream in Lake Mälaren as well as downstream in
Torsbyfjärden (Figs. 12 – 17 and table A7). The same type of gradient of decreasing
concentrations in sediments with distance from discharge points of wastewater, treated or
untreated, has been observed elsewhere (Lara-Martin et al., 2006). However, for the present
study the explanation for the much lower concentrations of AES compared to LAS
although the use of LAS is lower than AES, is the difference in octanol-water partition
coefficients (Kow). The log Kow value for a linear LAS, e.g. 68411-30-3 (11 – 13 carbon
atoms in the alkyl chain), is estimated to be 3.32 when calculated as having 11.6 carbon
atoms in the alkyl chain (HERA, 2009) For AES with 2.7 (average for household use)
ethoxylates the log Kow values are estimated to vary between 0.95 (12 carbon atoms) and
1.9 (14 carbon atoms) depending on carbon chain length (HERA, 2004). Around 95% of
all AES used has carbon chain lengths of 12 to 14 carbon atoms(HERA, 2004). There is a
small variation in log Kow values depending on the number of ethoxylates in the structure
of AES. The average for all AES produced is 2.4 which results in a estimated log Kow
values of 1.0 (12 carbon atoms) and 2.0 (14 carbon atoms) depending on carbon chain
length (HERA, 2004).
Vertical Profiles and aerobic versus anaerobic
degradation
Vertical concentration profiles of linear alkylbenzene sulfonates (LAS) in sediment have
previously been studied by León et al. (2001) and Lara-Martín et al. (2006). León et al.
(2001) found a strong decrease in LAS concentration in the upper oxic sediment zone,
whereas the decrease in concentration in the deeper anoxic sediment was less steep. The
decrease in LAS concentrations with sediment depth and the concurrent increase in
sulfophenyl carboxylic acids (SPCs) indicate the possibility of anaerobic degradation of
LAS in the sediments. SPCs are key intermediate metabolites in the anaerobic degradation
of LAS (Lara-Martin et al., 2007). However, Leon et al. (2001) could not rule out the
possibility of the increase in LAS at the sediment surface reflected the increased use of LAS
due to an increase in the population generating the urban wastewater effluents. LaraMartín et al. (2006) found a similar profile and also concluded that anaerobic degradations
of LAS producing SPCs takes place below the bioturbation depth, where the sediment is
anoxic.
Indeed, Lara-Martin et al. (2007) could conclude from a laboratory study with anaerobic
sediments spiked with LAS, that anaerobic degradation takes place and during this process
SPCs are produced. Lara-Martin et al. (2007) estimated a half-life of approximately 90 days
for LAS in anaerobic sediments, but significantly higher values, up to several years,
expected at concentrations above 20 ppm dw due to microbial toxicity.
29
.
IVL report
In the sediment cores from anoxic surface sediments in Saltsjön (S1) the peak in LAS
(Fig.8) concentrations at 10 – 15cm depth probably reflects a change in the use of LAS. As
espoused above, physical reworking of the sediments caused by erosion seems unlikely but
bioturbation by meifauna is likely to be present. If the mass transport caused by meiofauna
bioturbation would be the dominating process, the LAS concentration profile would mimic
that of Cs137 considering the high affinity to particulate organic carbon (Leon et al., 2001;
Lara-Martin et al., 2006). To some extent it does follow the curve of Cs137-activity but the
decline is much sharper closer to the sediment surface. This could reflect decreased use of
LAS in later years but also rapid degradation in a thin oxic layer close to the surface. In the
sediments below the peak in LAS concentrations, there could be anaerobic degradation of
LAS occurring (c.f. Lara-Martin et al., 2007) since this period, i.e. pre-1986, would
correspond to higher depositions of LAS.
A natural starting point for discussing the vertical distribution of LAS in the sediments of
the study area is to start close to the discharge points of the wastewater effluents since this
is the point where changes in the use of LAS will be most evident. In the oxic surface
sediments in Saltsjön (S2), the peak in LAS concentrations are in the region of 2 to 10 cm
of sediment depth (Fig.9). In this region poor correlation between LAS concentration and
Cs137-activity profiles could be attributed to discrimination of inorganic and organic matter
in mass transport processes (bioturbation) by meiofauna and macrofauna. In this case,
Monoporeia affinis, which was observed in the overlying water of some of the cores, cause
both upward and downward transport which could distribute compounds bound to organic
matter bound differently than more weakly adsorbed compounds to mineral surfaces
(Bradshaw et al., 2006). One example of the former could be LAS and one of the latter
could be Cs+. Below the zone of bioturbation. Considering the bioturbation depth of
Momoporeia affinis, which is would be expected to be less than 10cm (Bradshaw et al., 2006),
the position of the peak in LAS concentrations at S1 probably also reflects a dccrease in the
use of LAS during the last decade or so. The decreasing concentrations at S1 (Fig. 8)
probably also here reflects anaerobic degradation of LAS.
Using the interpretations of the sediment profiles of LAS from stations S1 and S2 in
Saltsjön (Fig. 8 and 9) , close to the discharge points of the WWTPs as indicatiors of LAS
use in the Stockholm region, the same processes can be seen in sediments from Eastern
Lake Mälaren (Fig. 6 and 7). At the anoxic site (M2), the sediments would be without any
pronounced redistribution of LAS due to bioturbation, an the vertical profile reflects a
decreasing use of LAS during the recent decade(s). At the oxic site (M1), bitoturbation by
meiofauna and macrofauna causes the typical vertical profile with a maximum just beneath
the sediment surface (Berner, 1980; Bradshaw et al., 2006), while the peak at 4-6 cm
sediment depths reflects an earlier higher deposition of and use of LAS.
The vertical profiles of LAS in Torbyfjärden, T2 (Fig. 10 and 11) are more difficult to
interpret since there are large relative variations between the cores at both sites and the
vertical resolution is in some instances very poor. The core with the highest concentration
of LAS, and hence highest deposition of organic particles, at the anoxic site (T1) seems to
clearly indicate on going bioturbation while the other one does not. Likewise the core with
the highest concentrations of LAS and deposition of organic matter at the oxic site (T2)
can be interpreted as bioturbation occurs in the top 6cm.
30
.
IVL report
Ecotoxicity
The toxicity of linear alkylbenzene sulphonate (LAS) towards freshwater and marine
benthic species has been studied by several authors (Tables 5 and 6). Reviews of available
data show that acute toxicity is greater for individual LAS homologues with longer alkyl
chain lengths and that biodegradation intermediates are significantly less toxic than the
parent LAS with L/EC50 values >1000 mg/L for fish and crustacean (Daphnia.magna)
(UNEP 2005).
Based on chronic toxicity data for three freshwater species of different taxonomic groups,
Comber et al. (2006) suggested a PNEC of 8.1 mg/kg dw, using an assessment factor of
10. For marine sediments Hampel et al. (2007) derived a PNEC of 4.9 mg/kg dw using an
assessment factor of 10 based on chronic ecotoxicity data for two mollusc and one fish
species. The endpoint studied resulting in the lowest value was lethality and that the study
duration was relatively short (9 days). Further, this PNEC-value is in the same range as
concentrations shown to cause acute effects on marine algae (Moreno-Garrido et al. 2003;
Moreno-Garrido et al. 2007). The PNECs derived are in the same range as site-specific
sediment quality values (SQVs)for the Gulf of Cádiz, derived by linking chemical
measurements and biological effects (sediment toxicity results for the clam Ruditapes
philippinarum and the amphipod Microdeutopus gryllotalpa;; DelValls et al., 2002).
In this study LAS, Pb and Ag were grouped together with the biological effects. For LAS
the analysis resulted in the SQV “Not polluted” (low or minimal biological effects) of <2.6
mg/kg dw, and the SQV “Highly polluted” (high biological effects) for concentrations
>8.7 mg/kg dw.
Table 5. Chronic toxicity of sediments spiked with LAS (mg/kg dw) towards benthic freshwater
species.mg/kg dw
Species
Duration
Effect Value
Lumbriculus variegatus
28 days
NOEC
Endpoint
81 survival/reproduction
Comment
Reference
1.7% OC
Comber et al., 2006
Caenorhabditis elegans 72 h
NOEC
100 egg production
2% OC
Chironomus riparius
10 days
NOEC
348 head capsule length
C12-LAS, 3.2% OC
Comber et al., 2006
Mäenpää &
Kukkonen 2006
Chironomus riparius
24 days
NOEC
319 emergence
4.2% OC
Pittinger et al., 1989
Table 6. Acute and chronic toxicity of sediments spiked with LAS (mg/kg dw) towards benthic marine
species.
Acute
Species
Duration
Effect
Value
Hydrobia ulvae
96 h
LC50
140.65 mortality
Hydrobia ulvae
96 h
LC50
101.77 mortality
Solea senegalensis
Cylindrotheca
closterium
Cylindrotheca
closterium
Phaeodactylum
tricornutum
96 h
LC50
2179.68 mortality
72 h
ErC50
17 growth rate
72 h
4.18 growth rate
Moreno-Garrido et al., 2003
72 h
EbC50
63%
inhib.
4.77 growth rate
Moreno-Garrido et al., 2007
Corophium volutator
5 days
LC50
295 mortality
31
Endpoint
Comment
Reference
0.618% OC
Hampel et al., 2007
Hampel et al., 2009
0.618% OC
Hampel et al., 2007
0.59% OC
Mauffret et al., 2010b
C12-LAS,
Rico-Rico et al., 2009
.
IVL report
1.38% OC
Chronic
C12-LAS,
0.06% OC
Rico-Rico et al., 2009
0.618% OC
Hampel et al., 2007
0.59% OC
Mauffret et al., 2010a
48.82 mortality
0.618% OC
Hampel et al., 2007
362.99 mortality
0.618% OC
Hampel et al., 2007
Corophium volutator
Ruditapes
philippinarum
5 days
LC50
162 mortality
30 days
LC10
560.52 mortality
Hydrobia ulvae
10 days
LC10
124 mortality
Hydrobia ulvae
9 days
LC10
Solea senegalesis
30 days
LC10
Solea senegalesis
30 days
NOEC
92.29 mortality
Hampel et al., 2008
Solea senegalesis
30 days
LOEC
222.66 mortality
Hampel et al., 2008
The concentration in the sediments in Lake Mälaren and Torsbyfjärden were considerably
lower than in Saltsjön, which indicate that a rapid sedimentation of LAS occurs near the
source, the WWTP. Exposure to benthic organisms is mainly relevant in oxic sediments,
provided that anoxic sediment does not become oxygenated. In Saltsjön the concentrations
in the surface, 3.9 – 7.9 mg/kg dw, was within the range of the PNEC suggested by
Hampel et al. (2007) but below the PNEC suggested by Comber et al. (2006). The highest
risk ratios (measured concentration/PNEC) for oxic sediments gives values in the range of
1.0 – 1.4 in the surface and a value up to 2.3 in the deeper layers. The highest risk ratios in
the anoxic sediments range between 1.5 - 2.5 in the surface and 8.9 in the deeper sections.
Thus, the results indicate that the presence of LAS in sediments in the area Saltsjön may
have a negative impact on benthic organisms today in the local area in the immediatie
vicinity of the discharge points of the two WWTPs of Stockholm city, Henriksdalsverket
and Brommaverket.The exact size of this area is uknown but is probably of a similar size to
that of nonylphenols, about 1 km2 (A. Jönsson unpublished results).
As pointed out by Woldegiorgis (2011), sediment toxicity data for the other anionic
surfactants alkyl sulphates (AS) and alkyl ether sulphates (AES) and amphoteric
cocoamidopropyl betaine (CAPB), are very scarce to non-existent but should be
considered to have similar toxicity compared to LAS based on toxicity data for water. In
the case of AS it could also be considered to have higher ecotoxicity for both water,
sediment and soil based on the limited data available (Fraunhofer Institut, 2003). The
proposed sediment predicted no effect concentrations (PNEC) in g/g dry weight for AS
varies depending on chain length: 0.0963 (12 carbon atoms), 0.0406 (13 carbon atoms),
0.0186 (14 carbon atoms), 0.22 (15 carbon atoms), 0,468 (16 carbon atoms) and 8.4 (18
carbon atoms). The three shortest chain lengths (C12 – C14) make up about 61% of the
AS used in EU, Norway and Switzerland, whereas chain lengths of 15, 16 and 18 carbon
atoms make up the remaining 39%. Thus, the concentrations found in the surface
sediments in Saltsjön at S1 (Fig. 14) and Torsbyfjärden at T1 (Fig. 16) indicate that there is
a risk for local negative environmental effects. The risk ratios can be as high as
0.322/0.0186 = 17 at T1 and 0.118/0.0186=6 at S1. These risk ratios are higher than the
ones for LAS identified in this study. Thus on a local scale, the environmental threat to the
benthic fauna can be larger from the surfactant AS than it is from LAS.
The findings of this study corroborate one of the hypotheses of Woldegiorigis (2011):
32
.
IVL report
The environmental risk for the benthic fauna on a local scale is equal or greater from other
surfactants, in this case AS, than that of LAS. This is true even when considering the
higher previous loads of LAS as indicated in the sediment record.
Conclusions

The main factor determining sediment concentrations of linear alkylbenzne
sulphonate (LAS) in the Stockholm area is where sedimentation of organic carbon
occurs and distance from the point of waste water discharge.

The redox environment (oxic/anoxic) of the surface sediments does not influence
the distribution of LAS on a kilometre wide scale.

There seems to be anaerobic degradation of LAS occurring in the sediment of,
Lake Mälaren and inner Stockholm archipelago (Saltsjön and Torsbyfjärden).

There are potential local environmental threats to the benthic fauna due to the
concentrations of LAS and alkyl sulphates (AS) in the sediments where the two
WWTPs of the city of Stockholm discharge their treated waste water. In the case of
AS, this threat also exists also about 20km further downstream in Torsbyfjärden,
close to the discharge point of the WWTP Käppala.

The study corroborates the two main hypotheses of Woldegiorgis (2011):
o Sediment concentrations of LAS may pose an environmental risk for the
benthic fauna only on a local scale close to the discharge points of WWTPs
where organic matter accumulates and this will not change after a return to
previous higher levels of LAS consumption
o If LAS is replaced by another surfactant the environmental threat to the
benthic fauna on this local scale will be shifted from LAS to the other
surfactant since the overall use of surfactants will not decrease. In the
present case, this threat comes from AS
33
.
IVL report
References
Acmite, 2010. Market report – World surfactant market. Acmite Market Intelligence,
Ratingen, Germany.
Appelo, C.A.J., Postma, D., 1999. A consisten model for surface complexation of
Birnessite (-MnO2) and its application to a column experiment. Geochimica et
Cosmochimica acta. 63: 3039-3048.
Aller, R.C., Hall, P. O. J., Rude, P.D., Aller, J.Y. 1998. Biogeochemical heterogeneity and
suboxic diagenesis in hemipelagic sediments of the Panama Basin. Deep-Sea Research
I 45:133-165.
Bester, K., Theobald, N., Schröder, H. 2001. Nonylphenol-ethoxylates, linear
alkylbenzenesulfonates (LAS) and bis (4-chlorophenyl)-sulfone in the German Bight of
the North Sea. Chemosphere, 45: 817-826.
Berner, R.A. 1980. Early diagenesis – A theoretical approach. Princeton University Press,
Princeton N. J.
Bradshaw, C., Kumblad, L., Fagrell, A. The use of tracers to evaluate the importance of
bioturbation in remobilising contaminants in Baltic sediments. Estuarine, Coastal and
Shelf Science. 66: 123-134.
Calvert, S.E., Pedersen, T.F., Karlin, R.E. 2001. Geochemical and isotopic evidence for
post-glacial palaeoceanographic changes in Saanich Inlet, British Columbia. Marine
Geology. 174: 287-305.
34
.
IVL report
Canfield, D.E., Thamdrup, B., Hansen, J. W., 1993. The anaerobic degradation of organic
matter in Danish coastal sediments: Iron reduction, manganese reduction, and
sulphate reduction. Geochimica et Cosmochimica acta. 57: 3867-3883.
Ceresana Research, 2012. Market Study: Surfactants. Ceresana, Konstanz, Germany.
Fischer, T.B., Heaney, P.J., Jang, J-H., Ross, D.E., Brantley, S.L., Post, J. E., Tien, M. 2008.
Continuous time-resolved X-ray diffraction of the biocatalyzed reduction of Mn oxide.
American Mineralogist. 93: 1929–1932.
Comber, S. D. W., Conrad, A. U., Höss, S., Webb, S., Marshall, S. 2006. Chronic toxicity of
sediment-associated linear alkylbenzene sulphonates (LAS) to freshwater benthic
organisms. Environ. Pollut. 144: 661-668.
DelValls, T.Á., Forja, J.M., Gómez-Parra, A. 2002. Seasonality of contamination, toxicity,
and quality values in sediments from littoral ecosystems in the Gulf of Cádiz (SW
Spain). Chemosphere, 46: 1033-1043.
Folke, J., Cassani, G., Ferrer, J., Lopez, I., Karlsson, M., Willumsen, B. 2003. Linear
alkylbenzene sulphonates, branched dodecylbenzene sulphonates and soap analysed in
marine sediments from the Baltic Proper and Little Belt. Tenside Surf. Det. 40: 17-24.
Fraunhofer Institut, 2003. Anaerobic biodegradation of detergent surfactants – final report.
Fraunhofer Institut fur Umwelt-, Sicherheits- und Energietechnik,. Oberhausen
Germany.
Gonzalez-Mazo, E., Forja, J. M., Gomez-Parra, A. 1998. Fate and distribution of linear
Alkylbenzene sulfonates in the litoral environment. Environ. Sci. Technol. 32:16361641.
Hampel, M., Gonzáles-Mazo, E., Vale, C., Blasco, J. 2007. Derivation of predicted no
effect concentrations (PNEC) for marine environmental risk assessment: Application
of different approaches to the model contaminant Linear Alkylbenzene Sulphonates
(LAS) in a site-specific environment. Environ. Int. 33: 486-491.
Hampel, M., Ortiz-Delgado, J.B., Sarasquete, C., Blasco, J. 2008. Effects of sediment
sorbed linear alkylbenzene sulphonate on juveniles of the Senegal sole, Solea senegalensis:
Toxicity and histological indicators. Histol. Histopathol. 23: 87:100.
Hampel, Canário, J., Branco, V., Vale, C., Blasco, J. 2008b. Environmental levels of linear
alkylbenzene sulfonates (LAS) insediments from the Tagus estuary (Portugal):
environmental implications. Environ. Monit. Assess. DOI 10.1007/s10661-008-0190-0
Hampel, M., Moreno-Garrido, I., Gonzáles-Mazo, E., Blasco, J. 2009. Suitability of the
marine prosobranch snail Hydrobia ulvae for sediment toxicity assessment: A case study
with the anionic surfactant linear alkylbenzene sulphonate (LAS). Ecotox. Environ.
Safety, 72: 1303-1308.
Hyacinthe, C., Anschutz, P., Carbonel, P., Jouanneau, J.-M., Jorissen, F.J. 2001. Early
diagenetic processes in the muddy sediments of the Bay of Biscay. Marine Geology
177: 111-128.
35
.
IVL report
Hyacinthe, C., Van Cappellen, P. 2004. An authigenic iron phosphate phase in estuarine
sediments: composition, formation and chemical reactivity. Marine Chemistry 91: 227–
251.
Human & Environmental Risk Assessments (HERA) on ingredients of European
household cleaning products – Alkyl Sulphates (AS) Environmental Risk Assessment.
2002.
Human & Environmental Risk Assessments (HERA) on ingredients of European
household cleaning products – Alcohol Ethoxysulphates (AES) Environmental Risk
Assessment. 2004.
Human & Environmental Risk Assessments (HERA) on ingredients of European
household cleaning products – Linear Alkylbenzene Sulphonate (LAS) Environmental
Risk Assessment. 2009.
Jensen, A., Gustavson, G. 2001. Havnesedimenters indhold af miljøfremmede organiske
forbindelser. Kortlægning af nuværende og fremtidige behov for klapning og
deponering. Miljøprojekt Nr. 627. Miljøstyrelsen, Miljø- og Energiministeriet.
Jönsson, A., Lindström, M., Carman, R., Mört, C-M, Meili, M and Ö Gustafsson. 2005.
Evaluation of the Stockholm Archipelago Northwestern Baltic Sea Proper, as a trap for freshwaterrunoff organic carbon. Journal of Marine Systems, 56: 167 – 178.
Jönsson, A. 2011. Ni, Cu, Zn, Cd and Pb in sediments in the city-centre of Stockholm,
Sweden Origins, deposition rates and bio-availability. IVL-B2013.
Kabata-Pendias, A. 2001. Trace elements in soils and plants, 3rd ed. CRC Press, Boca
Raton.
Kaj, L., Lilja, K., Remberger, M., Allard, A-S., Dusan, B., Brorström-Lundén, E. 2008.
Results from the Swedish National Screening Programme 2007. Subreport 4: Linear
alkyl benzene sulfonate (LAS). IVL Report B1808. October 2008.
Karlsson, M., Malmaeus, M., Rydin, E., Jonsson, P. 2010. Bottenundersökningar i Upplands,
Stockholms, Södermanlands och Östergötlands skärgårdar .2008-2009. IVL, report B1928, in
Swedish.
KEMI, 2012a. Flow analyses for chemical substances: Linear alkylbenzene sulfonates.
http://apps.kemi.se/flodessok/floden/flodessok.cfm?sokval
KEMI, 2012b. Flow analyses for chemical substances: Sodium dodecyl sulphate.
http://apps.kemi.se/flodessok/floden/flodessok.cfm?sokval
KEMI, 2012c. Flow analyses for chemical substances: sodium dodecyl eter sulphate.
http://apps.kemi.se/flodessok/floden/flodessok.cfm?sokval
KEMI, 2012d. Flow analyses for chemical substances: Coco amidopropyl betaine.
http://apps.kemi.se/flodessok/floden/flodessok.cfm?sokval
Kiirikki, M., Lehtoranta, J., Inkala, A., Pitkänen, H., Hietanen, S., Hall, P.O.J., Tengberg,
A., Koponen, J., Sarkkula, J. A simple sediment process description suitable for 3Decosystem modelling — Development and testing in the Gulf of Finland. Journal of
Marine Systems 61: 55–66.
36
.
IVL report
Lara-Martín, P.A., Gómez-Parra, A., González-Mazo, E. 2005. Determination and
distribution of alkyl ethoxylates and linear alkylbenzene sulfonates in coastal marine
sediments from the bay of Cadiz (southwest of Spain). Env. Toxicol. Chem. 24:21962202.
Lara-Martín, P. A., Petrovic, M., Gómez-Parra, A., Barceló, D., González-Mazo, E. 2006.
Presence of surfactants and their degradation intermediates in sediment cores and
grabs from the Cadiz Bay area. Environ. Pollut. 144: 483-491.
Lara-Martín, P. A., Gómez-Parra, A., Köchling, T., Sanz, J. L., Amils, R., González-Mazo,
E. 2007. Anaerobic degradation of linear alkylbenzene sulfonates in coastal marine
sediments. Environ. Sci. Technol. 41: 3571-3579.
León, V.M., Gonzáles-Mazo, E., Pajares, J.M.F., Gómez-Parra, A. 2001. Vertical
distribution profiles of linear alkylbenzene sulfonates and their long-chain intermediate
degradation products in coastal marine sediments. Environ. Toxicol. Chem. 20: 21712178.
Mäenpää, K., Kukkonen, J.V.K. 2006. Bioaccumulation and toxicity of 4-nonylphenol (4NP) and 4-(2-dodecyl)-benzene sulfonate (LAS) in Lumbriculus variegatus
(Oligochaeta) and Chironomus riparius (Insecta). Aquat. Toxicol. 77: 329-338.
Magen, C., Mucci, A., Sundby, B. Reduction rates of sedimentary Mn and Fe Oxides: An
incubation experiment with Arctic Ocean sediments. Aquat Geochem. 17: 629–643.
Mauffret, A., Moreno-Garrido, I., Blasco, J. 2010a. The use of marine benthic diatoms in a
growth inhibition test with spiked whole-sediment. Ecotox. Environ. Safety, 73: 262269.
Mauffret, A., Temara, A., Blasco, J. 2010b. Exposure of the marine deposit feeder Hydrobia
ulvae to sediment spiked with LAS homologs. Water Res. 44: 2831-2840.
Moreno-Garrido, I., Hampel, M., Lubián, L.M., Blasco, J. 2003. Marine benthic microalgae
Cylindrotheca closterium (Ehremberg) Lewin and Reimann (Bacillariophyceae) as a
tool for measuring toxicity of linear alkylbenzene sulfonate in sediments. Bull.
Environ. Contam. Toxicol. 70: 242-247.
Moreno-Garido, I., Lubián, L.M., Blasco, J. 2007. Sediment toxicity tests involving
immobilized microalgae (Phaeodactylum tricornutum Bohlin). Environ. Int. 33: 481-495.
Morford, J. L., Russel, A.D., Emerson, S. 2001. Trace metal evidence for changes in the
redox environment associatied with the transition from terrigenous clay to
diatomaceous sediment, Saanich Inlet, BC. Marine geology. 174: 355-369.
Mouret, A., Anschutz, P., Lecroart, P., Chaillou, G., Hyacinthe, C., Deborde, J., Jorissen, F.
J., Deflandre, B., Schmidt, S., Jouanneau, J-M., 2009.Benthic geochemistry of
manganese in the Bay of Biscay, and sediment mass accumulation rate. Geo-Mar. Lett.
29:133-149.
Mucci, A., Edelborn, H.M. 1992. Influence of an organic-poor landslide on the early
diagenesis of iron and manganese in a coastal marine sediment. Geochimica et
Cosmochimica Acta, 56: 3909 – 3921.
37
.
IVL report
Naturskyddsföreningen (2006) Kemiska Produkter.Version 2006:4
Nordic Ecolabelling (2012) Laundry detergents and stain removers. Version 7.1. 15
December 2011-31 December 2015.
Nerpin, L., Nordell, O., Burgdorf Nielsen, J., Hein, M., Carlsson, C., Bjerre, F., BrønsHansen, J. 2005. Miljögifter i Öresund, en översikt. Öresundsvattensamanbetet.
www.oresundvand.dk
Petrovic, M., Fernández-Alba, A.R., Borrull, F., Marce, R.M., González-Mazo, E., Barceló, D.
2002. Occurence and distribution of nonionic surfactants, their degradation products, and
linear alkylbenzene sulfonates in coastal waters and sediments in Spain. Env. Toxicol.
Chem. 21: 37-46.
Pittinger, C.A., Woltering, D.M., Masters, J.A. 1989. Bioavailability of sediment-sorbed and
aqueous surfactants to Chironomus riparius (midge). Environ. Toxicol. Chem. 8: 10231033.
Reed, D.C., Slomp, C.P., Gustafsson, B. G. 2011. Sedimentary phosphorus dynamics and
the evolution of bottom-water hypoxia: A coupled benthic–pelagic model of a coastal
system. Limnol. Oceanogr., 56(3): 1075–1092
Rico-Rico, Á., Temara, A., Hermens, J.L.M. 2009. Equilibrium partitioning theory to
predict the sediment toxicity of the anionic surfactant C12-2-LAS to Corophium
volutator. Environ. Pollut. 157: 575-581.
SSNC, 1999. Hushållskemikalier i förändring. En statistisk jämförelse mellan åren 1988 och
1996. Svenska Naturskyddsföreningen 1999. In Swedish.
Thiessen, O., Schmidt, M., Theilen, F., Schmitt, M., Klein, G. 2006. Methane formation
and distribution of acoustic turbidity in organic-rich sediments in the Arkona Basin,
Baltic Sea. Continental Shelf Research, 26: 2469-2483.
Tromp, T., K., Van Cappellen, P., Key, R.M., 1995. A global model for the early diagenesis
of organic carbonand organic phosphorus in marine sediments. Geochimica et
Cosmochimica Acta, 59 (7): 1259 – 1284.
Turner, A., Millward, G.E. Particle Dynamics and Trace Metal Reactivity in Estuarine
Plumes. Estuarine, Coastal and Shelf Science, 50: 761–774.
Wang, Y., van Cappellen, P. 1996. A multicomponent reactive transport model of early
diagenesis: Application to redox cycling in coastal marine sediments. Geochimica et
Cosmochimica Acta, 60 (16): 2993-3014.
Whiticar, M. J. 1999. Carbon and hydrogen isotope systematics of bacterial formation and
oxidation of methane. Chemical Geology, 161: 291-314.
Woldegiorgis, 2011. Environmental risk and hazard assessment of LAS compared to other
anionic surfactants- Scenario-based waiving of LAS in a Swedish perspective. . IVL,
report Bxxxx. In English.
Östlund, P., Sternbeck, J., Brorström-Lundén, E. 1998. Metaller, PAH, PCB och
totalkolväten i sediment runt Stockholm – flöden och halter. IVL, report B1297. In
Swedish.
38
.
IVL report
Appendix
Table A1. Cs137-activity data.
Lokal
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
M1
Skikt (cm)
0-2
2-4
4-6
6-8
8-10
10-11
11-12
12-13
13-14
14-15
15-16
16-17
17-18
18-19
19-20
20-21
21-22
22-23
23-24
24-25
25-26
26-27
27-28
28-29
CPM
20.7
25.7
30.7
23.9
17.6
13.9
12.5
13.5
16.3
18.0
12.8
14.1
13.2
13.8
11.5
8.8
8.9
7.1
8.8
15.8
22.3
22.7
19.4
23.1
Lokal
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
M2
Skikt (cm)
0-2
2-4
4-6
6-8
8-10
10-11
11-12
12-13
13-14
14-15
15-16
16-17
17-18
18-19
19-20
20-21
21-22
22-23
23-24
24-25
25-26
26-27
27-28
28-29
29-30
30-31
31-32
32-33
33-34
34-35
35-36
36-37
37-38
38-39
39-40
40-41
41-42
42-43
43-44
44-45
45-46
46-47
47-48
48-49
CPM
16.9
19.3
20.8
21.0
31.6
29.4
31.4
29.9
42.3
29.0
25.5
18.6
16.4
13.8
10.9
15.7
9.7
9.5
10.2
12.2
11.0
13.3
12.0
10.0
14.0
14.3
15.7
12.2
13.9
14.3
16.4
13.4
15.9
13.9
17.5
17.5
14.1
14.4
12.9
14.4
14.2
11.5
13.7
11.9
Lokal
S1
S1
S1
S1
S1
S1
S1
S1
S1
S1
S1
S1
S1
S1
S1
S1
Skikt (cm)
0-2
2-4
4-6
6-8
8-10
10-11
11-12
12-13
13-14
14-15
15-16
16-17
17-18
18-19
19-20
20-21
CPM
23.9
26.7
26.6
32.0
25.9
26.8
27.6
36.4
36.2
31.4
19.9
15.1
12.9
9.2
10.9
14.6
Lokal
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
S2
Skikt (cm)
0-2
2-4
4-6
6-8
8-10
10-11
11-12
12-13
13-14
14-15
15-16
16-17
17-18
18-19
19-20
20-21
21-22
22-23
23-24
24-25
25-26
26-27
27-28
28-29
29-30
30-31
31-32
CPM
18.9
14.5
13.9
15.9
17.3
11.0
7.0
9.0
8.9
8.2
6.3
5.5
6.6
8.0
9.4
6.9
6.7
7.9
8.1
10.2
8.1
7.9
10.9
9.2
12.7
12.1
3.9
Lokal
T1
T1
T1
T1
T1
T1
T1
T1
T1
T1
T1
T1
T1
T1
T1
T1
T1
Skikt (cm)
0-2
2-4
4-6
6-8
8-10
10-11
11-12
12-13
13-14
14-15
15-16
16-17
17-18
18-19
19-20
20-21
21-22
CPM
28.9
36.4
43.6
46.6
45.4
18.0
22.6
26.3
20.1
25.3
26.0
18.5
20.3
20.4
14.6
28.5
32.5
Lokal
T2
T2
T2
T2
T2
T2
T2
T2
T2
T2
T2
T2
T2
T2
T2
T2
T2
T2
Skikt (cm)
0-2
2-4
4-6
6-8
8-10
10-11
11-12
12-13
13-14
14-15
15-16
16-17
17-18
18-19
19-20
20-21
21-22
22-23
CPM
40.0
43.9
30.4
20.3
13.5
10.7
11.2
10.3
8.6
9.5
10.6
8.5
8.2
8.0
10.2
12.8
13.1
16.9
Table A2 Mn-speciation data: Mn-Asc and Mn-HCl concentrations and inventories in Lake
Mälaren sediment cores
Core
depth
Mn-Asc
Mn-HCl
(mmol/g1) (mmol/g1)
Mn2+
(mM)
% dm
39

I-Asc
I-HCl
(mmol/cm2) (mmol/cm2)
.
IVL report
(cm)
M1:0-1
49
46
16
0.84
21
20
M1:1-6
41
39
19
0.81
104
99
M1:6-11
25
28
24
0.76
79
86
M1:11-16
27
40
20
0.80
73
106
M2:0-5
127
36
20
10
0.90
176
50
25
20
16
0.84
M2:5-10
52
M2:10-15
64
30
21
18
0.82
157
74
M2:15-20
73
23
22
20
0.80
193
61
M2:20-28
51
30
21
20
0.80
214
125
Table A3. Mn-speciation data: Mn-Asc and Mn-HCl concentrations and inventories in
Saltsjön sediment cores
Core
Mn-Asc
Mn-HCl
(mmol/g) (mmol/g)
Mn2+
(mM)
%ts
40

I-Asc
(mmol/cm2)
I-Hcl
(mmol/cm2)
.
IVL report
depth
(cm)
S2: 0-2
12
10
12
53
0.47
35
28
S2:2-4
24
13
19
56
0.44
70
37
S2:4-6
11
14
8.2
34
0.66
20
25
S2:6-11
11
19
5.9
24
0.76
34
58
S2:11-16
9
17
4.9
28
0.72
32
62
S2:16-21
11
20
7.8
21
0.79
31
54
S2:21-26
9
20
9.8
26
0.74
31
68
S1:0-5
12
16
14
0.86
22
30
S1:5-10
8.4
18
20
0.80
23
48
S1:10-15
30
24
14
0.86
55
44
S1:15-20
31
24
12
0.88
51
39
S1:20-25
29
23
11
0.89
45
34
S1:25-30
39
28
13
0.87
66
48
41
.
IVL report
Table A4. Mn-speciation data: Mn-Asc and Mn-HCl concentrations and inventories in
Torsbyfjärden sediment cores
Core
depth
Mn-Asc
(mmol/g)
Mn-HCl Mn2+
(mmol/g) (mM)
%ts

I-Asc
I-Hcl
2
(mmol/cm ) (mmol/cm2)
(cm)
T2: 0-5
13
19
17
0.83
30
44
T2: 5-10
13
23
19
0.81
33
57
T2:10-15
16
21
15
0.85
32
42
T2:15-20
20
27
16
0.84
41
56
T2:20-25
21
29
17
0.83
47
67
T1:0-4
40
44
26
17
0.83
71
78
T1:4-9
17
34
19
17
0.83
40
79
T1:9-14
9.3
37
22
17
0.83
21
82
T1:14-19
32
43
33
19
0.81
82
109
T1:19-24
37
38
45
20
0.80
98
101
T1:24-29
29
37
57
21
0.79
80
103
T1:29-34
33
38
65
22
0.79
93
108
42
.
IVL report
Appendix A5. Measured concentration of the different homologs of linear alkyl benzene
sulfonate (LAS) as a function of sediment depth and oxic condition.
Sample id
Sampling site
Sedimentcor
e
Index
MnM1 1521
CsM1 0-2
CsM1 2-4
CsM1 4-6
MnM2 1825
MnM2 1318
MnM2 8-13
CsM2 0-4
CsM2 8-12
CsM2 12-17
Mälaren
Core 1
Mälaren
Mälaren
Mälaren
Core 2
Core 2
Core 2
Mälaren
Core 1
Mälaren
Core 1
Mälaren
Mälaren
Mälaren
Mälaren
Torsbyfjärde
n
Torsbyfjärde
MnT1 20-30
n
Torsbyfjärde
CsT1 0-2
n
Torsbyfjärde
CsT1 2-14
n
MnT1 30-34
Torsbyfjärde
n
Torsbyfjärde
MnT2 15-20
n
Torsbyfjärde
CsT2 0-2
n
Torsbyfjärde
CsT2 2-6
n
MnT2 20-25
Core 1
Core 2
Core 2
Core 2
Environmen
t
Oxic/Anoxi
c
Dept
h
10
Carbons
Concentration of LAS
11
12
13
Carbons
Carbons
Carbons
cm
ng/g dw
ng/g dw
ng/g dw
ng/g dw
ng/g dw
oxic
0-6
19
78
131
217
444
0-2
2-4
4-6
3.2
1.7
3.1
20
15
26
35
24
54
51
41
74
109
82
157
0-7
21
56
57
85
219
7-12
6.8
38
58
90
193
12-17
0-4
8-12
12-17
10
2.9
3.1
6.5
51
16
30
55
91
27
53
113
139
54
85
182
291
100
170
357
oxic
oxic
oxic
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
ΣHomolog
s
Core 1
Anoxic
0-4
33
90
82
64
270
Core 1
Anoxic
4-14
5.3
26
44
50
127
Core 2
Anoxic
0-2
0.5
<2.2
<5.9
<7.9
<11
Core 2
Anoxic
2-14
<0.4
3.3
<5.9
<7.9
<11
Core 1
Oxic
0-5
2.1
9.9
16
20
48
Core 1
Oxic
5-10
4.8
19
40
26
89
Core 2
Oxic
0-2
55
212
324
429
1020
Core 2
Oxic
2-6
4.5
16
26
24
71
MnS1 25-30
MnS1 20-25
MnS1 10-15
CsS1 0-2
CsS1 6-8
CsS1 12-15
CsS1 15-18
CsS1 18-21
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Core 1
Core 1
Core 1
Core 2
Core 2
Core 2
Core 2
Core 2
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
0-5
5-10
15-20
0-2
6-8
12-15
15-18
18-21
46
204
2723
121
1000
1013
480
1000
622
1592
15299
1699
10599
9432
3566
5499
2147
4264
10797
4164
8630
8030
6230
4530
4229
8296
14696
6096
11729
10996
8229
5396
7045
14356
43515
12080
31958
29472
18505
16425
MnS2 24-26
MnS2 20-24
MnS2 15-20
CsS2 0-2
CsS2 2-6
CsS2 6-10
CsS2 10-12
CsS2 14-16
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Core 1
Core 1
Core 1
Core 2
Core 2
Core 2
Core 2
Core 2
Oxic
Oxic
Oxic
Oxic
Oxic
Oxic
Oxic
Oxic
0-2
2-6
6-11
0-2
2-6
6-10
10-12
14-16
14
33
86
15
129
95
76
63
324
622
1062
244
1212
676
489
306
1934
2534
3430
1607
3497
2127
1814
1567
1673
2739
3729
5996
6496
4629
3463
4429
3944
5929
8308
7862
11335
7527
5841
6365
*Smaller than (<) values refers to lowest detection limit (LOD= S/N).
43
.
IVL report
Appendix A6. Measured concentration of alkyl sulphate (AS) and cocoamidopropyl
betaine (CAPB) as a function of sediment depth and oxic condition.
Sample id
Sampling site
Mälaren
Mälaren
Mälaren
Mälaren
Sedimentcore
Index
Core 1
Core 2
Core 2
Core 2
Environment
Oxic/Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
Depth
cm
0-6
0-2
2-4
4-6
MnM1 15-21
CsM1 0-2
CsM1 2-4
CsM1 4-6
MnM2 18-25
MnM2 13-18
MnM2 8-13
CsM2 0-4
CsM2 8-12
CsM2 12-17
Mälaren
Mälaren
Mälaren
Mälaren
Mälaren
Mälaren
Core 1
Core 1
Core 1
Core 2
Core 2
Core 2
Oxic
Oxic
Oxic
Oxic
Oxic
Oxic
0-7
7-12
12-17
0-4
8-12
12-17
<22
<22
<22
25
<22
<22
<11
<11
<11
<11
<11
<11
MnT1 30-34
MnT1 25-30
MnT1 20-25
CsT1 0-2
CsT1 2-4
CsT1 4-8
CsT1 8-14
Torsbyfjärden
Torsbyfjärden
Torsbyfjärden
Torsbyfjärden
Torsbyfjärden
Torsbyfjärden
Torsbyfjärden
Core 1
Core 1
Core 1
Core 2
Core 2
Core 2
Core 2
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
0-4
4-9
9-14
0-2
2-4
4-8
8-14
<22
<22
<22
94
322
<22
<22
<11
<11
<11
<11
<11
<11
<11
MnT2 20-25
CsT2 0-2
CsT2 2-4
CsT2 4-6
Torsbyfjärden
Torsbyfjärden
Torsbyfjärden
Torsbyfjärden
Core 1
Core 2
Core 2
Core 2
Oxic
Oxic
Oxic
Oxic
0-5
0-2
2-4
4-6
<22
<22
<22
<22
<11
<11
<11
<11
MnS1 25-30
MnS1 20-25
MnS1 10-15
CsS1 0-2
CsS1 6-8
CsS1 12-15
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Core 1
Core 1
Core 1
Core 2
Core 2
Core 2
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
0-5
5-10
10-15
0-2
6-8
12-15
118
37
66
70
42
22
<11
<11
<11
<11
<11
<11
MnS2 24-26
CsS2 0-2
Saltsjön
Saltsjön
Core 1
Core 2
Oxic
Oxic
0-2
0-2
<22
32
<11
<11
*Smaller than (<) values refers to lowest detection limit (LOD= S/N).
44
Concentration of AS
ng/g dw
<22
26
<22
<22
Concentration of CAPB
ng/g dw
<11
<11
<11
<11
.
IVL report
Appendix A7. Measured concentration of the different homologs of alkyl ether sulphate
(AES) as a function of sediment depth and oxic condition.
Concentration of AES
Environment Depth 1 Ethoxylate 2 Ethoxylates 3 Ethoxylates 4 Ethoxylates ΣHomologs
Oxic/Anoxic
cm
ng/g dw
ng/g dw
ng/g dw
ng/g dw
ng/g dw
Anoxic
0-6
<20
<26
<12
<4.5
<62
Anoxic
0-2
<20
<26
<12
4.8
<62
Anoxic
2-4
<20
<26
<12
<4.5
<62
Anoxic
4-6
<20
<26
<12
<4.5
<62
Sample id
Sampling site
MnM1 15-21
CsM1 0-2
CsM1 2-4
CsM1 4-6
Mälaren
Mälaren
Mälaren
Mälaren
Sedimentcore
Index
Core 1
Core 2
Core 2
Core 2
MnM2 18-25
MnM2 13-18
MnM2 8-13
CsM2 0-4
CsM2 8-12
CsM2 12-17
Mälaren
Mälaren
Mälaren
Mälaren
Mälaren
Mälaren
Core 1
Core 1
Core 1
Core 2
Core 2
Core 2
Oxic
Oxic
Oxic
Oxic
Oxic
Oxic
0-7
7-12
12-17
0-4
8-12
12-17
<20
<20
<20
<20
<20
<20
<26
<26
<26
<26
<26
<26
21
<12
<12
<12
<12
<12
4.6
<4.5
<4.5
<4.5
<4.5
5.5
<62
<62
<62
<62
<62
<62
MnT1 30-34
MnT1 25-30
MnT1 20-25
CsT1 0-2
CsT1 2-4
CsT1 4-8
CsT1 8-14
Torsbyfjärden
Torsbyfjärden
Torsbyfjärden
Torsbyfjärden
Torsbyfjärden
Torsbyfjärden
Torsbyfjärden
Core 1
Core 1
Core 1
Core 2
Core 2
Core 2
Core 2
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
0-4
4-9
9-14
0-2
2-4
4-8
8-14
<20
<20
<20
<20
<20
<20
<20
<26
<26
<26
<26
<26
<26
<26
21
16
<12
<12
<12
<12
<12
13
11
<4.5
<4.5
<4.5
<4.5
<4.5
<62
<62
<62
<62
<62
<62
<62
MnT2 20-25
CsT2 0-2
CsT2 2-4
CsT2 4-6
Torsbyfjärden
Torsbyfjärden
Torsbyfjärden
Torsbyfjärden
Core 1
Core 2
Core 2
Core 2
Oxic
Oxic
Oxic
Oxic
0-5
0-2
2-4
4-6
<20
<20
<20
<20
<26
<26
<26
<26
<12
<12
<12
<12
<4.5
<4.5
4.9
<4.5
<62
<62
<62
<62
MnS1 25-30
MnS1 20-25
MnS1 10-15
CsS1 0-2
CsS1 6-8
CsS1 12-15
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Saltsjön
Core 1
Core 1
Core 1
Core 2
Core 2
Core 2
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
Anoxic
0-5
5-10
10-15
0-2
6-8
12-15
56
21
56
32.2
22.2
<20
33
32
<26
<26
<26
<26
32
16
26
14
<12
<12
4.3
5.0
10
<4.5
5.2
8.4
126
74
92
<62
<62
<62
MnS2 24-26
CsS2 0-2
Saltsjön
Saltsjön
Core 1
Core 2
Oxic
Oxic
0-2
0-2
<20
<20
<26
<26
<12
<12
<4.5
<4.5
<62
<62
*Smaller than (<) values refers to lowest detection limit (LOD= S/N).
45