DOCTORA L T H E S I S

Department of Civil, Environmental and Natural Resources EngineeringDivision of Geosciences and Environmental Engineering

Water and Sediment Quality of Urban Water Bodies

in Cold Climates

Ralf Rentz

ISSN: 1402-1544 ISBN 978-91-7439-272-2

Luleå University of Technology 2011

ISSN: 1402-1544 ISBN 978-91-7439-XXX-X Se i listan och fyll i siffror där kryssen är

Ralf R

entz Water and Sedim

ent Quality of U

rban Water B

odies in Cold C

limates

1

Water and Sediment Quality

of Urban Water Bodies

in Cold Climates

Ralf Rentz

Division of Geosciences and Environmental Engineering Department of Civil, Environmental and Natural Resources Engineering Luleå University of Technology SE-97187 Luleå, Sweden Luleå 2011

Printed by Universitetstryckeriet, Luleå 2011

ISSN: 1402-1544 ISBN 978-91-7439-272-2

Luleå 2011

www.ltu.se

Cover Picture: Skutviken Panorama

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AbstractThe aim of this study was to investigate and quantify pollution impact on urban water

bodies in cold climates and to find out which complex processes and influencing factors cause trapping or spread of pollutants. In order to do that water, sediment and porewater samples from bays and stormwater ditches in Luleå, northern Sweden, and from an artificial stormwater pond in Sollentuna, south-central Sweden, were analysed for LOI, trace metals and PAHs. For surface water the particular, colloidal and truly dissolved element concentrations were determined by membrane filtration (0.22 μm pore size, 142 mm diameter, Millipore® mixed cellulose esters) and ultrafiltration in a Millipore® Prep/Scale system (manufacturer specified cut-off of 1 kDa and a filter membrane area of 0.54 m2).

Sediment and porewater samples from bays in Luleå, receiving stormwater discharge, showed enrichment of Cd, Cu, Pb and Zn. Also the PAH content was enriched, in particular for phenantrene, anthracene, fluoranthene and pyrene. Water volume and turnover rate in the water bodies with low or no surface runoff during wintertime, and ice covering, contribute to anoxic conditions in the water column and sediments. The enclosure of the bay Skutviken in 1962 illustrates for how reduced water circulation promotes the occurrence of anoxic conditions with sulphate reduction. As a consequence of these conditions, metals are trapped in the sediments as sulphides. The use of trace metal ratios could not indicate road runoff as main source for sediment pollution. The degree of pollution was higher in the sediments of the bays in Luleå than in a 1998 implemented, stormwater pond in Sollentuna, which receives highway runoff.

Water concentrations of Cd, Co, Cr, Cu, Fe, Mn, Na, Ni, Pb, S, and Zn showed seasonal variations in Sollentuna. In winter de-icing agents and use of studded tires cause higher metal concentrations of Co, Cr, Cu, Ni, Mn, Na, and Zn dominated by the truly dissolved phase. In Luleå depletion of oxygen under the thick ice cover can change the redox border from below sediment surface to above.

The sediment in stormwater draining ditches in Luleå showed seasonal variations in grain size, LOI and metal concentrations. Low runoff intensity in winter enables fine grain sediments to settle already in the ditches. A group of variables that had significant positive correlation between each other were Fe2O3 and LOI, Cd, Co, Ni and Zn.

Water and sediment quality of the investigated water bodies depends on catchment area characteristics and emission impact, from point sources in particular. At all sites, including the stormwater pond, retention of metals seems to be favoured by stagnant water and occurrence of organic material. Pollutants can be trapped due to sorption to organic material, and early diagenetic processes with formation of Mn- and Fe-hydroxides and sulphide reduction. In the stormwater pond this affects only a fraction of the metals in truly dissolved phase in the water column, while most of the dissolved concentrations will be released to the recipient

In Luleå postglacial land uplift implies continuous changes in the environment, which can lead to changing redox conditions which will necessitate new risk assessments. Future drainage of the buried sediments can result in oxidation and release of trapped pollutants.

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1

Preface

This thesis consists of the following five papers:

Rentz R., Widerlund A., Viklander M. and B. Öhlander (2011): Impact of urban stormwater

on sediment quality in an enclosed bay of the Lule River, northern Sweden. Water, Air and

Soil Pollution, vol. 218 (1), p. 651.

Rentz R. and B. Öhlander (2011): Urban impact on water bodies in the Luleå area, northern

Sweden, and the role of redox processes. Hydrology Research. In press.

Ralf Rentz; Fredrik Nordblad; Björn Öhlander (2011): Impact of urban stormwater on water

quality in an enclosed bay of the Lule river, northern Sweden. Manuscript.

Ralf Rentz; Godecke-Tobias Blecken; Charlotte Malmgren; Björn Öhlander; Maria Viklander

(2011): Stormwater impact on urban waterways: seasonal variations in sediment

concentrations in a cold climate. Submitted to Journal of Soils and Sediments.

Ralf Rentz; Magnus Westerstrand; Björn Öhlander (2011): Seasonal water and sediment

quality change in an artificial stormwater basin in cold climate receiving motorway runoff.

Manuscript.

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Contents

Introduction 1

The urban environment 1

Water bodies 3

Water cycle and water bodies in urban environments 4

Metals and polycyclic aromatic hydrocarbons (PAH) in the urban environment

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Protection of water bodies in urban areas 7

Landscape history and studied water bodies in the Luleå area 8

Objectives - Project and studies on water bodies in Luleå and in Sollentuna

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Sampling sites 13

Sampling and analysis 18

Main results and conclusions 22

Resume 36

Appendix - Abbreviations 44

Paper I

Paper II

Paper III

Paper IV

Paper V

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1

IntroductionThe title of this thesis “Water and sediment quality of urban water bodies in cold climates”

connects a number of fields in earth sciences. Water, that gave birth to all life on earth, and

sediments, which stand for the solid ground we walk on and which we use for cultivation.

Urban areas represent the impact we humans have on our environment in densely populated

areas, conscious or even unconsciously. In the use and design of our environment we still

have to adapt to natural premises, if it is climate or access of water or fertile soils. All of us

have an idea about how living in urban areas is like. In contrast, the countryside and

wilderness is often romanticised and awakes the wish to protect this natural environment.

Thereby it is easily forgotten that also the urban environment is worth efforts to be preserved

in a sustainable way, for their dwellers´ best and not at least, protecting the adjacent

countryside. To enable a successful, sustainable interacting with the environment we are

living in, we need to understand which components constitute the environment and how the

processes work, through which the elements in the environment are interconnected.

The urban environment Today urban areas are considered as environments in which natural processes take place.

Understanding the cycles of energy and matter, which includes transport and form of

chemical components on Earth, is one important task of geochemistry. We have methods and

tools to measure and quantify these effects, which enables us to estimate conditions of the

environment and potential risks. However, in urban areas there are greater risks for various

types of anthropogenic pollution. According to Endlicher & Simon (2005), the purpose of

urban ecology is the research on urban nature systems and their interaction with the urban

socioeconomic system. To enable sustainable development and better living conditions in

urban areas, interdisciplinary work by natural and social scientists and planners is necessary.

Endlicher & Simon (2005) point out that the ’urban natural system’ is becoming more and

more an interest for dwellers, and not at least for natural scientists. Furthermore, Endlicher

(2004) divided the urban natural system into the following most important spheres:

� urban atmosphere

� urban pedosphere

� urban hydrosphere

� urban biosphere (containing its flora and fauna)

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In these spheres geochemical processes take place (Figure 1). The urban natural system is

intensely affected by human activities. Supply of e.g. heat, water, particles and pollutants to

the spheres of the urban environment by human activities is evident (Arnfield, 2003). The

socioeconomic system provides opportunities for a feedstock of different pollutants. These

pollutants become part of the urban natural system and are exposed to geochemical processes

in the system.

Bolund and Hunhammar (1999) identify seven different urban ecosystems in Stockholm

City, namely street trees, lawns/parks, urban forests, cultivated land, wetlands, lakes/sea and

streams. These ecosystems are considered to be natural even though almost all areas in cities

are manipulated and managed by man. Bolund and Hunhammar (1999) point out the

contribution of urban ecosystems to public health and that they can increase the quality of life

for dwellers. For Stockholm they observe following ecosystem services: air filtration, micro

climate regulation, noise reduction, rainwater drainage, sewage treatment, and recreational

and cultural values.

Figure 1. The urban natural system and its subsystem, after Endlicher and Simon (2005).

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Since 2008 more than half the human population lives in urban areas. The urban population

may increase to 80% by 2030 (UNFPA, 2007). Realizing that in Sweden, 84% of the

population live in urban areas (Table 1) (Statistics Sweden, 2011), the importance of this

environment and its own “natural driving forces and patchwork patterns” (Endlicher and

Simon, 2005) for people becomes obvious. On the global level, all future population growth

will be in towns and cities (UNFPA, 2007), which requires reflection about consequences for

different resources in this environment. The future demand, use and preservation of water in

urban areas will become crucial for millions of people. Access to a supply of freshwater is

already an urgent problem in many regions of the world.

Population growth and the limited space in urban areas makes them focal points for

controversies in water use and water pollution control (Schirmer et al., 2007). If we can

understand which, how and when geochemical processes take place, we improve the

possibilities for efficient management of water and other natural resources.

Sweden Population Percentage of population

(%) In localities* 7 631 952 84.4 Outside of localities 1 415 800 15.6 Total population 9 047 752 100 * A locality (in Swedish “tätort”) consists of a group of buildings normally not more than 200 metres apart from each other, and must fulfil a minimum criterion of having at least 200 inhabitants (Statistics Sweden 2006).

Water bodies If we look at photographs that show Earth from space it is obvious why our planet is called

the blue planet. About 71% of its surface is covered by water. Water occurs in the

hydrosphere, atmosphere and lithosphere in gaseous, fluid or solid conditions of aggregation.

Water bodies are any significant accumulation of water occurring on Earth’s surface. ‘Body

of water’ refers to oceans, seas and lakes, but also smaller pools of water, like ponds, puddles

or wetlands are included. Geographical features where water moves from one place to

another, like rivers, streams and channels are not always considered bodies of water, but they

can be included as geographical formations featuring water.

Table 1. Number of people living in urban areas (localities) in Sweden in the year 2005. Source: Statistics Sweden (2011).

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The oceans comprise ~97% of the free water on Earth. About ~2% is found in glacier ice,

mostly on Greenland and Antarctica (Berner and Berner, 1987). Only 0.001% of world water

is found in the atmosphere and ~1% on continents. The small percentage of 0.01% free water

on Earth in lakes and 0.0001% in rivers plays a disproportionately large role in the natural

water cycle. River networks return the majority of surface and even subsurface runoff to the

oceans. At the same time, rivers transport eroded sediments, dissolved ions, nutrients and

organic matter. This also makes them an important supplier in the biogeochemical cycles. For

human societies, access to freshwater from groundwater, lakes and rivers is crucial for health,

agriculture and basic industries.

Of these types of water resources, impacts on rivers and lakes are targets in the present

research. Rivers are highly important for both natural systems and human societies (Simmons

1991, Hauer & Lamberti 2006). They form the physical environment, and are permanently

changing it. Rivers and lakes provide habitats for animal and plant species, and are suppliers

of water and food. They fill transport functions and they offer conditions for economical

business development. Rivers also are energy sources. The shores of rivers and lakes give

quality of life (waterfront development) and space for recreation in urban areas.

The functions of rivers and lakes, which intensely influence their environment, depend on

natural factors like geology, relief, climate and vegetation as well as on human activities.

Water cycle and water bodies in urban environments Water bodies in urban areas fulfil diverse functions. They are natural resources that provide

food, drinking water and process water for industry (Simmons, 1991, Hauer and Lamberti,

2006). Water surfaces improve the quality of life for dwellers, offering them space for

recreation and means of transportation, but they also offer space for habitats to plants and

animals. Water bodies in urban environments are exposed to emissions from manifold

sources. These emissions are integrated in a chain of natural processes affected by human

activities. The use of urban waters as sewers compromises their other functions (Walsh,

2000). Pollutants can reach water bodies in urban areas by airborne transport, infiltration, and

particularly by surface runoff (Figure 2).

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The urban water cycle describes the route of water from when it is collected for use in an

urban community or enters urban space, to when it is returned to the natural water cycle.

Water bodies and groundwater resources as well as sediments and soils will be affected by

discharged stormwater in the urban water cycle. Stormwater runoff represents a

contamination source with heavy metals, polycyclic aromatic hydrocarbons (PAH), mineral

Figure 2. Particle movement in the urban environment and overview of transport pathways, after Charlesworth and Lees (1999).

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oil hydrocarbons (MOH) and soluble salts for recipients (Karlsson and Viklander, 2008a,

Charlesworth and Lees, 1999, Westerlund, 2007, Brown and Peake, 2006, Schiff and Bay,

2003, Göbel et al., 2007). Increased supply of metals and organic pollutants to recipients can

pose risk for living organisms (Wildi et al., 2004, Munch Christensen et al., 2006).

The geomorphology and geochemistry of the water bodies and their catchment areas

determine which processes are important. The catchment characteristics of urban areas

contain a diversity of geochemical attributes which may have great impact on adjacent water

bodies and their sediments (Lindström, 2001).

Water transports suspended particles and dissolved compounds, and reacts with rocks, soils,

sediments and organisms, which makes it an important and powerful agent in the urban

natural system.

Figure 3. Reactions and processes of importance in the biogeochemical cycle of metals in the water-sediment recipient environment, after Benjamin and Honeyman (2000).

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When suspended particles and dissolved compounds transported by air and water finally

reach a recipient, reactions and processes in the biogeochemical cycle determine the

disposition of these compounds (Figure 3).

Metals and polycyclic aromatic hydrocarbons (PAH) in the urban environment

In urban environments metals are omnipresent. They occur in roofs, cars, street lamps, crash

barriers, gully covers, pipelines, cables, paints, computers, etc. Beyond the classic metal

working industries, many other industries are heavily reliant on metals. These industries and

their products constitute a large artificial source of metals. In the end, metals become part of

processes within the urban natural system. In the four spheres of this system (urban

atmosphere, urban pedosphere, urban hydrosphere, urban biosphere (Figure 1)), metals are

found in different compositions and species, which can show variable mobility. Based on

these premises, metals can have different effects on their environment. Before effects on

living organisms become noticeable, metals are transported and stored in some way in the

spheres and their components. Even if organisms need a certain amount of essential metals, an

excess of metals may be toxic for organisms.

PAHs originate in most cases from a number of different diffuse sources. Commonly,

pyrogenic sources are distinguished from petrogenic sources. The pyrogenic pollution comes

from combustion of fossil fuel or wood, and petrogenic pollution arises from petroleum

products in fluid or vapor form. Also, wear and leaching of asphalt and tire wear contribute to

the PAH content in stormwater. The most abundant PAHs in stormwater are phenantrene,

anthracene, fluoranthene and pyrene (Lau and Stenstrom, 2005, Viklander, 1998), and they

are often associated with particulate transport. The transport capability of stormwater for these

pollutants to receiving waters is affected by the particle size of the sediment load. Fine sand

fractions, and especially silt and clay fractions, were found to have the highest mass of metals

and PAHs (Menzie et al., 2002).

Protection of water bodies in urban areas The hazard of water and sediment pollution from nonpoint sources in urban areas demands

“Best Management Practices” (BMP) to prevent water bodies from quality degradation. BMP

describes any technique, measure or structure which controls stormwater quantity and/or

quality, as cost efficient as possible. Water bodies are most vulnerable for impact from

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nonpoint sources during runoff events which are storm- or snowmelt generated. Especially

runoff from sealed urban areas and roads can cause problems due to increased runoff

volumes, increased flashiness of runoff hydrographs and chemical contamination. Urban

stormwater can have an adverse impact on the ecology of the receiving water bodies, often

summarized under the term “urban stream syndrome” (Walsh et al., 2005). The intention of

BMP design is to function during and after runoff-generating events and to reduce the

generated runoff load or the delivery of material to a receiving water body (Ice, 2004). The

construction of stormwater basins (detention and retention ponds) as BMP is more and more

common to reduce negative effects on recipients. They are used to delay runoff, to reduce

peak discharges and to allow pollutants to settle out. In Sweden the function of stormwater

basins have received increased attention in a number of studies during the last years (Alm et

al., 2010, Falk, 2007; Färm, 2003). The role of pollution from nonpoint sources has been

recognized in the European Water Framework Directive (WFD) (European Parliament and

Council, 2000). The goal of the European WFD is to achieve “good surface water status” by

2015. This requires observation and measurement to control and prevent the discharges of

pollutants originating from both point and nonpoint sources. Urban runoff can be considered

to be an important component to deal with for reaching the designated “good surface water

status”.

Landscape history and studied water bodies in the Luleå area During the early Holocene deglaciation, the eastern parts of the county of Norrbotten in

northern Sweden were submerged up to 200 m by the Ancylus Lake, while the ice sheet

margin retreated towards the northwest (Björck, 1995). This flooding affected the present-day

30 - 40 km wide coastal plain. The hinterland plain, which now has up to 200 - 300 m high

hills, formed a deglacial archipelago, and the present river valleys of the Pite, Lule and Råne

Rivers were deeply incising bays (Hoppe, 1959, Björck, 1995). This region is near to the

centre of maximum isostatic recovery for the Scandinavian Ice Sheet. Therefore, land uplift

(isostatic rebound) was faster than the early water level rise of the expanding Ancylus Lake

(Lindén et al., 2006). The highest shoreline (HS) in the Luleå area is 230 m a.s.l. (metres

above sea level). Clear traces of wave erosion on till-covered slopes are found on e.g.,

Bälingeberget (Figure 4) with its cobble terraces (in Swedish: klapperstensfält). Snöberget

(Figure 4) is an example of a till-capped hill (in Swedish: kalottberg) that testifies to the

former HS (HS ca. 220 m). The shoreline impact on the hills is obvious with often wave-

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washed bare bedrock on the south and south-east weather sides, and beach sediment deposits

at falling altitudes on the leeward sides. In the valleys the soil-substrate consists of till,

glaciolacustrine/lacustrine and glaciofluvial/fluvial sediments.

Human settlement and cultivation of the landscape had to be adapted to this environment.

The valleys with the more productive soils became farmland, and the flood-endangered banks

were used as pastures. At the coast the navigable harbour was important for Luleå and its

hinterland. As a consequences of the land uplift, the old Luleå harbour became unnavigable.

For this reason, the whole town was moved in 1649 from its old location in Gammelstad to

the present-day location of Luleå.

Today, Luleå with its ~74,000 inhabitants, is situated at the mouth of the Lule River. The

river and former shallow bays of the brackish Bothnian Bay are the most characteristic

hydrodynamic patterns of Luleå (Figure 5).

Figure 4. View at a) cobble terraces at Bälingeberget, b) Snöberget with dense vegetation on the moraine-covered top and sparsely vegetated wave-washed slopes and c) the Råne River valley with productive soils.

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The 460 km long Lule River has a 25,240 km2 large catchment with an annual average

discharge of 498 m3/s (Raab & Vedin 1995). It rises in the mountain area close to the

Norwegian border, where vegetation of tundra type occurs. Downstream, coniferous and birch

forests dominate, covering 58% of the total catchment area. Also, lakes and mires are

common, covering 11% of the total catchment area. Since the beginning of the 20th century,

the river has been regulated and today there are 15 power stations along the river (Drugge,

2003).

The system of former shallow bays of the brackish Bothnian Bay (innerfjärdar) are partially

enclosed (Figure 5) due to the post-glacial rebound (8-9 mm/a (Lindén et al., 2006)) or

Figure 5. Water bodies in the Luleå area. Gammelstadsviken (GV), Notviken (NV), Ytterviken (YV), Skutviken (S), the spit Gültzauudden (G), Björsbyfjärden (BF), Sinksundet (SS), Sörfjärden (SF), Mulöfjärden (M), Inre Skurholmsfjärden (IS), Lövskataviken (L), Bredviken (B), Inre Hertsöfjärden (IH); D1-3: Watergates

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artificial banks. However, these water bodies, situated in and around the town of Luleå, are

affected by local catchments, which contain urbanized and industrial areas as well as rural and

forested areas. Consequences of the ongoing land uplift are decreasing water surfaces (and

volumes) in the shallow bays. Silting-up processes are accompanied by increasing vegetation

in the former shallow bays (Erixon, 1996). To preserve the shallow bays for recreation, they

were dammed up at their two connections with the Bothnian Bay. Also, the water level in the

Lule River, and especially in the Bothnian Bay, affects the water level and water quality in the

shallow bays (Erixon, 1996).

The bay Skutviken is located close to the centre of Luleå, and is enclosed by a road bank

constructed in 1962. It is still connected to the Luleå River via a channel that is approximately

8 m broad, 3 to 4 m deep and 35 m long. Several stormwater pipes discharge into the bay

from a sewer drainage area comprising 0.53 km2 industrial area and 0.73 km2 housing area.

Hertsöfjärden is a bay especially affected by the outlets of the steel plant SSAB Tunnplåt

AB (formerly Norrbottens Järnverk and SSAB) since the 1940s. Due to plans to build a new

steel plant, Stålverk 80, the outer part of the bay was infilled in 1975-76 and an artificial bank

divided the bay in two parts. The water in the inner part is dammed up (Timner, 1994).

Lövskataviken and Inre Skurholmsfjärden, in central Luleå, are water bodies in the

innerfjärdar system with more than 100 years of industrial history on their banks. The urban

catchment area contains industrial and housing areas with parks. A road bank, built in the

1960s, separates the two water bodies, but they are still connected via road culverts (Olofsson,

2002).

Gammelstadsviken is an enclosed bay which until 1649 sheltered the harbour of the old

town Luleå. Today the bay is part of a nature reserve and is on the UN list of wetlands worthy

of protection. For being so far north the area has become an outpost habitat for a number of

southern plant and animal species. It is biologically similar to a flatlands lake in southern

Sweden. Buckbean, water plantain and arrowhead cover primarily the lake, but the total

number of flora species in the area probably exceeds 30. The bird life has over 200 species

and the lake is a valuable breeding ground (Öberg, 2006).

Objectives - Studies on stormwater receiving water bodies in Luleå and in Sollentuna

In Luleå, research on river geochemistry and heavy-metal contamination in different sites is

well established and numerous investigations in natural waters and on stormwater processes

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have been conducted (Widerlund and Ingri, 1996, Öhlander et al., 1991, Westerlund and

Viklander, 2006, Viklander, 1994, Drugge, 2003). With the post-glacial geomorphologic

history and persisting processes, the water bodies in the Luleå area represent a unique,

naturally changing environment affected by human settlement. In Luleå are no stormwater

basins implemented yet, so that the main objectives of this study in Luleå were to give a

description of the current water and sediment status of certain sites, and to identify important

geochemical and geomorphological processes and possible pollution sources for the water

bodies in that area (paper 1, 2, 3 & 4). A review of previous works on sediment and water

quality in the Luleå area, and comparison with newly collected data, helps to determine the

geochemical conditions in water bodies in the Luleå area (paper 2). From this, information

about dominating processes in these water bodies can be obtained, and different impact

factors for sediment and water quality can be identified. Previous studies of stormwater and

gully pot sediments in the Luleå area (Westerlund, 2007, Karlsson and Viklander, 2008a)

indicated particle-related transport of metal and organic pollutants with seasonal variations. A

further aim of this study was to look into stormwater impact from the surrounding urban area

on an enclosed bay of the Lule River, Skutviken, near the centre of Luleå by investigating

heavy metal and PAH concentrations in bottom sediments (paper 1), and the speciation of

elements in the water column in summer and wintertime (paper 3). The current water,

sediment and porewater geochemistry was described and possible pollution sources tracked,

with the aim of quantifying environmental effects of urban stormwater. The geochemistry of

the water, sediment and porewater in Skutviken was investigated and compared with a

reference site unaffected by stormwater discharge. Ways of transport, the transported particles

form (species) and amount of transported substances were analyzed (paper 3). The results

may aid our understanding of the consequences enclosures (natural or man-made) may have

for the geochemical processes taking place in bays, sediments and water. Determining the

species in which heavy metals occur in the stormwater transport system, with focus on a catch

basin like bay of the Lule River, can contribute to a better understanding and estimation of the

effects of, and dangers posed by, these pollutants. Questions about the benefits of the water

management and the costs of curtailing impact on the natural environment can be discussed

on the basis of the results. In addition heavy metal concentrations in bottom sediments of

three different recipients in front of storm sewer outlets in Luleå were investigated in autumn

(before the snow season) and in spring (after snowmelt) (paper 4). The aims were to evaluate

if there is an impact of stormwater discharges on sediment metal concentrations, if there are

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seasonal metal variations and how the geomorphology and vegetation influences the

distribution of discharged stormwater sediments and associated metals.

To get a clear view of the effects of runoff from highways with heavy traffic, geochemical

processes in an artificial stormwater basin were studied and sediment and water samples were

taken in a stormwater basin in Sollentuna, close to Stockholm (paper 5). At this site heavy

metal and PAH concentrations in bottom sediments were investigated, and the speciation of

elements in the water column in summer and wintertime. The aims were to evaluate if impact

of stormwater discharges on sediment is detectable, if there are seasonal elemental variations

in water and sediments and if the stormwater basin does function as a trap for pollutants.

Clear effects of road traffic can then be compared with the total pollution in the Luleå area.

Sampling sites Sampling was conducted in several water bodies in Luleå in northern Sweden and at one

site in Sollentuna, central east Sweden. The annual precipitation in the Luleå area is about 500

mm of which 40 to 50 % falls as snow between November and April/May (Hernebring,

1996), and is thus discharged during snowmelt. From November until May the Lule River and

the bays close to the city centre are ice covered. In Luleå a site in the Lule River ahead the

spit Gültzauudden served as reference site. The site Gültzauudden is located beside the main

stream bed of the Lule River, and the water depth is ca 6 m. The bay Skutviken is a sampling

site close to the city centre of Luleå (74,000 inhabitants) (Figure 5, 6). The surface area of the

bay Skutviken is ~12 ha, and the mean and maximum depths of the bay are 1.6 m and 3.4 m,

respectively. It is mainly separated from the Lule River by a road bank constructed in 1962,

and only connected trough a channel (8 m in width, 3 to 4 m in depth, 35 m in length). These

physical conditions give the bay similarities with the shallow naturally enclosed bays. Besides

the road bank, the bay is mostly surrounded by two highly frequented roads with traffic

intensities of 22900 and 13600 vehicles per day, respectively (Luleå Kommun, 2010). The

sewer drainage area contains 0.53 km2 industrial area and 0.73 km2 housing area. Since the

road bank runoff and six stormwater channels enter the bay, it almost functions as a large

stormwater pond where a high amount of stormwater sediment is trapped, resulting in a

reduced sediment supply to the Lule River. All outlets are located below the water surface,

except during periods of very low water level.

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Sediment samples in front of storm sewer outlets were taken at three sites in Luleå (Figure

7). At Notviken, stormwater from a 67 ha large catchment area including an industrial area

with 5 ha roads and 18 ha parking lots is discharged through a 600 mm pipe into a ditch

having a length of ca 250 m before opening into the bay Notviken The bay has an area of ca

256 ha and is connected to the delta of the Lule River. The southward open water surface

Figure 6. Location of the study area Skutviken (A) with its stormwater sewer catchment area and the reference sample site at Gültzauudden (B) in Luleå, Northern Sweden. (Terrängkarta: Lantmäteriet, 2011)

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allows that waves affect the mouth of the ditch and cause redeposition of sediment along the

local banks. The ground in front of the ditches mouth shows ripple marks. Also ground

freezing and ice floes affect the deposited sediments along the shallow banks. The banks of

the ditch are partly fixed with stones. At Gammelstadsviken stormwater from a 67 ha

catchment area is discharged. Of the whole catchment, 29 ha are an industrial area, 8 ha

residential area, 23 ha roads, and 13 ha parking lots. The sewer (800 mm diameter) opens into

a 30 m long ditch ending in Gammelstadsviken. This recipient is densely overgrown by

mainly Typha spec. and Carex spec. communities. At Ytterviken four sewers (680 mm, 1150

mm, 1350 mm and 210 mm in diameter) lead into a ditch with a length of 230 m. At this site,

water from a 70 ha large catchment is discharged (thereof 20 ha roads and 4 ha car parks;

remaining: industrial area and university campus).

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Figure 7. Sampling sites Notviken, Gammelstadsviken, and Ytterviken. Sampling stations A, B, C; outlet pipe OP; ditch: gray line.

(aerial photographs: Lantmäteriet, 2011)

Gammelstadsviken

Ytterviken

Notviken

17

The studied stormwater basin in Sollentuna (Figure 8) is situated west of the highway E4 at

the highway intersection Häggvik 15 km north of central Stockholm, and has been in full

operation since 1998. The facility consists of a “3-step system” with a pump station and two

sedimentation basins followed by an overflow surface. The system receives highway runoff

from Häggviksleden (6.8 ha) and the E4 highway (1.9 ha), totally a sealed road area of 8.7 ha

(ALcontrol Laboratories 2005). Häggviksleden connects the E4 with the main road

Danderydsvägen in Edsberg. The runoff from Häggviksleden and parts of the E4 is led via a

pump into the first basin. At the pumping station separation of oil is conducted. A second

inflow adds only water from the E4. The first basin is elongated with a maximum size of 100

x 50 m and its depth varies between 2 - 2.5 m with a capacity between 4,500 m3 and nearly

6,000 m3 depending on the lowest or highest water level. On the opposite side of the pump

station inlet at the basin ground, an outlet tube with a diameter of 800 mm (D 800) leads the

water over a distance of 55 m to the second basin. The second basin is ca 70 x 60 m in size

and its depth varies between 2 – 2.5 m. The volume at highest water level is about 8000 m3

and at lowest water level 6000 m3. At both basins the banks are stone covered between

highest and lowest water level to prevent erosion. Groundwater infiltration is averted by the

use of a bentonite carpet covered with macadam. The water leaves the second basin trough a 2

chamber gully which function is to extend the water retention time in the basin. At a 2-years

rain the retention time is calculated to 36 hours in both basins. From the gully the water leads

over a 35 m long and 120 m wide grass covered overflow-area slope before it reaches a ditch

which ends in Lake Ravalen after ca 1000 m.

18

Sampling and analysis

Sediment and porewater samplingSediment sampling was conducted with a Kajak gravity corer (Blomqvist and

Abrahamsson, 1985) with a core tube diameter of 64 mm (Figure 9). In winter sampling was

done from ice and in summer from a boat or wading into the water. The sediment core

surfaces were judged to be undisturbed. Cores were sectioned in subsamples (0.5 cm thick for

the uppermost 3 centimetres and 1 cm thick until the core ends). The subsamples were stored

in plastic containers or bags. For PAH analyses sediment samples were placed in glass

containers with Teflon lined caps. For porewater analyses the sediment samples were put into

plastic bags directly after sectioning. All air was pressed out of the bag before it was placed in

an Ar-filled container to keep the sediments in an oxygen free environment until the

porewater was extracted within the following six hours. The porewater was separated by

0 250 km

Helsinki

DENMARK

NORWAY

Oslo

Stockholm

North Atlantic

Baltic Sea

SWEDENFINLAND

Gulf of Bothnia

Figure 8. Stormwater basins for Häggviksleden, 15 km north of central Stockholm, with pumping station (P), stormwater basin 1 and 2 (B1, B2), grass overflow area (O-A) and sediment and water sampling station (X) in basin 2. (arial photograph: Lantmäteriet, 2011)

19

vacuum filtration (0.22 µm Millipore® membrane filters) arranged in an Ar-flushed glove

box. The porewater samples were collected in 60 ml acid washed polyethylene bottles and

refrigerated until further analysis. Bottom water was sampled from the core tube immediately

after retrieval, 3 cm above the sediment surface. The water was drawn with a small plastic

tube fixed on a syringe and filtered through a 0.22 µm Millipore® membrane filter.

Water sampling and membrane filtration/ultrafiltration Surface water was sampled with a tube 50 cm below the surface or 50 cm below the

underside of the ice, respectively. Water was pumped by a peristaltic pump (Masterflex®

L/S®) trough the tube into 25L poly-ethylene (PE) containers.

Membrane filtration (0.22 µm pore size, 142 mm diameter, Millipore® mixed cellulose

esters) was carried out inside a laboratory within the next 6 hours. The principle of membrane

filtration is similar to sieving, just that the pore sizes of membrane filters are several orders of

magnitude smaller than sieves. In this study membrane filtration was used to separate

particulate and soluble fractions. The first membrane filter was used until it was clogged

Figure 9. Kajak sediment sampler (“HTH Sediment Corer” to the right) and extruding device for sub-sampling of the sediment core.

20

completely; the filtered volume was measured and then discarded. For the actual sample new

filters were used, trough which half the clogging volume was allowed to pass the filter. This

was performed to decimate discrimination of colloids that is caused by clogging of filters

(Morrison and Benoit, 2001). The filtrate was collected in a 25L PE container from which

subsamples were taken for analyses. The membrane filtered water was then ultrafiltrated in a

Millipore® Prep/Scale system. Ultrafilters separate solids based on their molecular weight

rather than physical size. The filter had a manufacturer specified cutoff of 1 kDa and a filter

membrane area of 0.54 m2. The filter material was regenerated cellulose. The cross-flow

filtration (CFF) system was connected with a Watson Marlow peristaltic pump. In that system

water is recirculated parallel (tangential) to the filter membrane at a high flow rate (Figure

10). After the ultafiltration (Cheryan 1998), subsamples were taken from the retentate and the

filtrate. Subsamples were collected in 60 ml acid-washed polyethylene bottles and refrigerated

until further analysis. All used tubing, bottles and containers were acid-cleaned in 5% HCl

with subsequent wash in MilliQ water (Millipore, 18.2 M�) before sampling.

Ultrafiltration is an applicable technique for determination of the size distribution of

components in natural water samples. The method is often applied for studies of the colloidal

and truly dissolved species of metals and organic matter in natural waters (Guéguen and

Dominik 2003; Ingri et al. 2004). The enrichment of species concentrations in the retentate

Figure 10. The principle of cross flow filtration (CFF).

21

facilitates the determination of low-abundance species (e.g. colloidal concentrations).

Ultrafiltration techniques have previously been described and evaluated by several workers

(Guéguen et al. 2002; Wilding et al. 2004). Two critical aspects when applying the method for

natural water samples are, the mass balance recovery and the accuracy of determination of the

species concentrations in the retentate. Larsson et al. (2002) found that a cross-flow ratio

above 15 was necessary to achieve mass balance recoveries close to 100 %. The cross-flow

ratio CFR is defined as:

perm

ret

QQCFR �

Qret and Qperm denote the retentate- and the permeate flow rate, respectively. It was also

found that an enrichment factor (total feed water volume : final retentate volume) larger than

10 was required for accurate determination of the colloidal species. The enrichment factor EF

and the colloidal concentration Ccoll can be calculated using:

ret

retperm

VVV

EF�

�

EFCC

C permretcoll

��

Where Vperm, Vret denote the volumes of the permeate and the retentate. Cperm, Cret and Cfeed

denote the concentrations of the permeate, the retentate and the feed sample, respectively.

Finally, the mass balance recovery R in percent units may be determined as:

feed

retperm

CCC

R�

�

The truly dissolved phase constitutes the fraction <1kDa and the colloidal fraction contains

particles >1 kDa and <0.22 μm.

Sediment and water analyses Total Carbon (TC) and Total Nitrogen (TN) of the sediment was analysed by Umeå Marine

Sciences Centre (paper 1). Analyses of carbon and nitrogen in sediments were performed with

a Carlo Erba model 1108 high temperature combustion elemental analyzer, using standard

procedures and a combustion temperature of 1030 ºC. For standardization Acetanilide was

utilized.

Cesium-137 of the sediment from Skutviken was analyzed by high resolution gamma

spectrometry at Risø DTU, Radiation Research department (paper 1).

22

Detailed particle size analyses was performed with a CILAS 1064 laser diffraction particle

size analyser in wet mode for 4 samples from a profile at Skutviken and a profile at

Gültzauudden (paper 1). The grain size fractions of clay/silt (<63 μm) sand (63 μm - 2 mm)

and gravel (>2 mm) of sediments from ditches were determined by wet sieving according to

the Swedish standard method SS-EN 933-1 (paper 4).

Metal and PAH analyses were accomplished in the concerned papers by the accredited

laboratory ALS Scandinavia AB in Luleå. The water, sediment and porewater was analysed

for major elements and trace metals. The water samples were analysed by inductively coupled

plasma atomic emission spectrometry (ICP-AES) and inductively coupled plasma sector field

mass spectrometry (ICP-SFMS). For instrument operation details see (Rodushkin and Ruth,

1997). To the porewater samples 1 ml nitric acid (suprapur) was added per 100 ml. Sediment

samples for determination of As, Cd Co, Cu, Hg, Ni, Pb, S and Zn were dried at 50 °C

digested in a microwave oven in closed Teflon bowls with a nitric acid : water ratio of 1:1.

For other elements 0.125 g dried matter (DM) was melted with 0.375 g LiBO2 and dissolved

in HNO3. Metal determinations of the sediments were made by ICP-AES and ICP-SFMS. The

following 16 PAHs were analysed in sediments: Naphthalene (NAP), Acenaphthylene (ACY),

Acenaphthene (ACE), Fluorene (FL), Phenanthrene (PHEN), Anthracene (ANT),

Fluoranthene (FLR), Pyrene (PYR), Benzo(a)anthracene (BaA), Chrysene (CHY),

Benzo(b)fluoranthene (BbF), Benzo(k)fluoranthene (BkF), Benzo(a)pyrene (BaP),

Dibenz(a,h)anthracene (DBA), Benzo(ghi)perylene (BPY) and Indeno(1,2,3-cd)pyrene (INP).

The PAH sediment samples were leached with acetone : hexan : cyclohexan (1:2:2) and

measurements were done with gas chromatography mass spectrometry (GC-MS).

The dissolved oxygen in the water column was determined with a Hach LDO™ sensor

mounted on a Hydrolab® MS5 sonde. Also pH was measured with this sonde.

Main results and conclusions

Paper I Sediment and porewater samples from an enclosed bay receiving stormwater discharge

(Skutviken) near the centre of Luleå, northern Sweden and a local reference site

(Gültzauudden) were taken. In the surface water at Skutviken the oxygen saturation 10 cm

above the sediment surface was close to 0% in wintertime, when the bay was ice-covered. In

contrast, the water column was well oxygenated (saturation 85-90%) during the ice free

23

season. This variation can cause changing redox conditions in the surface sediment

determining release or accumulation of pollutants through formation/dissolution of Fe-Mn

oxyhydroxides.

The particle size analyses of the sediment showed that the 2-3 cm and 5-6 cm layers at

Skutviken and Gültzauudden had very similar particle size distribution, where the main

components (60% cumulative volume) in these layers had a grain size from 10-30 μm. At

Skutviken the particle size from 10-11 cm depth contained only 15% >10 μm.

At Gültzauudden the high Mn content in the sediment top layers can be related to the oxic

environment at this site where Mn occurs mostly as Mn oxyhydroxides (Davison, 1993)

(Figure 11). The decomposition of organic material and increasingly anoxic environment with

sediment depth results in reduction of Mn oxyhydroxides and increased porewater

concentration of Mn(II). At 4 cm depth the MnO content stabilises (0.2 wt%) together with

the increasing porewater Mn concentration. This suggests that anoxic conditions predominate

below 4 cm. The porewater profile indicates Mn(II) flux upward, resulting in the oxidation of

Mn(II) to Mn(IV) in the oxic parts of the sediment (Davison 1993; Wehrli 1991). The Fe2O3

peak in the Gültzauudden sediment profile is situated below the MnO peak (Figure 11). In

oxic sediment Fe occurs as Fe(III) in iron oxyhydroxides, explaining the solid Fe peak at 3 cm

depth. Below 5 cm the solid Fe content declines continuously. The porewater Fe

concentration indicates that reduction of solid Fe(III) to the soluble Fe(II) occurs when

porewater becomes more anoxic (Davison, 1993, Wehrli, 1991).

At Skutviken the MnO content in the sediment is much lower than at Gültzauudden in the

upper part of the sediment. The geochemical conditions where Mn(IV) is reduced to Mn(II)

appear to be reached already in the bottom water above the sediment surface. During winter,

when the bay is ice covered, the oxygen concentration in the bottom water is <0.42 mg l-1.

The presence of a solid Fe2O3 maximum at the sediment surface at Skutviken indicates that the

redox conditions permit precipitation of Fe(III) hydroxides at the sediment-water interface,

corresponding to the processes at the depth of 3 cm in the sediment at Gültzauudden. At

Skutviken reductive dissolution of Fe hydroxides take place already at the sediment surface.

The decrease of total S in porewater at Skutviken suggests that reduction of SO42- occurs

immediately below the sediment-water interface (0-2 cm). The simultaneous increase of solid

S indicates precipitation of solid sulphides in the sediment. The solid S concentration at 0.5 –

11 cm depth (2500 – 4200 mg kg-1) exceed that at Gültzauudden by a factor of 5-7.

24

The element/Al ratios for the major elements Ti, Ca, Mg, Na, and K are similar at Skutviken

and Gültzauudden. Only small deviations from local till ratios for Ca/Al, Na/Al and K/Al

indicate that both sediments mainly are composed of local minerogenic matter. The higher

Fe/Al and the Mn/Al ratios in the 1-7 cm section at Gültzauudden, suggest precipitation of Fe-

Figure 11. MnO and Fe2O3 in sediment (wt%) and Mn and Fe in porewater (�g l-1) at Skutviken and Gültzauudden. The top value for the “porewater” represents the bottom water (3cm above sediment surface).

25

Mn oxyhydroxides in a more oxic environment (Davison 1993). The Si/Al ratio is at both

sampling sites lower than that of local till (Öhlander et al 1991), suggesting a negligible

content of diatoms in both sampled cores (Stabel 1985).

The TC/TN molar ratio indicates a change in sediment composition at Skutviken from 7 to

11 cm depth, where the TC/TN ratio decreases from 19 to 11 cm. Below 11 cm depth the

TC/TN ratio at both sites are similar. Above 11 cm the concentration of organic material is

enriched at Skutviken, which is consistent with low oxygen saturation above the sediment in

wintertime. The TC/TN molar ratio is thereby higher than the C/N ratio of 6.6 in the Redfield

empirical formula ((CH2O)106(NH3)16(H3PO4)) (Redfield 1958), which indicates an

anthropogenic impact.

For the sediment section 0-2 cm a comparison with deviation values from the Swedish

Environmental Protection Agency (Swedish EPA, 1999) indicate a significant influence of

stormwater sediment for Cd, Cu, Pb and Zn at Skutviken, while at Gültzauudden no effect can

be seen for any of the studied elements. Cadmium, Cu, Pb and Zn concentrations at Skutviken

show almost identical depth profiles with the highest concentrations in the 0.5 to 6 cm

section. The concentrations are for Cd and Pb 3 times higher and Cu and Zn 6 times higher

than at Gültzauudden. Porewater minima for elements from 0.5 to ~5 cm at Skutviken indicate

a sink in the sediment. From 0.5 to ~5 cm depth Cd, Cu, Pb and Zn show maxima in the solid

sediment, coinciding with maxima for solid S, TC and the TC/TN ratio. The change in

concentrations of Cd, Cu, Pb and Zn at Skutviken around 6 cm depth accompanies a change in

the composition of the sediment with coarser particles and higher TC in the upper sediment.

This suggests that higher element contents in the upper sediment column may be more related

to organic components than mainly to clay minerals. The S decline in porewater in the upper

sediment at Skutviken signifies sulphate reduction and coeval sulphide formation in the solid

sediment. The enrichment of Cd, Cu, Pb and Zn in the sediment at 0.5 to ~5 cm depth may

thus be related to sulphide formation in the organic rich 1-7 cm section of the sediment.

Correlation of the trace elements Cd, Cu, Pb and Zn with S at Skutviken shows a uniform

pattern where the trace element content increases with higher S content. The trace elements

Cd, Cu, Pb and Zn are also positively correlated with TC. It is unclear whether organic matter

is a carrier for Cd, Cu, Pb and Zn, or whether this pattern reflects a coupling between organic

matter and sulphide formation in the sediment.

The most abundant PAHs in stormwater, phenantrene, anthracene, fluoranthene and pyrene

(Brown, 2002, Gonzalez et al., 2000), are found in high-very high concentrations in the 0-2

cm sediment layer at Skutviken. At Gültzauudden the PAH contents do not exceed moderately

26

high contents. The particle size analysis at Skutviken for the 2-3 cm and 5-6 cm layers showed

a range from fine to coarse silt, offering conditions for light and heavy PAHs to be associated

with the sediment particles.

Dating with the radionuclide 137Cs was conducted. The activity of the radionuclide 137Cs

showed 2 peaks, whereof the upper peak is interpreted to represent the Chernobyl fallout from

the reactor accident in April 1986 (Ilus & Saxén 2005), while the lower peak is interpreted to

be caused by the fallout from nuclear weapons testing in the early 1960s (Appleby 2002).

However, this peak should be concurrent with the construction of the road bank in 1962, and

may be displaced slightly downward in the sediment due to reworking of sediments during

construction works. Caesium-137 data indicate that changes in sediment characteristics

(particle size, concentrations of TC, TN, metals and PAHs) from 11 cm and upwards became

apparent in the early 1960s.

Characteristic metals in stormwater like Cu, Cd, Pb, and Zn (Hvitved-Jacobsen and Yousef

1991) are significantly enriched at Skutviken compared with the reference sampling site at

Gültzauudden. The concentrations of Cu, Pb, Zn in the sediment at Skutviken are in the range

of the metal concentration reported in street sediment on the road bank that separates

Skutviken from the Lule River (Viklander, 1998), while the metal concentrations reported in

the gully pots are lower than in the Skutviken bay (Karlsson and Viklander, 2008b). A reason

for this might be that most metals, which concentration is higher in the Skutviken sediment

than in the gully pots, are attached to smaller particles

Assuming that the sediment above a depth of 6-7 cm represents the time period after

construction of the road bank, stormwater impact appears to have increased the concentrations

of Cd, Cu, Pb and Zn by a factor of 3-4. However, these metals are probably present as

relatively immobile metal sulphides.

Trace element ratios show that in the upper 5 cm the Pb/Zn ratio follows the ratio for gully

pot sediment from a road in Luleå. The pollutants that are linked to the clay and silt fraction

pass through gully pots and eventually reach the bay. These particle fractions also offer

surfaces for PAHs to bind to (Evans et al. 1990). The PAH profiles at Skutviken resemble

those of Cd, Cu, Pb and Zn, with high concentrations in the upper sediment and lower

beneath. This suggests a common stormwater origin for PAHs and trace metals.

The bay Skutviken has functioned as a large stormwater pond since the road bank was

constructed in 1962, with calm conditions within the bay and a limited water exchange with

the Lule River. This has resulted in a spatial arrangement of the sediment supply, with coarse

sand near the stormwater channels and in particular silt and clay in the deeper central parts of

27

the bay. The stormwater contaminations have resulted in increased concentrations of Cd, Cu,

Pb and Zn in the upper 7 cm of the sediment. Also the PAH concentrations are very high for

Pyrene and high for Phenanthrene, Anthracene, Fluranthene, Benzo(a)anthracene, Chrysene,

Benzo(k)fluoranthene and Benzo(a)pyrene in the surface sediment at Skutviken. An increased

settling of particulate matter and seasonal occurrence of anoxic bottom waters leading to

sulphate reduction appear to be the main effects of the road bank. Sedimentation of pollutant

carriers and the sulphate reduction result in an increased fixation of metals and PAHs in the

sediment. Skutviken appears to be an efficient trap for stormwater contamination, since the

sediment at Gültzauudden is almost unpolluted. The analysis of the trace element and PAH

concentrations in the sediment of a stormwater-receiving bay and a reference sampling site

compared to road run-off sediment enabled to identify the stormwater as an impact factor on

the bay. The sediment shows increased contamination of pollutants which most likely

originate from stormwater. Fixation of pollutants in the sediment occurred for the last ~50

years after the building of a road bank.

This study suggests that enclosed bays with restricted water circulation may be efficient

traps for urban pollutants. As a consequence, the present-day input of pollutants to the sea are

reduced.

Paper II Sediments from urban water bodies in the Luleå area, northern Sweden, were studied to

determine the degree of contamination from metals and PAHs (polycyclic aromatic

hydrocarbons). Beside Skutviken and the reference site Gültzauudden, the partly enclosed

bays Lövskataviken, Skurholmsfjärden (Olofsson, 2002), Bredviken and Inre Hertsöfjärden

(Timner, 1994) were compared.

The sediment profiles for solid Mn at Lövskataviken, Skurholmsfjärden and Bredviken

resemble the characteristics at Skutviken with constant low concentrations of MnO over the

whole depth. Only at Inre Hertsöfjärden, does an increase of MnO in the uppermost 5 cm in

the sediment indicate more oxic conditions in the sediment top. A high concentration of solid

Fe(III) already at the sediment surface at Inre Hertsöfjärden suggests that the oxic conditions

are low compared with Gültzauudden but higher than at the other sites. At Lövskataviken,

Skurholmsfjärden and Inre Hertsöfjärden the change from high concentrations to low

background concentrations is abrupt for Fe2O3. At Lövskataviken the S concentration in the

sediment and porewater indicates similar conditions as at Skutviken. Visible are, in particular

28

at Lövskataviken, Inre Hertsöfjärden and Bredviken, increasing concentrations of solid S at

sediment depths below 15 cm, simultaneously with apparent unchanged low S concentration

in the porewater. The LOI content at all sites is consistently highest in the uppermost section

of the sediment columns.

Gültzauudden has the lowest metal concentrations (Cd, Cr, Cu, Ni, Pb, Zn) in comparison

with the same depth sections of the other sites. Of all sites Inre Hertsöfjärden exhibits the

highest concentrations of all metals except for Ni. The concentrations at Inre

Skurholmsfjärden and Bredviken resemble those at Skutviken. Traffic and urban stormwater

are most possible sources for metal pollution at Skutviken, Inre Skurholmsfjärden and

Lövskataviken, while particularly Inre Hertsöfjärden is exposed to spill water from a steel

plant. To a minor degree Bredviken is exposed to the same spill water besides urban

stormwater.

The high PAH concentrations in the sediment top suggest that the PAH enrichment is

generated from sources in the catchment areas of Skutviken, Inre Skurholmsfjärden and

Lövskataviken. The concentrations at Inre Skurholmsfjärden exceed those of the other sites,

and the sediment at Gültzauudden contains the lowest concentrations for each PAH. The

comparison with the Swedish EPA classification (Swedish EPA, 1999) for organic pollutants

shows clearly increased concentrations at Inre Skurholmsfjärden, Skutviken and

Lövskataviken.

Buried metal pollutants in the sediments at present can become a future risk if they get

mobilized with land uplift (Lindén et al., 2006). Future drainage of the buried sediments can

lead to oxidation and release of trapped pollutants. Metal release from sulphate soils of local

catchments has led to temporally decreasing water quality in the Luleå area before (Erixon,

2009). Human impact on the water levels, such as damming up the partially enclosed bays,

can slow down the long-term processes which result in oxidation of soils and further transport

of pollutants. Erixon (2009) showed that besides urban stormwater, sulphate soils also have to

be considered as an influential factor for disturbance of local water bodies in Luleå.

The investigated water bodies in the Luleå area show clear urban impact on sediment

quality. The metals Cd, Cu, Pb and Zn, which are of main concern in urban stormwater, are

enriched in all investigated bays (Table 2). Metals can bind to surfaces of settling organic and

small inorganic particles. In the sediment they can become part of sulphide formation and are

thus fixed in the sediment.

Water and sediment quality in the Luleå area are dependent on catchment area (size, natural

premises and exploitation) and emission impact, especially from point sources. Important

29

factors are water volume and turnover rate in the water bodies with low water levels and no

surface runoff wintertime, and ice covering during winter, which also contributes to anoxic

conditions in water column and sediment. The redox status in the sediments is crucial for

release or bonding of pollutants in the sediments. The bays do have the capacity to retain

pollutants in their sediment, but there is still a potential risk of release if the redox conditions

change. Postglacial land uplift implies continuous changes in the environment, which can lead

to changing redox conditions. This will necessitate new risk assessments.

Element Depth in cm Skutviken Gültzauudden Lövskataviken

Inre Skurholmsfjärden

Inre Hertsöfjärden Bredviken

0-4 0.7 0.3 0.6 0.8 2 0.7 Cd 4- * 0.4 0.3 0.4 0.3 0.8 0.7 0-4 83 68 80 80 319 122 Cr 4- * 87 67 66 62 98 78 0-4 60 17 56 68 92 37 Cu 4- * 37 24 30 30 41 33 0-4 23 19 47 46 44 34 Ni 4- * 22 19 24 25 24 31 0-4 66 13 39 55 236 69 Pb 4- * 47 26 28 20 101 64 0-4 284 97 302 357 1733 343 Zn 4- * 180 118 166 127 392 283

*core end: Skutviken & Gültzauudden 21 cm, Lövskataviken 30 cm, Inre Skurholmsfjärden 38 cm, Inre Hertsöfjärden 22.5 cm, Bredviken 24.5 cm

Paper III Membrane- and ultafiltration were used to determine different speciation (truly dissolved

phase <1kDa; colloidal fraction >1 kDa and <0.22 μm) of element concentrations in surface

water samples from Skutviken and the reference site Gültzauudden. Sampling was conducted

in winter and summer.

The elemental concentrations of the dissolved phase (<0.22 μm) at Skutviken, Gültzauudden

and the Boden power station show seasonal and spatial variations. The water at Gültzauudden

resembles the Lule River water. In contrast, element concentrations at Skutviken show

stronger seasonal variations. Furthermore, Skutviken is characterized by high concentrations

Table 2. Average element concentration (mg/kg DM) in sediment sections 0-4 cm and 4 cm to core end at Skutviken, Gültzauudden, Lövskataviken, Inre Skurholmsfjärden, Inre Hertsöfjärden and Bredviken.

30

(<0.22 μm) of Ca, Fe, K, Na, Co compared with Gültzauudden, especially in late-winter. Data

from catch basins in Luleå (Karlsson et al., 2009) show clearly highest concentrations for

nearly all elements. Just at Skutviken especially in winter the concentrations of Fe and Mn

can exceed catch basin concentrations.

In Skutviken the high winter concentrations of Fe, S, K, Mg, Mn, Na, and Ca are found in

the truly dissolved fraction, except for Fe. The concentrations of Al, Ba, Co, Cu, are clearly

0

1

2

3

Fe (m

g/l)

unfiltered

<0.22 µm

colloidal

<1kDa

0

100

200

300

400

Mn

(µg/

l)

0

1

2

3

S (m

g/l)

SUMMERWINTERSUMMERWINTER

SUMMERWINTER

Figure 12. Speciation of Fe, Mn and S at Skutviken and Gültzauudden.

31

higher in the catch basin water. In Skutviken, Co and Cu are mainly found in the particulate

phase. Skutviken shows higher unfiltered concentrations for Mn, Fe, S, Co, Cr, K, Ni and Zn

in winter. The seasonal variation of dissolved oxygen in the bay Skutviken can be an

influential factor on concentration of the trace metal species of Mn, Fe, S in the water column

(Figure 12). Skutviken shows higher unfiltered concentrations for Mn, Fe, S, Co, Cr, K, Ni

and Zn in winter, and higher than the reference site but still lower than catch basins. Except

for Fe, these elements were mostly dissolved in winter. The winter conditions at Skutviken can

enhance the fraction of dissolved Mn and other metals in the bay when oxygen in the water

column is depleted under an ice cover. However, the amount of release from the sediment is

not determined. The stormwater is a source for elevated metal concentrations, even though the

dissolved, concentrations in Skutviken are still distinct lower than concentrations in catch

basins.

Paper IV Sediment samples at eight sampling points from three stormwater draining ditches at

Notviken (N), Gammelstadsviken (G) and Ytterviken (Y) in Luleå and their downstream

recipient (Figure 7) were taken in autumn and spring after snowmelt. Comparing the metal

concentrations from all eight sampling points with northern Sweden background values

(Swedish EPA 2000), showed especially high deviations from the background values for Cr

and Cu, while Cd, Pb, Ni, and Zn showed at most sampling points only no or slight deviation.

Large or very large deviation was detected for Cr at 6 sampling points, for Cu at 4 points, and

for Ni at 1 point. Two sampling points, both located downstream the mouth of a ditch, show

the highest LOI concentrations and exhibited also the highest concentrations of fine grain

fractions (<0.063 mm). Wave impact on the sediment is at these sampling points decreased by

vegetation and a deeper water column, favouring settling of fine grains. For all sample points

(except point GC where only one autumn sample was taken) seasonal changes in particle size

composition are observed with a higher content of fine grain sizes (<0.125 mm) in May after

snowmelt. That indicates varying stream conditions in the ditches. Changed runoff intensity

causes change in sediment loads. It is likely that the low runoff during winter/snowmelt with

its lower velocity only has the capacity to transport fine particles. Since there are no intense

runoff events during a stable winter season, fine particles settle in the ditches or the recipient

itself. Along the shores in the relatively open bay of Notviken the sedimentation conditions

32

can vary strongly due to ice covering and wave activity. During the ice free season fine grain

sediments are kept in the water column or redistributed by wave activities.

For all samples the SiO2 and Al2O3 concentration were almost identical compared to the

sediment from a bay at the Lule River mouth (non-stormwater affected reference point

Gültzauudden). Differences in other element concentrations were, however, noticed for all

sampling stations with deviations being especially high at Gammelstadsviken and Ytterviken.

The three sample points at Gammelstadsviken stood out with spatial differences and

seasonal variation in element concentrations. The sample point GA nearest to the stormwater

outlet showed less seasonal variation in trace metal concentrations than GB, even though they

had seasonal variation in grain size in common. GA exhibited less LOI in spring when the

grain size fractions complied with GB. Fe2O3 and MnO showed significant positive

correlation in the sediments suggesting their common occurrence.

The first component on the Score Scatter plot seems likely to point out geographical

similarities in element concentrations along the ditches. Related to the geochemical conditions

along a ditch, enrichment of elements can occur, where the water column is relatively stable

and organic material is present. The second component is in particular affected by the

elements Hg, Cu, and Ca, which have the highest concentrations at GA and GB in spring.

Similarities between GA and GB are also shown in the Loading Scatter plot. A nearby road

and bridge construction site can have affected the increase of Ca in the ditch sediment related

to concrete works at this site. The Hg and Cu concentrations can be related to the construction

site too, or to the nearby railway (Malawska & Wio�komirski, 2001). The third component is

notably affected by the differentiation in particle sizes smaller and larger then 0.125 mm, and

with that it mostly represents seasonal variation in particle transport and sedimentation at the

sampling sites.

The sample points YC and GC, downstream the mouth of a ditch, are characteristic for the

surface sediment of brackish-lacustrine bays along coasts of the Bothnian bay. Due to the

standing body of water and decomposition of the high organic content, suboxic/anoxic

conditions can exist already in the surface sediment. Bacterial sulphate reduction can in that

case account for enrichment of FeS, FeS2 and other metal sulphides in the sediment (Boman,

Fröjdö et al. 2010). Accompanied to that the increased concentrations of Cd, Co, Cr, Cu, Ni,

Pb and Zn can be caused by desorption to organic complex builders and fixation with

sulphides. So the organic material and the fine grained mineral fraction can exhibit adsorption

surfaces for metals, but also formation of FeS and further on FeS2 may lead to metal fixation.

33

Paper IV A stormwater pond that receives highway runoff in Sollentuna, had elevated heavy metal

concentrations in the water column and elevated heavy metal concentrations and PAH

concentrations in the sediment. For the surface water samples the speciation of Ca, Cd, Co,

Cr, Cu, Fe, K, Mg, Mn, Na, Ni, Pb, S, and Zn was determined with membrane filtration and

ultrafiltration (particulate phase >0.22 μm; colloidal phase <0.22 μm and >1 kDa; truly

dissolved phase <1kDa). Sediment and porewater concentrations of Al2O3; Al, Cd, Co, Cr,

Cu, Fe2O3; Fe, Na2O; Na, Ni, Mn, Pb, S, SiO2; Si, and Zn were determined.

The elements Cd, Co, Cr, Cu, Fe, Mn, Na, Ni, Pb, S, and Zn showed seasonal variations in

element concentrations in the water column. The concentrations in the water column of Ca, K,

Mg, Mn, Na and S were found truly dissolved to 100 % during both summer and winter. For

these elements the concentrations in winter are higher than the summer concentrations, most

obvious for Na with a 5 times higher concentration in winter.

Higher concentrations in the water column at wintertime were observed also for Cd, Co, Cr,

Cu, Fe, Ni, Pb and Zn, which occur in different speciations than the truly dissolved phase

only. For Cu and Ni the dissolved and particulate phases increase most in winter. Iron and Pb

are in both seasons dominated by the particulate phase.

The seasonal variation of total concentrations of Cd, Cr, Co, Cu, Ni, Pb and Zn in water

from summer 2009 to winter 2010 can be a consequence of road salt applied as a de-icing

agent and increased street wear due to use of studded tires in winter. Previous studies have

shown the relation that use of de-icing agents in combination with use of studded tires result

in higher metal concentrations in road runoff (Hvitved-Jacobsen & Yousef 1991; Legret and

Pagotto 1999; Bäckström et al. 2003). Even if seasonal variation in the metal concentrations is

in accordance with other studies, the total concentrations measured in basin 2 are low in

comparison with Legret and Pagotto (1999), Bäckström, Nilsson et al. (2003) and Karlsson et

al. (2010).

In the sediment a concentration change is present for LOI and more or less all elements (Si,

Al, Na, Mn, Fe, S, Cd, Co, Cr, Cu, Ni, Pb, and Zn) at 3-5 cm depth in both summer and

winter. The concentration change for the elements at a depth of 3-5 cm shows the boundary

between the collected stormwater sediment and the macadam ground of the constructed

stormwater basin. For LOI the concentration in the upper 3 cm is constant around 25 % DM,

and then it decreases to less than 3 % DM at 5 cm depth, and it is constant around 2 % DM in

the sediment deeper than 5 cm. The clear change in sediment composition allows the

34

estimation that the upper 3.5 cm of the sediment have settled since the stormwater facility was

taken in use in 1998.

Si and Al concentrations in the sediment have similar characteristics in their concentration

profile with increase in the solid phase from 3-5 cm depth downwards. The Na concentrations

in the solid sediment resemble the profiles of Si and Al. For the concentrations in porewater,

Na shows high variation between summer and winter. In summer, porewater and surface

water concentrations for Na are constantly close to 100 mg/l or below. In winter the

concentrations rises in the surface water to the fivefold. The porewater concentration drops

from concentrations >500 mg/l in the sediment top to 168 mg/l at 11 cm depth. Manganese

shows a little higher concentration in winter in the sediment top than in summer (Figure 13).

The porewater has a Mn minimum in winter and lower concentrations than the bottom near

water and surface water in the basin. The peak of solid Mn at the sediment surface in winter

indicates oxic conditions with formation of Mn oxyhydroxides (Davison 1993). The Fe

concentrations in the solid sediment have a relative peak in the sediment top (3.4 %DM

summer; 3.8 winter %DM), but the concentrations in the upper 5 cm are in general lower than

concentrations in the deeper sediment. The relative peak of Fe in the solid phase at the

sediment surface in winter indicates that also Fe-oxyhydroxides have formed, but low Fe

concentration in the surface water and porewater do not show dynamics at the sediment water

interface (Figure 13). This suggests that Fe reaches the sediment mostly in particulate form,

which is supported by the Fe speciation in the water column. The sulfur concentrations in the

solid sediment have a peak at ca 3 cm sediment depth (Figure 13). In the sediment deeper than

5 cm, the S concentrations are more than 90% lower with the exception of a relative peak at

6.5 cm depth. Especially in winter the porewater profile matches the solid S profile. Sulfur

peaks in porewater are placed just about 1 cm above the peak in the solid sediment. Thus, the

porewater concentration increases in winter from the top (with concentration similar to the

surface water) until the peak at 2 cm depth. From there the concentration decreases until the

relative minimum at 3 cm depth were the solid sediment has a peak. Below 3 cm the

porewater concentration increases until 5-6 cm depth, from where the concentration drops

continuously with depth. The S-enriched layer at 3 cm depth indicates precipitation of solid

sulphides in the stormwater sediment just above the border to macadam. At the same depth

depletion in the porewater concentration of Fe can be observed which indicate Fe-sulphide

formation (Fortin et al. 1993).

35

0 0.02 0.04 0.06 0.08MnO, DM %

200

-15

-10

-5

0

5

Dep

th in

cm

0 20 40 60 80 100Mn, µg/l

0 0.02 0.04 0.06 0.08MnO, DM %

200

-15

-10

-5

0

5

0 20 40 60 80 100Mn, µg/l

summer winter

0 2 4 6Fe2O3, DM %

200

-15

-10

-5

0

5

Dep

th in

cm

0 2 4 6 8Fe, mg/l

0 2 4 6Fe2O3, DM %

200

-15

-10

-5

0

5

0 2 4 6 8Fe, mg/l

summer winter

0 1000 2000 3000S, mg/kg DM

200

-15

-10

-5

0

5

Dep

th in

cm

0 20 40 60 80S, mg/l

0 1000 2000 3000S, mg/kg DM

200

-15

-10

-5

0

5

0 20 40 60 80S, mg/l

summer winter

surface water

solid sedimentporewatersediment surface

Figure 13. MnO, Fe2O3, S in the stormwater basin sediment and Mn, Fe, S in porewater and surface water (both 0.22 �m filtered).

36

A concentration increase in the solid sediment at 5 to 3 cm depth upward is most evident for

Cu and Zn. Also Cd, Co, Cr, Ni, and Pb have higher concentrations in the solid sediment in

the upper sediment section (0-3 cm) than just below. For the metals Cd, Co, Cr, Cu, Ni, and

Zn the concentrations in porewater at 1-2 cm depth in winter are lower than in the surface

water, which means that diffusion of these elements into the sediment is likely. The organic

material offers precipitation surfaces and coating on Mn- and Fe-hydroxides or bonding under

anoxic conditions on sulfates is most likely.

In the sediment pollutants can be trapped due to precipitation on organic material, and early

diagenesis processes with formation of Mn- and Fe-hydroxides and sulphide reduction. This

will just affect a fraction of the concentrations of metals in truly dissolved phase while most

of the dissolved concentrations are most likely not retained in the stormwater facility. Most of

the dissolved concentrations are supposedly transported further on leaving the stormwater

facility. A technical solution could be the application of a peat-filter to bind metal cations. A

fraction of the concentrations of metals in truly dissolved phase can also diffuse into the

sediment. The precipitation on organic material, and early diagenesis processes with

formation of Mn- and Fe-hydroxides and sulphide reduction are capable to trap pollutants.

The PAH concentrations in the stormwater basin sediment are in general higher in the

surface (0-2 cm) than in the deeper part (6-7 cm). According to the Swedish EPA

classification for coast sediments (Swedish EPA 1999), the sum of 11 PAHs are on a

moderate level at both sediment depths.

Resume The investigated water bodies in the Luleå area show clear urban impact on sediment

quality. The metals Cd, Cu, Pb and Zn, which are of main concern in urban stormwater, are

enriched in all investigated bays. Metals can bind to surfaces of settling organic material and

small inorganic particles. In the sediment they can become part of sulphide formation and

thus be fixed in the sediment.

Water and sediment quality in the Luleå area are dependent on catchment area (size, natural

premises and exploitation) and emission impact, especially from point sources. Important

factors are water volume and turnover rate in the water bodies with low water levels and no

surface runoff wintertime, and ice covering during winter, which also contributes to anoxic

conditions in water column and sediment. The redox status in the sediments is crucial for

release or bonding of pollutants in the sediments. The bays do have the capacity to retain

37

pollutants in their sediment, but there is still a potential risk of release if the redox conditions

change. Postglacial land uplift implies continuous changes in the environment, which can lead

to changing redox conditions. This will necessitate new risk assessments.

Skutviken was investigated in detail. It has functioned as a large stormwater pond since the

road bank was constructed in 1962, with calm conditions within the bay and a limited water

exchange with the Lule River. Stormwater contamination has resulted in increased

concentrations of Cd, Cu, Pb and Zn in the sediment deposited after 1962. Also the PAH

concentrations are very high for Pyrene and high for Phenanthrene, Anthracene, Fluranthene,

Benzo(a)anthracene, Chrysene, Benzo(k)fluoranthene and Benzo(a)pyrene. An increased

settling of particulate matter and seasonal occurrence of anoxic bottom waters leading to

sulphate reduction appear to be the main effects of the road bank. Sedimentation of pollutant

carriers and the sulphate reduction result in an increased fixation of metals and PAHs in the

sediment. Skutviken appears to be an efficient trap for stormwater contamination. The use of

trace metal ratios could not identify road runoff as main source for sediment pollution.

Especially in winter the water column in Skutviken is enriched in metals such as Co, Cr,

Ni and Zn compared with the reference site, but the enrichment is not very strong. About half

the content of Co, Ni and Zn is truly dissolved, but only a fifth for Cr.

The study of stormwater ditches and associated sediments showed that stormwater

discharge has impact on the metal concentrations in the analyzed surfaces sediment. The

recipients GC and YC clearly exceed metal concentrations from the reference point for Cd,

Co, Cr, Ni, Pb and Zn. Crucial for retention of metals is a calm water column, and occurrence

of organic material in the recipient. Redox conditions, which can lead to metal sulphide

formation, are likely controlled by decomposition of organic material at the studied sites.

The higher contents of the fractions <0.063 mm (silt and clay) in spring of all surface

samples (0-2 cm) in the ditches is based on seasonal variation in runoff. Changed runoff

intensity causes change in sediment loads. The spring sampling was conducted after the main

snowmelt but before the first intense rain event in Luleå. During winter runoff transport

capacity is lowered. Snowmelt occurred relatively constant with daily variation in contrast to

flush floods during a heavy rain event. A lower runoff velocity comprises sediment transport

of fine grain sizes only.

The case of the stormwater pond at Sollentuna allowed the study of the influence of

highway runoff without the influence of other pollution sources. Elevated heavy metal

concentrations in the water column and elevated heavy metal concentrations and PAH

concentrations in surface sediment were found. Seasonal variations in element concentrations

38

are most evident for Cd, Co, Cr, Cu, Fe, Mn, Na, Ni, Pb, S, and Zn in the water column.

Especially in winter the metal concentrations of Co, Cr, Cu, Ni, Mn, Na, and Zn were

dominated by the truly dissolved phase. Most of the dissolved concentrations will be

transported further on leaving the stormwater facility to the recipient. A technical solution

could be the application of some sort of filter, for instance a peat filter, to bind metal cations.

A fraction of the concentrations of metals in truly dissolved phase can also diffuse into the

sediment. Sorption to organic material, and early diagenesis processes with formation of Mn-

and Fe-hydroxides, and sulphur reduction further down are capable to trap pollutants. Road

salts can affect the partitioning of metals leading to an increased fraction of the more

environmentally harmful dissolved phase, and use of studded tires in winter is a potential

pollution source.

The degree of pollution in the sediment in the stormwater pond at Sollentuna is lower than

in the most of the bays in the Luleå area. This indicates that road runoff is not the only

explanation to pollution in the bays in the Luleå area, probably not even the most important.

39

AcknowledgementsI would like to thank my supervisors Björn Öhlander and Anders Widerlund for contributing with their knowledge and support, and spending time in discussions beside their other tasks. I will not forget how Fredrik Nordblad helped with sampling and analysis, no matter what time or weather it was. Thanks! Also Magnus Westerstrand proved that he stands out with rain and big waves in a small boat, and of course with me. The latter did even Kristin Karlsson and Godecke Blecken and all my colleges at the Division of Applied Geology and the Urban Water Research Group, which were there with a helping hand when I needed it. Thanks to Milan Vnuk and Kent Bergström, who worked with the layout. The thesis has been financed mainly by Luleå University of Technology and the Swedish Research Council for Environment, Agriculture Sciences and Spatial Planning (FORMAS). I also have to thank my family, kombos and friends for the support since I came to Luleå.

40

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Appendix

AbbreviationsACE Acenaphthene ACY Acenaphthylene ANT Anthracene BaA Benzo(a)anthracene BaP Benzo(a)pyrene BbF Benzo(b)fluoranthene BkF Benzo(k)fluoranthene BMP Best management practice BPY Benzo(ghi)perylene Ccoll Colloidal concentration Cfeed Concentration of the feed sample CFF Cross flow filtration CFR Cross flow ratio CHY Chrysene Cperm Concentration of permeate Cret Concentration of retentate DBA Dibenz(a,h)anthracene DM Dried matter EF Enrichment factor FL Fluorene FLR Fluoranthene GC-MS Gas chromatography mass spectrometry HS Highest shoreline ICP-AES Inductively coupled plasma atomic emission spectrometry ICP-SFMS Inductively coupled plasma sector field mass spectrometry INP Indeno(1,2,3-cd)pyrene MOH Mineral oil hydrocarbons NAP Naphthalene PAH Polycyclic aromatic hydrocarbons PE Poly-ethylene PHEN Phenanthrene PYR Pyrene Qperm Permeate flow rate Qret Retentate flow rate R Mass balance recovery SSAB Svenskt stål aktiebolag TC Total carbon TN Total nitrogen TSS Total suspended solids UNFPA United Nations Fund for Population Activities Vperm Volume of permeate Vret Volume of retentate WFD Water framework directive wt weight

Paper I

Impact of Urban Stormwater on Sediment Qualityin an Enclosed Bay of the Lule River, Northern Sweden

Ralf Rentz & Anders Widerlund &

Maria Viklander & Björn Öhlander

Received: 29 April 2010 /Accepted: 20 October 2010# Springer Science+Business Media B.V. 2010

Abstract Sediment and porewater samples from anenclosed bay receiving stormwater discharge (Skutviken)near the centre of Luleå, northern Sweden wereanalysed for major and trace elements and 16polycyclic aromatic hydrocarbons (PAHs). Amongthe studied metals Cd, Cu, Pb and Zn wereenriched at Skutviken. Also, the PAH content wasenriched, in particular for phenantrene, anthracene,fluoranthene and pyrene which are regarded ascommon constituents in stormwater. The use oftrace metal ratios provided indications about pol-lutant sources for the sediment. Cs-137 dating wasused to determine historical changes in metal andPAH fixation in the sediment. The bay Skutviken isenclosed through the construction of a road banksince 1962. The enclosure led to reduced watercirculation in the bay that promotes the occurrence ofanoxic conditions with sulphate reduction within thebay. As a consequence of these conditions, metals aretrapped in the sediments as sulphides. This studysuggests that enclosed bays with restricted watercirculation may be efficient traps for urban pollutants,

reducing the present-day input of pollutants to the sea.In areas with postglacial land uplift, where such bays arecommon, bay sediments are a potential future source ofpollutants when uplift results in erosion and oxidation ofthe sediments.

Keywords Stormwater . Sediment quality .

Trace metals . PAH

1 Introduction

Urban hydrosphere and pedosphere are parts of anurban natural system (Endlicher 2004), which isintensely affected by human activities. In 2008, forthe first time in history, more than half the humanpopulation in the world lives in urban areas, possiblyincreasing to 80% in 2030 (UNFPA 2007). InSweden, today 84% of the population already livesin urban areas (Statistics Sweden 2006), and thus thisenvironment and its own “natural driving forces andpatchwork patterns” (Endlicher and Simon 2005) forthe society are very important.

Rivers are important for both natural systems andhuman societies (Simmons 1991). Hauer and Lam-berti (2006) use the term riverscape to describe the“expansive view of a stream or river and itscatchment, including natural and cultural attributesand interactions”, which may change with time.

In a riverscape, surface waters and groundwater aswell as sediments and soils will be affected by

Water Air Soil PollutDOI 10.1007/s11270-010-0675-7

R. Rentz (*) :A. Widerlund :B. ÖhlanderDivision of Geosciences, Luleå University of Technology,SE-97187 Luleå, Swedene-mail: ralf.rentz@ltu.se

M. ViklanderDivision of Architecture and Infrastructure,Luleå University of Technology,SE-97187 Luleå, Sweden

stormwater discharge, which is an important contam-ination source for trace metals and polycyclic aro-matic hydrocarbons (PAH) (Brown and Peake 2006).Accumulation of metals and organic pollutants inrecipients are a risk for living organisms (Wildi et al.2004; Munch Christensen et al. 2006). Kayhanian etal. (2008) report grab and composite samples fromurban highway runoff in Los Angeles to be toxic onfreshwater and marine species, where in general thefirst samples taken during a storm event were foundmore toxic than those collected later. Previous studiesof stormwater and gully pot sediments in Luleå innorthern Sweden (Westerlund 2007; Karlsson andViklander 2008a) indicated particle-related transportof metal and organic pollutants with seasonal varia-tions. McKenzie et al. (2008) point out that tracemetals from anthropogenic sources were enrichedtogether with stormwater transported particles, whereenrichment increased with decreasing particle size.

The objective of this study was to investigatehow an enclosed bay of the Lule River in northernSweden affects the transport of urban metal andorganic pollutants to the nearby Lule River estuary.The objective is based on two hypotheses: (1) thatmetals are trapped as sulphides in the bay sedimentand (2) that sediment grain size may be importantfor the sequestering of organic pollutants. To definethe urban impact in Skutviken, its sediment andporewater geochemistry was compared with areference sampling site unaffected by stormwaterdischarge.

2 Materials and Methods

2.1 Sampling Site

The bay Skutviken (Fig. 1) is located north of the citycentre of Luleå (73,000 inhabitants) in northernSweden. The most characteristic hydrodynamic pat-terns of Luleå are the Lule River and former shallowbays of the brackish Bothnian Bay, which are partiallyenclosed due to the postglacial rebound (8–9 mm a−1

(Lindén et al. 2006)) or artificial banks. The LuleRiver enters the Bothnian Bay passing the centre ofLuleå. The 25,000-km2 large catchment area of the460-km-long river has an annual average discharge ofaround 500 m3 s−1 (Raab and Vedin 1995). However,the water bodies situated close to Luleå are also

affected by smaller local catchments, which con-tain urbanised and industrial areas as well as ruraland forested areas (Erixon 1996; Hübinette 1998;Olofsson 2002).

The surface area of Skutviken is ∼12 ha, and themean and maximum depths of the bay are 1.6 and3.4 m, respectively. It is separated from the LuleRiver by a ca 360-m-long road bank constructed in1962 (Fig. 1). At the southern end of the road bank, achannel (8 m in width, 3 to 4 m in depth and 35 m inlength) through the bank permits a limited waterexchange with the Lule River. These physicalconditions give the bay similarities with shallownaturally enclosed bays in the region. The bay issurrounded by the road bank and one more highlyfrequented road with traffic intensities of 23,500 and13,600 vehicles per day, respectively (Luleå Kommun2007). The sewer drainage area contains 0.53-km2

industrial area and 0.73-km2 housing area (Fig. 1).Since parts of the road bank runoff and six

stormwater channels enter the bay, it almost functionsas a large stormwater pond where a high amount ofstormwater sediment is trapped, resulting in a reducedsediment supply to the Lule River. All outlets arelocated below the water surface, except duringperiods of very low water level. A reference samplingsite was chosen beside the main stream of the LuleRiver in front of the spit Gültzauudden (Fig. 1).

The annual precipitation in the Luleå area is about500 mm of which 40% to 50% falls as snow betweenNovember and April/May (Hernebring 1996), and isthus discharged during snowmelt. From Novemberuntil May the Lule River and the bays close to the citycentre are ice covered.

2.2 Sampling

The sampling station in Skutviken was chosen in thedeeper part of the bay with fine grained sediment.Sediment cores (25–30 cm long) were collected fromSkutviken and Gültzauudden in March 2007 and 2008using a Kajak gravity corer with a core tube diameterof 64 mm. Sampling was performed from the winterice, and the sediment core surfaces were judged to beundisturbed (no resuspended sediment in core tubesand apparently undisturbed surface sediment struc-tures). The cores were sectioned in subsamples(0.5 cm thick for the uppermost 3 and 1 cm thickuntil the core ends). For porewater analyses, the

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sediment samples were put into plastic bags immedi-ately after core collection and sectioning in the field.All air was pressed out of the bag before it was placedin an Ar-filled container to keep the sediments in anoxygen free environment until the porewater wasextracted within the following 6 h. The porewater wasseparated by vacuum filtration (0.22 μm Millipore®membrane filters) arranged in an Ar-flushed glovebox. The porewater samples were collected in 60-mlacid-washed polyethylene bottles and refrigerateduntil further analysis. Bottom water was sampledfrom the core tube immediately after retrieval, 3 cmabove the sediment surface. The water was drawnwith a small plastic tube fixed on a syringe andfiltered through a 0.22 μm Millipore® membranefilter.

2.3 Analytical and Chemical Analyses

The total carbon (TC) and total nitrogen (TN) of thesediment was analysed by Umeå Marine SciencesCentre. Analyses of carbon and nitrogen in sedimentswere performed with a Carlo Erba model 1108 hightemperature combustion elemental analyzer, using

standard procedures and a combustion temperatureof 1,030°C. For standardisation Acetanilide wasutilised.

Metal and PAH analyses were accomplished by theaccredited laboratory ALS Scandinavia AB in Luleå.The sediment and porewater was analysed for majorelements and trace metals. Sediment samples fordetermination of As, Cd Co, Hg, Ni, Pb and S weredried at 50°C and digested in a microwave oven inclosed Teflon bombs with a nitric acid: water ratio of1:1. For other elements, 0.125 g dried matter wasmelted with 0.375 LiBO2 and dissolved in HNO3.Metal determinations were made by inductivelycoupled plasma atomic emission spectrometry (ICP-AES) and inductively coupled plasma mass spectrom-etry. To the porewater samples 1 ml nitric acid(suprapur) was added per 100 ml. Analyses weremade with ICP-AES and inductively coupled plasmasector field mass spectrometry. The following 16PAHs were analysed in the sediment: naphthalene,acenaphthylene, acenaphthene, fluorene, phenan-threne, anthracene, fluoranthene, pyrene, benzo(a)anthracene, chrysene, benzo(b)fluoranthene, benzo(k)fluoranthene, benzo(a)pyrene, dibenz(a,h)anthracene,

Fig. 1 Location of the study area Skutviken (a) and the reference sampling site at Gültzauudden (b) in Luleå, Northern Sweden andthe stormwater sewer catchment area at Skutviken

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benzo(ghi)perylene and indeno(1, 2, 3-cd)pyrene. ThePAH sediment samples were leached with acetone/hexan/cyclohexan (1:2:2), and measurements wereperformed with gas chromatography mass spectrometry.

Particle size analyses were performed with a Cilas1064 laser diffraction particle size analyser in wetmode for four samples from a profile at Skutvikenand a profile at Gültzauudden.

Water fraction and porosity were determinedthrough weighing before and after drying the sedi-ment at 50°C for at least 7 days. The dissolvedoxygen in the water column was determined with aHydrolab® MiniSonde 5 water quality probe.

Radionucleide activity of 137Cs (mean standarddeviation±5%) was determined by gamma spectrom-etry at Risø National Laboratory for SustainableEnergy, Denmark.

3 Results and Discussion

3.1 Sediment Characteristics

3.1.1 Particle Size and Sedimentation Rate

The particle size analyses showed that the 2–3 and 5–6 cm layers at both sites had very similar particle sizedistribution (Fig. 2). The main components (60%cumulative volume) in these layers had a grain sizefrom 10 to 30 μm. At Skutviken, the 10–11 cmsample contains the overall finest sediment with 70%accumulated volume in particle size 2–10 μm. The15–16-cm-layer particle size distribution at Skutvikenfalls between the two uppermost and third layer withrespect to particle size. At Gültzauudden the 10–11-cm layer contains the finest material at this site with60% cumulative volume containing grain size 5–11 μm.The deepest sample (15–16 cm) shows the coarsestgrain composition with 60% cumulative volume con-sisting of material with the grain size 20–100 μm.

The activity of the radionuclide 137Cs shows 2peaks (Fig. 3). The upper peak 4 cm upwards isinterpreted to represent the Chernobyl fallout from thereactor accident in April 1986 (Ilus and Saxén 2005),while the lower peak is interpreted to be caused by thefallout from nuclear weapons testing in the early1960s (Appleby 2002). However, this peak should beconcurrent with the construction of the road bank in1962, and may be displaced slightly downward in the

sediment due to reworking of sediments during con-struction works. Caesium-137 data indicate that changesin sediment characteristics (particle size, concentrationsof TC, TN, metals and PAHs) from 11 cm and upwardsbecame apparent in the early 1960s.

Fig. 2 Particle size distribution at Skutviken and Gültzauuddenfor sediment samples at 2–3, 5–6, 10–11 and 15–16 cm

Fig. 3 Plotted 137Cs (Becquerel per kg (Bq kg−1)) versussediment depth at Skutviken

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3.1.2 Redox Conditions

Dissolved oxygen was measured in the water column toprovide information on the redox conditions at thesediment-water interface. At Skutviken the oxygensaturation in the water 10 cm above the sedimentsurface is close to 0% in wintertime, when the bay is icecovered. In contrast, the water column is well oxygen-ated (saturation 85–90%) during the ice free season.These changing redox conditions can affect release oraccumulation of pollutants through formation/dissolu-tion of Fe–Mn oxyhydroxides in the surface sediment.

Sediment cores contain information about past andpresent processes in the sediment. It is possible tofollow element concentrations back in time, assumingthe stratigraphy is undisturbed. At Gültzauudden, thehigh Mn content in the sediment top layers (Fig. 4)

can be related to the oxic environment at this sitewhere Mn occurs mostly as Mn oxyhydroxides(Davison 1993). The decomposition of organicmaterial and increasingly anoxic environment withsediment depth results in reduction of Mn oxyhydr-oxides and increased porewater concentration of Mn(II). A breaking point for the Mn in the solid phase isreached at 4 cm depth where the MnO contentstabilises at 0.2%. Together with the increasingporewater Mn concentration, this suggests that anoxicconditions predominate below 4 cm. The porewaterprofile indicates Mn(II) flux upward, resulting in theoxidation of Mn(II) to Mn(IV) in the oxic parts of thesediment (Davison 1993; Wehrli 1991) (Fig. 4). Thesediment content and porewater concentration of Fe atGültzauudden comply with the Mn observations. TheFe2O3 peak in the sediment profile is situated below

Fig. 4 MnO and Fe2O3 insediment (wt.%) and Mnand Fe in porewater (μg l−1)at Skutviken and Gült-zauudden. The top value forthe “porewater” representsthe bottom water (3 cmabove sediment surface)

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the MnO peak. In oxic sediment Fe occurs as Fe(III)in iron oxyhydroxides, resulting in a solid Fe peak at3 cm depth. Below 5 cm the solid Fe content declinescontinuously. The porewater Fe concentration indi-cates that reduction of solid Fe(III) to the soluble Fe(II) occurs when porewater becomes more anoxic(Davison 1993, Wehrli 1991).

At Skutviken the Mn and Fe sediment and pore-water concentrations differ from those at Gültzauud-den (Fig. 4). The MnO content in the sediment ismuch lower than at Gültzauudden in the upper part ofthe sediment. The geochemical conditions where Mn(IV) is reduced to Mn(II) appear to be reached alreadyin the bottom water above the sediment surface.During winter, when the bay is ice covered, theoxygen concentration in the bottom water is<0.42 mg l−1. In the porewater, Mn concentrationsincrease with depth but never reach as high concen-trations as at Gültzauudden.

The presence of a solid Fe2O3 maximum at thesediment surface at Skutviken indicates that the redoxconditions permit precipitation of Fe(III) hydroxidesat the sediment-water interface. The same anoxicconditions that occur at a depth of 3 cm in thesediment at Gültzauudden seem to occur alreadyabove the sediment column in Skutviken, withreductive dissolution of Fe hydroxides taking placealready at the sediment surface. The decrease of totalS in porewater at Skutviken suggests that reduction ofSO4

2− occurs immediately below the sediment-waterinterface (0–2 cm). The simultaneous increase of solidS indicates precipitation of solid sulphides in thesediment (Fig. 5). The solid S concentration at 0.5–

11 cm depth (2,500–4,200 mg kg−1) exceed that atGültzauudden by a factor of 5–7.

3.1.3 Element/Al Ratios in the Sediment Profiles

Regional element/Al ratios have been found to berelatively constant in sediment, also when sedimentgrain size changes and sedimentation rates vary (Hirst1962; Loring 1991; Ebbing et al. 2002). In thesampled sediments the element/Al ratios for the majorelements Ti, Ca, Mg, Na and K are similar atSkutviken and Gültzauudden, with only small devia-tions from local till ratios for Ca/Al, Na/Al and K/Al(Table 1). This indicates that both sediments mainlyare composed of local minerogenic matter. In the1–7 cm section, the Fe/Al and the Mn/Al ratios arehigher at Gültzauudden (Table 1), suggesting pre-cipitation of Fe–Mn oxyhydroxides in a more oxicenvironment (Davison 1993).

Peinerud et al. (2001) used the Si/Al ratio of lakesediments as a measure of the diatom concentration.In the two sampled cores the Si/Al ratio is even lowerthan that of local till (Öhlander et al. 1991), suggest-ing a negligible content of diatoms at both samplingsites (Table 1).

3.2 Total Carbon and Nitrogen in Sediments

TC at both sites shows high concentrations in thesurface sediment and a decrease with depth (Fig. 6).At Skutviken the concentration in the upper sedimentsegment (1–7 cm) is 4–5%, which is significantlyhigher than at Gültzauudden (1–2.5%). At Skutviken

Fig. 5 Sulphur in solid sed-iment (mg kg−1) and S inporewater (mg l−1) at Skut-viken and Gültzauudden.The top value for “pore-water” represents the bot-tom water (3 cm abovesediment surface)

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TC decreases sharply below 7 cm depth to ca 1% at10 cm depth, from where on the TC concentration isapproximately constant. The content of TN follows asimilar pattern as for TC at both sample sites (Fig. 6).The TC/TN molar ratio indicates a change insediment composition at Skutviken from 7 to 11 cmdepth, where the TC/TN ratio decreases from 19 to11. Below 11 cm depth the TC/TN ratio of both sitesare similar. Above 11 cm, the concentration of organicmaterial is enriched at Skutviken which is consistentwith low oxygen saturation above the sediment inwintertime. The TC/TN molar ratio is therebyhigher than the C/N ratio of 6.6 in the Redfieldempirical formula ((CH2O)106(NH3)16(H3PO4))(Redfield 1958), which indicates an anthropogenicimpact.

3.3 Trace Elements in Sediments Comparedwith Reference Values

For the sediment section 0–2 cm the detected contentsof As, Cd, Co, Cr, Cu, Hg, Ni, Pb and Zn can becompared with reference values for coastal sediment

from the Swedish Environmental Protection Agency(Swedish EPA 1999) and a deviation value can bedetermined by dividing the sediment content valuewith the reference value (Table 2). According toSwedish EPA (1999), the deviation values for Cu(3.63) and Zn (2.98) at Skutviken are classified as“large”, while the deviation at Gültzauudden onlyshows “slight” difference from the reference value.Cadmium (3.09) and Pb (1.87) appear with a“significant” deviation at Skutviken. Cadmium, Cu,Pb and Zn are of main concern in urban stormwater(Hvitved-Jacobsen and Yousef 1991). Thus, a signif-icant influence of stormwater sediment can beassumed for these four metals in Skutviken, while atGültzauudden no effect can be seen for any of thestudied elements.

3.4 Trace Elements in the Sediment and Porewater

Cadmium, Cu, Pb and Zn concentrations in porewaterand sediment are shown in Figs. 7 and 8. AtSkutviken, these elements show almost identicalsediment profiles with the highest concentrations

Depth (cm) Site Ti/Al Fe/Al Mn/Al Ca/Al Mg/Al Na/Al K/Al P/Al Si/Al

1–7 Gültzauudden 0.06 0.84 0.08 0.27 0.17 0.33 0.38 0.02 3.96

1–7 Skutviken 0.07 0.76 0.01 0.29 0.20 0.32 0.36 0.02 3.86

10–21 Gültzauudden 0.06 0.58 0.02 0.27 0.17 0.35 0.39 0.01 3.99

10–21 Skutviken 0.06 0.62 0.01 0.25 0.18 0.32 0.38 0.02 3.77

Local till 0.08 0.62 0.01 0.35 0.17 0.39 0.30 0.01 4.02

Cont. rock 0.05 0.48 0.01 0.31 0.18 0.30 0.28 0.01 3.81

Table 1 Mean element/Alweight ratios in differentsediment sections at Skut-viken and Gültzauuddencompared with mean weightratios of local till (Öhlanderet al. 1991) and continentalrock (Rudnick and Gao2003)

Fig. 6 Total carbon (TC), total nitrogen (TN), and mol ratio TC/TN in the sediment at Skutviken and Gültzauudden (TN value at21 cm depth at Gültzauudden <0.05%)

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above 6 cm depth, with exception of the surfacesediment. At Gültzauudden, the Cd, Cu, Pb and Znprofiles are different. The contents of Cd and Pb arethree times higher and Cu and Zn six times higher inthe 0.5 to 6 cm section at Skutviken compared withGültzauudden.

At Skutviken, low values were detected for Cd,Cu, Pb and Zn in the uppermost layer in the solidsediment (0–0.5 cm). The bottom water contents ofthese elements are below the porewater contents inthe uppermost sediment. Porewater maxima at orbelow the sediment surface indicate element transferfrom the solid sediment to the porewater for Cd, Pband Zn (Figs. 7 and 8). The porewater minima for theelements from 0.5 to ∼5 cm for the elements indicatea sink in the sediment. From 0.5 to ∼5 cm depth Cd,Cu, Pb and Zn show maxima in the solid sediment,coinciding with maxima for solid S, TC and the TC/TN ratio (Figs. 5 and 6). The change in concentrationsof Cd, Cu, Pb and Zn at Skutviken around 6 cm depthaccompanies a change in the composition of thesediment. The particle size distribution at Skutvikenwas similar for the upper two analysed layers (2–3and 5–6 cm). For both layers, the content of particles>10 μm is about 60%, while for the sample from 10to 11 cm depth the content >10 μm is 15%. Coarserparticles in the upper sediment column and higherTC suggest that elements with higher contents inthe upper sediment column may be more related toorganic components than mainly to clay minerals.Also the TC/TN ratio indicates a change insediment composition at Skutviken between 7 and11 cm depth. The high fraction of TC representsmostly organic compounds which decompose slow-

ly in the upper 7 cm of the sediment at Skutviken,since at this depth, anoxic conditions exist in thesediment column.

The S decline in porewater in the upper sediment atSkutviken signifies sulphate reduction and coevalsulphide formation in the solid sediment (Fig. 5).The enrichment of Cd, Cu, Pb and Zn in thesediment at 0.5 to ∼5 cm depth may thus be relatedto sulphide formation in the organic rich 1–7 cmsection of the sediment. Below 6 cm the sedimentcontents of Cd, Cu, Pb and Zn decline rapidly, andstabilise at a much lower value than in the 0.5 to∼5 cm section (Figs. 7 and 8). If organic compoundsact as carriers of trace elements, they can alsocontribute to the enrichment of Cd, Cu, Pb and Znin the upper 7 cm of the Skutviken sediment(Charlesworth and Lees 1999).

At Gültzauudden, the sediment and porewaterprofiles of As resemble those of Fe (Fig. 4), andappear to be coupled to the redox cycling of Fe.Porewater concentrations of As are low in theoxidised surface layer (0–2.5 cm), and a solid Asmaximum of ∼40 mg kg−1 occurs at 3.5 cm depth inthe sediment. At Skutviken, where anoxic conditionsprevail in the sediment, only a slight increase inporewater As up to 5–8 μg l−1 occurs below 2 cmdepth, and no solid maximum of As occurs in thesediment (Fig. 8).

The correlation of the trace elements Cd, Cu, Pband Zn with S shows a uniform pattern where thetrace element content increases with higher S content(excluding two samples from 6 to 11 cm depth thecorrelation coefficient is 0.98 for Cd, Cu, Pb and Zn)(Fig. 9). Two points with high S concentrations

Table 2 Comparison of trace element contents of the 0–2 cmsediment layer of Skutviken and Gültzauudden with the EPAcoastal and sea reference values for total analysis (Swedish

EPA 1999) and their deviation value for coastal sedimentscalculated as element concentration divided by EPA referencevalues

As Cd Co Cr Cu Hg Ni Pb Zn

Skutviken (mg kg−1) 7.47 0.62 11.33 82.8 54.45 0.08 22.48 57.95 253.1

Gültzauudden(mg kg−1)

18.53 0.31 15.23 71.63 19.03 0.07 21.1 13.3 106.38

EPA r.v. (mg kg−1) 10 0.2 14 80 15 0.04 33 31 85

Skutviken (d.v.) 0.75 3.09 0.81 1.04 3.63 1.95 0.68 1.87 2.98

Gültzauudden (d.v.) 1.85 1.53 1.09 0.9 1.27 1.74 0.64 0.43 1.25

The “deviation value” is calculated as “sediment content” divided by “reference value”

r.v. reference values, d.v. deviation values

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deviate from the main trend. These are situated in the6–11 cm depth section, where the Cd, Cu, Pb and Znconcentrations change rapidly. The trace elements Cd,Cu, Pb and Zn are also positively correlated with TC(Fig. 10) for the samples from 0.5 cm to 21 cm(correlation coefficient for Cd, 0.99; Cu, 0.98; Pb,

0.97; and Zn, 0.99). Only the 0–0.5-cm layer with thehighest TC content does not fit into this pattern. It isunclear whether organic matter is a carrier for Cd, Cu,Pb and Zn, or whether this pattern reflects a couplingbetween organic matter and sulphide formation in thesediment.

Fig. 7 Cd, Cu and Pb insediment (mg kg−1) and Cd,Cu and Pb in porewater(μg l−1) at Skutviken andGültzauudden. The top val-ue for “porewater” repre-sents the bottom water(3 cm above sediment sur-face). At Gültzauudden Cdwas only detectable inporewater at 0–0.5 cm sed-iment depth (detection level0.01 μg l−1)

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3.5 PAH Content in the Sediment

In general the most abundant PAHs in stormwater arephenantrene, anthracene, fluoranthene and pyrene(Gonzalez et al. 2000; Brown 2002), which areclassified as priority pollutants by the US Environ-mental Protection Agency (US EPA) (ATSDR 1995).All of them are found in high to very high concen-trations in the 0–2 cm sediment layer at Skutviken. Inthe 14–16 cm, only pyrene shows high contents. AtGültzauudden the PAH contents do not exceed moder-ately high contents (Tables 3). As found by Marsalek(1997) and Gonzalez et al. (2000), PAHs are correlatedto suspended solids and according to Krein andSchorer (2000), heavy PAHs (four to six benzo rings)are enriched in the fine and fine-middle silt phase ofroad runoff and light PAHs correlated with fine sand.

At Skutviken the particle size analysis for the 2–3 and 5–6 cm layers showed a range from fine tocoarse silt, offering conditions for light and heavy

PAHs to be associated with the sediment particles.In the upper 7 cm sediment section at Skutvikenthe TC content is permanently high around 5%suggesting a possible coupling to the presence ofPAHs (Menzie et al. 2002).

3.6 Stormwater Impact and Possible Sourcesof Contamination

Characteristic metals in stormwater like Cu, Cd, Pband Zn (Hvitved-Jacobsen and Yousef 1991) aresignificantly enriched at Skutviken compared withthe reference sampling site at Gültzauudden. Themean concentrations of Cu, Pb and Zn are with 60, 67and 287 mg kg−1, respectively, in the uppermost 6 cmof the sediment at Skutviken in the range of the metalconcentrations reported in street sediment on the roadbank that separates Skutviken from the Lule River(Viklander 1998) while the metal concentrationsreported in the gully pots are lower than in the

Fig. 8 Zn and As in sedi-ment (mg kg−1) and Zn andAs in porewater (μg l−1) atSkutviken and Gültzauud-den. The top value for“porewater” represents thebottom water (3 cm abovesediment surface)

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Skutviken bay (Karlsson and Viklander 2008b). Areason for this might be that most metals, withconcentrations higher in the Skutviken sediment thanin the gully pots, are attached to smaller particles.Gully pots are relatively poor in retaining smallparticles (Sartor and Boyd 1972). Compared withthe Swedish EPA (2000), the Skutviken sediment isclassified as class 3 for Cd, and Pb (biological effectscan be found), and class 4 for Cu and Zn (enhancedrisk for biological effects). The concentration ofmetals in the sediment at Skutviken was higher thanfound by Schiff and Bay (2003), in Santa Monica bay,USA, while it was in the same range as in an urbanstream in Denmark, where Munch Christensen et al.(2006) found that sediment and porewater were toxic toalgae. Assuming that the sediment above a depth of 6–

7 cm represents the time period after construction of theroad bank, stormwater impact appears to have increasedthe concentrations of Cd, Cu, Pb and Zn by a factor of3–4 (Figs. 7 and 8). However, these metals are probablypresent as relatively immobile metal sulphides.

The use of trace element ratios can help to identifythe potential sources of these contaminants. The ratiosfor Pb/Zn, Hg/Zn, Cd/Zn, Cu/Zn, Ni/Zn and As/Zn inthe Skutviken sediment are comparatively constantwith depth from 0.5 to 5 cm. Except for Hg, all ratioschange below 5 cm sediment depth (Fig. 11). In theupper 5 cm the Pb/Zn ratio follows the ratio for gullypot sediment from a road. For the Cr/Zn ratio achange below 5 cm depth to higher Cr impact for theSkutviken sediment can be noticed, while the 0.5 to5.5 cm section has a ratio close to both gully pot

Fig. 9 Element/S correla-tion in the Skutviken sedi-ment (mg kg−1)

Water Air Soil Pollut

ratios. Even though the gully pot sediment containsmore coarse particles than the Skutviken sediment,similarities for the trace element ratios are evident inFig. 11. If gully pots are an interim storage also forclay and silt (Morrison et al. 1988), similar ratios canindicate the stormwater particle transport chain. Thepollutants that are linked to the clay and silt fractionpass through gully pots and eventually reach the bay.These particle fractions also offer surfaces for PAHsto bind to (Evans et al. 1990).

In floodplain sediments from the Rhine Valleydeposited over the last 170 years, the verticaldistribution profiles of PAHs are similar to those ofthe heavy metals Cr, Cu, Pb and Zn (Gocht et al.2001). Even though the analysed sediment at Skut-

viken was accumulated over a shorter time period, thePAH profiles resemble in this case those of Cd, Cu,Pb and Zn, with high concentrations in the uppersediment and lower beneath. This suggests a commonstormwater origin for PAHs and trace metals. Theanoxic conditions in the Skutviken sediment hamperbiological activity and reduce the degradation oforganic matter, which results in accumulation oforganic matter (Canfield et al. 1993). Dissipation ofPAHs is less efficient (and limited to three-ring PAHs)in anoxic sediments when oxidation of organic matteris coupled with the microbial reduction of manganese,iron and sulphur (Quantin et al. 2005). As aconsequence, PAHs can accumulate with organicmatter. PAH affinity to fine particles is known from

Fig. 10 Element/TC corre-lation in the Skutviken sed-iment (total C in wt.%;metal contents in mg kg−1)

Water Air Soil Pollut

other studies (Budzinski et al. 1997; Krein andSchorer 2000) and seems certain for this study wherethe sediments are mostly covering the silt fraction atSkutviken.

4 Conclusions

Skutviken has functioned as a large stormwater pondsince the road bank was constructed in 1962, withcalm conditions within the bay and a limited waterexchange with the Lule River. This has resulted in aspatial arrangement of the sediment supply, with

coarse sand near the stormwater channels and inparticular silt and clay in the deeper central parts ofthe bay.

The stormwater contaminations have resulted inincreased concentrations of Cd, Cu, Pb and Zn in theupper 7 cm of the sediment. Also the PAH concen-trations are very high for pyrene and high forphenanthrene, anthracene, fluoranthene, benzo(a)anthracene, chrysene, benzo(k)fluoranthene and benzo(a)pyrene in the surface sediment at Skutviken.

An increased settling of particulate matter andseasonal occurrence of anoxic bottom waters leadingto sulphate reduction appear to be the main effects of

Table 3 Concentrations (μg kg−1) of 16 PAHs in the sediment from Skutviken and Gültzauudden compared with gully pot sedimentfrom a housing area and road in Luleå (Karlsson and Viklander 2008a)

Depth (cm) Skutviken (0–2) Gültzauudden (0–2) Skutviken (14–16) Gültzauudden (14–16) Housing area (mean) Road (mean)

PHENa 89f 22e 37e 21e 400 1,300

ANTa 24f <10d <10d <10d 90 300

FLRa 130f 28e 64e 52e 600 1,200

PYRa 240g 20e 56f 43e 300 700

BaAa 59f 13e 19e 18e 50 70

CHYa 69f <10d 23e 18e 30 40

BbFb 180f 10d 31e 25e 80 2

BkFb 44e <10d 14d 11d 20 0.3

BaPb 74f <10d 21e 21e 20 7

BPYb 89e <10d 15d 16d 40 100

INPb 99e <10d 19d 21d <340 30

Σ 11 PAHs 1,097f 93d 299e 246d

DBAb 30 <10 <10 <10 10 20

NAPa 39 <10 <10 <10 1,800 12,000

ACYa 11 <10 <10 <10 <250 <250

ACEa <10 <10 <10 <10 2 800

FLa 15 <10 <10 <10 200 600

Σ 16 PAHs 1,200 93 300 250 3,800 17,000

PAH concentrations (μg kg−1 ) in the Skutviken and Gültzauudden sediment at 0–2 and 14–16 cm depth judged after the Swedish EPAguidelines for 11 PAHs

NAP naphthalene, ACY acenaphthylene, ACE acenaphthene, FL fluorene, PHEN phenanthrene, ANT anthracene, FLR fluoranthene,PYR pyrene, BaA benzo(a)anthracene, CHY chrysene, BbF benzo(b)fluoranthene, BkF benzo(k)fluoranthene, BaP benzo(a)pyrene,DBA dibenz(a,h)anthracene, BPY benzo(ghi)perylene, INP indeno(1, 2, 3-cd)pyrenea Light PAHbHeavy PAHcClass 1, no contentd Class 2, low contente Class 3, moderately highf Class 4, highg Class 5, very high

Water Air Soil Pollut

the road bank. Sedimentation of pollutant carriers andthe sulphate reduction result in an increased fixationof metals and PAHs in the sediment. Skutvikenappears to be an efficient trap for stormwatercontamination since the sediment at Gültzauudden isalmost unpolluted.

The analysis of the trace element and PAHconcentrations in the sediment of a stormwater-receiving bay and a reference sampling site compared

with road runoff sediment enabled to identify thestormwater as an impact factor on the bay. Thesediment shows increased contamination of pollutantswhich most likely originate from stormwater. Fixationof pollutants in the sediment occurred for the last∼50 years after the building of a road bank.

This study suggests that enclosed bays withrestricted water circulation may be efficient traps forurban pollutants. As a consequence, the present-day

Fig. 11 Trace element/Znratios of Skutviken sedi-ment compared with gullypot sediment (<2,000 μm)from a housing area androad in Luleå (Karlsson andViklander 2008b)

Water Air Soil Pollut

input of pollutants to the sea are reduced. In areaswith postglacial land uplift, where such bays arecommon, bay sediments are a potential future sourceof pollutants when uplift results in erosion andoxidation of the sediments.

Acknowledgements This study was financed by LuleåUniversity of Technology (LTU) and the Swedish ResearchCouncil for Environment, Agriculture Sciences and SpatialPlanning (FORMAS). This support is gratefully acknowl-edged. For their help with analytical work we like to thankPer Roos at the Radiation Research Division at RisøNational Laboratory for Sustainable Energy, TechnicalUniversity of Danmark (DTU), Erik Lundberg at UmeåMarine Sciences Centre and Bertil Pålsson at the Division ofMineral Processing at LTU. We also thank Kristin Karlsson,Fredrik Nordblad and Magnus Westerstrand for assistanceduring the field work and contributing with their knowledgein discussions.

References

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Water Air Soil Pollut

Paper II

1

Urban impact on water bodies in the Luleå area, northern Sweden, and the role of redox processes Ralf Rentz a* and Björn Öhlander a

a Division of Geosciences and Environmental Engineering, Luleå University of Technology,

SE-97187 Luleå, Sweden

* tel. +46 (0)920-492193, fax +46 (0)920-491199, e-mail: ralf.rentz@ltu.se

Abstract Sediment and water from urban water bodies in the Luleå area, northern Sweden, were

studied to determine the degree of contamination from metals and PAHs (polycyclic aromatic hydrocarbons). The heavy metals Cd, Cu, Pb and Zn, which are of main concern in urban stormwater, are enriched in all investigated bays. PAH concentrations were also found to be enriched. The water and sediment quality of the investigated water bodies depends on catchment area characteristics and emission impact, from point sources in particular. Water volume and turnover rate in the water bodies with low water levels and no surface runoff during wintertime, and ice covering during winter, contribute to anoxic conditions in the water column and sediments. The present redox conditions in the water bodies predominantly cause fixation of pollutants in the sediment due to formation of sulphides and slow oxidation of organic pollutants. Postglacial land uplift implies continuous changes in the environment, which can lead to changing redox conditions, thereby necessitating new risk assessments.

Keywords

Redox, sediment, trace metals, urban, water

Introduction

Urban impact on water bodies

Water bodies in urban areas fulfil diverse functions. They are natural resources offering

food, drinking water and process water for industries (Simmons, 1991, Hauer and Lamberti,

2006). Water surfaces enhance quality of life for the dwellers and offer them space for

recreation and transportation. However, water bodies in urban environments are exposed to

emissions from manifold sources. These emissions are integrated in a chain of natural

processes affected by human activities. Urban waters are also a medium for sewage transport

2

(Walsh, 2000). Pollutants can reach water bodies in urban areas by airborne transport,

infiltration and, particularly, by surface runoff.

The diversity of urban environments with residential, commercial and industrial areas, roads

and parks affect adjacent water bodies differently. Without passing through any treatment

facility, stormwater can have great impact on water bodies and groundwater resources as well

as on sediments and soils.

Investigations of stormwater treatment that not only consider the question of efficient

drainage from urban areas but also minimization of pollution effects became more common in

scientific circles in the late-1960s in Sweden. This is reflected by works on stormwater

quality by Söderlund and Lehtinen (1970, 1971), who pointed out that stormwater transports

toxic substances in larger quantities to receiving water bodies than treated wastewater. Lisper

(1974) also concluded that the heavy-metal content in stormwater was as high as in

wastewater. Malmqvist (1983) gave a detailed picture of urban stormwater pollutant sources

for better prediction and control of stormwater runoff. Recent investigations deal with

prediction and simulation of stormwater flows (Björklund et al., 2011), or the efficiency of

stormwater treatment systems (Blecken et al., 2009), often related to finding Best

Management Practises (BMP).

Human impact and contributions of pollutants from urban areas to the environment have

been widely studied (Menzie et al., 2002; Förstner and Müller, 1981; Brown and Peake, 2006;

Gocht et al., 2001). Commonly investigated pollutants in stormwater are metals and

polycyclic aromatic hydrocarbons (PAHs), due to the potential risks they pose for living

organisms (Wildi et al., 2004; Munch Christensen et al., 2006).

The sources of the pollutants are as many as there are utilizations of their components, and

release mechanisms are complex. In urban environments metals occur in roofing materials,

cars, street lamps, crash barriers, gully covers, pipelines, cables, paints, computers, etc

(Brown and Peake, 2006). Exposure of these urban constituents to weathering processes

makes them a large artificial source of metals. PAHs originate generally from pyrogenic

sources, like fossil fuel or wood combustion, and petrogenic sources, such as petroleum

products. Also, wear and leaching of asphalt and tire wear contribute to the PAH content in

stormwater (Gonzalez, 2000).

The transport capability of stormwater for these pollutants to receiving waters is affected by

the particle size of the sediment load. Fine sand fractions, and especially silt and clay

fractions, were found to have the highest mass of metals and PAHs (Menzie et al., 2002). The

most abundant PAHs in stormwater are phenantrene, anthracene, fluoranthene and pyrene

3

(Lau and Stenstrom, 2005; Viklander, 1998). Previous studies of stormwater and gully pot

sediments in the Luleå area (Westerlund, 2007; Karlsson and Viklander, 2008), indicated

particle-related transport of metal and organic pollutants with seasonal variations. Pollutants

in dissolved form and associated with small suspended particles are not retained effectively in

their transport chain from urban surfaces to receiving waters. These pollutants can become

enriched in sediments under the right geochemical conditions when reaching standing water

bodies. Larger particle sizes have also been found to transport high metal contents (Brown,

2002; Gonzalez et al., 2000). The geomorphology and geochemistry of the water bodies and

their catchment area determine which and how processes take place in the water bodies. The

characteristics of different catchments in the Luleå area may have great impact on

geochemical processes in water bodies and sediments.

It is, therefore, of interest to study sediments affected by stormwater and to determine

whether water and sediment quality differs from other water bodies with less stormwater

impact. The main objectives of this study were to describe the water and sediment status of

urban waters in the Luleå area with its shallow bays with brackish water, and to identify

important geochemical and geomorphological processes and possible sources of pollution.

Other aims were to investigate the role of redox processes for fixation and release of metals in

local sediments, and to identify potential risks for dwellers and the environment that may

arise from the current environmental status and ongoing processes.

Water bodies in the Luleå area

The town of Luleå, with ~73,000 inhabitants, is situated at the mouth of the Lule River in

Norrbotten, Sweden. The river and former shallow bays of the brackish Bothnian Bay are the

most characteristic hydrodynamic patterns of Luleå. These bays, called innerfjärdar, are the

result of postglacial land uplift (8-9 mm/a), or the construction of artificial banks, often

partially enclosed (Lindén et al., 2006). Consequences of the ongoing land uplift are

diminishing water surfaces (and volumes) in the shallow bays. Increasing vegetation

accompanies silting-up processes in the Luleå innerfjärdar (Erixon, 1996). To preserve the

shallow bays for recreation, they were dammed up at their two connections with the Bothnian

Bay (Fig. 1). Also, the water level in the Lule River and Bothnian Bay affects the turnover

rate and water quality in the shallow bays (Erixon, 1996). Luleå’s innerfjärdar are situated in

and around the town of Luleå, and are affected by local catchments, which contain urbanized

4

and industrial areas as well as rural and forested areas. They are used for many recreational

purposes. Large parts of the catchment areas of the innerfjärdar are covered by sea bottom

sediments characterised as acid sulphate soils (Erixon, 2009).

The bay Skutviken, located close to the centre of Luleå, is enclosed by a road bank

constructed in 1962. Skutviken is still connected to the Lule River via a channel. Several

stormwater pipes discharge into the bay from a sewer drainage area with industrial and

housing areas (Rentz et al., 2011).

Hertsöfjärden is a bay that has been especially affected by the outlets of the steel plant

SSAB Tunnplåt AB (formerly Norrbottens Järnverk and SSAB) since the 1940s. Due to plans

to build a new steel plant, Stålverk 80, the outer part of the bay was infilled in 1975-76 and an

artificial bank divided the bay in two parts. The water in the inner part was dammed up

(Timner, 1994).

Figure 1. Water bodies in the Luleå area. S: sample point Skutviken; G: sample point ahead Gültzauudden; IS: sample point Inre Skurholmsfjärden; L: sample point Lövskataviken; B: sample pointBredviken; IH: sample point Inre Hertsöfjärden; D1-3: Watergates;

5

Lövskataviken and Inre Skurholmsfjärden are water bodies in the innerfjärdar system in

central Luleå. Industrial activities have taken place on their banks for more than 100 years

(Olofsson, 2002). Petrol stations have been located in the catchment area of Inre

Skurholmsfjärden since 1954. The urban catchment area contains industrial and housing areas

with parks. A road bank built in the 1960s separates the two water bodies, which are still

connected via road culverts (Olofsson, 2002).

The Lule River, with its 25,263 km2 large catchment area, has an annual average discharge

of 506 m3/s (SMHI 2010). The Lule River rises in the mountain area in the west, close to the

Norwegian border, where vegetation of tundra type occurs. Downstream, coniferous and birch

forest dominate, covering 58% of the total catchment area. Lakes and mires are also common,

accounting for 11% of the total catchment area. Since the beginning of the 20th century, the

river has been regulated and today there are 15 power stations along the river (Drugge, 2003).

6

Site

Sk

utvi

ken

Gül

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ken

Inre

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1994

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94)

Dru

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(200

3)

Met

als i

n se

dim

ent

Mn,

Fe,

S, C

d,

Cr,

Cu,

Ni,

Pb,

Zn

Mn,

Fe,

S, C

d,

Cr,

Cu,

Ni,

Pb,

Zn

Mn,

Fe,

S, C

d,

Cr,

Cu,

Ni,

Pb,

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Mn,

Fe,

S, C

d,

Cr,

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Ni,

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Zn

Mn,

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S, C

d,

Cr,

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Mn,

Fe,

S, C

d,

Cr,

Cu,

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--

16 P

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M

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C

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Pb, Z

n M

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, Cd,

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Met

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, M

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a,

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Sr,

Zn, A

s, C

d, C

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i, P,

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-- --

-- --

Mn,

Fe,

S,

K, M

g, N

a,

Si, A

l, B

a,

Co,

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Ca,

M

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,

Sedi

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7

Materials and methods

Sampling sites

The surface area of Skutviken is ~12 ha, and the mean and maximum depths of the bay are

1.6 m and 3.4 m, respectively. The bay is separated from the Lule River by a road bank

constructed in 1962, and is connected to the river via a single channel (8 m in width, 3 to 4 m

in depth, 35 m in length). These physical conditions make the bay similar to other shallow

bays in this region. The bay is almost completely enclosed by two heavily trafficked roads

with averages of 23,100 and 13,600 vehicles per day, respectively (Luleå Kommun, 2009).

The sewer drainage area contains 0.53 km2 industrial area and 0.73 km2 housing area. Since

surface runoff and six stormwater channels enter the bay, it almost functions as a large

stormwater pond where a large amount of stormwater sediment is trapped, resulting in a

reduced sediment supply to the Lule River. All channel outlets are located below the water

surface, except during periods of very low water level.

To compare sediment quality, a reference sampling site with less-affected conditions was

chosen, situated beside the main streambed of the Lule River in front of the spit Gültzauudden

(Fig. 1). The sites Hertsöfjärden (Timner, 1994), Lövskataviken & Inre Skurholmsfjärden

(Olofsson, 2002) and the Lule River (Drugge, 2003), described in previous studies, were

compared with the Skutviken and Gültzauudden sites.

The annual precipitation in the Luleå area is about 500 mm, of which 40 to 50% falls as

snow between November and April/May, and thus is discharged during snowmelt

(Hernebring, 1996). From November until May the Lule River and the bays close to the city

centre are ice-covered.

Previous studies in the Luleå area

Previous studies have examined geochemical characteristics (Table 1) of water bodies in

Luleå (Fig. 1). Sediment data from Skutviken and Gültzauudden were compared with data

from Timner (1994) and Olofsson (2002), who both took sediment samples with a Kajak-

corer and analysed metal contents in sediment and porewater. Timner (1994) found that

sediments can bind trace metals coming from the catchment area. Compared with the

sediment in the main basin of Inre Hertsöfjärden, the sediment in Bredviken showed a thinner

oxidized sediment top layer as a result of less water turnover and ice covering in winter. Parts

of the main basin stay ice-free even in wintertime because of the warm water outflow from

8

adjacent industry. The impact of the SSAB steel plant is noted for the main basin in terms of

increased concentrations of As, Cd, Co, Hg, Ni, Pb, V, Zn, Fe and Mn in the sediment

deposited after 1946. However, also in Bredviken, the values in the sediment top were

increased for As, Cd, Cr, Pb, V, Zn and Fe after 1946. A large part of the metal discharge

from SSAB, as calculated by Timner (1994), is accumulated in the sediment. Possible

secondary movements of Fe, Mn, Zn and other trace elements make it difficult to see

changing contamination levels in the sediments and to relate them to the time of

sedimentation. Olofsson (2002) showed enrichment of trace metals (As, Cd, Cr, Cu, Hg, Ni,

Pb, Zn) and PAHs in the sediments caused by stormwater impact from local industrial areas.

For the Lövskataviken sediment, Olofsson (2002) points out that stormwater supply from the

industrial areas in the west and south of the bay imports pollutants, as stormwater from rain

and melted snow on the road bank. From the road bank, stormwater even reaches Inre

Skurholmsfjärden, which is mostly affected by stormwater from an industrial area in the east

and an outflow from the housing area Skurholmen. The spreading conditions for the enhanced

contents of heavy metals and organic pollutants at Lövskataviken and Inre Skurholmsfjärden

are considered to be low, because of the relatively sheltered location of the bays and their low

water turnover rates. Owing to the fact that there are adjacent recreation areas, the risk of

spreading these pollutants was estimated as low under current conditions. Water data from

Skutviken and Gültzauudden were compared with Drugge’s (2004) datasets from the Luleå

River at Boden power station, ca 30 km upstream from Luleå, and with data from gully pot

catch basins published by Karlsson & Viklander 2008 (Table 2).

9

Tabl

e 2.

Ele

men

t con

cent

ratio

ns (µ

g/l)

in th

e fil

tere

d ph

ase

(<0.

22 µ

m) i

n th

e Lu

le R

iver

at B

oden

pow

er st

atio

n, G

ültz

auud

den,

Sku

tvik

en a

nd

from

3 c

atch

bas

in m

ixtu

res i

n Lu

leå

with

diff

eren

t ann

ual a

vera

ge d

aily

traf

fic (v

ehic

les/

day

(v/d

)) .

B

oden

pow

er st

atio

n(a)

Lule

å

Lul

e R

iver

G

ültz

auud

den

Sk

utvi

ken

Cat

ch b

asin

sf)

Elem

ent

Win

ter

Sum

mer

Sp

ring

-w

inte

rb)

Sum

mer

c)Sp

ring

-w

inte

rd)

Sum

mer

e)

"1"

"2

"

"3"

M

n 3

5.1

6.57

3.

4 32

4 1.

6 18

6 82

36

5 Fe

71

215

52.9

68

56

9 16

5 53

00

60

100

S64

7 56

5 85

0 15

70

2660

19

50

3200

28

00

2600

K

440

523

<400

73

5 22

50

977

6000

34

00

3900

M

g58

2 57

1 70

7 19

70

3190

17

10

2000

28

00

3300

N

a 11

39

890

1090

11

700

6680

68

50

6000

0 24

800

3360

0 Si

1285

13

19

1370

15

70

3100

60

1 -

- -

Al

8.7

18.6

5.

27

9.65

5.

49

6.67

73

8 42

24

Ba

6.4

5,3

6.49

5.

53

23.1

8.

14

38

48

73

Co

0.01

0.

01

0.02

0.

02

0.28

0.

01

1.2

0.5

2.3

Cu

0.67

0.

52

0.44

0.

41

0.36

0.

64

22

3.1

2.3

Ca

2785

24

10

3340

30

50

1560

0 60

40

1460

0 18

900

2200

0 M

o0.

29

0.24

0.

21

0.29

0.

3 0.

31

10

2.4

12

Sr10

.7

9.9

12.1

18

.7

53.7

24

.7

47

31

58

Zn

2.5

1.2

0.44

0.

45

3.42

0.

48

82

5.8

14

As

- -

0.21

<0

.4

0.47

0.

22

5.1

0.8

3 C

d-

- <0

.002

<0

.002

<0

.002

<0

.002

0.

2 0.

02

0.02

C

r -

- 0.

05

0.08

0.

11

0.05

4

0.2

0.1

Hg

- -

<0.0

02

<0.0

02

<0.0

02

<0.0

02

0.00

3 0.

003

0.00

3 N

i -

- 0.

26

0.18

0.

69

0.36

5.

5 1

3.9

P-

- 1.

17

1.1

3.11

4.

18

78

16

12

Pb-

- 0.

01

0.02

0.

04

0.19

70

0.

7 0.

3 a)

wee

kly

sam

plin

g fr

om D

rugg

e, (2

003)

, Sum

mer

: ave

rage

June

to A

ugus

t, W

inte

r: av

erag

e D

ecem

ber t

o M

arch

; b) s

ampl

ing

date

: 200

7-03

-07;

c) s

ampl

ing

date

: 20

07-0

7-09

; d) s

ampl

ing

date

: 200

7-03

-05;

e) s

ampl

ing

date

: 200

7-07

-04;

f) d

ata

from

Kar

lsso

n et

al.,

(200

9), "

1" (5

00 v

/d),

"2" (

13,8

00 v

/d),

"3" (

25,5

00 v

/d)

10

Water sampling at Skutviken and Gültzauudden

The surface water was sampled 50 cm below the surface and 50 cm below the ice in winter,

respectively. Water was pumped by a peristaltic pump (Masterflex® L/S®) into a 25-litre

polyethylene (PE) container. Membrane filtration (0.22 μm pore size, 142 mm diameter,

Millipore® mixed cellulose esters) was carried out in a laboratory within 6 hours of sampling.

During filtration each filter was used only until half of its filtration capacity. This was done to

decimate discrimination of colloids that is caused by clogging of filters (Morrison and Benoit,

2001). The filtrate was collected in a 25-litre PE container from which subsamples were

taken for analyses. Subsamples were collected in 60-ml acid-washed polyethylene bottles and

refrigerated until further analysis. All used tubing and containers were pre-cleaned with 5%

HCl and rinsed with MQ water (Millipore, 18.2 Mohm).

Sediment and porewater sampling at Skutviken and Gültzauudden

The sampling station in Skutviken was located in the deeper parts of the bay with fine-

grained sediment. At Skutviken the water depth was 2.2 m and at Gültzauudden it was 6.1 m.

Sediment samples from Skutviken and Gültzauudden were taken in March 2007 from the ice

using a Kajak gravity corer with a core tube diameter of 64 mm. The sediment core surfaces

were judged to be undisturbed. Cores were sectioned in subsamples (0.5 cm thick for the

uppermost 3 centimetres and 1 cm thick for the remainder of the core).

For porewater analyses the sediment samples were put into plastic bags directly after

sectioning. All air was pressed out of the bag before it was placed in an Ar-filled container to

keep the sediments in an oxygen-free environment until the porewater was extracted within

the following eight hours. The porewater was separated by vacuum filtration (0.22 μm

Millipore® membrane filters) arranged in an Ar-flushed glove box. The porewater samples

were collected in 60-ml acid-washed polyethylene bottles and refrigerated until further

analysis.

Near-bottom water was sampled inside the Kajak-corer tube 3 cm above the sediment surface.

The water was drawn with a small plastic tube fixed on a syringe and filtered through a 0.22

�m Millipore® membrane filter.

11

Analyses

The 0.22 μm membrane filtered surface-water samples were analyzed for major and trace

elements by inductively coupled plasma atomic emission spectrometry (ICP-AES) and

inductively coupled plasma sector field mass spectrometry (ICP-SFMS). For instrument

operation details, see Rodushkin and Ruth (1997).

The sediment was analyzed for loss on ignition (LOI) and together with porewater for major

elements and trace metals. Sediment samples for determination of As, Cd, Co, Hg, Ni, Pb and

S were dried at 50°C digested in a microwave oven in closed Teflon bowls with a nitric acid :

water ratio of 1:1. For other elements, 0.125 g dried matter (DM) was melted with 0.375 g

LiBO2 and dissolved in HNO3. Metal determinations were made by ICP-AES and inductively

coupled plasma mass spectrometry (ICP-MS). To the porewater samples, 1 ml nitric acid

(suprapur) was added per 100 ml sample water. Analyses were done with ICP-AES and ICP-

SFMS. The following 16 PAHs were analyzed in the sediment: Naphthalene (NAP),

Acenaphthylene (ACY), Acenaphthene (ACE), Fluorene (FL), Phenanthrene (PHEN),

Anthracene (ANT), Fluoranthene (FLR), Pyrene (PYR), Benzo(a)anthracene (BaA), Chrysene

(CHY), Benzo(b)fluoranthene (BbF), Benzo(k)fluoranthene (BkF), Benzo(a)pyrene (BaP),

Dibenz(a,h)anthracene (DBA), Benzo(ghi)perylene (BPY) and Indeno(1,2,3-cd)pyrene (INP).

The PAH sediment samples were leached with acetone : hexan : cyclohexan (1:2:2) and

measurements were done with gas chromatography mass spectrometry (GC-MS).

Results and discussion

Water column

Dissolved oxygen concentration in the water column at Skutviken varies from summer to

winter. The oxygen saturation in the bottom water is close to 0% in wintertime, when the bay

is ice-covered. In contrast, the water column is well oxygenated (saturation 85-90%) during

the ice-free season (Rentz et al., 2011).

The elemental concentrations of the dissolved phase (<0.22 μm) at the 3 sites (Skutviken,

Gültzauudden and the Boden power station on the Lule River) show seasonal and spatial

variations. Seasonal variations in the element concentrations are less distinct in the regulated

Lule River as compared with the pristine Kalix River (Drugge, 2003). It is evident that the

water at Gültzauudden is much like the Lule River water (Table 2). In contrast, element

concentrations at Skutviken show stronger seasonal variations. Late-winter concentrations of

12

K, Mg, As, Cr, Ni and Sr are twice as high as in summer. The concentrations of Ca are 2.6, Fe

3.4, Si 5, Zn 7, Co 33 and Mn 200 times higher in late-winter than in summer. The

concentrations of Na, S, Al, Mo and P do not show much variation. The late-winter

concentration of Cu is just half the summer concentration, and for Pb a fifth. Furthermore,

Skutviken is characterized by high concentrations of Ca, Fe, K, Na, Co compared with the

other sites, especially in late-winter.

Redox conditions and LOI in sediments

The sediment core at Gültzauudden shows the typical concentration profile of freshwater

sediments for Mn, Fe and S (Fig. 2, 3 & 4). Oxic conditions in the top of the sediment core

imply occurrence of Mn oxyhydroxides (Song and Müller 1999). Decomposition of organic

material leads to increasing anoxic conditions with depth, and results in reduction of Mn

oxyhydroxides and increased porewater concentration of Mn(II). Anoxic conditions

predominate below 4 cm, where the MnO content stabilises at 0.2%, probably occurring in

silicate minerals. From that point the Mn concentration increases in porewater. This indicates

Mn(II) flux upward, resulting in the oxidation of Mn(II) to Mn(IV) in the oxic parts of the

sediment (Davison, 1993; Wehrli, 1991) (Fig. 2). The Mn observations comply with the

sediment content and porewater concentration of Fe at Gültzauudden. A Fe2O3 peak in the

sediment profile is situated below the peak of MnO. The solid Fe2O3 peak at a depth of 3 cm

depends on the oxic sediment conditions, where Fe occurs as Fe(III) in iron oxyhydroxides.

Below 5 cm the solid Fe content declines continuously. When porewater becomes more

anoxic with depth the Fe concentration indicates that reduction of solid Fe(III) to the soluble

Fe(II) occurs (Davison, 1993; Wehrli, 1991) (Fig. 3).

At Skutviken the MnO content in the sediment is much lower than at Gültzauudden in the

upper parts of the sediment. It appears that the geochemical conditions where Mn(IV) is

reduced to Mn(II) are reached already in the near-bottom water above the sediment surface.

During winter, when the bay is ice-covered, the oxygen concentration in the bottom water is

<0.42 mg/l (Rentz et al., 2011). The Mn concentrations in the porewater increase with depth,

but never reach as high concentrations as at Gültzauudden.

13

0 0.5 1 1.5 2MnO (%DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 4000 8000 12000Mn (µg/l)

0 0.5 1 1.5 2MnO (%DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 0.5 1 1.5 2MnO (%DM)

-40

-30

-20

-10

0

Dep

th (c

m)

Skutviken Gültzauudden

Lövskataviken Skurholmsfjärden

Inre HertsöfjärdenBredviken

0 0.5 1 1.5 2MnO (%DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 4000 8000 12000Mn (µg/l)

0 0.5 1 1.5 2MnO (%DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 4000 8000 12000Mn (µg/l)

0 0.5 1 1.5 2MnO (%DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 4000 8000 12000Mn (µg/l)

Mn (µg/l)

MnO (%DM)

Figure 2. MnO in sediment (%DM) and Mn in porewater (µg/l) at Skutviken, Gültzauudden, Lövskataviken, Skurholmsfjärden, Inre Hertsöfjärden and Bredviken. The top value for “porewater” represents the bottom near surface water at Skutviken, Gültzauudden and Lövskataviken. Porewater was not analyzed at Skurholmsfjärden and Bredviken.

14

5 10 15 20 25 30Fe2O3 (%DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 10 20 30 40Fe (µg/l)

5 10 15 20 25 30Fe2O3 (%DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 10 20 30 40Fe (µg/l)

5 10 15 20 25 30Fe2O3 (%DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 10 20 30Fe2O3 (%DM)

-40

-30

-20

-10

0

Dep

th (c

m)

SkutvikenGültzauudden

Lövskataviken

5 10 15 20 25 30Fe2O3 (%DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 10 20 30 40Fe (µg/l)

5 10 15 20 25 30Fe2O3 (%DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 10 20 30 40Fe (µg/l)

Skurholmsfjärden

Inre HertsöfjärdenBredviken

Fe (µg/l)

Fe2O3 (%DM)

Figure 3. Fe2O3 in sediment (%DM) and Fe in porewater (µg/l) at Skutviken, Gültzauudden, Lövskataviken, Skurholmsfjärden, Inre Hertsöfjärden and Bredviken. The top value for “porewater” represents the bottom near surface water at Skutviken, Gültzauudden and Lövskataviken. Porewater was not analyzed at Skurholmsfjärden and Bredviken.

15

0 2000 4000 6000 8000 10000S (mg/kg DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 2 4 6 8 10S (mg/l)

0 2000 4000 6000 8000 10000S (mg/kg DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 20 40 60 80 100S (mg/l)

0 2000 4000 6000 8000 10000S (mg/kg DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 2000 4000 6000 8000 10000S (mg/kg DM)

-40

-30

-20

-10

0

Dep

th (c

m)

Skutviken Gültzauudden

Lövskataviken

0 2000 4000 6000 8000 10000S (mg/kg DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 2 4 6 8 10S (mg/l)

0 2000 4000 6000 8000 10000S (mg/kg DM)

-40

-30

-20

-10

0

Dep

th (c

m)

0 2 4 6 8 10S (mg/l)

Skurholmsfjärden

Inre HertsöfjärdenBredviken

S in mg/l

S (mg/kg DM)

Figure 4. S in sediment (mg/kg DM) and S in porewater (mg/l) at Skutviken, Gültzauudden, Lövskataviken, Skurholmsfjärden, Inre Hertsöfjärden and Bredviken. The top value for “porewater” represents the bottom near surface water at Skutviken, Gültzauudden and Lövskataviken. Porewater was not analyzed at Skurholmsfjärden and Bredviken.

16

The redox conditions at Skutviken permit precipitation of Fe(III) hydroxides at the

sediment-water interface, indicated by the presence of a solid Fe2O3 maximum at the sediment

surface. The anoxic conditions occurring at Gültzauudden at a sediment depth of 3 cm seem

in winter to occur already above the sediment column at Skutviken. Therefore, reductive

dissolution of Fe hydroxides takes place already at the sediment surface. The decrease in total

S in porewater at Skutviken suggests that reduction of SO42- occurs immediately below the

sediment-water interface (0-2 cm). Precipitation of solid sulphides in the sediment is indicated

by the simultaneous increase in solid S (Fig. 4).

The sediment profiles for solid Mn at Lövskataviken, Skurholmsfjärden and Bredviken

resemble the characteristics at Skutviken with constant low concentrations of MnO over the

whole depth. Only at Inre Hertsöfjärden, does an increase of MnO in the uppermost 5 cm in

the sediment indicate more oxic conditions in the sediment top. A high concentration of solid

Fe(III) already at the sediment surface at Inre Hertsöfjärden suggests that the oxic conditions

are low compared with Gültzauudden but higher than at the other sites.

A solid Fe2O3 maximum in the sediment top is common for Lövskataviken,

Skurholmsfjärden, Inre Hertsöfjärden and Bredviken and corresponds to Skutviken. However,

the very high Fe2O3 concentrations in the uppermost 5 cm at Lövskataviken, Skurholmsfjärden

and Inre Hertsöfjärden are notable. The change from high concentrations to low background

concentrations is abrupt at these sites.

At Lövskataviken the S concentration in the sediment and porewater indicates similar

conditions as at Skutviken. Here, the content of S in the porewater decreases and the solid S in

the sediment increases in the same sediment layer where iron is enriched. Visible are, in

particular at Lövskataviken, Inre Hertsöfjärden and Bredviken, increasing concentrations of

solid S at sediment depths below 15 cm, simultaneously with apparent unchanged low S

concentration in the porewater.

The LOI content at all sites is consistently highest in the uppermost section of the sediment

columns (Fig. 5). After a thin layer with constant, relatively high LOI content, the values

decline radically at Skutviken, Inre Skurholmsfjärden, Inre Hertsöfjärden and Bredviken. At

Lövskataviken the LOI content is relatively low already in the top layer compared with the

other sites, and shows no strong decrease with depth. At Gültzauudden the LOI content

declines directly from the sediment top, which suggests a minor input of organic material but

also relatively high oxygen concentration that helped to decompose organic material. That can

be ascribed to continuous circulation in the water column due to the nearby main streambed

of the Lule River. Also, at Inre Hertsöfjärden the sediment top seems to be more oxic than at

17

the other sites. The surface water in this bay does not always freeze, since warm water enters

the bay via outlets from the nearby industry. The water surface at Lövskataviken, Inre

Skurholmsfjärden and Bredviken do freeze regularly, as does Skutviken. Below the ice cover

the oxygen is consumed as a result of decomposition of organic material. Anoxic conditions

slow down further decomposition. The content of organic material in the sediment of the

shallow bays shows a high input of organic components from the surrounding catchment

areas, as indicated by LOI (Fig. 5). The low water turnover rate during wintertime at these

sites excludes the inflow of fresh oxygenated water, which can hamper decomposition.

0 4 8 122 6 10LOI (%DM)

-40

-30

-20

-10

0

-35

-25

-15

-5

Dep

th (c

m)

Skutviken

Gültzauudden

LövskatavikenInre Skurholmsfjärden

Inre HertsöfjärdenBredviken

Metal concentrations in sediments

The average concentrations of the selected metals Cd, Cr, Cu, Ni, Pb, Zn in the upper

sediment section (0-4 cm) generally exceed the average concentrations of the deeper section

from the same core (Table 3). Only Gültzauudden deviates from the others, since it has the

Figure 5. Loss on ignition (LOI) versus sediment depth at Skutviken, Gültzauudden, Lövskataviken, Inre Skurholmsfjärden, Inre Hertsöfjärden and Bredviken.

18

lowest metal concentrations in comparison with the same depth sections at the other sites. Of

all sites, Inre Hertsöfjärden exhibits the highest concentrations of all metals except for Ni.

Inre Hertsöfjärden shows pollution concentrations for Cd, Ni, and Pb in the range of

sediments in central Stockholm (Sternbeck et al., 2003), the Cr and Zn concentrations exceed

the averages from central Stockholm by more than three times, and Cu is half the Stockholm

average. The Ni concentrations at Inre Skurholmsfjärden and Lövskataviken exceed the

Stockholm average slightly. Inre Hertsöfjärden is exposed to spill water from a steel plant. To

a minor degree, besides urban stormwater, Bredviken is exposed to the same spill water. This

can explain the higher Cr concentrations than at Skutviken, Skurholmsfjärden and

Lövskataviken. The concentrations at Inre Skurholmsfjärden and Lövskataviken resemble

those at Skutviken for Cd and Cu, with less than half the concentration of the Stockholm

average. The Cr concentrations are in the range of the central Stockholm sediment. For all

sites in Table 3, except Gültzauudden, the catchment areas exhibit possible sources for the

enrichment of metals in the sediment. Catch basin mixtures from gully pots in Luleå showed

high concentrations of Cd, Cr, Cu, Ni, Pb and Zn, which suggests that traffic and urban

stormwater are probable sources at Skutviken, Inre Skurholmsfjärden and Lövskataviken

(Karlsson et al., 2009) (Table 2).

Element Depth in cm Skutviken Gültzauudden Lövskataviken

Inre Skurholmsfjärden

Inre Hertsöfjärden Bredviken

0-4 0.7 0.3 0.6 0.8 2 0.7 Cd 4- * 0.4 0.3 0.4 0.3 0.8 0.7 0-4 83 68 80 80 319 122 Cr 4- * 87 67 66 62 98 78 0-4 60 17 56 68 92 37 Cu 4- * 37 24 30 30 41 33 0-4 23 19 47 46 44 34 Ni 4- * 22 19 24 25 24 31 0-4 66 13 39 55 236 69 Pb 4- * 47 26 28 20 101 64 0-4 284 97 302 357 1733 343 Zn 4- * 180 118 166 127 392 283

*core end: Skutviken & Gültzauudden 21 cm, Lövskataviken 30 cm, Inre Skurholmsfjärden 38 cm, Inre Hertsöfjärden 22.5 cm, Bredviken 24.5 cm

Table 3. Average element concentration (mg/kg DM) in sediment sections 0-4 cm and 4 cm to core end at Skutviken, Gültzauudden, Lövskataviken, Inre Skurholmsfjärden, Inre Hertsöfjärden and Bredviken.

19

PAH in sediments

The high PAH concentrations in the sediment top (Table 4) suggest that the PAH

enrichment is generated from sources in the catchment areas of Skutviken, Inre

Skurholmsfjärden and Lövskataviken. The concentrations at Inre Skurholmsfjärden exceed

those of the other sites, and the sediment at Gültzauudden contains the lowest concentrations

for each PAH. The comparison with the Swedish EPA classification (Swedish EPA, 1999) for

organic pollutants shows clearly increased concentrations at Inre Skurholmsfjärden, where the

light PAHs PHEN, PYR, BaA and CHY reach Class 5, the highest of five contamination

classes. At Skutviken only PYR reaches Class 5. However, the total PAH contamination (all

11 PAHs) at Inre Skurholmsfjärden and Skutviken reaches Class 4, the second highest

contamination class, while at Lövskataviken total PAH concentrations reach Class 3. Even if

the high concentrations of PHEN, PYR, BaA and CHY at Skurholmsfjärden and PYR at

Skutviken reach Class 5, the concentrations are distinctly lower than the average from 7

sampling stations in central Stockholm (Sternbeck et al., 2003), PHEN 3.5, PYR 3.8, BaA

4.5, and CHY 1.4 times lower.

Skutviken Gültzauudden Lövskataviken Inre Skurholmsfjärden

n=1 n=1 n=3 n=3 depth in cm 0-2 0-2 0-4 0-4 ^PHEN 89**** 22*** <80 137***** ^ANT 24**** <10** <80 <80 ^FLR 130**** 28*** 102**** 217**** ^PYR 240***** 20*** 106**** 240***** ^BaA 59**** 13*** <80 120***** ^CHY 69**** <10 116**** 363***** ^^BbF 180**** 10** 32** 193**** ^^BkF 44*** <10 28** 147**** ^^BaP 74**** <10 <80 123**** ^^BPY 89*** <10 <80 180**** ^^INP 99*** <10 <80 70*** � 11 PAH 1097**** 93** 384*** 1790**** ^^DBA 30 <10 <80 <80 ^NAP 39 <10 <80 <80 ^ACY 11 <10 <80 <80 ^ACE <10 <10 <80 <80 ^FL 15 <10 <80 <80 � 16 PAH 1200 93 384 1790

Table 4. Concentrations (µg/kg DM) of 16 PAHs in the sediment in 0-2 cm depth at Skutviken and Gültzauudden and 0-4 cm depth at Lövskataviken and Inre Skurholmsfjärden. Eleven PAHs are included by the Swedish EPA guidelines. ^light PAH, ^^heavy PAH. Swedish EPA guidelines for 11 PAHs: class 1, no content *; class 2, low content **; class 3, moderately high ***; class 4, high ****; class 5, very high *****.

20

Future scenarios

The future risk of enriched metal pollutants in the sediments at present is conditional on

whether they can be retained in the sediments or mobilized. Mobile dissolved pollutants are

made available for uptake by living organisms (Munch Christensen et al., 2006). If the

geochemical processes in shallow bays in the Luleå area lead to fixation of metals in anoxic

sediments, metal mobility may be impeded as long as these sediments do not become

oxygenated. However, anoxic conditions limit PAH degradation due to the fact that biological

activity is hampered. Only 3-ring PAHs where found to become degraded under anoxic

conditions (Quantin et al., 2005). Conditions that benefit the decomposition of PAHs will

cause higher risk of secondary release of metal pollutants. Present land uplift (Lindén et al.,

2006) can implicate future drainage of the buried sediments, which today accumulate on the

bottom of coast-near narrow bays. If the submerged soils become oxidized when they are no

longer water-covered, release of trapped pollutants occurs. Metal release from sulphate soils

of local catchments has led to temporally decreasing water quality (Erixon 2009). Several

studies from Finnish areas, concerning sulphate soils and metal release (Boman et al., 2008;

Österholm and Aström, 2008; Åström, 1998), indicate the need for investigation of related

risks. In postglacial land uplift areas, ditching of sulphate soils and seasonal variations in

precipitation can imply changes of redox conditions in the soil profile (Österholm and

Åström, 2008; Erixon 2009). Human impact on the water levels, such as damming up the

partially enclosed bays, can slow down the long-term processes which result in oxidation of

soils and further transport of pollutants. For the year 2004, with low precipitation causing

extremely low groundwater levels, Erixon (2009) calculated mass transport of Zn, Ni, Co and

Mn from catchment areas with sulphate soils in the Luleå area. The catchment areas of

Holmsundet (60 km2) and Persöfjärden (402 km2), with a runoff 0.3 x 108 m3/a and 1 x 108

m3/a, respectively, can release 8.5, 3.5, 2.4, 396 tons Zn, Ni, Co and Mn per year. This is

more than from the Kalix River (Zn 8.5, Ni 3.2, Co 0.6, Mn 194 t/a) with 23,600 km2

catchment area and a 100*108 m3/a runoff. The catchment areas of Holmsundet and

Persöfjärden contain 50% and 20%, respectively, marine and lacustrine clay sediments.

Besides urban stormwater, sulphate soils also have to be considered as an influential factor for

disturbance of local water bodies.

21

Conclusions

The investigated water bodies in the Luleå area show clear urban impact on sediment

quality. The metals Cd, Cu, Pb and Zn, which are of main concern in urban stormwater, are

enriched in all investigated bays. Metals can bind to surfaces of sedimenting organic and

small inorganic particles. In the sediment they can become part of sulphide formation and are

thus fixed in the sediment.

In Skutviken, which is an efficient trap for particulate stormwater pollution, the dissolved

and particulate pollutants may be enriched and more concentrated. Concentrations are in

general higher during wintertime, which may be due to the reduced inflow of fresh river water

and lack of surface runoff, whereby the water turnover in the bay is reduced. The same

principle applies for Lövskataviken, Inre Skurholmsfjärden and Bredviken due to their

sheltered position. Lövskataviken and Inre Skurholmsfjärden receive stormwater from nearby

industrial areas. Inre Skurholmsfjärden is especially contaminated by PAHs, probably from

leakage from a former nearby petrol station. For Inre Hertsöfjärden, the impact of the water

inflow from the steel plant contributes to the more oxygenated sediment conditions because

the warm water prevents the bay from freezing during winter. Even here, pollutant transport

to the sediment is a result of the water inflow from the industrial area. All sediment samples

comprised mainly particles of the silt and clay fractions, which offer good conditions for

bonding on particle surfaces. The high LOI values could be caused by a combination of

organic pollutants and natural organic matter. At Skutviken, Bredviken, Lövskataviken and

Inre Skurholmsfjärden, decomposition of natural and anthropogenic organic material

consumes the oxygen and causes reduced conditions in the bottom-near water and the

sediment during winter.

Water and sediment quality in the Luleå area are dependent on catchment area (size, natural

premises and exploitation) and emission impact, especially from point sources. Important

factors are water volume and turnover rate in the water bodies with low water levels and no

surface runoff wintertime, and ice covering during winter, which also contributes to anoxic

conditions in water column and sediment. The redox status in the sediments is crucial for

release or bonding of pollutants in the sediments. The bays do have the capacity to retain

pollutants in their sediment, but there is still a potential risk of release if the redox conditions

change. Postglacial land uplift implies continuous changes in the environment, which can lead

to changing redox conditions. This will necessitate new risk assessments.

22

Acknowledgements

This study was financed by Luleå University of Technology and the Swedish Research

Council for Environment, Agriculture Sciences and Spatial Planning. This support is

gratefully acknowledged. Metal, LOI and PAH analyses were performed by the accredited

laboratory ALS Scandinavia AB in Luleå. We would like to thank Kristin Karlsson, Fredrik

Nordblad and Magnus Westerstrand for assistance during the field work and contributing their

knowledge in discussions.

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25

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Paper III

Impact of urban stormwater on water quality in an enclosed bay of the Lule River, northern Sweden

R. Rentz1*, F. Nordblad1, B. Öhlander1

1 Division of Geosciences and Environmental Engineering, Department of Civil, Environmental and Natural Resources Engineering, Luleå University of Technology, 97187

Luleå, SWEDEN

* Corresponding author: ralren@ltu.se

Abstract Membrane- and ultafiltration were used to determine different speciation (truly dissolved

phase <1kDa; colloidal fraction >1 kDa and <0.22 μm) of element concentrations in surface

water samples from Skutviken, an enclosed bay with stormwater impact, and a reference site

in the Lule River. Sampling was conducted in winter and summer. Skutviken shows higher

unfiltered concentrations for Mn, Fe, S, Co, Cr, K, Ni and Zn in winter, and higher than the

reference site but still lower than catch basins or constructed stormwater ponds. Except for Fe,

these elements were mostly dissolved in winter. The winter conditions at Skutviken can

enhance the fraction of dissolved Mn and other metals in the bay when oxygen in the water

column is depleted under an ice cover. Stormwater is the probable source for elevated metal

concentrations.

Keywords Ultrafiltration; speciation; stormwater; sesonal variation

IntroductionWater bodies in urban areas fulfil diverse functions (Hauer and Lamberti 2006; Simmons

1991), and are exposed to emissions from manifold sources. Stormwater represents an

important contamination source of heavy metals in urban areas (Charlesworth and Lees 1999;

Schiff and Bay 2003). Metals are commonly investigated pollutants in stormwater, and they

may pose potential risks for living organisms (Munch Christensen et al. 2006; Wildi et al.

2004). Because stormwater often, like in Luleå, reaches receiving waters without passing

through any treatment facility, discharged stormwater can have a great impact on water bodies

and groundwater resources as well as on sediments and soils.

Previous studies on stormwater and sediments in gully pots and stormwater receiving water

bodies in the Luleå area (Karlsson and Viklander 2008; Rentz et al. 2010; Westerlund 2007),

indicated particle-related transport of metal and organic pollutants with seasonal variations.

Since pollutants, in dissolved form and associated with small suspended particles, are not

retained effectively in their transport chain from urban surfaces to receiving waters, these

pollutants reach recipients and can become enriched in sediments in the end. Larger particle

sizes have also been found to transport high metal contents (Brown 2002; Gonzalez et al.

2000). Depending on geochemical conditions, pollutants can be buried in sediments or will be

released from sediments.

Sediment samples and associated porewater in an enclosed bay (Skutviken) affected by storm

water discharge near the centre of Luleå, northern Sweden, were analyzed for major and trace

elements and 16 polycyclic aromatic hydrocarbons (PAHs), and compared to a reference site

(Rentz et al. 2010). Among the studied metals Cd, Cu, Pb and Zn were particularly enriched

in the sediment at Skutviken. Also the PAH content was enriched, in particular phenantrene,

anthracene, fluoranthene and pyrene, which are common constituents in storm water (Brown,

2002, Gonzalez et al., 2000). Skutviken was enclosed trough the construction of a road bank

since 1962, and has after that functioned as an efficient trap for urban pollutants. The

enclosure led to decreased water circulation in the bay, which promoted the occurrence of

anoxic conditions with sulphate reduction below the ice during winter. As a consequence of

these conditions, metals are trapped in the sediments as sulphides. Seasonally varying redox

conditions result in variations in trace element concentrations in the porewater and the

overlying bottom water. It is, therefore, of interest to study interactions of contaminated

sediments and the overlaying water body.

To study the water quality in an enclosed bay with polluted sediments, the water columns at

Skutviken and at a reference site were sampled at 0.5 m depth and filtered both by membrane

filtration (<0.22 μm) and ultrafiltration (“truly dissolved fraction”, <1 kD). The major aims

were to determine and quantify the degree of pollution as a result of storm water impact or

release of contaminants from the polluted sediments to the water column. Winter and summer

sampling should enable to evaluate if seasonal variation exist. We applied

membrane/ultrafiltration to determine if the contaminants occur in suspended particles, in

colloids or in the dissolved fraction (suspended particles >0.22 μm; colloidal fraction <0.22

μm and >1 kDa; truly dissolved phase <1kDa).

Materials and methods

Sampling sites The surface area of Skutviken is ~12 ha, and the mean and maximum depths of the bay are 1.6

m and 3.4 m, respectively. The bay is mainly separated from the Lule River by a road bank

constructed in 1962, and is connected to the river via a single channel (8 m in width, 3 to 4 m

in depth, 35 m in length). These physical conditions make the bay similar to other shallow

bays in this region. The sampling station in Skutviken was chosen in the deeper parts of the

bay with fine grained sediment. The bay is surrounded by the road bank and another highly

frequented road with traffic intensities of 22900 and 13600 vehicles per day, respectively

(Luleå Kommun 2010). The sewer drainage area contains 0.53 km2 industrial area and 0.73

km2 housing area. Since surface runoff and six stormwater channels enter the bay, it almost

functions as a large stormwater pond where a large amount of stormwater sediment is trapped,

resulting in a reduced sediment supply to the Lule River. All channel outlets are located

below the water surface, except during periods of very low water level.

To compare water quality, a reference sampling site with less affected conditions was chosen,

situated beside the main streambed of the Lule River in front of the spit Gültzauudden (Figure

1).

The annual precipitation in the Luleå area is about 500 mm (SMHI 2009a) of which 35 to 40

% falls as snow between November and April/May (SMHI 2009b). Thus, relatively high

amounts of surface runoff are discharged during snowmelt mainly in April and May. From

November until May the Lule River and the bays close to the city centre are ice-covered.

Sampling The surface water was sampled in May and July 2007 50 cm below the surface and 50 cm

below the ice underside, respectively. Water was pumped by a peristaltic pump (Masterflex®

L/S®) through a polyethylene (PE) tube into 25-litre PE containers. Membrane filtration (0.22

μm pore size, 142 mm diameter, Millipore® mixed cellulose esters) was carried out in a

laboratory within 6 hours of sampling. The first filter was used until it was clogged

completely; the filtered volume was measured and then discarded. For the actual sample, new

filters were used, through which half the clogging volume was allowed to pass. This was done

to decimate discrimination of colloids that is caused by clogging of filters (M. A. Morrison

and Benoit 2001). The filtrate was collected in a 25-litre PE container from which

subsamples were taken for analyses. Subsamples were collected in 60 ml acid-washed PE

bottles and refrigerated until further analysis. The membrane filtered water was then

ultrafiltrated in a Millipore® Prep/Scale system. The filter had a manufacturer specified cutoff

of 1 kDa and a filter membrane area of 0.54 m2. The filter material was regenerated cellulose.

The system was connected with a Watson Marlow peristaltic pump. All used tubing and

Figure 1. Location of the study area Skutviken (A) and the reference sample site at Gültzauudden (B) in Luleå, Northern Sweden and the stormwater sewer catchment area at Skutviken.

containers were acid-cleaned in 5% HCl with subsequent wash in MilliQ water (Millipore,

18.2 M�) prior sampling.

Determination of colloidal and truly dissolved phase

For determination of the size distribution of components in natural water samples

ultrafiltration is used as an applicable technique. Ultrafiltration is often applied for studies of

the colloidal and truly dissolved species of metals and organic matter in natural waters

(Guéguen and Dominik 2003; Ingri et al. 2004). Low-abundance species (e.g. colloidal

concentrations) can be determined more precisely with help of the retentate where species

concentrations are enriched. Ultrafiltration techniques have previously been described and

evaluated by several workers (Guéguen et al. 2002; Wilding et al. 2004). Two critical aspects

when applying the method for natural water samples are the mass balance recovery and the

accuracy of determination of the species concentrations in the retentate. To achieve mass

balance recoveries close to 100 %, Larsson et al. (2002) found that a cross-flow ratio above

15 was necessary. The cross-flow ratio CFR is defined as:

perm

ret

QQCFR �

Qret and Qperm denote the retentate- and the permeate flow rate, respectively. For accurate

determination of the colloidal species, it was also found that an enrichment factor (total feed

water volume : final retentate volume) larger than 10 was required. The enrichment factor EF

and the colloidal concentration Ccoll can be calculated using:

ret

retperm

VVV

EF�

�

EFCC

C permretcoll

��

Where Vperm, Vret denote the volumes of the permeate and the retentate. Cperm, Cret and Cfeed

denote the concentrations of the permeate, the retentate and the feed sample, respectively.

Finally, the mass balance recovery R in percent units may be determined as:

feed

retperm

CCC

R�

�

The truly dissolved phase constitutes the fraction <1kDa and the colloidal fraction contains

particles >1 kDa and <0.22 μm.

AnalysesDissolved oxygen in the water column was determined with a Hydrolab® MiniSonde 5 water

quality probe. Unfiltered water samples, 0.22 μm and 1kDa filtrate were analyzed for major

and trace elements in inductively coupled plasma atomic emission spectroscopy (ICP-AES)

and inductively coupled plasma sector field mass spectrometry (ICP-SFMS). For instrument

operation details, see (Rodushkin and Ruth 1997).

ResultsThe elemental concentrations of the dissolved phase (<0.22 μm) at Skutviken, Gültzauudden

and the Boden power station show seasonal and spatial variations. The water at Gültzauudden

resembles the Lule River water (Table 1). In contrast, element concentrations at Skutviken

show stronger seasonal variations. Late-winter concentrations (<0.22 μm) of K, Mg, As, Cr,

Ni and Sr are twice as high as in summer. The concentrations of Ca are 2.6, Fe 3.4, Si 5, Zn 7,

Co 33 and Mn 200 times higher in late-winter than in summer. The concentrations of Na, S,

Al, Mo and P do not show much variation. The late-winter concentration of Cu is just half the

summer concentration, and for Pb a fifth. Furthermore, Skutviken is characterized by high

concentrations (<0.22 μm) of Ca, Fe, K, Na, Co compared with Gültzauudden, especially in

late-winter. The catch basins show clearly highest concentrations for nearly all elements. Just

at Skutviken especially in winter the concentrations of Fe and Mn can exceed catch basin

concentrations.

Comparison of trace metal concentration in Skutviken with concentrations in catch basin

water offers the picture of similar <0.22 μm concentrations in Skutviken in winter of Mn

(except catch basin 3), Fe, S, K, Mg, Na, and Ca (Table 1). Except for Fe, these are elements

found in the truly dissolved fraction in Skutviken. The concentrations of Al, Ba, Co, Cu, are

clearly higher in the catch basin water. In Skutviken, Co and Cu are mainly found in the

particulate phase. The speciation of analysed elements in the water column is shown in figures

2-5. Fe shows seasonal variation at Skutviken with ca 6 times higher unfiltered concentration

in winter than in summer (Figure 2). This variation applies to the large concentration Fe in the

suspended phase, which is found in winter. Skutviken shows higher concentration of Fe in all

fractions than Gültzauudden at both sampling occasions.

Tabl

e 1.

Ele

men

t con

cent

ratio

ns (µ

g/l)

in th

e fil

tere

d ph

ase

(<0.

22 µ

m) i

n th

e Lu

le R

iver

at B

oden

pow

er st

atio

n, G

ültz

auud

den,

Sku

tvik

en a

nd

from

3 c

atch

bas

in m

ixtu

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Also Mn variations at Skutviken from winter to summer is large, indicated by 1000 time’s

higher truly dissolved concentration of Mn in winter, when the Mn concentration consists

mainly of the truly dissolved phase (Figure 2). However, in summer the solid phase has 14

times higher concentration than the <0.22 filtered phase. The 10 times lower unfiltered Mn

content at Skutviken in summer consists mostly of Mn in the particlulate phase. Gültzauudden

shows in winter 1/51 and summer 1/27 of unfiltered Mn concentration as there is in

wintertime at Skutviken.

Figure 2. Speciation of Fe, Mn and S at Skutviken and Gültzauudden.

0

1

2

3

Fe (m

g/l)

unfiltered

<0.22 µm

colloidal

<1kDa

0

100

200

300

400

Mn

(µg/

l)

0

1

2

3

S (m

g/l)

SUMMERWINTERSUMMERWINTER

SUMMERWINTER

The S concentration at Skutviken shows in winter a high content of truly dissolved S and is

about double the total S concentration at Gültzauudden (Figure 2). In summer both sites have

more similar, lower, concentrations with mainly truly dissolved S.

The concentrations for unfiltered Cu and Pb comprise half or more than half of the particulate

phase and show higher concentration at Skutviken in both seasons.

The K concentration at Skutviken in wintertime is 10 times higher than at Gültzauudden. In

summer both sites have less than half the K concentration which is found at Skutviken

wintertime, and for both sites the truly dissolved phase is dominant (Figure 3).

In winter and summer the Na concentration at Skutviken is about the same concentration

while at Gültzauudden more variation can be seen (Figure 3). In winter the Na concentration

at Gültzauudden is low.

0

0.5

1

1.5

2

2.5

K (m

g/l)

0

4

8

12

Na

(mg/

l)

unfiltered

<0.22 µm

colloidal

<1kDa

SUMMERWINTER SUMMERWINTER

Figure 3. Speciation of K and Na at Skutviken and Gültzauudden.

The highest unfiltered concentrations of Co, Cr, Ni and Zn are measured in wintertime at

Skutviken, when about half the Co, Ni and Zn content is truly dissolved, while for Cr only

about a fifth of the total content is truly dissolved (Figure 4).

0

1

2

3

4

Zn (µ

g/l)

0

0.1

0.2

0.3

0.4

0.5

Co

(µg/

l)

0

0.05

0.1

0.15

0.2

0.25

Cr (

µg/l)

unfiltered

<0.22 µm

colloidal

<1kDa

0

0.2

0.4

0.6

0.8

Ni (

µg/l)

SUMMERWINTER

SUMMERWINTER

SUMMERWINTER

SUMMERWINTER

Pb and Cu have the highest unfiltered concentrations in summertime at Skutviken followed by

high winter concentrations there (Figure 5). The total Pb concentrations are about 8 times

higher at Skutviken compared to Gültzauudden in seasons. The Cu concentrations at both

sites show little seasonal variation, and variations in speciation.

Figure 4. Speciation of Co, Cr, Ni and Zn at Skutviken and Gültzauudden.

0

0.1

0.2

0.3

0.4

Pb

(µg/

l)unfiltered

<0.22 µm

colloidal

<1kDa

0

0.2

0.4

0.6

0.8

1

Cu

(µg/

l)SUMMERWINTER SUMMERWINTER

DiscussionThe element concentrations of the analyzed filtered phase (<0.22 μm) show that the sources

for Skutviken must be different from Gültzauudden. That the water at Gültzauudden resembles

typical Lule River water which is natural due to the location of the sampling point near the

main stream bed of the Lule River. The stronger variation in concentrations at Skutviken can

be influenced by stormwater. Hallberg et al. (2007) points out that Co, Cr, Mn, and Ni belong

to a group of elements in road runoff which have a significant higher dissolved concentration

in winter than in summer. The same seasonal variation is found in Skutviken. Lead and Cu, in

contrast, are predominantly in the particulate phase, and we found with higher concentrations

in summer. Whether the particular fraction of Pb and Ni are bound to Fe-hydroxides or

organic material is not clear. The increase of Pb and Cu should be connected to a carrier.

Possibly a high amount of organic matter in summer can increase their concentration.

Stormwater impact on the elemental concentrations in Skutviken is probable due to the high

concentrations from catch basins (Karlsson et al. 2009). Skutviken receives water from sewers

with similar catch basins. The different number of vehicles passing a catch basin does not

affect the concentrations consistently. The high concentrations <0.22 μm of Mn, S, K, Mg,

Na, and Ca (Table 1) in Skutviken in winter occur in the truly dissolved fraction, while high

Figure 5. Speciation of Pb and Cu at Skutviken and Gültzauudden.

concentrations of Fe mainly occur as particles. This reflects the low retaining capability of

catch basins for truly dissolved species (G. M. Morrison et al. 1988). It also suggests that Co

and Cu are mostly attached to colloidal particles like Fe-oxide or even organic particles.

The strong Mn variation at Skutviken from winter to summer points out a clear change in

water geochemistry, which can be related to changes in oxygen saturation in the water column

causing a shift of location of redox boundaries. Dissolved oxygen concentration in the water

column at Skutviken varies so that the oxygen saturation in the bottom water is close to 0% in

wintertime, when the bay is ice-covered. In contrast, the water column is well oxygenated

(saturation 85-90%) during the ice-free season (Rentz et al., 2011). Wintertime the redox

boundary of Mn is located in the water column which leads to release of dissolved Mn from

the sediment, while summertime this boundary will be found in the sediment. During both

seasons water exchange seems to be more efficient at Gültzauudden, from the point of view

that water exchange favours higher oxygen saturation in the water column. This is supported

by the sediment and pore water profiles of Mn at Skutviken and Gültzauudden (Figure 6). At

Gültzauudden enrichment of likely Mn-hydroxide is probable, while the Mn peak in solid

sediment at Skutviken is missing.

0 0.5 1 1.5 2 2.5

MnO(%DM)

-25

-20

-15

-10

-5

0

5

Dep

th (c

m)

0 2000 4000 6000 8000 10000Mn (µg/l)

Skutviken

Mn in porewaterMnO in Sediment

0 0.5 1 1.5 2 2.5MnO(%DM)

-25

-20

-15

-10

-5

0

5

0 2000 4000 6000 8000 10000Mn (µg/l)

Gültzauudden

Figure 6. Concentrations of Mn in porewater and MnO in Sediment at Skutviken and Gültzauudden wintertime. The top value for the “porewater” represents the bottom water (3cm above sediment surface) (Rentz et al., 2011).

In winter the Na concentration at Gültzauudden is low because of the constant flow from the

regulated Lule River. The ice covered Bothnian Sea has usually a low water level with no

major variations. This may change in summer when wind from southern direction presses sea

water upward the mouth of the Lule River and can cause higher Na concentrations. The

higher concentrations of S, Mg in summer and lower Ca concentrations suggest impact of

Bothnian bay water. At Skutviken other factors influence the concentrations of Na, S, Mg, and

Ca. From stormwater basins in cold climates and road runoff, high winter concentrations of

Na, Mg and K are reported in context with use of de-icing agents (Bäckström et al. 2003). The

use of sodium chloride as de-icing agent is, however not applied in Luleå, resulting in that

winter concentrations in Skutviken are not in class with stormwater pond concentrations

(Karlsson et al. 2010).

Compared with Swedish EPAs reference values (Swedish EPA 2000), the concentrations of

Cr, Ni, Zn are very low and low for Cu and Pb.

ConclusionsSkutviken shows higher unfiltered concentrations for Mn, Fe, S, Co, Cr, K, Ni and Zn in

winter. The seasonal variation of dissolved oxygen in the bay Skutviken can be an influential

factor on concentration of the trace metal species of Mn, Fe, S in the water column.

Wintertime the ice cover prevents the water column from wind mixing, and under the ice

cover the available oxygen will be consumed. The inflow of water to the bay is reduced

during wintertime due to low temperatures and cannot add more oxygenated water. The little

inflow that occurs, can have relatively high concentrations of dissolved trace metals from the

sewer system. A source for enrichment of the trace metals in the water column at wintertime

is the sediment in the bay. The consumption of oxygen in the water column at Skutviken raises

the redox barrier from inside the sediment or from the sediment surface to the water column.

Therefore trace metals already bound to the sediment may be released again to the water

column. At Gültzauudden where the redox conditions are relatively constant over the year,

trace metals become trapped in the sediment without release, while at Skutviken release of

elements bond to Mn-oxides may occur wintertime. However, the amount of release from the

sediment is not determined. The stormwater is a source for elevated metal concentrations,

even though the dissolved, concentrations in Skutviken are still distinct lower than

concentrations in catch basins or stormwater ponds.

References Brown, J. N. (2002), 'Partitioning of chemical contaminants in urban stormwater', dissertation

(University of Otago). Bäckström, M., et al. (2003), 'Speciation of Heavy Metals in Road Runoff and Roadside Total

Deposition', Water, Air, & Soil Pollution, 147 (1), 343-66. Charlesworth, S. M. and Lees, J. A. (1999), 'Particulate-associated heavy metals in the urban

environment: their transport from source to deposit, Coventry, UK.', Chemosphere, 39 (5), 833-48.

Gonzalez, A., et al. (2000), 'Determination of Polycyclic Aromatic Hydrocarbons in Urban Runoff Samples from the "Le Maraisâ" Experimental Catchment in Paris Centre', Polycyclic Aromatic Compounds, 20 (1), 1-19.

Guéguen, C. and Dominik, J. (2003), 'Partitioning of trace metals between particulate, colloidal and truly dissolved fractions in a polluted river: the Upper Vistula River (Poland)', Applied Geochemistry, 18 (3), 457-70.

Guéguen, C., Belin, C., and Dominik, J. (2002), 'Organic colloid separation in contrasting aquatic environments with tangential flow filtration', Water Research, 36 (7), 1677-84.

Hallberg, M., Renman, G., and Lundbom, T. (2007), 'Seasonal Variations of Ten Metals in Highway Runoff and their Partition between Dissolved and Particulate Matter', Water, Air, & Soil Pollution, 181 (1), 183-91.

Hauer, F. R. and Lamberti, G. A. (2006), Methods in Stream Ecology (2 edn.; Amsterdam: Elsevier).

Ingri, J., et al. (2004), 'Size distribution of colloidal trace metals and organic carbon during a coastal bloom in the Baltic Sea', Marine Chemistry, 91 (1-4), 117-30.

Karlsson, K. and Viklander, M. (2008), 'Trace Metal Composition in Water and Sediment from Catch Basins', Journal of Environmental Engineering, 134 (10), 870-78.

Karlsson, K., et al. (2009), 'Physicochemical Distribution of Metals in the Water Phase of Catch Basin Mixtures', Water Quality Research Journal of Canada, 44 (2), 151 - 60

Karlsson, K., et al. (2010), 'Heavy metal concentrations and toxicity in water and sediment from stormwater ponds and sedimentation tanks', Journal of Hazardous Materials, 178 (1-3), 612-18.

Larsson, J., Gustafsson, Ö., and Ingri, J. (2002), 'Evaluation and Optimization of Two Complementary Cross-Flow Ultrafiltration Systems toward Isolation of Coastal Surface Water Colloids', Environmental Science & Technology, 36 (10), 2236-41.

Luleå Kommun (2010), 'Trafikmängder för Luleå Kommun -2010', (Luleå). Morrison, G. M., et al. (1988), 'Transport mechanisms and processes for metal species in a

gullypot system', Water Research, 22 (11), 1417-27. Morrison, M. A. and Benoit, G. (2001), 'Filtration Artifacts Caused by Overloading

Membrane Filters', Environmental science & technology, 35 (18), 3774-79. Munch Christensen, A., Nakajima, F., and Baun, A. (2006), 'Toxicity of water and sediment

in a small urban river (Store Vejlea, Denmark)', Environmental Pollution, 144 (2), 621-25.

Rentz, R., et al. (2010), 'Impact of Urban Stormwater on Sediment Quality in an Enclosed Bay of the Lule River, Northern Sweden', Water, Air, & Soil Pollution, 1-16.

Rodushkin, I. and Ruth, T. (1997), 'Determination of Trace Metals in Estuarine and Sea-water Reference Materials by High Resolution Inductively Coupled Plasma Mass Spectrometry', Journal of analytical atomic spectrometry, 12 (10), 1181.

Schiff, K. and Bay, S. (2003), 'Impacts of stormwater discharges on the nearshore benthic environment of Santa Monica Bay', Marine Environmental Research, 56 (1-2), 225-43.

Simmons, I. G. (1991), Earth, air, and water : resources and environment in the late 20th century (London Edward Arnold).

SMHI (2009a), 'Normal uppmätt nederbörd 1961-1990', (Swedish Meteorological and Hydrological Institute), Klimatkarta som illustrerar uppmätt nederbörds medelvärde i oktober för den av WMO definierade normalperiod 1961-90

SMHI (2009b), 'Klimatkarta som illustrerar andelen snö av årsnederbörden, medelvärde för den av WMO definierade normalperioden 1961-1990', (Swedish Meteorological and Hydrological Institute).

Swedish EPA (2000), 'Bedömningsgrunder för sjöar och vattendrag', (Stockholm: Swedish Environmental Protection Agency).

Westerlund, C. (2007), 'Road Runoff Quality in Cold Climates'. Wildi, W., et al. (2004), 'River, reservoir and lake sediment contamination by heavy metals

downstream from urban areas of Switzerland', Lakes & Reservoirs: Research and Management, 9 (1), 75-87.

Wilding, A., Liu, R., and Zhou, J. L. (2004), 'Validation of cross-flow ultrafiltration for sampling of colloidal particles from aquatic systems', Journal of Colloid and Interface Science, 280 (1), 102-12.

Paper IV

1

Stormwater impact on urban waterways: seasonal variations in sediment metal concentrations in a cold climate

Rentz, R. 1, Blecken, G.-T. 2*, Malmgren, C. 2, Öhlander, B. 1, Viklander, M. 2

1 Division of Geosciences and Environmental Engineering, Department of Civil, Environmental and Natural Resources Engineering, Luleå University of Technology, 971 87 Luleå, SWEDEN 2 Urban Water, Department of Civil, Environmental and Natural Resources Engineering, Luleå University of Technology, 971 87 Luleå, SWEDEN

* Corresponding author:

godble@ltu.se

Phone +46 (0)920 491394,

Fax: +46 (0)920 492818

2

Abstract

Purpose Stormwater discharges cause discharge of contaminated sediments which accumulate in the recipient. It is thus important to investigate sediment and pollutant pathways from the catchment to the recipient and within it. Those processes may be influenced by seasonal changes. The aim of this study was to investigate the stormwater impact on recipients in the Luleå area, Northern Sweden, seasonal changes in contamination loads in the recipient and factors influencing the pollutant pathways in the recipient.

Materials and Methods In front of three storm sewer outlets in Luleå, bottom sediment samples from the connecting ditches and the downstream recipient were taken in autumn and spring (before and after the snow season). The characteristics of the recipients differed in (inter alia) geomorphology and vegetation. The sediment was analyzed for LOI, SiO2, Al2O3, CaO, Fe2O3, MnO, Na2O, P2O5, TiO2, As, Cd, Co, Cr, Cu, Hg, Ni, Pb, S, V, and Zn. The sediment contamination was compared to a stormwater unaffected reference point in Luleå and with Swedish environmental quality guidelines. Pearson’s correlation and a principal component analysis were used to further explain the results.

Results and Discussion Compared to the reference point, at the sampling stations elevated trace metal concentrations were detected. For two sampling points a clear seasonal difference was also observed. Those seasonal variations in grain size, LOI, and chemical concentrations in the ditches originate in stormwater sediment. Changes in runoff intensity cause changes in sediment loads. The retention of metals seems to be due to low turbulence water and the presence of organic material.

Conclusions Stormwater discharge has an impact on the concentrations of contaminant concentrations in the analyzed bottom sediments. The observed seasonal variation of contaminants indicate that a relatively high amount of contaminants is discharged during snowmelt and then reallocated within the recipient either directly or after some temporal retention, depending on the characteristics of the recipient. A calm water column and the presence of organic material in the recipient are crucial for the retention of metals.

Keywords.

Stormwater contamination; Snowmelt; Heavy metals; Environmental impact; Seasonal variation; Cold climate

3

1. Introduction Anthropogenic activities affect the ecology of urban waterways in terms of both hydrology and water chemistry,

e.g., due to urban and agricultural wastewater discharge, urban stormwater runoff, industrial waste disposal, and

atmospheric deposition (Duda 1993). Recently, the importance of urban stormwater discharge on urban stream

quality has been recognized as a significant problem (Chocat et al. 2001). Owing to increased runoff volumes,

increased flashiness of runoff flow, and chemical contamination, urban stormwater can have an adverse impact

on the ecology of the receiving water bodies, often summarized under the term ‘urban stream syndrome’ (Walsh

et al. 2005).

Stormwater might be polluted with a wide range of substances, for example sediments, heavy metals, nutrients,

oil and grease, and salt, originating from manifold sources like car traffic, building materials, construction sites,

winter road maintenance, and fertilizers (Makepeace 1995). A wide range of heavy metals have been detected in

stormwater; some of those most commonly reported are Cd, Cu, Pb, and Zn (Makepeace 1995). Stormwater

metal contamination is of particular concern regarding the environment since metals have been shown (inter

alia) to accumulate in the bottom sediments of urban water bodies (Rentz et al. 2010), possibly causing toxicity

(Karlavi�ien� et al. 2009). Increased supply of metals and organic pollutants to recipients can pose risks for

living organisms (Wildi et al. 2004; Munch et al. 2006).

Special problems might occur in regions with cold climates since snow and snowmelt runoff often show far

higher metal concentrations than stormwater (Marsalek 1991; Engelhard et al. 2007). Reasons for this include

the accumulation of pollutants in the snowpack and the increased presence of metals during winter; the latter

being due to less efficient combustion processes and the increased corrosion caused by road salts applied as a de-

icing agent (Viklander 1998). Salt from winter road maintenance can furthermore affect the partitioning of

metals, leading to an increased fraction of the more environmentally harmful dissolved phase (Bäckström et al.

2004). When (instead of or in combination with salt) sand or fine gravel is used as an anti-slip agent, the total

suspended solids (TSS) concentration in the runoff water may be elevated (Viklander 1999). The fine TSS

fractions (<0.25 mm) are especially important carriers of other (particle-bound) contaminants (e.g., heavy

metals, phosphorus). Thus, winter or spring runoff might have a particularly high impact on the recipient waters

and their bottom sediments.

4

Previous studies of stormwater and gully-pot sediments in the Luleå area in Northern Sweden have indicated

particle-related transport of metal and organic pollutants with seasonal variations (Westerlund and Viklander

2006; Karlsson and Viklander 2008). Large amounts of the pollutants are trapped in the sediments of urban

water bodies in Luleå (Rentz and Öhlander 2011). Furthermore, in Northern Scandinavia, secondary release of

pollutants can be caused by changing redox conditions originating from the postglacial land uplift (Boman et al.

2010).

Given the environmental problems owing to stormwater discharges (Walsh et al. 2005) which include the

sediment and pollutant accumulation in the recipient (Rentz and Öhlander 2011), the increased sediment and

pollutant load in stormwater during winter (Viklander 1998), and the lack of studies monitoring the pollutant

behaviour in stormwater recipients (Taylor and Owens 2009), ther is a need to further investigate sediment and

contaminant pathways from the catchment to the recipient and within the recipient. In regions with cold winters,

seasonal changes are an important factor determining sediment fluxes and cycling. Thus, we investigated the

heavy metal concentrations in bottom sediments of three different recipients in front of storm sewer outlets in the

Luleå area, and their variation between autumn (before the snow season) and spring (after snowmelt). The aim

was to evaluate (1) if there is an impact of stormwater discharges on sediment metal concentrations, (2) if there

are seasonal metal variations, and (3) how the geomorphology and vegetation influences the distribution of

discharged stormwater sediments and associated metals.

2. Materials and Methods.

2.1 Area description

Sediment samples were taken at the three sites Gammelstadsviken, Ytterviken, and Notviken in Luleå, Northern

Sweden (Figure 1, Supplementary Material: aerial photo). At Gammelstadsviken, stormwater from a 67-ha

catchment area is discharged. Of the whole catchment, 29 ha is industrial area, 8 ha residential area, 23 ha roads,

and 13 ha parking lots. While conducting this study a road and bridge construction site was located close to the

storm sewer outlet. The sewer (800 mm in diameter) opens into a 30-m long ditch ending in Gammelstadsviken.

This recipient is densely overgrown by mainly Typha spp. and Carex spp. communities. It is a Swedish nature

reserve and part of the natura 2000 network. At Ytterviken, four sewers (680 mm, 1150 mm, 1350 mm, and 210

mm in diameter) lead into a ditch with a length of 230 m. Of the catchment area, 20 ha is roads and 4 ha car

5

parks; industrial area and the university campus comprise the remainder. At Notviken, stormwater from a 67-ha

large catchment area, including an industrial area with 5-ha of roads and 18-ha of parking lots, is discharged

through a 600-mm pipe into a ditch having a length of ca. 250 m before opening into the bay Notviken. The bay

has an area of ca. 256 ha and is connected to the delta of the Lule River. The southward open-water surface

allows waves to affect the mouth of the ditch and cause redeposition of sediment along the local banks. The

ground in front of the ditch’s mouth shows ripple marks. Also ground-freezing and ice floes affect the deposited

sediments along the shallow banks. The banks of the ditch are partly fixed with stones.

The annual precipitation in the Luleå area is about 500 mm (SMHI 2009a) of which 35 to 40% falls as snow

between November and April/May (SMHI 2009b). Thus, relatively high amounts of surface runoff are

discharged during snowmelt (mainly in April and early May) through a separated sewer system. From November

until May the Lule River and the bays close to the city centre are ice-covered.

2.2 Sampling method

The sediment samples were taken in front of stormwater discharge points at three sites which differed in the

characteristics of the ditch and recipient, as described above. At Notviken, two sampling stations were chosen

(Supplementary Material.), the first (NA) ca. 30 m downstream from the sewer outlet (i.e., in the ditch) and the

second (NB) in shallow water (depth <0.5 m) in front of the mouth of the ditch. At Gammelstadsviken, three

sampling stations were chosen (Supplementary Material), the first (GA) ca. 12 m downstream from the sewer

outlet and the second (GB) ca. 29 m from the pipeline outlet in the ditch. The ditch ends in an open-water pool

framed by dense vegetation. The third sampling station (GC) was in the middle of the pool (depth ca. 1.2 m).

Unfortunately, at this sampling point in December, it was not possible to take a sample in May. At Ytterviken,

three sampling stations were chosen (Supplementary Material), the first (YA) ca. 5 m downstream from the

sewer outlets and the second (YB) ca. 100 m downstream the ditch. The third sampling station (YC) was situated

in front of the mouth of the ditch (depth ca. 1.1 m).

At the three sampling sites, surface-sediment samples (depth: 0-2 cm) were taken in December 2009 and May

2010. The sampling in May was performed at the end of the snowmelt before the first intense rain of the season.

The sediment sampling was done with a Kajak gravity corer. At each sampling station, three undisturbed

sediment cores (64 mm in diameter; 15-25 cm in length) were taken at a distance of about 15 cm from each

6

other. The upper 2 cm of the cores were removed and placed in plastic bags. Before analyses, the three sub-

samples from each sampling point were homogenized and, from that, the sample for analysis was taken.

2.3 Sample Analyses

The sediment was analyzed for LOI, SiO2, Al2O3, CaO, Fe2O3, MnO, Na2O, P2O5, TiO2, As, Cd, Co, Cr, Cu, Hg,

Ni, Pb, S, V, and Zn. Sediment samples for the determination of As, Cd, Co, Cu, Hg, Ni, Pb, S and Zn were

dried at 50°C and digested with HNO3:H2O (1:1) in a microwave oven in closed Teflon bowls. For the other

chemicals, 0.125 g dried matter was melted with 0.375 g LiBO2 and dissolved in HNO3. Determination of

chemical concentrations was made by inductively coupled plasma-atomic emission spectroscopy (ICP-AES) and

inductively coupled plasma-sector field mass spectroscopy (ICP-SMS). Furthermore, the percentages of weight

of grain size fractions were determined by wet sieving according to the Swedish standard method SS-EN 933-1.

All analyses were performed by a SWEDAC accredited (www.swedac.se) laboratories.

2.4 Data analyses To measure the influence of snowmelt runoff on the bottom sediment metal concentration, the concentrations

observed in autumn and spring were compared with each other. Furthermore, they were compared with the metal

concentrations at the reference point Gültzauudden in Luleå which is a bay at the mouth of the Lule River

unaffected by stormwater (as described by Rentz et al., 2010) The samples were compared with Swedish

environmental quality guidelines (Swedish EPA 2000), in order to evaluate the environmental significance of the

metals in the sediment. The characteristics of the catchment and the recipients were used to explain the results.

Pearson’s correlation coefficients were calculated for the concentrations of SiO2, Al2O3, Fe2O3, MnO, Na2O,

LOI, Cd, Co, Cr, Cu, Ni, Pb, S, Zn, and the percentage by weight of the grain size fractions. A principal

component analysis (PCA) was performed (using Umetrics SIMCA-P+ 12.0.1.0) of the concentrations of SiO2,

Al2O3, CaO, Fe2O3, MnO, Na2O, P2O5, TiO2, LOI, As, Cd, Co, Cr, Cu, Hg, Ni, Pb, S, V, Zn and the percentage

of the grain size fractions. The loading scatter plot derived from the PCA, helps to identify relationships among

the variables and groups of samples with similar geochemical behaviour can be visualized.

7

3. Results The results of the sediment analyses are presented in Tables 1 and 2. Of all sediment samples (Table 1), the

highest trace metal and S concentrations were found at the sampling points B and C in both Ytterviken and

Gammelstadsviken. A comparison of these concentrations at all sampling sites with northern Sweden

background levels (Swedish EPA 2000) showed especially high deviations from the background levels for Cr

and Cu, while the concentrations of Cd, Pb, Ni, and Zn at most of the sampling points deviated only slightly or

not at all from the background levels (Table 2). Large or very large deviation was detected for Cr at six sampling

points, for Cu at four points, and for Ni at one point (out of eight points in total). The three sample points at

Gammelstadsviken stood out, with spatial differences and seasonal variation in trace element concentrations. The

sample point GA, nearest to the stormwater outlet, showed less seasonal variation in trace metal and S

concentrations than did GB, even though they had seasonal variation in grain size in common (Tables 1, 3).

Also, the LOI in spring was lower at GA than at GB, although the grain-size distributions were similar. At

Ytterviken in particular, the samples taken in May after snowmelt had similar grain-size distributions at all three

sample points. It is also noticeable that the samples from YC, which had high LOI, had distinctly higher

concentrations of trace metals and S. Even though there was some seasonal variation in grain-size distribution at

both sampling points at Notviken, LOI and the concentrations of trace metals and S showed little variation. The

chemical composition of the sediment samples at Notviken resembles those of YA and YB (Figure 2).

For all samples, the SiO2 and Al2O3 concentrations were almost identical to those found in the sediment at the

non-stormwater affected reference point Gültzauudden. However, differences in other mineral and trace element

concentrations were noticed for all sampling stations, with deviations being especially high at Gammelstadsviken

and Ytterviken (Fig. 2). At all sampling stations the MnO concentration was lower than at the reference point.

For the other chemicals, the differences were most obvious at YC and at all three sampling points at

Gammelstadsviken. The concentrations of Fe2O3, S, Cd, Co, Cr, Cu, Ni, Pb, and Zn and LOI were higher at GC

and YC than at the reference point.

The sampling points YC and GC exhibited the highest percentage of fine grain fractions (<0.063 mm; silt and

clay), which at YC was 87% in spring and 70% in winter, and at GC was 92% in winter (Table 3). The wet

sieving showed that the mean percentage of the finest-grain fractions over all samples was up to 27% in winter

and up to 71% in spring after snowmelt. At all sample points seasonal changes in particle size composition were

8

observed, with a higher content of fine grains (<0.125 mm) observed in May after snowmelt at all sampling

points except YC (Table 3; and it can be assumed that GC would not show much variation owing to the

conditions being similar to those at YC). At YA, YB, GA, and GB the seasonal variation in particle-size

distribution was highest, while the least variation was at YC.

The results from GA and (even more apparent) GB stick out, with the highest seasonal differences in dry weight

(DW), LOI, and trace element concentrations. At these two sampling points, the concentrations were especially

elevated in spring. High variation between sampling points at the same sampling site on the same date was

observed for both Ytterviken and Gammelstadsviken, where, in each case, the sediment composition at sample

point C differs from that in the ditches. The results show that the concentrations of SiO2 and Al2O3 follow each

other if seasonal changes are observed, even though these changes are small (Table 1).

Pairwise correlation coefficients for all chemical concentrations, LOI and grain-size percentages are given in

Table 4. There were significant positive correlations between SiO2, Al2O3 and Na2O. Also Fe2O3 and MnO

showed significant positive correlation. A group of variables that had significant positive correlation witch each

other were LOI and the concentrations of Fe2O3, Cd, Co, Ni, and Zn.

The loading scatter plot (Figure 3) derived from the PCA, showed that in the first and second components the

concentrations of Co, Ni, Fe2O3, Zn, Cd, As, and S were grouped close together. LOI and the concentration of

P2O5 were close in the first and second components and close to the aforementioned group in the first

component. Separate from these two groups and grouped close together were the concentrations of SiO2, Al2O3

and Na2O. The percentage of finest grains (<0.063 mm) had the opposite sign to the percentage of coarser grains

(0.125 – 0.5 mm) in the first, second, and third components. The score scatter plot (Figure 4) showed, in the first

component, a division into two groups. On one side was the group YC in both seasons, together with GB in

spring and GC in winter, and on the other side were the remaining samples. In the second component, the spring

samples from GA and GB were separated from the remaining samples, while, in the third, the samples were

distributed by season, except for GC (for which there was only a winter sample) and GB in spring.

9

4. Discussion The Luleå area contains a range of sites where stormwater is discharged untreated to its recipients. If a

stormwater outlet does not end in the recipient directly, the stormwater often flows through a ditch to the

recipient (as was the case at all three sampling points). The stormwater impact on the recipient’s bottom

sediment is coupled for the most part with the runoff pattern and the ability to transport and settle fine-grain

particles. How the transported metals can be fixed in the sediment also depends on the impact of vegetation

leading to relatively stable conditions in the water column, and decomposing organic material.

In the analyzed ditches and recipients, both seasonal and geographical variations were detected. Seasonal

variations in the percentage of fine grains indicate that the stream conditions in the ditches and the water bodies

in front of the ditches’ mouths vary. It is likely that the low runoff flow during winter and snowmelt, with its

lower velocity, only has the capacity to transport fine particles. Coarse grains will not flush away from surfaces

in the catchment area. Since there are no intense runoff events during a stable winter season, fine particles settle

closely to the storm sewer outlet in the ditches or in the recipient itself. Snowmelt runoff is characterized by high

concentrations of total suspended solids (TSS) (Sansalone 1996; Westerlund et al. 2003). Even though there are

no intense rainfall-runoff events during winter, the snowmelt from the catchment areas provides relatively large

water volumes. These may lead to high water levels in the recipients possibly causing ponding of water in the

ditches. This causes a lower runoff velocity in the ditches, facilitating the sedimentation of fine particles. Along

the shores in the relatively open bay of Notviken, the sedimentation conditions can vary strongly due to ice

covering and wave activity. During the ice-free season, fine-grain sediments are retained in the water column or

redistributed by wave activities. At Ytterviken and Gammelstadsviken, wave impact on sediment is decreased by

vegetation and a deeper water column (depth at YC and GC >1 m; depth at NB <0.5 m). YC and GC are both

situated downstream of the mouths of the ditches, with dense vegetation and low turbulences in the water

column. Over a year, the variations in flow conditions are relatively low at both sample points. These conditions

facilitate the sedimentation of fine grains (<0.063 mm), thus also accumulating pollutants (bound to fine

particles). This pollutant accumulation is of special concern for Gammelstadsviken given its status as a nature

reserve being an important birdlife habitat included in the natura 2000 network.

The correlation coefficients between chemical concentration, chemical concentrations and LOI, and chemical

concentrations and grain size distributions indicate similar geochemical behavior and/or possibly common

10

sources of these compounds (Table 4). The significant positive correlation between the concentrations of Al2O3

and SiO2 suggest their common occurrence in aluminosilicates, which is also indicated by the significant positive

correlation of these two compounds with Na2O. The occurrence of Fe and Mn oxides as common constituents of

stormwater sediment can lead to the significant positive correlation between the concentrations of those oxides

(Stone and Marsalek 1996). The significant positive significant correlation of LOI with the concentrations of

Fe2O3, Cd, Co, Ni, and Zn is due to the extend of metal sorption to organic complex builders. That the

concentrations of Cu and LOI do not show a similar correlation is because the samples from GC had the second

highest LOI but a relatively low Cu concentration. At the same time, Cu concentrations were the highest at GA

and GB in spring, which points to recent contamination which can be related to a nearby construction site.

The first component of the score scatter plot seems to capture geographical similarities in the variables along the

ditches. Beneficial geochemical conditions for enrichment of elements can occur along a ditch, where the water

column is relatively stable and organic material is present. The second component was in particular affected by

the concentrations of Hg, Cu, and CaO, which were highest at GA and GB in spring. That the second component

differentiates the spring samples at GA and GB from the rest indicates that Hg, Cu, and Ca contaminations are

relatively recent. Similarities between GA and GB are also shown in the loading scatter plot. Concrete works at a

nearby road and bridge construction sites (Supplementary Material) may have led to the increase in Ca levels in

the ditch sediment at this site. The Hg and Cu concentrations may be due to the construction site too, or to the

nearby railway; Malawska and Wio�komirski (2001) found Hg and Cu, among other heavy metal concentrations,

elevated in soil and plant samples in the area of a railway junction. The third component was noticeably affected

by the proportion of particles sizes smaller than 0.125 mm, and as such it mostly represents seasonal variation in

particle transport and sedimentation at the sampling sites.

At the sites GB (in spring), GC, and YC (which had high LOI), the high P2O5 concentrations may be due top

organic sources of P instead of inorganic P bound to iron (III) oxide-hydroxide (FeOOH). Phosphorus may be

partly bound to Fe-oxides too, which are correlated with LOI. In snow samples taken along roads in Luleå high P

concentrations were observed (Viklander 1999), which also might explain the high P concentrations at those

sites (which are those mostly affected by stormwater).

11

That the MnO concentration is much higher at the reference point Gültzauudden than at the three sampling

points is due to the redox conditions in the surface sediment and the Mn-cycling (Rentz et al. 2010). Under oxic

conditions in the sediment top, Mn occurs mostly as Mn oxyhydroxides and becomes enriched in the sediment

surface during early diagenesis (Davison 1993). In the ditches and at NB, the sediment accumulation is less

continuous (truly with phases of erosion), so that early diagenesis cannot occur to the same extent.

The samples from YC and GC were characteristic of the surface sediment of brackish-lacustrine bays along the

Swedish and Finnish coasts of the Bothnian bay. Owing to the standing body of water and decomposition of the

high organic content, suboxic or anoxic conditions can exist already in the surface sediment possibly causing S

enrichment at these two sampling points (cf. Urban et al. 1999). Bacterial sulphate reduction can in that case

account for the enrichment of FeS, FeS2, and other metal sulphides in the sediment (Boman et al. 2010). In

addition, the increased concentrations of Cd, Co, Cr, Cu, Ni, Pb, and Zn can be caused by desorption to organic

complex builders and fixation with sulphides. So the organic material and the fine-grained mineral fraction can

exhibit adsorption surfaces for metals, but also the formation of FeS and, further on, FeS2 may lead to metal

fixation.

That GA had lower LOI and lower concentrations of the heavy metals Cd, Co, Ni, Pb, and Zn in spring, when

the grain-size distribution was similar to that of GB, indicates that organic matter functions as a carrier or

complex builder for trace metals in this environment.

5. Conclusions Stormwater discharge has an impact on the concentrations of metals, other elements and their oxides in the

analyzed sediments. The seasonal variations in grain size, LOI, and chemical concentrations in the ditches must

originate from stormwater sediment. The recipients GC and YC clearly had higher concentrations of Cd, Co, Cr,

Ni, Pb, and Zn than did the reference point. The highest metal concentrations observed were in the Swedish EPA

(2000) deviation classes 4 and 5 for Cr and Cu in 17 of 32 samples. The temporary impact of a nearby

construction site on the sediment concentrations of CaO, Hg and Cu was likely to have affected the GA and GB

samples in spring. It appears that a calm water column (low flow velocity, low wave impact, dense vegetation)

and the presence of organic material in the recipient are crucial for the retention of metals. Redox conditions,

12

which can lead to metal sulphide formation, are likely controlled by the decomposition of organic material at the

studied sites.

The proportion of particles <0.063 mm (silt and clay) in spring of all surface samples (0-2 cm) from the ditches

was due to seasonal variation in runoff. Changes in runoff intensity and high sediment loads in snowmelt cause

changes in sediment loads. The spring sampling was conducted after the main snowmelt but before the first

intense rain event in Luleå. Snowmelt runoff variations are commonly relatively low in contrast to flash floods

due to intense rain event. A lower runoff velocity results in sediment transport of fine grains only which then are

likely to accumulate in the ditches and recipients.

The seasonal variation in flow in the ditches causes variation in the surface sediment (of grain size, LOI, and

chemical concentrations). At YC and GC, three important factors contribute to the accumulation of trace metals.

Beside conditions which are beneficial for sedimentation and a supply of dead organic matter, the high sulphur

content of the coastal sediment deposits can also contribute to the fixation of metals. These three factors increase

the depletion of O2 in the sediment (perhaps already in the bottom near surface water), resulting in good

conditions for FeS and FeS2 formation.

The observed seasonal variation of contaminants indicate that a relatively high amount of contaminants is

discharged during snowmelt and then reallocated within the recipient either directly (Notviken) or after some

temporal retention (Ytterviken, Gammelstadsviken), depending on the characteristics of the recipient.

Along the ditches, light particles of dead organic matter are more likely to be transported downstream during

flash floods which prevents the long-term accumulation of trace metals. Also, varying water levels and streaming

water will oxygenate the surface sediment in the ditches at times.

Acknowledgements

The authors thank Monica Olofsson for her help with the sampling.

13

References Boman, A., S. Fröjdö, et al. (2010). Impact of isostatic land uplift and artificial drainage on oxidation of brackish-water sediments rich in metastable iron sulfide. Geochimica Cosmochimica Acta 74(4): 1268-1281. Bäckström, M., S. Karlsson, et al. (2004). Mobilisation of heavy metals by deicing salts in a roadside environment. Water Res. 38: 720-732. Chocat, B., P. Krebs, et al. (2001). Urban drainage redefined: from stormwater removal to integrated management. Water Sci. Technol. 43(5): 61-68. Davison, W. (1993). Iron and manganese in lakes. Earth-Science Reviews 34(2): 119-163. Duda, A. M. (1993). Addressing nonpoint sources of water-pollution must become an international priority. Water Sci. Technol. 28(3-5): 1-11. Engelhard, C., S. De Toffol, et al. (2007). Environmental impacts of urban snow management - The alpine case study of Innsbruck. Sci. Total Environ. 382(2-3): 286-294. Karlavi�ien�, V., Švedien�, S., Mar�iulionien�, D. E., Randerson, P., Rimeika, M. et al. (2009). The impact of stormwater runoff on a small urban stream. J. Soils & Sediments 9(1): 6-12. Karlsson, K. and M. Viklander (2008). Trace metal composition in water and sediment from catch basins. J. Environ. Eng. 134(10): 870-878. Makepeace, D. (1995). Urban stormwater quality - summary of contaminant data. Crit. Rev. Environ. Sci. Technol. 25(2): 93-139. Malawska, M. and B. Wio�komirski (2001). An analysis of soil and plant (Taraxacum officinale) contamination with heavy metals and polycyclic aromatic hydrocarbons (PAHs) in the area of the railway junction I�awa G�ówna, Poland. Water Air Soil Pollut. 127(1): 339-349. Marsalek, J. (1991). Urban drainage in cold climate: problems solutions and research needs. International Conference on Urban Drainage and New Technologies 1991, Dubrovnik, Yugoslavia. Munch Christensen, A., F. Nakajima, et al. (2006). Toxicity of water and sediment in a small urban river (Store Vejlea, Denmark). Environ. Pollut. 144(2): 621-625. Rentz, R., A. Widerlund, et al. (2010). impact of urban stormwater on sediment quality in an enclosed bay of the Lule River, Northern Sweden. Water Air Soil Pollut. 1-16. in press. Rentz, R. and B. Öhlander (2011). Urban impact on water bodies in the Luleå area, northern Sweden, and the role of redox processes. Hydrol. Res. in press. Sansalone, J. J. (1996). Characterization of metals and solids in urban highway winter snow and spring rainfall-runoff. Transit 1523(1): 147. SMHI (2009a). Klimatkarta: Uppmätt nederbörd 1961-1990, månadsvis. Swedish Meteorological and Hydrological Institute. (In Swedish). SMHI (2009b). Klimatkarta: Andel snö av årsnederbörden, medelvärde för den av WMO definierade normalperioden 1961-1990. Swedish Meteorological and Hydrological Institute. (In Swedish). Stone, M. and J. Marsalek (1996). Trace metal composition and speciation in street sediment: Sault Ste. Marie, Canada. Water Air Soil Pollut. 87(1): 149-169. Swedish EPA (2000). Environmental quality criteria - lakes and watercourses. Report 5050. Swedish Environmental Protection Agency (Naturvårdsverket), Stockholm, Sweden. .

14

Taylor, K. G., Owens, P. N. (2009). Sediments in urban river basins: a review of sediment-contaminant dynamics in an environmental system conditioned by human avtivities. J. Soils & Sediments 9(4): 281-303. Urban, N. R., K. Ernst, et al. (1999). Addition of sulfur to organic matter during early diagenesis of lake sediments. Geochimica Cosmochimica Acta 63(6): 837-853. Walsh, C. J., A. H. Roy, et al. (2005). The urban stream syndrome: current knowledge and the search for a cure. J. N. Am. Benthol. Soc. 24(3): 706-723. Westerlund, C., M. Viklander, et al. (2003). Seasonal variation in road runoff quality in Luleå, Sweden. Water Sci. Technol. 48(9): 93-101. Westerlund, C., M. Viklander, et al. (2006). Particles and associated metals in road runoff during snowmelt and rainfall. Sci. Total Environ. 362:143-156. Viklander, M. (1998). Snow quality in the city of Luleå, Sweden - time-variation of lead, zinc, copper, and phosphorus. Sci. Tot. Environ. 216: 103-112. Viklander, M. (1999). Dissolved and particle-bound substances in urban snow. Water Sci. Technol. 39(12): 27-32. Wildi, W., J. Dominik, et al. (2004). River, reservoir and lake sediment contamination by heavy metals downstream from urban areas of Switzerland. Lakes Reservoirs: Res. Manage. 9(1): 75-87.

15

Tables

16

Tab

le 1

. DW

, LO

I, an

d tra

ce e

lem

ent c

once

ntra

tions

in th

e se

dim

ent s

ampl

es a

nd a

t the

refe

renc

e si

te (G

ültz

auud

den,

cf.

Ren

tz, e

t al.

2010

).

Sam

ple

Seas

on

DW

L

OI

SiO

2 A

l 2O3

Fe2O

3 M

nO

Na 2

O

Cd

Co

Cr

Cu

Ni

Pb

S Zn

%

%

DW

%

DW

%

DW

%

DW

%

DW

%

DW

m

g kg

-1

DW

mg

kg-1

D

W

mg

kg-1

D

W

mg

kg-1

D

W

mg

kg-1

D

W

mg

kg-1

D

W

mg

kg-

1 DW

mg

kg-1

D

W

NA

12

/09

75.6

1.

2 64

.7

12.9

4.

9 0.

07

3.5

0.21

6.

2 30

7 29

18

13

35

8 10

4

05/1

0 69

.9

1.9

70.8

13

.4

4.4

0.06

3.

5 0.

26

5.8

99

35

16

11

455

98

NB

12

/09

73.8

0.

9 67

.4

12.9

3.

6 0.

08

3.6

0.36

3.

8 10

3 21

11

12

46

9 89

05/1

0 66

.1

1.3

69.8

12

.9

5.3

0.09

3.

5 0.

15

4.5

105

17

13

8 39

9 73

G

A

12/0

9 80

.0

1.3

63.4

13

.0

5.3

0.08

3.

4 0.

16

6.1

95

77

16

14

859

76

05

/10

62.2

3.

0 67

.9

13.0

5.

2 0.

07

3.4

0.26

6.

2 10

0 14

5 15

21

12

40

96

GB

12

/09

73.3

1.

0 69

.4

13.4

3.

5 0.

06

3.7

0.15

4.

4 67

10

1 10

12

51

3 68

05/1

0 41

.3

10.2

59

.1

12.3

6.

9 0.

09

3.0

0.65

11

.2

82

263

24

40

4570

26

8 G

C

12/0

9 40

.7

11.2

54

.5

11.8

4.

7 0.

08

2.8

0.58

21

.0

75

41

39

17

1600

0 13

0

05/1

0 -

- -

- -

- -

- -

- -

- -

- -

YA

12

/09

79.4

1.

4 64

.7

13.0

5.

6 0.

15

3.4

0.30

8.

3 66

27

18

22

31

1 15

6

05/1

0 75

.6

1.7

69.4

13

.4

5.4

0.12

3.

5 0.

30

6.6

64

24

14

9 32

4 15

1 Y

B

12/0

9 75

.9

1.1

65.7

13

.1

6.1

0.09

3.

4 0.

24

5.0

92

21

16

12

1090

13

8

05/1

0 71

.6

1.6

70.9

13

.4

4.4

0.06

3.

5 0.

32

5.2

83

23

13

10

1020

13

0 Y

C

12/0

9 34

.1

9.6

53.8

11

.2

10.6

0.

13

2.6

1.04

20

.7

129

75

49

42

8000

45

2

05/1

0 21

.2

13.3

53

.3

11.4

11

.6

0.13

2.

4 1.

23

25.0

14

5 71

50

29

95

70

470

Ref

eren

ce

site

03

/07

99.0

7.

3 56

.1

12.4

8.

5 1.

74

2.9

0.31

15

.2

72

19

21

13

498

106

17

Table 2. Comparison of trace element deviation from reference values for sediments (Swedish EPA 2000) for 0-2 cm sediment depth of the ditches and the reference site Gültzauudden. The deviation is calculated as sediment content divided by reference value (a = no deviation, b = slight deviation, c = significant deviation, d = large deviation, e = very large deviation).

Cd Cr Cu Pb Ni Zn

YA 12/09 0.4 a 4.4 c 1.8 b 0.4 a 1.8 b 1.0 b

05/10 0.4 a 4.2 c 1.6 b 0.2 a 1.4 b 1.0 b

YB 12/09 0.3 a 6.1 c 1.4 b 0.3 a 1.6 b 0.9 a

05/10 0.4 a 5.5 c 1.5 b 0.2 a 1.3 b 0.9 a

YC 12/09 1.3 b 8.6 d 5.0 d 0.8 a 4.9 d 3.0 c

05/10 1.5 b 9.7 d 4.7 d 0.6 a 5.0 d 3.1 c

NA 12/09 0.3 a 20.5 e 1.9 b 0.3 a 1.8 b 0.7 a

05/10 0.3 a 6.6 d 2.3 c 0.2 a 1.6 b 0.7 a

NB 12/09 0.5 a 6.9 d 1.4 b 0.2 a 1.1 b 0.6 a

05/10 0.2 a 7.0 d 1.1 b 0.2 a 1.3 b 0.5 a

GA 12/09 0.2 a 6.4 d 5.2 d 0.3 a 1.6 b 0.5 a

05/10 0.3 a 6.7 d 9.7 e 0.4 a 1.5 b 0.6 a

GB 12/09 0.2 a 4.5 c 6.7 d 0.2 a 1.0 b 0.5 a

05/10 0.8 a 5.5 c 17.5 e 0.8 a 2.4 c 1.8 b

GC 12/09 1.2 b 6.8 d 9.6 e 1.0 a 3.4 c 2.6 b

Reference site 0.4 a 4.8 c 1.3 b 0.4 a 2.1 c 0.7 a

18

Table 3. Particle-size distributions in the sediment samples. Sample Season Particle size (mm)

0-0.063 0.063-0.125

0-0.125 0.125-0.25

0.25-0.5

% % % % %

NA 12/09 29.9 4.1 34 19 30 05/10 62.5 25.5 88 12 0

NB 12/09 4.1 19.9 24 20 36 05/10 70.8 25.2 96 2 2

GA 12/09 10.7 3.3 14 9 21 05/10 68.1 24.9 93 7 0

GB 12/09 5.7 3.3 9 18 43 05/10 67.9 23.1 91 7 2

GC 12/09 92 4 96 3 0 05/10 - - - - -

YA 12/09 1.3 0.7 2 7 42 05/10 66.6 20.4 87 13 0

YB 12/09 4.6 1.4 6 11 39 05/10 73.5 20.5 94 6 0

YC 12/09 70.2 13.8 84 11 3 05/10 87.4 8.6 94 2 0

Tab

le 4

. Pea

rson

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**

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-0.8

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* 0.

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1

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**

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**

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**

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**

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* 0.

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0.59

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0.36

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orre

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ifica

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Cor

rela

tion

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he 0

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l (2-

taile

d).

20

Figures

Figure 1. Sampling sites Notviken (N), Gammelstadsviken (G), Ytterviken (Y), and the reference site Gültzauudden (R) in the Luleå area, Northern Sweden.

21

Figure 2. Trace element concentrations, LOI, and fine grain-size fraction (<0.063 mm) at Ytterviken (Y), Gammelstadsviken (G), and Notviken (N); normalized to sediment from the reference point Gültzauudden. Si = SiO2; Al = Al2O3; Mn = MnO; P = P2O5.

22

Figure 3. Loading scatter plot of chemical concentrations, LOI, and grain-size fractions.

Figure 4. Score scatter plot of the 15 samplings (sample date: 05 = May; 12 = December).

23

Supplementary Material

Supplementary Figure 5. Sampling site Gammelstadsviken (arial photograph Digitala Kartbiblioteket I 2010/0046). Sampling points (A,B,C); outlet pipe (OP).

Supplementary Figure 6. Sampling site Ytterviken (arial photograph Digitala Kartbiblioteket I 2010/0046). Sampling points (A,B,C); outlet pipe (OP), course of the ditch (light grey line).

24

Supplementary Figure 7. Sampling site Notviken (arial photograph Digitala Kartbiblioteket I 2010/0046). Sampling points (A,B,C); outlet pipe (OP), course of the ditch (light grey line).

Paper V

1

Water and sediment quality in an artificial stormwater basin receiving highway runoff

Ralf Rentz1*, Magnus Westerstrand1, Björn Öhlander1

1 Division of Geosciences and Environmental Engineering, Department of Civil, Environmental and Natural Resources Engineering, Luleå University of Technology,

97187 Luleå, SWEDEN

* Corresponding author: ralren@ltu.se

2

Abstract

Water and Sediment samples were taken from a stormwater pond that receives

highway runoff. For the surface water samples the speciation of Ca, Cd, Co, Cr, Cu,

Fe, K, Mg, Mn, Na, Ni, Pb, S, and Zn was determined with membrane

filtration/ultrafiltration (truly dissolved phase <1kDa; colloidal fraction >1 kDa and

<0.22 μm). For sediment and porewater concentrations of Al2O3; Al, Cd, Co, Cr, Cu,

Fe2O3; Fe, Na2O; Na, Ni, Mn, Pb, S, SiO2; Si, and Zn were determined. Elevated

heavy metal concentrations in the water column of a stormwater basin and elevated

heavy metal concentrations and PAH concentrations in surface sediment of the

stormwater basin were found. The elements Cd, Co, Cr, Cu, Fe, Mn, Na, Ni, Pb, S,

and Zn showed seasonal variations in element concentrations in the water column.

Especially in winter, the metal concentrations of Co, Cr, Cu, Ni, Mn, Na, and Zn are

dominated by the truly dissolved phase. In the sediment pollutants can be trapped due

to sorption on organic material, and early diagenesis processes with formation of Mn-

and Fe-hydroxides and sulphate reduction. This will just affect a fraction of the

concentrations of metals in truly dissolved phase, while most of the dissolved

concentrations are most likely not retained in the stormwater facility. The PAH

contamination in the sediment is relatively low.

Keywords

Highway runoff; stormwater pond; seasonal variation, water quality; sediment quality

Introduction

Runoff from highways is identified as a major agent for pollutant transport. A

common technique to prevent recipients from pollution and damage is the

implementation of stormwater basins. Their main functions are to limit risk for

flooding and to achieve pollutant removal before the stormwater reaches a recipient.

In addition, they can have a landscaping value and take on habitat functions. Lee et al.

(1997) found that in a stormwater basin along a motorway in France, particles <20 �m

settled, and the sediment showed heavy-metal enrichment. Traffic density, vehicle

wear, road construction materials, road wear and road management are some factors

that impact the loading of highway runoff. Important sources identified by Davis et al.

3

(2001) are vehicle brake emissions for copper and tire wear for zinc. In regions with

cold climates, snow and snowmelt runoff often show far higher metal concentrations

than stormwater (Marsalek 1991; Engelhard, De Toffol et al. 2007). Accumulation of

pollutants in the snowpack and the increased presence of metals wintertime due to less

efficient combustion processes and increased corrosion due to road salts applied as a

de-icing agent etc., are reasons for higher metal concentrations in winter (Viklander

1998). Road salts can furthermore affect the partitioning of metals, leading to an

increased fraction of the more environmentally harmful dissolved phase (Bäckström,

Karlsson et al. 2004). Bäckström et al. (2003) ascribed seasonal changes with

increased dissolved metal concentrations (Al, Cd, Co, Cr, Mn, Ni) in winter runoff

from 2 highways in mid Sweden to use of studded tires causing increased pavement

wear. Studies of stormwater and gully pot sediments in northern Sweden, Luleå area

(Westerlund 2007; Karlsson and Viklander 2008), indicated particle-related transport

of metal and organic pollutants with seasonal variations. Tire wear is a source for Cr,

Cu, Fe, Ni, Zn and polycyclic hydrocarbons (PAHs) in particular (Aryal, Vigneswaran

et al. 2010). Suspended solids in road runoff were found to be important carriers for

metal contents (Tuccillo 2006) and to affect PAH concentration in runoff (Aryal,

Furumai et al. 2005). Stormwater basins where particle sedimentation is possible can

effectively reduce metal concentrations in the water column (Färm 2003).

In this study we investigated the heavy metal and PAH concentrations in bottom

sediments of a stormwater basin, and the speciation of elements in the water column

in summer and winter. The aims were to evaluate (1) if impact of stormwater

discharges on sediment is detectable, (2) if there are seasonal elemental variations in

water and sediments, and (3) if the stormwater basin can function as a trap for

pollutants.

Material and methods

Sampling site

The studied stormwater basin (Fig. 1) is situated west of the highway E4 at the

highway intersection Häggvik, 15 km north of central Stockholm, and has been in full

operation since 1998. The facility consists of a “3-step system” with a pump station

4

and two sedimentation basins followed by an overflow surface. The system receives

highway runoff from Häggviksleden (6.8 ha) and the E4 highway (1.9 ha), totally a

sealed road area of 8.7 ha (ALcontrol Laboratories 2005). Häggviksleden connects the

E4 with the main road Danderydsvägen in Edsberg. The runoff from Häggviksleden

and parts of the E4 is led via a pump into the first basin. At the pumping station

separation of oil is conducted. A second inflow adds only water from the E4. The first

basin is elongated with a maximum size of 100 x 50 m and its depth varies between 2

- 2.5 m with a capacity between 4,500 m3 and nearly 6,000 m3 depending on the

lowest or highest water level. On the opposite side of the pump station inlet at the

basin ground, an outlet tube with a diameter of 800 mm (D 800) leads the water over a

distance of 55 m to the second basin. The second basin is approximately 70 x 60 m in

area and its depth varies between 2 – 2.5 m. The volume at highest water level is

about 8,000 m3 and at lowest water level 6,000 m3. At both basins the banks are stone-

covered between highest and lowest water level to prevent erosion. Groundwater

infiltration is averted by the use of a bentonite carpet covered with macadam. The

water leaves the second basin through a 2-chamber gully whose function is to extend

the water retention time in the basin. At a 2-year rain the retention time is calculated

to 36 hours in both basins. From the gully the water runs over a 35 m long and 120 m

wide grass-covered overflow-area slope before reaching a ditch which discharges into

Lake Ravalen after about 1000 m.

5

0 250 km

Helsinki

DENMARK

NORWAY

Oslo

Stockholm

North Atlantic

Baltic Sea

SWEDENFINLAND

Gulf of Bothnia

Figure 1. Stormwater basins for Häggviksleden, 15 km north of central Stockholm, with pumping station (P), stormwater basin 1 and 2 (B1, B2), grass overflow area (O-A) and sediment and water sampling station (X) in basin 2. (arial photograph Digitala Kartbiblioteket I 2010/0046)

Sediment, porewater and basin water sampling

Water and sediment sampling was conducted in the second sedimentation basin (B2,

Fig. 1). In the end of August 2009 and in March 2010, a Kajak gravity corer with a

core tube diameter of 64 mm was used to receive sediment cores from a boat and from

the ice. The sediment sampling was conducted in the centre of the second basin with a

water depth of > 2 m. The sediment core surfaces were judged to be undisturbed. The

cores were sectioned in subsamples (0.5 cm thick for the uppermost 3 centimetres and

1 cm thick for the remainder of the core).

For porewater analyses the sediment samples were put into plastic bags directly after

sectioning. All air was pressed out of the bags before they were placed in an Ar-filled

container to keep the sediments in an oxygen-free environment until the porewater

was extracted within the following six hours. The porewater was separated by vacuum

filtration (0.22 μm Millipore® membrane filters) arranged in an Ar-flushed glove box.

The porewater samples were collected in 60-ml acid-washed polyethylene bottles and

6

refrigerated until further analysis. The remaining sediment was collected in

polyethylene boxes until further preparation for the metal analyses. For PAH analyses

a separate sediment core was sampled and the sediment was placed in glass containers

with Teflon-lined caps.

Bottom water was sampled from the core tube immediately after retrieval, 3 cm above

the sediment surface. The water was drawn with a small plastic tube fixed on a

syringe and filtered through a 0.22 μm Millipore® membrane filter.

The surface water was sampled 50 cm below the water surface and 50 cm below the

ice underside, respectively. Water was pumped by a peristaltic pump (Masterflex®

L/S®) through the tube into 25-litre polyethylene (PE) containers. Membrane

filtration (0.22 μm pore size, 142 mm diameter, Millipore® mixed cellulose esters)

was carried out indoors within 6 hours from sampling. The first filter was used until it

was completely clogged; the filtered volume was measured and then discarded. For

the actual sample, new filters were used, through which half the clogging volume was

allowed to pass. This was done to decimate discrimination of colloids that is caused

by clogging of filters (Morrison and Benoit, 2001). The filtrate was collected in a 25-

litre PE container from which subsamples were taken for analyses. The membrane

filtered water was then ultrafiltrated in a Millipore® Prep/Scale system. The filter had

a manufacturer specified cutoff of 1 kDa and a filter membrane area of 0.54 m2. The

filter material was regenerated cellulose. The system was connected with a Watson

Marlow peristaltic pump. After the ultafiltration (Cheryan 1998), subsamples were

taken from the retentate and the filtrate. Subsamples were collected in 60-ml acid-

washed polyethylene bottles and refrigerated until further analysis. All used tubing,

bottles and containers were acid-cleaned in 5% HCl with subsequent wash in MilliQ

water (Millipore, 18.2 M�) before sampling.

The pH and dissolved oxygen in the water column were determined with a Hydrolab®

Mini Sonde 5 water quality probe.

In summer unfiltered and membrane-filtered (0.22 μm) water samples were also taken

in stormwater basin 1 (B1, Fig. 1).

Analyses

Metal analyses were performed by the accredited laboratory ALS Scandinavia AB in

Luleå. The surface- and porewater samples were analyzed for major and trace

elements in inductively coupled plasma atomic emission spectroscopy (ICP-AES) and

7

inductively coupled plasma with sector field mass spectrometry (ICP-SFMS). To the

water samples, 1 ml nitric acid (suprapur) was added per 100 ml. For instrument

operation details, see Rodushkin and Ruth (1997). Sediment samples for

determination of As, Cd Co, Hg, Ni, Pb and S were dried at 50°C digested in a

microwave oven in closed Teflon bowls with a nitric acid : water ratio of 1:1. For

other elements 0.125 g dried matter (DM) was melted with 0.375 g LiBO2 and

dissolved in HNO3. Metal determinations were made by ICP-AES and ICP-SFMS.

The following 16 PAHs were analyzed in the sediment: Naphthalene (NAP),

Acenaphthylene (ACY), Acenaphthene (ACE), Fluorene (FL), Phenanthrene (PHEN),

Anthracene (ANT), Fluoranthene (FLR), Pyrene (PYR), Benzo(a)anthracene (BaA),

Chrysene (CHY), Benzo(b)fluoranthene (BbF), Benzo(k)fluoranthene (BkF),

Benzo(a)pyrene (BaP), Dibenz(a,h)anthracene (DBA), Benzo(ghi)perylene (BPY) and

Indeno(1,2,3-cd)pyrene (INP). The PAH sediment samples were leached with acetone

: hexan : cyclohexan (1:2:2) and measurements were performed with gas

chromatography mass spectrometry (GC-MS).

Determination of colloidal and truly dissolved phase

Ultrafiltration is an applicable technique for determination of the size distribution of

components in natural water samples. The method is often applied for studies of the

colloidal and truly dissolved species of metals and organic matter in natural waters

(Guéguen and Dominik 2003; Ingri, Nordling et al. 2004). The enrichment of species

concentrations in the retentate facilitates the determination of low-abundance species

(e.g. colloidal concentrations). Ultrafiltration techniques have previously been

described and evaluated (Guéguen, Belin et al. 2002; Wilding, Liu et al. 2004). Two

critical aspects when applying the method for natural water samples are the mass

balance recovery and the accuracy of determination of the species concentrations in

the retentate. Larsson et al. (2002) found that a cross-flow ratio above 15 was

necessary to achieve mass balance recoveries close to 100%. The cross-flow ratio

CFR is defined as:

perm

ret

QQCFR �

Qret and Qperm denote the retentate- and the permeate flow rate, respectively. It was

also found that an enrichment factor (total feed water volume : final retentate volume)

8

larger than 10 was required for accurate determination of the colloidal species. The

enrichment factor EF and the colloidal concentration Ccoll can be calculated using:

ret

retperm

VVV

EF�

�

EFCC

C permretcoll

��

Where Vperm, Vret denote the volumes of the permeate and the retentate. Cperm, Cret and

Cfeed denote the concentrations of the permeate, the retentate and the feed sample,

respectively. Finally, the mass balance recovery R in percent units may be determined

as:

feed

retperm

CCC

R�

�

The truly dissolved phase constitutes the fraction <1kDa and the colloidal fraction

contains particles >1 kDa and <0.22 μm.

Results

The probe measurements indicated a seasonal difference in dissolved oxygen of 89%

saturation in summer and 29% in winter, while pH was 7.0 in winter and 7.9 in

summer.

The total concentrations of Cd, Cr, Cu, Ni, Pb and Zn in water from the stormwater

basins showed higher concentrations in March 2010 compared to summer 2009 (Table

1). For the earlier studies the seasonal variation was not so clear. Copper and Zn have,

at least in two of three winter-summer cycles, higher measured winter concentrations.

The concentrations in the water column of Ca, K, Mg, Mn, Na and S were found truly

dissolved to 100% during both summer and winter. For these elements the

concentrations in winter are higher than the summer concentrations, most obvious for

Na with a 5 times higher concentration in winter.

9

Table 1. pH and total concentrations for Cd, Cr, Cu, Ni, Pb, and Zn in stormwater basin B1 and B2 in summer 2008 and late winter 2009 compared with concentrations measured at the inlet at B1 in earlier years (ALcontrol Laboratories 2001; ALcontrol Laboratories 2005).

month/year pH Cd Cr Cu Ni Pb Zn

μg/l μg/l μg/l μg/l μg/l μg/l

03/10 B2 7.0 0.02 1.23 6.19 1.51 0.33 55.05

08/09 B2 7.9 <0.002 0.14 1.98 0.60 0.21 1.73

08/09 B1 7.9 <0.002 0.57 4.29 0.64 0.21 6.09

03/05 inlet

B1 7.95 0.03 11.13 5.38 <0.1 0.20 8.50

06/05 inlet

B1 8.13 0.05 4.13 4.65 1.95 1.08 5.00 12/99 &

02/00 inlet

B1 8.20 <1 <3 <20 2.75 <3 0.02 06/00 &

08/00 inlet

B1 8.35 <1 3.25 3.00 <6 3.25 0.01 12/00 &

01/01 inlet

B1 7.85 1.25 1.25 7.50 <6 2.50 0.11 06/01 &

08/01 inlet

B1 8.65 <0.1 1.25 4.00 <5 <1 0.21

Higher concentrations in the water column in wintertime were observed also for Cd,

Co, Cr, Cu, Fe, Ni, Pb and Zn (Fig. 2-4), which occur in different speciations than the

truly dissolved phase only. For Cd the unfiltered phase is about 16 times higher in

winter, while the phase <0.22 �m is just 4 times the summer concentration. Cobalt

shows only a small colloidal contingent in both seasons, while the truly dissolved

phase is dominant. For Cr the colloidal contingent is very small and in both seasons

the dissolved phase dominates over an existing particulate phase. The speciation of Cu

clearly shows a colloidal contingent in both seasons. In winter dissolved Cu stands for

most of the increase of the unfiltered phase. For Ni the dissolved and particulate

phases increase most in winter, in a similar way as Cu. Iron and Pb are in both seasons

dominated by the particulate phase.

10

sampling date

0

40

80

120C

a (m

g/l)

2009-08-27 2010-03-04 sampling date

0

4

8

12

16

K (m

g/l)

2009-08-27 2010-03-04

sampling date

0

4

8

12

Mg

(mg/

l)

2009-08-27 2010-03-04 sampling date

0

10

20

30

40

Mn

(µg/

l)

2009-08-27 2010-03-04

sampling date

0

200

400

600

Na

(mg/

l)

2009-08-27 2010-03-04 sampling date

0

5

10

15

20

25

S (m

g/l)

2009-08-27 2010-03-04

unfiltered <0.22 µm colloidal <1kDa Figure 2. Seasonal speciation of Ca, K, Mg, Mn, Na and S in the surface water (depth 0.5 m) of B2.

11

sampling date

0

0.004

0.008

0.012

0.016

Cd

(µg/

l)

2009-08-27 2010-03-04 sampling date

0

0.5

1

1.5

2

2.5

Co

(µg/

l)

2009-08-27 2010-03-04

sampling date

0

0.4

0.8

1.2

Cr (

µg/l)

2009-08-27 2010-03-04 sampling date

0

2

4

6

Cu

(µg/

l)

2009-08-27 2010-03-04

sampling date

0

0.02

0.04

0.06

0.08

0.1

Fe (m

g/l)

2009-08-27 2010-03-04 sampling date

0

0.4

0.8

1.2

1.6

Ni (

µg/l)

2009-08-27 2010-03-04

unfiltered <0.22 µm colloidal <1kDa Figure 3. Seasonal speciation of Cd, Co, Cr, Cu, Fe and Ni in the surface water (depth 0.5 m) of B2.

12

sampling date

0

0.1

0.2

0.3P

b (µ

g/l)

2009-08-27 2010-03-04 sampling date

0

20

40

60

Zn (µ

g/l)

2009-08-27 2010-03-04

unfiltered <0.22 µm colloidal <1kDa Figure 4. Seasonal speciation Pb and Zn in the surface water (depth 0.5 m) of B2.

13

LOI in sediment and metal concentrations in sediment and porewater

In the sediment a concentration change is present for LOI and most elements at 3-5

cm depth in both summer and winter (Fig. 5-10). For LOI the concentration in the

upper 3 cm is constant at around 25% DM, and then it decreases to less than 3% DM

at 5 cm depth, and it is constant at around 2% DM in sediment deeper than 5 cm.

0 10 20 30LOI, DM %

-15

-10

-5

0

Dep

th in

cm

summerwinter

Figure 5. Loss on ignition (LOI) of the stormwater basin sediment.

Si and Al concentrations in the sediment have similar characteristics in their

concentration profile. There is the characteristic increase in the solid phase from 3-5

cm depth for both elements, and also the porewater concentrations follow each other

in both profiles. The Na concentrations in the solid sediment resemble the profiles of

Si and Al. For the concentrations in porewater, Na shows high variation between

summer and winter. While in summer, porewater and surface water concentrations for

Na are constantly close to 100 mg/l or below, the concentrations increase fivefold in

the surface water in winter. The porewater concentrations drop from fivefold

concentration in the sediment top to 168 mg/l at 11 cm depth.

Manganese shows a little higher concentration in winter in the sediment top than in

summer. The porewater has a Mn minimum in winter and lower concentrations than

the bottom-near water and surface water in the basin. In the deeper sediment below 5

14

cm, solid Mn has relatively high concentrations, while the porewater concentrations

are relatively low. From 3-5 cm the porewater reaches a relative peak for Mn in both

seasons where the sediment concentrations have decreased.

The Fe concentrations in the solid sediment have a relative peak in the sediment top

(3.4% DM summer; 3.8% DM winter), but the concentrations in the upper 5 cm are in

general lower than concentrations in the deeper sediment. Especially in the sediment

from the winter profile, Fe concentrations vary more.

The sulphur concentrations in the solid sediment have a peak at ca 3 cm sediment

depth. In the sediment deeper than 5 cm, the S concentrations are more than 90%

lower with the exception of a relative peak at 6.5 cm depth. Especially in winter, the

porewater profile matches the solid S profile. Sulphur peaks in porewater are placed

just about 1 cm above the peak in the solid sediment. Thus, the porewater

concentration increases in winter from the top (with concentration similar to the

surface water) until the peak at 2 cm depth. From there the concentration decreases

until the relative minimum at 3 cm depth, where the solid sediment has a peak. Below

3 cm the porewater concentration increases until 5-6 cm depth, from where the

concentration drops continuously with depth.

The metal concentrations of Cd, Co, Cr, Cu, Ni, and Zn in porewater at 1-2 cm depth

in winter have in common that they are lower than bottom-near water in the basin and

even lower than the surface water. For Cu and Zn, this is also observed in summer. A

concentration increase in the solid sediment at 5 to 3 cm depth upward is most evident

for Cu and Zn. Also Cd, Co, Cr, Ni, and Pb have higher concentrations in the solid

sediment in the upper sediment section (0-3 cm) than just below.

15

0 4 8 12 16Al2O3 , DM %

200

-15

-10

-5

0

5D

epth

in c

m

0 2000 4000 6000 8000 10000Al, mg/l

0 4 8 12 16Al2O3 , DM %

200

-15

-10

-5

0

5

0 2000 4000 6000 8000 10000Al, mg/l

summer winter

0 20 40 60 80SiO2, DM %

200

-15

-10

-5

0

5

Dep

th in

cm

0 5 10 15 20 25Si, mg/l

0 20 40 60 80SiO2, DM %

200

-15

-10

-5

0

5

0 5 10 15 20 25Si, mg/l

summer winter

0 1 2 3 4Na2O, DM %

200

-15

-10

-5

0

5

Dep

th in

cm

0 200 400 600Na, mg/l

0 1 2 3 4Na2O, DM %

200

-15

-10

-5

0

5

0 200 400 600Na, mg/l

summer winter

Figure 6. Al2O3, SiO2, Na2O in the stormwater basin sediment and Al, Si, Na in porewater and surface water (both 0.22 μm filtered).

surface water

solid sedimentporewatersediment surface

16

0 0.02 0.04 0.06 0.08MnO, DM %

200

-15

-10

-5

0

5

Dep

th in

cm

0 20 40 60 80 100Mn, µg/l

0 0.02 0.04 0.06 0.08MnO, DM %

200

-15

-10

-5

0

5

0 20 40 60 80 100Mn, µg/l

summer winter

0 2 4 6Fe2O3, DM %

200

-15

-10

-5

0

5

Dep

th in

cm

0 2 4 6 8Fe, mg/l

0 2 4 6Fe2O3, DM %

200

-15

-10

-5

0

5

0 2 4 6 8Fe, mg/l

summer winter

0 1000 2000 3000S, mg/kg DM

200

-15

-10

-5

0

5

Dep

th in

cm

0 20 40 60 80S, mg/l

0 1000 2000 3000S, mg/kg DM

200

-15

-10

-5

0

5

0 20 40 60 80S, mg/l

summer winter

Figure 7. MnO, Fe2O3, S in the stormwater basin sediment and Mn, Fe, S in porewater and surface water (both 0.22 μm filtered).

surface water

solid sedimentporewatersediment surface

17

0 0.1 0.2 0.3 0.4Cd, mg/kg DM

200

-15

-10

-5

0

5D

epth

in c

m

0 0.02 0.04 0.06 0.08Cd, µg/l

0 0.1 0.2 0.3 0.4Cd, mg/kg DM

200

-15

-10

-5

0

5

0 0.04 0.08 0.12Cd, µg/l

summer winter

0 4 8 12Co, mg/kg DM

200

-15

-10

-5

0

5

Dep

th in

cm

0 1 2 3 4Co, µg/l

0 4 8 12Co, mg/kg DM

200

-15

-10

-5

0

5

0 1 2 3 4Co, µg/l

summer winter

0 40 80 120 160Cr, mg/kg DM

200

-15

-10

-5

0

5

Dep

th in

cm

0 2 4 6 8 10Cr, µg/l

0 40 80 120 160Cr, mg/kg DM

200

-15

-10

-5

0

5

0 2 4 6 8 10Cr, µg/l

summer winter

Figure 8. Cd, Co, Cr in the stormwater basin sediment and Cd, Co, Cr in porewater and surface water (both 0.22 μm filtered).

surface water

solid sedimentporewatersediment surface

18

0 20 40 60 80 100Cu, mg/kg DM

200

-15

-10

-5

0

5

Dep

th in

cm

0 4 8 12 16 20Cu, µg/l

0 20 40 60 80 100Cu, mg/kg DM

200

-15

-10

-5

0

5

0 4 8 12 16 20Cu, µg/l

summer winter

0 4 8 12 16 20Ni, mg/kg DM

200

-15

-10

-5

0

5

Dep

th in

cm

0 2 4 6 8 10Ni, µg/l

0 4 8 12 16 20Ni, mg/kg DM

200

-15

-10

-5

0

5

0 2 4 6 8 10Ni, µg/l

summer winter

0 5 10 15 20 25Pb, mg/kg DM

200

-15

-10

-5

0

5

Dep

th in

cm

0 4 8 12 16 20Pb, µg/l

0 5 10 15 20 25Pb, mg/kg DM

200

-15

-10

-5

0

5

0 5 10 15 20 25Pb, µg/l

summer winter

Figure 9. Cu, Ni, Pb in the stormwater basin sediment and Cu, Ni, Pb in porewater and surface water (both 0.22 μm filtered).

surface water

solid sedimentporewatersediment surface

19

0 50 100 150 200 250Zn, mg/kg DM

200

-15

-10

-5

0

5D

epth

in c

m

0 10 20 30 40 50Zn, µg/l

0 50 100 150 200 250Zn, mg/kg DM

200

-15

-10

-5

0

5

0 20 40 60 80 100Zn, µg/l

summer winter

Figure 10. Zn in the stormwater basin sediment and Zn in porewater and surface water (both 0.22 μm filtered).

surface water

solid sedimentporewatersediment surface

20

PAH in sediment

The PAH concentrations in the stormwater basin sediment are in general higher at the

surface (0-2 cm) than in the deeper part (6-7 cm) (Table 2). According to the Swedish

EPA classification for coast sediments (Swedish EPA 2000), the sum of 11 PAHs are

on a moderate level at both sediment depths. Compared with an enclosed bay of the

Lule River, which receives stormwater from roads, industrial and residential areas, the

levels for each of the 11 PAHs are lower in the stormwater basin (Rentz, Widerlund et

al. 2010).

Table 2. Concentrations (μg kg-1) of 16 PAHs (^light PAH, ^^heavy PAH ) in the sediment from stormwater basin 2 (B2) compared with Skutviken (Rentz et al 2010), judged after the Swedish EPA (2000) guidelines for 11 PAHs in coast sediments (class 1, no content *; class 2, low content **; class 3, moderately high ***; class 4, high ****; class 5, very high *****). Half the value for limit of detection is used to calculate � PAH. B2 B2 Skutviken

Depth in cm 0-2 6-7 0-2

^PHEN 28*** 14*** 89****

^ANT 11**** <10 24****

^FLR 65*** 40*** 130****

^PYR 85**** 32*** 240*****

^BaA 31*** 21*** 59****

^CHY 35*** 21*** 69****

^^BbF 65*** 28** 180****

^^BkF 31*** 14** 44***

^^BaP 53*** 25*** 74****

^^BPY 13** 37*** 89***

^^INP 80*** 37** 99***

� 11 PAH 497*** 464*** 1097****

^^DBA 26 10 30

^NAP 82 <10 39

^ACY <10 <10 11

^ACE <10 <10 <10

^FL <10 <10 15

� 16 PAH 620 513 1197

21

Discussion

The lower oxygenation of the water column in winter is a consequence of ice covering

and reduced inflow. The lower oxygen concentration in the water column in winter

seems not to have a direct effect on redox conditions in the water column and the

sediment near the water sediment interface. The seasonal variation in total

concentrations of Cd, Cr, Co, Cu, Ni, Pb and Zn in water from summer 2009 to winter

2010 (Table 1) may be a consequence of road salt applied as a de-icing agent and

increased street wear due to use of studded tires in winter. This becomes clear looking

at the related speciation diagrams, which show that the higher winter concentrations

are to a large extent caused by higher truly dissolved concentrations of Cr, Co, Cu, Ni,

and Zn. Compared with Swedish EPA guidelines for estimation of current conditions

of metals in freshwater (Swedish EPA 2000), the winter concentrations of Cd, Cu, and

Zn show moderate high concentrations (risk class 3) while Cr, Ni, and Pb reach low

concentrations (risk class 2). In summer these elements have concentrations which are

classified as low or very low (risk class 2 or 1). Previous studies have shown the

relation that use of de-icing agents in combination with use of studded tires results in

higher metal concentrations in road runoff (Hvitved-Jacobsen and Yousef 1991;

Legret and Pagotto 1999; Bäckström, Nilsson et al. 2003). Even if seasonal variation

in the metal concentrations is in accordance with other studies, the total

concentrations measured in Basin 2 are low in comparison with Legret and Pagotto

(1999), Bäckström, Nilsson et al. (2003) and Karlsson et al. (2010).

Inflow occurs in cold winters , mostly due to use of road salt lowering the melting

point of water. The higher concentrations of Ca, K, Mg, Mn, Na and S in the water

column can be caused by reduced (flush) runoff through snow removal and temporary

water storage as snow and ice on the roadsides. In particular, the use of road salt

(NaCl) increases the Na concentration in wintertime. The low runoff and, especially

runoff in the phase of early melting, mainly transports the truly dissolved elements

(Oberts, Marsalek et al. 2000). The salt use and insufficient combustion in cold

climate and combustion products which get enriched in snow layers, may result in

higher S concentration in winter, but compared with main road and motorway runoff

the concentrations in the water are low (Göbel, Dierkes et al. 2007).

The concentration change for most elements at a depth of 3-5 cm shows the boundary

between the collected stormwater sediment and the macadam ground of the

22

constructed stormwater basin. Especially the Al, Si and Na concentrations (Fig. 6) in

the solid sediment indicate higher feldspar concentration in the macadam ground. The

high LOI content in the sediment is ascribed to settling of stormwater-transported

organic particles, algae in the basin and local vegetation. The clear change in sediment

composition allows the estimation that the upper 3.5 cm of the sediment have settled

since the stormwater facility was in use in 1998. That would result in an annual

sedimentation rate of approximately 3 mm at the sampling point. The high Na

concentration in the surface water in winter affects the porewater, but the Na

concentrations in the sediment phase do not vary. High elemental concentrations in

the water column implicate possible diffusion into the sediment via porewater. This is

the case for higher Mn concentration in the water column in winter (Fig. 7). In the

sediment surface (0-1 cm) Mn shows enrichment in the solid phase in winter. At the

same level the porewater has a relatively low Mn concentration, which rises with

depth until about 3 cm. At that depth the solid Mn shows a relatively low

concentration. The peak of solid Mn at the sediment surface in winter indicates oxic

conditions with formation of Mn oxyhydroxides (Davison 1993). Due to

decomposition of organic material, conditions in the sediment become more anoxic

with depth. This results in reduction of Mn oxyhydroxides and increased porewater

concentration of Mn(II) until 3-4 cm depth. In the upper oxic parts of the sediment

Mn(II) is oxidised to Mn(IV). In the sediment below 4 cm the Mn concentration is

determined by the relative high Mn contents in the macadam bottom. Via porewater,

Mn diffusion occurs from a sediment level with high content at 6.6 cm depth) to

levels above and below, where Mn in solid phase is relatively depleted.

The relative peak of Fe in the solid phase (Fig. 7) at the sediment surface in winter

also indicates that Fe-oxyhydroxides have formed, but low Fe concentration in the

surface water and porewater do not show dynamics at the sediment water interface.

This suggests that Fe reaches the sediment mostly in particulate form, which is

supported by the Fe speciation in the water column (Fig. 3). However, for S an S-

enriched layer at 3 cm depth has formed, which indicates precipitation of solid

sulphides in the stormwater sediment just above the border to macadam. At the same

depth depletion in the porewater concentration of Fe can be observed, which indicates

Fe-sulphide formation (Fortin, Leppard et al. 1993). The porewater shows an increase

of S concentration just above and below the peak of solid S. The lower peak in 5 cm

depth results from S-release of the macadam ground. The relatively high porewater

23

concentrations of S in the macadam are above the surface-water concentrations and

prompt an upward diffusion of S in the porewater. When the S in porewater reaches

the layer of the organic rich stormwater sediment the contamination decreases. In the

stormwater sediment sulphate reduction is possible and a higher content of organic

material (LOI, Fig. 5) offers precipitation surfaces. The LOI concentration in the

stormwater sediment creates beneficial conditions for sulphate-reducing bacteria

(SRB). Fortin (2000) suggests that SRB can be either oxygen-tolerant or live in

anoxic microenvironments within oxic sediments. Most likely, the oxic-anoxic border

in the sediment is located in 1-3 cm depth, interpreting the Fe-Mn-S profiles (Fig. 7).

The activity of SRB can even accomplish S-reduction in colder months (Fortin,

Goulet et al. 2000), but the increase in S in the porewater above 3 cm depth in winter

suggests that not all S can be bound in sulphates. This may be explained by less SRB

activity in winter.

For the metals Cd, Co, Cr, Cu, Ni, and Zn the concentrations in porewater at 1-2 cm

depth in winter are lower than in the surface water, which means that diffusion of

these elements into the sediment is likely. The organic material offers precipitation

surfaces and coating on Mn- and Fe-hydroxides or bonding under anoxic conditions

on sulphates is most likely. That causes enrichment of these metals in the stormwater

sediment (Fig. 8, 9, 10). Compared with stormwater pond sediment from Karlsson et

al. (2010) the concentrations in the stormwater sediment at B2 are in general lower for

Cd, Cu, Pb, and Zn, while Cr and Ni are in the same range.

The PAH concentrations in the sediment of B2 show that stormwater impacts the

relatively higher concentrations at the surface (0-2 cm) compared with the deeper part

(6-7 cm). The fact that there is a filter for organic pollutants installed at the pump

station before reaching the inlet for water from Häggviksleden can have a positive

effect on the PAH contamination. The sum of 11 PAHs is on a moderate level,

according to Swedish EPA (2000). This is also suggested by comparison with an

enclosed bay of the Lule River (Rentz, Widerlund et al. 2010), which receives

untreated stormwater from roads, industrial and residential areas.

24

Conclusions

In this study elevated heavy metal concentrations in the water column of a stormwater

basin and elevated heavy metal concentrations and PAH concentrations in surface

sediment of the stormwater basin were found. Seasonal variations in element

concentrations are most evident for the elements Cd, Co, Cr, Cu, Fe, Mn, Na, Ni, Pb,

S, and Zn in the water column. Especially in winter, the metal concentrations of Co,

Cr, Cu, Ni, Mn, Na, and Zn are dominated by the truly dissolved phase. Most of the

dissolved concentrations are supposedly transported further on leaving the stormwater

facility. A technical solution could be the application of a peat-filter to bind metal

cations. A fraction of the concentrations of metals in truly dissolved phase can also

diffuse into the sediment. The precipitation on organic material, and early diagenesis

processes with formation of Mn- and Fe-hydroxides and sulphide reduction are able to

trap pollutants. The PAH contamination in the sediment is relative low, most likely

due to filtration of incoming stormwater.

Acknowledgments

This study was financed by Luleå University of Technology. This support is gratefully

acknowledged. We also thank our colleague Fredrik Nordblad for his assistance in

Luleå. For contribution of information regarding the stormwater basin facilities we

thank Magnus Billerberger and Martin Larsson (Vägverket), Joakim Börefeldt (YIT),

Ann-Christine Granfors (Sollentuna Kommun), and Anne Hafez (Trafikverket).

25

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DOCTORA L T H E S I S

Department of Civil, Environmental and Natural Resources EngineeringDivision of Geosciences and Environmental Engineering

Water and Sediment Quality of Urban Water Bodies

in Cold Climates

Ralf Rentz

ISSN: 1402-1544 ISBN 978-91-7439-272-2

Luleå University of Technology 2011

ISSN: 1402-1544 ISBN 978-91-7439-XXX-X Se i listan och fyll i siffror där kryssen är

Ralf R

entz Water and Sedim

ent Quality of U

rban Water B

odies in Cold C

limates