Distribution and Sources of Polycyclic Aromatic Hydrocarbons
in Sediments, Suspended Particulate Matter and Waters from
the Siak River System, Estuary and Coastal Area of Sumatra,
Indonesia
A dissertation submitted for the degree of
- Doktor der Naturwissenschaften -
(Dr. rer. nat.)
at the Faculty of Biology/Chemistry
the University of Bremen, Germany
Presented by:
Muhammad Lukman
Distribution and Sources of Polycyclic Aromatic Hydrocarbons in
Sediments, Suspended Particulate Matter and Waters from the Siak River
System, Estuary and Coastal Area of Sumatra, Indonesia
A dissertation submitted for the degree of Doktor der Naturwissenschaften (Dr. rer. nat.) at the Faculty of Biology/Chemistry, the University of Bremen, Germany
Presented by:
Muhammad Lukman
Referees : 1. Professor Dr. Wolfgang Balzer
2. Professor Dr. Wolfram Thiemann
Examiners : 1. Professor Dr. Gerhard Kattner
2. Professor Dr. Venugopalan Ittekkot
ACKNOWLEDGMENTS
First of all, I would like to express my sincerely and great gratitude to Prof. Dr. Wolfgang Balzer (FB 2, Marine Chemistry, University of Bremen) for his remarkable role in pouring me with lots of insight, motivation, and supervision throughout my PhD work. I do very much appreciate for his willingness to offer and provide me the possibility to do PhD in his working group as well as to join the SPICE Cluster 3.1. (Science for the Protection of Indonesian Coastal Ecosystem) Project in Riau, Sumatra, Indonesia. Secondly, I am grateful to DAAD (Deutscher Akademischer Austausch Dienst or German Academic Exchange Service) in providing me a great support to do my PhD in Germany during the period of 2004 – 2007.
I would like to thank Prof. Dr. Wolfram Thiemann as the second referee, Prof. Dr. Gerhard Kattner and Prof. Dr. Venugopalan Ittekkot as the examiners, and all the colleagues and those previous colleagues in the marine chemistry working group, University of Bremen: Dr. Uwe Schüßler, Dr. Wolfgang Barkmann, Immo Becker, Olaf Wilhelm, Timo Daberkow, Xiaoliang Tang, Jun Fu, Björn Bach, Dominique Schobes, Sonia Tambou for their valuable assistances and discussion in laboratory and analytical aspects, as well as Mrs. Ute Wolpmann who faithfully helps me with administrative matters during my stay in Bremen.
I would like to extent my appreciation to all my SPICE Cluster 3.1. colleagues: Dr. Tim Rixen, Dr. Antje Baum (ZMT Bremen), Dr. Herbert Siegel (IOW Warnemuende), Dr. Thomas Pohlmann (University of Hamburg), Dr. Ralf Woestmann (Terramare, Wilhemshaven), and to ZMT staffs Dorothee Dasbach and Matthias Birkicht for all their countless assistances, advices and critics, as well as to Nathan Giles for improving the English.
Also, I would like to thank to Prof. Dr. Gerd Liebezeit (Terramare, University of Oldenburg) to all insights, advices and discussion. Also, to all Indonesian SPICE colleagues in University of Riau, Riau during sampling campaigns, particularly Dr. Joko Samiaji, Dr. Christine Jose, Dewi Kristina, friends at Hasanuddin University, Makassar, and many others. I thank you so much for your supports.
Last but not least, very special thanks I dedicate to my wife, Rahmawati Yusuf, to my family – my Mother, Father, Sister and Brothers -, and to my all relatives - who always inspire me during hard time. Finally, I would like to dedicate this work to my country Indonesia and to those who are fond of better environment.
Kurzfassung
Die vorliegende Arbeit untersucht Ursprung und Verteilung von Polycyclischen Aromatischen Kohlenwasserstoffen (PAKs) als Indikator für anthropogene Verschmutzung in den Küsten- und Flussregionen der Insel Sumatra in Indonesien. Im Vordergrund steht dabei die Analyse der 16 PAK-Prioritätsverbindungen, von denen Referenzmaterial gemäß der USA-EPA priority pollutants Liste vorliegt. Untersucht wurden Proben in der Lösung und von Oberflächensedimenten und Schwebstoffen (SPM) des Siak-Flusses, seiner Flussmündung und des Riau Küstenbereichs in Sumatra.
Die Quantifizierung der PAKs wurde unter Einsatz eines Hochleistungs-Flüssigkeits-Chromatographen (RP-C18-HPLC) mit UV- und Fluoreszenz-Detektion durchgeführt. Die Analysenmethode beinhaltete eine Reihe von Probenahmetechniken für die individuellen Phasen (Sediment, SPM, Lösung), Probenaufbereitung, Extraktion, Aufarbeitung und HPLC-Quantifizierung. Die Untersuchung der PAKs im Sediment konzentrierte sich auf zwei Größenklassen: Grobfraktion (Sand) 2 mm – 63 μm und Feinfraktion (Schlick) < 63 μm. Die Schwebstoffe wurden über 0,7 μm Glasfaserfilter (GF/F) herausgefiltert. Die PAKs der Lösung wurden dann mittels eines Octadecyl Festphasen Extraktionssystems (SPE) extrahiert. Die Qualitätskontrolle beinhaltete den Gebrauch von Blindwerten und Ersatzstandards, um Genauigkeit und Effizienz der Analyse und die Reproduzierbarkeit der Ergebnisse zu gewährleisten. Die Aufteilung in die verschiedenen Stoffquellen der PAK-Verbindungen wurde unter Einbeziehung bekannter Indexe der Molekulargewichte und spezifischen Isomer-Verhältnissen ausgeführt.
Die Untersuchungen ergaben, dass sowohl der Flusslauf als auch Flussmündung und Küstenbereich des Siak-Flusses erheblich mit PAKs belastet sind. Die Untersuchungsergebnisse weisen auf kräftige pyrogene Stoffquellen hin, insbesondere auf die Verbrennung von Biomasse und Erdöl. Sie können daher als Nachweis für großräumiges, länger anhaltendes und intensives Verbrennen landwirtschaftlicher Nutzflächen im Zusammenhang mit kräftigen Wald- und Torf-Feuern, die über die letzten Jahrzehnte stattfanden, angesehen werden. In diesen von Buschfeuern heimgesuchten Gegenden bilden die PAK-Verteilungen zwischen Grob- und Feinfraktion an der Küste und in der Flussmündung ein deutliches Muster, das von den Verteilungen üblicher Küstenbereiche deutlich abweicht. Ein Vergleich der PAK-Verteilungen in den beiden Größenfraktionen in den Sedimenten der Siak-Küste mit den Küstenbereichen von Wenchang und Wanquan in China deutet darauf hin, dass die PAKs der Küstengewässer um Sumatra hauptsächlich mit den hohen kohleartigen Materialien wie Ruß und verbranntem Torf assoziiert sind. Wie die Untersuchungen zeigen, können auch andere relevante Stoffquellen wie andauernde Erdölverschmutzung in den Gewässern um die Städte, in den industriellen Vororten von Perawang, der Ölstadt Dumai und der Erdölraffinerien im Gebiet der Flussmündung für die Belastung mit PAKs verantwortlich sein.
Als Zusammenfassung der Ergebnisse der Einzeluntersuchungen wurden die folgenden drei Manuskripte erstellt, die an begutachtete wissenschaftliche Zeitschriften zu versenden sind.
PAKs im Sediment (Manuskript – I, Kapitel IV)
Die PAK-Gehalte (Summe der 16 Standard-Verbindungen) in der Sedimentfraktion aller beprobten Gebiete bewegen sich zwischen 0,13 und 5,47 μg/g Trockengewicht (TG). In der Grobfraktion (Sand) wurden mit Werten zwischen 0,16 bis 5,47 μg/g TG (median m = 0,84) weitaus höhere Konzentrationen gefunden (etwa um einen Faktor 2) als in der Feinfraktion (Schlick) mit Werten zwischen 0,13 und 1,31 μg/g TG (m = 0,52). In der Grobfraktion ist die Anreicherung unerwartet, da diese in der Regel wegen der größeren Oberfläche pro Masseneinheit in der Feinfraktion zu erwarten ist. Ein ähnliches Muster wurde für das organische Material beobachtet. So variiert der Anteil des organischen Kohlenstoffs an der Gesamtmasse der Grobfraktion zwischen 0,01% und 24%, während
sich der Anteil in der Feinfraktion zwischen 0,34% und 3,7% bewegt. Ebenfalls war ein nahezu linearer Zusammenhang zwischen PAK und organischem Kohlenstoff nur in der Grobfraktion zu erkennen. Aus diesen Untersuchungsergebnissen kann geschlossen werden, dass eine bestimmte Sorte organischen Materials für die Affinität zwischen Kohlenstoff und PAKs verantwortlich ist, nämlich vaskuläre Pflanzenreste, Torf und Ruß, wie auch in ähnlichen Untersuchungen festgestellt wurde. Entlang des Flusslaufes in Richtung Flussmündung konnte kein klares Muster in den PAK-Gehalten festgestellt werden, auffallend sind nur die Anreicherungen in den urbanen und industriellen Gebieten. Die weitgehend hohen Molekulargewichte und die Molekularverhältnisse lassen auf pyrogene Spurenstoffquellen schließen, insbesondere auf die Verbrennung von Biomasse und Erdöl. Die PAK-Gehalte gelangen daher über den Land- und Luftweg in die Gewässer.
PAKs in der Lösung und in den Schwebstoffen (Manuskript – II, Kapitel V)
Die gemessenen PAK-Gehalte in der Lösung bewegen sich zwischen 0,13 und 5,14 μg/L im Flusslauf, zwischen 0,32 und 0,62 μg/L in dem Ästuar und zwischen 0,12 und 0,13 μg/L im Küstenbereich. In Richtung Küste nimmt die mittlere Konzentration um einen Faktor 3 ab. Die höchsten PAK-Gehalte wurden an der Einmündung des Mandau Flusses in den Siak Fluss gemessen. Die PAKs wurden durch 2-, 3- und 4-Ring-Aromate dominiert. Die PAK-Gehalte in den Schwebstoffen variieren zwischen 1,48 bis 59.1 μg/g im Flusslauf, zwischen 0,16 und 7,67 μg/g in der Flussmündung und zwischen 0,33 und 10,2 μg/g in der Küstenregion. Auf das Volumen bezogen bewegen sich die PAK-Gehalte im SPM jeweils zwischen 0,06 und 0,69 μg/L, 0,03 und 0,29 μg/L und zwischen 0,01 und 0,15 μg/L. Die PAK-Gehalte nehmen im Allgemeinen in Richtung Küste ab, was auf eine Ablagerung im Sediment und/oder Verdünnung mit Seewasser zurückzuführen ist. Eine Anreicherung mit PAKs findet sowohl in der Trockenzeit als auch in der Regenzeit statt, die durch verschiedene Ring-Größen gekennzeichnet sind. Ebenfalls lassen sich dadurch die verschiedenen Transportwege der PAKs erschließen. Die Anreicherungen der PAKs in der Regenzeit können durch eine Zunahme des Oberflächenabflusses entstehen, während in der Trockenzeit die Anreicherung wahrscheinlich durch die Atmosphäre stattfindet.
Eine Partitionierung der PAKs zwischen Schwebstoffen und Lösung wurde durchgeführt, um ein besseres Verständnis des Verlaufs der Anreicherung dieser Schadstoffe in den Gewässern zu bekommen. Die gemessenen Partitionskoeffizienten (KD) zeigen eine beträchtliche Streuung zwischen
den beprobten Stationen. Der mittlere KD-Wert im Flusslauf und in der Flussmündung variiert
zwischen 4 und 5 auf der logarithmischen Skala, an der Küste hat er dann einen Wert von 6 erreicht. Normalisiert auf den organischen Kohlenstoff bewegt sich der Partitionskoeffizient (KOC) zwischen 2
und 4 im Fluss und an der Mündung und zwischen 3 und 5 an der Küste. Diese Variation deutet auf eine unterschiedliche Qualität des partikulären organischen Materials hin, in der Rußablagerungen eine bedeutende Rolle spielen können. Anscheinend erhöht die Zunahme des Salzgehalts und eventuell des pH-Wertes den Koeffizienten KD. Auf der anderen Seite hat die Zunahme des gelösten
organischen Kohlenstoffs (DOC) einen umgekehrten Effekt. Dies deutet auf eine wichtige Rolle des DOC für die Erhaltung schwerer PAKs in der Lösung hin, was auch eine Erleichterung ihres Transports beinhaltet.
Ein Vergleich zwischen den Messungen am Siak-Ästuar (torfhaltiger Boden) und denen an den Wenchang und Wanquan Flussmündungen (sandiger Boden) (Manuskript – III, Kapitel VI)
Die Verteilung der PAK-Konzentrationen in den Sedimentfraktionen der Siak-Flussmündung und des Küstengebietes wurde mit den Messungen an den Wenchang (WW)- und Wanquan (WQ)-Flussmündungen in Hainan, China verglichen. Die Gebiete zeichnen sich durch ihre unterschiedliche Bodenformation aus: torfhaltig am Siak-Fluss und eher mineralisch oder huminstoffarmer Boden auf Hainan. Wie oben erwähnt, sind die PAK-Konzentrationen in den Siak-Sedimenten durch einen hohen Anteil von PAKs und organischem Material in der Grobfraktion charakterisiert. Dieses Muster wurde nicht in den WW/WQ-Sedimenten gefunden, in denen die PAKs hauptsächlich in der Feinfraktion zu finden sind. Die gemessenen PAK-Konzentrationen der 15 US EPA priority pollutants (ausgenommen Acenaphthylen) bewegen sich in der Siak-Flussmündung zwischen 0,13 μg/g TG und
1,83 μg/g TG (median m = 0,69 μg/g TG) in der Grobfraktion und zwischen 0,09 μg/g TG und 0,43 μg/g TG (m = 0,20 μg/g TG) in der Feinfraktion. Dagegen wurden PAK-Werte in der Grobfraktion der WW/WQ-Sedimente zwischen 0,11 μg/g TG und 0,39 μg/g TG (m = 0,19 μg/g TG) und in der Feinfraktion zwischen 0,09 μg/g TG und 0,68 μg/g TG (m = 0,47 μg/g TG) gefunden. Die relativ hohe PAK-Konzentration in der Grobfraktion der torfreichen Bodenformation von Sumatra konnte den hohen Konzentrationen der kohlenstoffhaltigen Materialien wie Ruß, Torf und Pflanzenablagerungen zugeordnet werden, welche zur Anreicherung der PAKs an den Materialien beitragen. In den WW/WQ-Sedimenten wurden diese hohen Kohlenstoffkonzentrationen nicht gefunden. In beiden Gebieten gab es keine Unterschiede in den Molekularverteilungen zwischen Grob- und Feinfraktion, was als Hinweis gewertet werden kann, dass die PAK-Verunreinigungen von ähnlichen Quellen stammen. Dagegen deuten die Isomerenverhältnisse, die für die Quellenbestimmung hinzugezogen wurden, darauf hin, dass die PAKs der Siak-Flussmündung durch Buschfeuer und ähnliche Verbrennungen entstanden sind, wohingegen die PAKs der WW/WQ-Sedimente aus der Kohle- und Ölverbrennung stammen. Die Konzentration der gelösten PAKs in den küstennahen Ästuaren von WW/WQ war relativ gering. Die PAKs waren zwischen 7,36 und 16,2 ng L-1. Im Gegensatz dazu war die Konzentration im Siak-Ästuar sowie an der Küste erheblich höher. Sie betrug zwischen 121 und 619 ng L-1. Huminstoffe in den überliegenden Wasserschichten spielen ebenfalls eine Rolle für die Verteilung der PAKs in den Sedimentfraktionen. So enthält die Siak-Flussmündung und die Küstenregion signifikante Mengen an DOC, die in den beprobten WW/WQ-Gebieten nicht gefunden wurden (Balzer, unveröffentlichte Daten). Die Verteilungskoeffizienten Sediment-Wasser der einzelnen PAKs im Siak-Ästuar ergaben Werte (auf der logarithmischen Skala) zwischen 2,22 und 5,58 (Median m = 3,37) für die Grobfraktion und zwischen 1,53 und 5,03 (m = 3,01) in der Feinfraktion. So liegen die KD-Werte im Siak-Ästuar niedriger als in den
WW/WQ-Ästuaren, zwischen 1,12 und 5,89 (m = 3,93) für die Grobfraktion und zwischen 2,58 und 5,85 (m = 4,40) für die Feinfraktion. Dies lässt vermuten, dass die hohe DOC-Konzentration in der Wassersäule des Siak eine Adsorption an die organischen Sedimentbestandteile behindert.
SUMMARY
This study examined the overall distribution and sources of polycyclic aromatic hydrocarbons (PAHs), particularly the 16 parent PAHs of the US Environmental Protection Agency priority pollutant, as an indicator for anthropogenic pollution, in the surface sediments, suspended particulate matter (SPM) and water solution of the Siak river system, its estuary and the Riau coast, Sumatra, Indonesia.
The PAHs were determined by high performance liquid chromatography with reverse phase octadecyl column (RP-C18-HPLC) using ultraviolet and programmable fluorescence detectors. Method analysis included various sampling techniques for the individual phases, sample preparation, extraction, work-up procedures and HPLC quantification. Analysis of PAHs in sediment was focused on the content distribution among two size-fractions: sand/coarse (2 mm - 63 m) and mud/fine (< 63 m). The particulate PAHs were those embedded in suspended materials retained by 0.7 μm glass fiber filter (GF/F). Dissolved PAHs were obtained from the filtered water-solution, using an octadecyl solid phase extraction (SPE) system. Quality control measures included the use of procedural blanks and surrogate standards in order to optimize and validate procedural accuracy, efficiency and the reproducibility of results. Source apportionment of PAHs was carried out by applying existing indices of molecular weights and specific isomer ratios.
In general, the results show that PAHs significantly impact the Siak river, the estuary and the coastal waters. Source apportionment indicated intense signatures of pyrogenic sources, particularly biomass burnings and petroleum combustion. The results might be evidence of the effects of widespread, long-term and intense agricultural burnings coupled with multitudinous forest/peat swamp fires which have occurred frequently over the last decades. In such burning-affected estuaries and coastal waters, distribution of PAH between the size-fraction in sediments showed distinctive patterns to those of other coastal areas. Comparison of distribution of PAH in the coarse and fine fractions between the Siak Sumatra and the Wenchang and Wanquan coastal estuaries of Hainan China indicates that PAH transferred to the coastal waters of Sumatra were mostly associated with high carbonaceous materials – the burning product particles – such as black carbon and peat. However, the apportionment also showed that another relevant source of PAHs was chronic petroleum pollution centred in the waters around cities, the industrial estates of Perawang, oil city of Dumai, and the oil refinery located in the estuary area.
Summarizing the different results the following three manuscripts were produced which are sent to scientific journals with a referee system.
Sedimentary PAHs (Manuscript – I, Chapter IV)
The PAHs for the sedimentary fractions in all sampled areas ranged between values of 0.13 to 5.47 μg g-1 dry weight (d.w.) sediment. The PAHs in the sand fraction ranged from 0.16 to 5.47 μg g-1 d.w. (median m = 0.84). In general, the sand fraction contained PAH levels higher by a factor of ±2 as compared to those found in the mud fraction that ranged from 0.13 to 1.31 μg g-1 d.w. (m = 0.52). The enrichment of PAHs in the sand fraction was quite astonishing, especially since we assumed that the fine fraction would generally evidence much higher levels of contaminants, due to its large surface area per unit mass for adsorption. The same pattern of enrichment was shown by the organic matters. The organic carbon (OC) contents varied greatly from 0.01% to 24% in the sand, but only slightly in the mud from 0.34% to 3.70%. A linear relationship between the PAH and
the OC was shown only by the sand fraction. These all evidences lead us to the assumption that a specific kind of organic matter should be responsible for high affinity for PAHs. Thus, it is most likely materials such as black carbon, vascular plant debris, and peat as figured out by many other studies. The spatial distribution showed no clear pattern in distance to the river mouth. But, increased content of the PAHs was centred at urban and industrial areas. The high molecular weight compounds were widespread predominant and the molecular ratios for source apportionment provide further evidences for pyrogenic sources: biomass and petroleum combustions. PAHs were delivered to those aquatic systems by both land-water and air-water transports.
Dissolved and Particulate PAHs (Manuscript – II, Chapter V)
Dissolved PAHs ranged from 0.13 to 5.14 μg L-1, 0.32 to 0.62 μg L-1, and 0.12 to 0.13 μg L-1 in the Siak River, its estuary and the coastal areas, respectively. The mean concentration decreased by a factor of 3 towards the coast. The highest concentration was observed in the water at the confluence of the black water Mandau River, a Siak tributary. The PAHs were dominated by 2-, 3-, and 4-ring compounds. The PAHs in the SPM varied greatly from 1.48 to 59.1 μg g-1, 0.16 to 7.67 μg g-1, and 0.33 to 10.2 μg g-1 for the Siak river, its estuary and the coast, respectively. In volume basis, the concentration of particulate PAHs ranged from 0.06 to 0.69 μg L-1, 0.03 to 0.29 μg L-1, and 0.01 to 0.15 μg L-1 in the River, the estuary and the coast, respectively. The PAHs generally decreased towards the coast suggesting an entrapment and/or dilution effect of sea water. PAH enrichment occurred in both wet and dry seasons characterized by different ring-size dominance. It suggests different mode of transport by which PAH were integrated into the aquatic environments. The accumulation of PAHs in the river during rainy season could be attributed to an increasing land-water surface runoff. Meanwhile, in the dry season, the enrichment was most likely caused by atmospheric deposition.
Partitioning of the PAHs between SPM and water solution was evaluated to understand the fate of these contaminants in the given aquatic systems. Measured partition coefficient (KD) showed
a considerable variation between the sampling locations. The mean KD values in the River and the
estuary ranged from 4 to 5 on the logarithmic scale, while in the coast was 6. The mean values of organic-carbon normalized partition coefficient (KOC) ranged from 2 to 4 on the log scale in the
River and the estuary, whereas from 3 to 5 in the coast. This variation suggests a different quality of particulate organic matter, in which black carbon might play a significant role. The increase in salinity and possibly also in the pH apparently turns to increase the KD, but enriched DOC affects
negatively. It indicates an important role of DOC in sustaining heavier PAHs in the dissolved phase, including a facilitation of their transport.
A comparison between peatland aquatic system of Siak Estuary and a non-peatland aquatic system of Wenchang and Wanquan Coastal Estuaries (Manuscript – III, Chapter VI)
Distribution of PAHs in the two grain-size fractions of the Siak Estuary and the Coast was compared with those from non-peatland systems of Wenchang and Wanquan (WW/WQ) Coastal Estuaries, Hainan, China. As described earlier, the PAHs in the Siak sediments were generally characterized by high content of PAHs and organic matter in the coarse fraction. This kind of distribution was not confirmed by the WW/WQ sediments which in most cases PAHs enriched in the fine sediments. The level of the total 15 US EPA priority pollutants excluding acenaphthylene in
the Siak estuary and the coast ranged from 0.13 μg/g d.w. to 1.83 μg/g d.w. (median m = 0.69 μg/g d.w.) in the coarse fraction, and from 0.09 μg/g d.w. to 0.43 μg/g d.w. (m = 0.20 μg/g d.w.) in the fine fraction, while the the WW/WQ sediments PAHs ranged from 0.11 μg/g to 0.39 μg/g d.w. (median m = 0.19 μg/g d.w.), and from 0.09 μg/g to 0.68 μg/g d.w. (m = 0.47 μg/g d.w.) in the coarse and the fine fractions, respectively. The high content of PAHs in the coarse fraction of the degraded peatland system of Sumatra was due to the existence of high carbon content carbonaceous materials i.e. black carbon, peat, and plant debris acting as strong sorbents for PAHs, which were not found in the WW/WQ sediments of Hainan. In both compared areas, there were no differences in molecular distribution between the fractions suggesting that PAH contamination stemmed from similar sources. Furthermore, the isomeric ratios used for source apportionment indicated that the PAHs found in the Siak basin had mostly been generated through biomass burnings, whereas PAHs analyzed in the WW/WQ sediments from Hainan Island stemmed from a mixture of coal and petroleum combustion. The concentration of dissolved PAHs in WW/WQ coastal estuary was relatively low. The PAHs ranged from 7.36 to 16.2 ng L-1. In contrast, the level of the PAHs dissolved in the Siak estuary and the coast ranged from 121 to 619 ng L-1. Humic substance in the overlaying water plays a role in distribution of PAH in the sediment fractions. The Siak estuary and the coast contained significant amount of DOC compared to the Wenchang and Wanquan coastal estuaries (Balzer, unpublished data). DOC in the Siak water may sustain the PAHs in the water column impeding their association onto the sediment organic matter as shown by sediment-water distribution coefficient (KD). The sediment-water distribution coefficients (KD) values of the
individual PAH in the Siak estuary ranged in logarithmic value from 2.22 to 5.58 (median m = 3.37) for the coarse fraction, and from 1.53 to 5.03 (m = 3.01) for the fine fraction. The KD values in the
Siak estuary are generally lower than those of WW/WQ estuaries, which greatly ranged from 1.12 to 5.89 (m = 3.93) in the coarse fraction, and from 2.58 to 5.85 (m = 4.40) in the fine fraction. It suggests that high DOC in the Siak water may sustain the PAHs in the water column impeding their association with the sedimentary organic matter.
TABLE OF CONTENTS
I. INTRODUCTION ... 1
1.1. Polycyclic aromatic hydrocarbons: definition and physio-chemical characteristics ... 2
1.1.1. Definition ... 2
1.1.2. Selected physio-chemical characteristics ... 3
1.2. PAH contamination in the aquatic environments: background to environmental problems 5 1.2.1. Toxicity of PAHs ... 5
1.2.2. Bioconcentration and magnification in aquatic food webs ... 7
1.2.3. Persistence, low degradation rates and pollution indicators ... 9
1.3. The relevance of rivers, estuaries and coastal areas in Indonesia ... 10
1.3.1. A Perspective on the global pollution dispersal ... 10
1.3.2. Environmental settings of the study areas: An Overview ... 13
1.4. Study Objectives ... 19
II. SOURCES, DISTRIBUTION AND FATE OF PAHs IN AQUATIC COMPARTMENTS: A SHORT REVIEW ... 20
2.1. Introduction ... 20
2.2. Source and Signatures ... 20
2.2.1. Natural PAHs ... 20
2.2.2. Pyrogenic PAHs ... 22
2.2.3. Petrogenic PAHs ... 29
2.2.4. Source Apportionment ... 32
2.3. Distribution in Aquatic Compartments ... 39
2.3.1. Surface Sediment and Grain Size Fractions ... 39
2.3.2. Suspended Particulate Matter and Water ... 42
2.3.3. Water Solution as dissolved PAHs ... 43
2.4. The fate of PAHs in the water: a partitioning concept and the role of natural organic matter ... 44
III. METHODS OF ANALYSIS ... 47
3.1. Introduction ... 47
3.2. Sample Collection and Treatments for PAH ... 47
3.2.1. Surface Sediment and Size Fractionation ... 47
3.2.2. Suspended Particulate Matter (SPM) ... 48
3.2.3. Solid phase extraction (SPE) for pre-concentration of dissolved PAHs ... 48
3.3. Determination of polycyclic aromatic hydrocarbons using high performance liquid chromatography coupled with ultraviolet and fluorescence detectors (HPLC UV/FLD) .. 50
3.3.1. Soxhlet Extraction of sediment and SPM ... 50
3.3.2. Extract Working-Up ... 51
3.3.3. Elution of the SPE Cartridges for dissolved PAHs ... 52
3.3.4. PAH determination: High Performance Liquid Chromatography with ultraviolet and fluorescence detectors (HPLC UV/FLDs) ... 53
3.4. Quality Controls ... 57
References (Chapter I – III) ... 59
IV. DISTRIBUTION AND SOURCE OF POLYCYCLIC AROMATIC HYDROCARBONS (PAHs) IN SURFACE SEDIMENTS FROM THE SIAK RIVER, ITS ESTUARY AND THE ADJACENT COASTAL AREA OF RIAU PROVINCE, INDONESIA ... 69
Abstract ... 69
4.1. Introduction ... 70
4.2.1. Study Area & sampling locations ... 71
4.2.2. Sample Collection and Fractionation ... 72
4.2.3. Analytical Methods ... 72
4.3. Results & Discussion ... 74
4.3.1. Geochemistry of sediment fractions ... 74
4.3.2. Content and Distribution of PAH ... 75
4.3.3. PAH & OC Relationship ... 77
4.3.4. Relative Composition of PAHs ... 79
4.3.5. Source Apportionment ... 80
4.4. Conclusion ... 82
V. POLYCYCLIC AROMATIC HYDROCARBONS IN SURFACE WATERS OF THE SIAK RIVER, ITS ESTUARY AND THE COASTAL AREAS OF RIAU PROVINCE, INDONESIA: DISTRIBUTION AND SOURCES ... 86
Abstract ... 86
5.1. Introduction ... 87
5.2. Materials and Methods ... 88
5.2.1. Study Areas and Sampling Locations ... 88
5.2.2. Sample Collection and Treatments ... 89
5.2.3. Extraction & Work-Up Procedures for particulate PAH ... 89
5.2.4. Solid Phase Extraction (SPE) system for dissolved PAH ... 89
5.2.5. Determination of PAH by HPLC UV/FLD ... 90
5.3. Results and Discussion ... 91
5.3.1. Dissolved PAHs ... 91
5.3.2. PAHs in the SPM ... 92
5.3.3. Distribution Coefficient of PAHs between SPM and Water Solution ... 96
5.3.4. Source apportionment ... 98
5.4. Conclusion ... 100
VI. A COMPARISON OF POLYCYCLIC AROMATIC HYDROCARBONS (PAHs) IN PEATLAND AND NON-PEATLAND AQUATIC SYSTEM SURFACE SEDIMENTS: A STUDY OF THE SIAK ESTUARY, SUMATRA, INDONESIA AND OF THE WENCHANG AND WANQUAN ESTUARIES, HAINAN ISLAND, CHINA ... 104
Abstract ... 104
6.1. Introduction ... 105
6.2. Materials and Methods ... 106
6.2.1. Study areas, Sample Collection and Fractionation ... 106
6.2.2. Determination of PAHs ... 108
6.2.3. Determination of Sedimentary Organic Matter ... 110
6.3. Results ... 110
6.3.1. PAHs and Organic Matter in Sedimentary Size Fractions... 110
6.3.2. Relative Composition of PAHs ... 112
6.3.3. Source Apportionment ... 113
6.3.4. Sediment-Water Distribution Coefficient ... 114
6.4. Discussion ... 115
6.5. Conclusion ... 119
1
I. INTRODUCTION
An examination the distribution and sources of polycyclic aromatic hydrocarbons (PAHs) in aquatic systems (rivers, estuaries and coastal waters) tackles key factors, which concern both the overall health of the environment and also these pollutants' behavior before they are integrated into the open ocean. During the last three decades, there has been a mounting body of scientific literature about PAHs in aquatic environments from many parts of the world. These studies indicate substantial and widespread concerns about these compounds as many persistent organic pollutants (POP) such as pesticides and polychlorinated biphenyls (PCBs). It also confirms a significant breakthrough in analytical methods such as sample collection, working-up procedure, and determination of PAH with analytical instruments such as capillary GC, GC-MS, and high performance liquid chromatography with ultraviolet and fluorescence detectors. Unfortunately, research in developing countries, in particular Indonesia and other tropical regions of the globe, has up to the present been sorely lacking. On the other hand, the relevance of tropical environments for the global redistribution of organic pollutants has been increasingly recognized by scientists. For instance, Iwata et al. (1994) showed that the discharge of persistent and semi-volatile compounds taking place in eastern and southern Asia such as India, Thailand, Vietnam, Malaysia, Indonesia and Oceania had a significant effect on pollutant redistribution on a global scale, particularly important through air and water phases.
This study attempts to refine the scientific understanding of the distribution and magnitude of PAH contamination in such aquatic environments, especially those surrounded by large areas of peat. It attempts to comprehensively examine the concentrations of PAHs in the sediments, suspended particulate matter (SPM), and water solution of Siak River, its estuary and the coastal areas of the Riau province, Sumatra, Indonesia. To start off, the definition and physico-chemical of selected PAHs are concisely introduced to give a general understanding on the subject of the compounds investigated (Chapter I). Also, in this chapter, the relevance of the PAH contamination in the aquatic environment is presented for some points of motivation i.e. toxic/carcinogenic effects, bioconcentration/biomagnification as well as persistence and degradation aspects of these compounds. In response to those motivations, the distribution and the sources of PAHs in the aquatic environment are shortly reviewed to examine the key environmental problems underlying this PAH investigation (Chapter II). Then, the PAH determination is presented for three aquatic compartments: sediment, SPM and water as for dissolved PAHs. Methods of analysis are given in Chapter III. The results and discussion are elaborated in three manuscripts for publication. The first manuscript deals with the distribution and source of PAHs in the sediment of the Siak river system (Chapter IV). This manuscript examines the levels and sources of anthropogenic PAH contamination in two sediment fractions
2 (sand/coarse: 2 mm – 63 m and mud/fine: <63 m on the Wentworth scale). The second manuscript evaluates the concentration, spatial distribution and sources of anthropogenic PAH contamination on SPM and dissolved PAHs in surface waters of the given study areas (Chapter V). The last manuscript deals with the character differentiation of PAHs in surficial sediments from peat- and non-peatland aquatic systems, which is a comparative study of the Siak estuary of Sumatra and Wenchang/Wanquan estuaries of Hainan, China (Chapter VI).
The work undertaken in this study was an integrated part of Cluster 3.1. of the SPICE Project (German-Indonesian Science for the Protection Indonesia Coastal Ecosystem) - a research collaboration on marine biogeoscience carried out between 2004 and 2007. The general theme of Cluster 3.1 was "coastal ecosystem health: the transfer of natural and anthropogenic materials from land to the coastal sea", focusing in particular on the Siak River in Riau Province, Sumatra, Indonesia. It is fully recognized that there has been no analogous study undertaken for the particular areas mentioned above up to the present. This, of course, brings with it the added difficulty of comparing the magnitude of contaminants in Indonesia with those in other areas of the world. The nature of this study is important, because it presents initial scientific data as part of the first comprehensive study on PAH distribution for this region. In this respect, the results can serve as a starting point for discussion and reference for the future analogous studies attempting to further refine the knowledge in this area.
1.1. Polycyclic aromatic hydrocarbons: definition and physio-chemical
characteristics 1.1.1. Definition
Polycyclic aromatic hydrocarbons (PAHs), also called polyaromatic hydrocarbons or polynuclear aromatic hydrocarbons, refer to a group of organic arene compounds composed of two or more fused aromatic benzene rings. These compounds contain various configurations of molecular ring structures, however the base-unit rings do not have heteroatoms within the rings. PAHs are therefore explicitly associated with their parent compounds. A fused ring results from the sharing of two or three specific carbon atoms by two/three connected aromatic rings. The carbon-sharing rings lead to the formation of a virtually single-planed structure composed entirely of carbon and hydrogen atoms (Neff, 1979). This planar structure allows for large and highly-diverse molecules, which can be constructed with different numbers and positions of the aromatic rings.
There are possibly hundreds of PAH compounds occurring in an extremely complex mixture in the environment. For the purposes of this study, however, we restricted ourselves to the sixteen parent PAH compounds listed by the US Environmental Protection Agency on its priority pollutant list (the 16 US EPA). These compounds are among those which have been frequently used for the purposes of environmental quality assessments. The base structures of
3 the sixteen parent compounds are composed of 2-6 aromatic rings with molecular masses ranging from 128 Dalton to 278 Dalton (Fig. 1.1). They include naphthalene (NAPH), acenaphthylene (ACYN), acenaphthene (ACEN), fluorene (FLU), phenanthrene (PHEN), anthracene (ANTH), fluoranthene (FLA), pyrene (PYR), benzo(a)anthracene (BaA), chrysene (CHRY), benzo(b)fluoranthene (BbFLA), benzo(k)fluoranthene (BkFLA), benzo(a)pyrene (BaP), dibenzo(a,h)anthracene (DANTH), benzo(g,h,i)perylene (BPERY) and indeno(1,2,3-c,d)pyrene (IPYR).
1.1.2. Selected physio-chemical characteristics
These unsubstituted aromatic compounds are non-polar, lipophilic and highly hydrophobic in nature. Selected physio-chemical characteristics of given PAHs are summarized in Appendix 1 (see for details and references). Briefly, the compounds are highly soluble in certain organic solvents with logarithmic values of the octanol-water partition coefficient (Log Kow) for each of the compounds ranging from 3.45 to 6.75 (Williamson et al., 2002). Their water
solubility is very low and ranges between values of only 0.3 g/L and 30.2 mg/L (Williamson et al., 2002). Low molecular weight compounds i.e. naphthalene, acenaphthene and acenaphthylene, have the highest water solubility with values of 30.2, 16.1, and 3.93 mg/L, respectively. The solubility decreases with increasing molecular mass. PAHs generally tend to be more easily adsorbed onto organic matter. In the environment, PAHs are readily associated with other natural, organic substances such as biopolymers (e.g. polysaccharides, lipids, protein and polynucleic acids), humic substances (e.g. humic acids and fulvic acids), kerogens and black carbon. One important review of the interactions between contaminants and these geo-sorbents was given by Weber et al. (2001).
The vapor pressure of PAHs is quite low, ranging from 8.89 ·10-2 – 2.10 ·10-11 mmHg, which often leads to their classification as semi-volatile compounds. Their boiling points ranges from 218 oC to 542oC, and the melting point span from 80 oC to 279 oC. At ambient conditions, they usually occur as almost colorless solids, normally being white or pale yellow in color e.g. IPYR. In the aquatic environment, however, they occur either as free molecules or associated with dissolved organic matter, particulate phases and sediments.
The benzene ring structures of these compounds are rigid. PAH molecules are appreciably stable and prefer substitution reactions to additions. With respects to molecular structure, there is a different degree of thermodynamic stability between peri-condensed PAH compounds to cata-condensed ones (Grope, 2001). Peri-condensed PAHs such as fluoranthene, pyrene, benzofluroanthene, benzo(a)pyrene, benzo(g,h,i)perylene, indeno(1,2,3-c,d)pyrene, are less stable. On the other hand, within the cata-condensed structure, those with linear structure are less stable than their angular isomers such as anthracene to its isomer phenanthrene (Grope,
4 2001). PAH compounds are characterized by broad ultraviolet absorbance spectra. In addition, most PAHs are fluorescent and emit light when excited, with the exception of acenaphthylene.
Fig. 1.1. The molecular structures and masses of the 16 parent PAHs as listed in the US EPA priority pollutant list. The structures are listed by the number of increasing ring groups (from 2 to 6 rings).
BPERY Benzo(g,h,i)perylene C22H12 MW=276,34 IPYR Indeno(1,2,3-c,d)pyrene C22H12 MW=276,34 BbFLA Benzo(b)flouranthene C20H12 MW=252,32 BkFLA Benzo(k)fluoranthene C20H12 MW=252,32 BaP Benzo(a)pyrene C20H12 MW=252,30 DANTH Dibenzo(a,h)anthracene C22H14 MW=278,35 FLA Fluoranthene C16H10 MW=202,26 PYR Pyrene C16H10 MW=202,26 Ba A Benzo(a)anthracene C18H12 MW=228,29 CHRY Chrysene C18H12 MW=228,29 NAPH Naphthalene C10H8 MW=128,17 ACYN Acenaphthylene C12H8 MW=152,20 ACEN Acenaphthene C12H10 MW=154,21 FLU Fluorene C13H10 MW=166,22 PHEN Phenanthrene C14H10 MW=178,22 ANTH Anthracene C14H10 MW=178,23
5
1.2. PAH contamination in the aquatic environments: background to
environmental problems
PAH contamination is one of the primary environmental problems facing humanity at present. This is due to the fact that PAHs are toxic, mutagenic and/or carcinogenic to both humans and other organisms. They also are subject to bioaccumulation and concentration in the aquatic food web. Additionally, they are relatively persistent in the environment. Mounting literature has provided sufficient evidence of their global distribution and high concentrations in atmospheric, terrestrial and aquatic systems. For example, Zhang & Tao (2009) published the global atmospheric emission inventory of 16 PAH priority pollutants from 37 countries. It was estimated that PAH global emission in 2004 accounted for 520 giga grams per year, with 55% of the total emissions coming from Asia. China, India and the United State were the top three countries with the highest PAH emission. However, as environmental investigation increases worldwide it can be assumed that the number of and places in which significant amounts of PAHs are found may also increase. This implies an increasing potential of human exposure to PAHs and a growing number of possible sources for these compounds. Particularly important in this respect are aquatic systems, which include rivers, estuaries and coastal waters, upon which tens or even hundreds of millions of people rely upon for food and other resources. Both water for drinking and hygiene and also the vast number of both freshwater and marine fish species and other organisms used as food sources for humans are directly tied to such ecosystems. This is especially true for a country like Indonesia where people rely on river, estuary and coastal resources.
1.2.1. Toxicity of PAHs
PAHs pose toxic, mutagenic and/or carcinogenic threats to humans and other organisms. A number of toxicity and cancer cases in both human and aquatic organisms have partly been associated with increasing chronic and acute exposure to high concentrations of PAHs from the ambient environment or specific contaminated sites (e.g. Brasseur et al., 2007; Cachot et al., 2006; Chiang et al., 2009; Hu et al., 2007; Smith et al., 2000). Many molecular epidemiological studies have shown evidence of increased levels of several biomarkers indicating PAH genotoxicity, for example PAH-DNA adducts and oxidative DNA damage in populations exposed to increasing levels of PAH (e.g. Shou et al., 1996; Hussain et al., 1998; Liu et al., 2007; Singh et al., 2007). A relationship between PAHs and cancer causes of lung, skin, bladder (Bofetta et al., 1997) and prostate (Rybicki et al., 2006) was shown to be conclusive.
The mutagenic/carcinogenic effects of PAHs are mainly exerted through electrophilic metabolic activation of the compounds due to their planar, highly conjugated aromatic structures. PAH metabolites are then capable of modifying DNA, which is the key to carcinogenesis. Several mechanisms of metabolite activation have been widely proposed by
6 scientists, including the formation of dihydrodiol epoxide. This mechanism turns out to be the most frequent pathway, which is often called the "bay region dihydrodiol epoxides pathway" (Xue & Warshawsky, 2005). In this fashion, PAHs are oxidized by P450 enzymes in the initial step in the activation process. This then produces reactive electrophilic metabolites which are capable of interacting with cellular macromolecules, in particular with nucleic acids and proteins. A review of the mechanisms of metabolic activation of PAHs exerting carcinogenicity was provided by Xue & Warshawsky (2005).
Although it is still an area of intensive epidemiological study, toxicological profiles of PAHs have already been introduced (e.g. ATSDR, 1995 see www.atsdr.crc.gov). The International Agency for Research on Cancer (IARC), an institute within the World Health Organization, has classified the carcinogenicity of individual PAH compounds in several of the parent groups (Table 1.1.). This classification was based on the strength of evidence derived from studies in humans and experimental animals (http://monographs.iarc.fr). However, even those individual substances which could not be positively identified as carcinogens may act synergistically (Wenzl et al., 2006). High molecular mass (HMW) PAHs tend to be more carcinogenic, but less acutely toxic than their cousins with lower molecular masses (LMW). Benzo(a)pyrene (BaP) in particular has revealed itself to be the most prominent carcinogen (Group 1 of IARC) and is often used as a hazard index for PAH exposure (Bofetta et al., 1997; Ravindra et al., 2008, Rappaport et al., 2004). However, even the LMW compounds like naphthalene and acenaphthene have been shown to exert carcinogenic effects on animals and may also carry potential risk for humans (Long et al., 1995; Rappaport et al., 2004).
As a consequence of increasing concentrations of PAHs in the environment, the likelihood of human exposure could raise. Human exposure can occur through inhalation, absorption/adsorption (skin), and ingestion (Bofetta et al., 1997; ATSDR, 1995). Inhalation and skin contact have been proven to be important pathways for atmospheric PAH to enter the human organism. Both of these routes of entry into the body are strongly related to specific occupations, such as aluminum and coke production, coal gasification, iron and steel foundry work, tar distillation, petroleum cracking, shale oil extraction, wood impregnation, roofing, road paving and carbon production (Bofetta et al., 1997). However, ingestion through food and drinking remains the most significant route for PAH contamination in humans. As much as 90% of the total daily intake of persistent pollution compounds into the human organism could result solely from diet (Binelli & Provini, 2004; Wenzl et al., 2006). In this context, animal food sources play a significant role, particularly fish and seafood.
7 Table 1.1. List of priority PAHs and their carcinogenicity classification, water and sediment quality guidelines (content in g/L for water and ng/g dry sediment weight).
Compounds Carcinogenicity
Classification*
Water Quality
Standard** Sediment Quality Guidelines***
(group) ERL ERM
Acenaphthene b,c,d 3 16 500 Acenaphthylene b,c,d NI 44 640 Anthracene a,b,c,d 3 0.006** 85.3 1100 Benz[a]anthracene b,c,d 2B 261 1600 Benzo[a]pyrene a,b,c,d 1 0.001e 430 1600 Benzo[b]fluoranthene a,b,c,d 2B 0.001e Benzo[e]pyrene d 3 Benzo[g,h,i]perylene a,b,c,d 3 0.001e Benzo[j]fluoranthene d 2B Benzo[k]fluoranthene a,b,c,d 2B 0.001e Chrysene b,c,d 2B 384 2800 Dibenz[a,h]anthrancene b,c,d 2A 63.4 260 Fluoranthene a,b,c,d 3 0.01** 600 5100 Fluorene b,c,d 3 19 540 Indeno[1,2,3-c,d]pyrene a,b,c,d 2B 0.001e Naphthalene a,b,c 2B 0.01** 160 2100 Phenanthrene b,c,d 3 240 1500 Pyrene b,c,d 3 665 2600
a EU WFD, 2000/60/EC in Annex X (http://europa.eu/scadplus/leg/en/lvb/l28108.htm);
b the US EPA (http://www.epa.gov/epaoswer/hazwaste/minimize/chemlist.htm);
c NPI Australia;
d Canadian NPRI Substance 2007;
e Priority & hazardous substances based on the decision # 2455/2001/CE of the European Parliament
* IARC: http://monographs.iarc.fr/ENG/Classification/index.php. Group 1: Carcinogenic to humans; Group 2A: probably carcinogenic to humans; Group 2B: possibly carcinogenic to humans; Group 3: not classifiable as to carcinogenicity to humans; Group 4: probably not carcinogenic to humans;
** Maggi et al 2008;
*** adopted from Long et al., 1995: ERL= the effects range-low, ERM= the effects range-median NI = information not available
1.2.2. Bioconcentration and magnification in aquatic food webs
PAHs are bioconcentrated and biomagnified by marine organisms. Bioconcentration and bioaccumulation of PAHs in organisms occurs via all three chemical exposure routes, including dietary absorption, transport across respiratory surfaces and dermal absorption (Mackay & Fraser, 2000). Of these possibilities, the first two routes are the most important. The degree of bioconcentration is determined by tissue lipid content (Kayal & Connell, 1995). An appreciable increase in PAH concentration has been observed in various marine organisms as compared to their environment (water or sediment). For instance, Barbour et al. (2008) observed appreciable bioaccumulation factors in oysters harvested from a war-induced oil spill zone in the Eastern Mediterranean Sea, which extend from 242 to 3700.
Along with bioconcentration, biomagnification has been recorded from lower to higher trophic levels of organisms, e.g. plankton (Carls et al., 2006), benthic amphipods (Viganò et al., 2007); mussels (Okay et al 2000; Richardson et al., 2003; Pérez-Cadahía et al 2004; Hellou et al., 2005), other bivalve organisms (Oros & Ross 2005), oysters (Mondon et al., 2001), eels (Ribeiro et al 2005), feral finfish (Hellou et al 2006), and other fish species (Liang et al., 2007).
8 It has been suggested that the higher the trophic level, the more and various food types are consumed. As a result, higher trophic levels are more susceptible to pollutant magnification. However, in aquatic systems benthic organisms such as mussels and clams accumulate more PAHs than the carnivorous fish which use these organisms as prey. Martí-Cid et al. (2007) found that shellfish (mussels & clams) and shrimp can contain higher concentrations of PAHs than fish such as tuna, mackerel and salmon, due to a low capacity to biotransform contaminants. Physiologically, mollusks do not metabolize PAHs as quickly as fishes. Instead, they tend to accumulate these toxins.
Bioconcentration/biomagnification is determined by bioavailability of the compounds in the water phase. Low molecular mass PAHs (those with 2-3 rings) indicated high levels of bioavailability in benthic organisms such as mussels and clams when compared to high molecular mass compounds (e.g. Baumard et al., 1999a; Baumard et al., 1999b; Kayal & Connell, 1995). This is due to the fact that those compounds are highly water-soluble. On the other hand, HMW PAHs (4-6 rings) have been proven to be relatively non-bioavailable when compared to the lighter compounds. Baumard et al. (1998) found low concentrations of heavier PAHs in mussel, despite bottom sediments containing high levels of various pyrogenic-PAHs. Thorsen et al. (2004) observed that petrogenic PAHs are more bioavailable than pyrogenic PAHs.
However, the bioavailability of the compounds is also determined by water and sediment properties i.e. organic matter content, SPM and sediment grain-size. Organic matter found in both water and sediment affects the bioavailability of PAHs to a large extent. One notable review of the effects of dissolved organic matter (DOM) on the bioconcentration of several organic contaminants (PAHs, chlorinated hydrocarbons, and TBT) in aquatic organisms, including water fleas, mussels, amphipods and fish, was given by Haitzer et al. (1998). The authors reviewed the lack of bioavailability of DOM-bound chemicals, which would in turn lead to a decrease in their bioaccumulation. However, at low level of DOM (~10 mg/L), bioconcentration of pollutant could be enhanced up to 300% (Haitzer et al., 1998). Moreover, the extent to which DOM affects PAHs is further determined by factors including the quality and quantity of DOM, the contact time between humic substances and the chemicals before exposure, and the overall exposure time. Data might therefore differ from place to place according to environment-specific characteristics.
Further factors such as the presence of suspended particulate matter and the fineness of sediment grain-sizes are also important for the bioaccumulation of organic pollutants, especially in filter/suspension-feeding and detritivorous organisms such as bivalves. Baumard et al. (1998) found a significant bioconcentration of PAHs on organisms living in close contact with sediments compared to carnivorous organisms. The latter might be exposed to a much lesser extent to sedimentary particles. Menon & Menon (1999) found in a 10-day experiment that the
9 bioaccumulation of PAHs in clams exposed to sediment in suspension increased almost two-fold when compared to clams tested in undisturbed sediments. This indicates that higher sediment levels in aqueous suspensions can result in an increased PAH content in the surrounding water. Such conditions are important in dynamic systems such as those found in rivers, estuaries and coastal waters. High turbidity should be expected to increase the concentration of carcinogenic, high molecular weight PAHs in benthic organisms (e.g. Baumard et al., 1999b). Such benthic invertebrates are important prey (food) for many carnivorous fishes. Therefore, increased bioaccumulation can also be extrapolated for higher trophic levels of organisms found in the food webs of these aquatic systems.
1.2.3. Persistence, low degradation rates and pollution indicators
PAHs are persistent over longer periods of time when found in bottom sediments. Once bound to aquatic sediment particles, PAHs can effectively survive for years. This fact stems from their relatively stable chemical structures, particularly under anaerobic conditions (Neff, 1979; Mihelcic & Luthy, 1988). Due to this, PAH signatures in sediments are often used as indicators in identification of pollutant sources. PAHs have also been tested by geochronicle studies looking at dated sediment cores (e.g. Gevao et al., 1998; Yunker et al., 1999; Yamashita et al., 2000; Fabbri et al., 2003; Ricking et al., 2005). Thirty years ago, Hites et al (1977) published an experiment which showed the relative stability of PAH composition over hundreds of years. Their analysis documented unsubstituted PAHs and their alkyl homologues in three sections of core sediments taken from Buzzards Bay, Massachusetts, USA. The researchers found that the distribution of PAHs remained qualitatively constant, even though the intensity of the pollution source had increased considerably.
Degradation of PAHs might, however, potentially affect the relative levels and makeup of persistent PAHs in aquatic environments. Such degradation might occur by photo-/chemical oxidation (e.g. Behymer & Hites, 1985; Lehto et al., 2000; Shemer & Linden, 2007) or biodegradation (Poeton et al., 1999; Kanaly et al., 2000; Kot-Wasik et al., 2004). Miller and Olejnik (2001) studied the photolysis of water-borne PAHs (BaP, Chry, Flu) via UV radiation. The study revealed that the photo-degradation of PAHs in water involves rather complicated mechanisms involving oxygen, the function of pH, and scavengers (organic materials). Degradation of BaP and Chry (high molecular weight PAHs) was found to be retarded in water with a lack of oxygen, high pH values (alkalinity) and an increase in the level of organic scavengers. Conversely, FLU (low molecular weight) proved itself to be independent of those parameters, except for the fact that FLU was eliminated more quickly when oxygen concentrations were lower. Microbial degradation is often enhanced when chemicals are found in the dissolved phase (high bioavailability). Low molecular weight PAHs (2 – 3 rings) are more susceptible to microbial degradation due to bioavailability than high molecular mass
10 compounds (>4 rings). The latter are more recalcitrant when it comes to undergoing chemical changes (Juhaz & Naidu, 2000). However, a great deal of research has confirmed the microbial degradation in sediment-bound PAHs. Xia et al. (2006) concluded that an increase in the population of PAH-degrading bacteria and desorption of PAHs from the solid phase increases the rate of contact between PAH and bacteria. This can subsequently enhance biodegradation rates. Some of PAH-degrading bacteria are mentioned in the work of Samanta et al. (2002). However, it is important to note that biodegradation is generally a relatively slow process (Männistö et al., 1996).
In addition to being used as indicators for anthropogenic pollutants, the presence and amount of PAHs often serve as measures for estimating overall environmental quality. This includes factors such as the degree of toxicity (e.g. Bihari et al., 2006; Cachot et al., 2006), environmental and human health risk assessment (e.g. Galloway 2006) and petroleum pollution (e.g. Requejo et al., 1996).
In conclusion, PAHs are regulated not only for food and drink, but also for environmental aspects, due to their toxicity and carcinogenic properties, confirmed bioaccumulation and biomagnification and their persistence over longer periods of time. Examples of legal directives regulating PAHs are: the European Union Water Framework Directory (the EU WFD, 2000/60/EC in Annex X), the US EPA list of priority pollutants, the National Pollutant Inventory (NPI) for Australia, and the European Scientific Committee on Food (2002) (see Stolyhwo & Sikorski, 2005). Despite these efforts, PAHs have still not been specifically mentioned in the list of twelve POPs (Persistent Organic Pollutants) under the Stockholm Convention, which has so far been signed and ratified by more than 150 countries including Indonesia (http://chm.pops.int/).
1.3. The relevance of rivers, estuaries and coastal areas in Indonesia
1.3.1. A Perspective on the global pollution dispersal
Rivers and their tributaries, estuaries and coastal areas represent a transitional boundary between the terrestrial and marine aquatic system. These bodies of water act as the "front line" with respect to receiving the majority of land-based loading materials, including organic pollutants. Therefore, their role in the distribution and transport of chemicals into the ocean reservoir is crucial. Chester (2003) pointed out that river runoff is one of the primary, global-scale sources for material to enter the oceans when taken together with atmospheric deposition and hydrothermal activity. With regard to this, rivers play a large and important role as the main carriers of numerous chemical signatures into the ocean.
A variety of surface runoff types exists. The flows from diverse landscapes, e.g. municipalities, various sorts of heavy, medium and light industry, and agriculture, groundwater seeps into riverine systems, and into the ocean. Atmospheric deposition also enriches the
11 magnitude of chemicals found in aquatic systems. These sources become especially important for aquatic systems in regions where economic development has been significantly and rapidly taking place, areas of the world such as Southeast Asia. Accordingly, the presence of pollution in the rivers, estuaries and coastal waters of these regions represents one of the key environmental challenges facing humanity today. Schwarzenbach et al. (2006) estimated that anthropogenic fluxes of organic pollutants stemming from fertilizers, pesticides, synthetic organic chemical productions and accidental oil spills annually contribute ca. 450 million tons worldwide to aquatic systems.
Scientists have established the global significance of the rivers draining southern Asia, which flush large amounts of terrestrial materials into the world ocean. On a global scale, recent estimation of the annual total sediment load flowing from rivers into the global ocean is ca. 20 Gigatons, suspended sediment load contributes 90%, and the rest is mainly from bed sediment load (Syvitski et al., 2003).
Asia and Oceania are the largest producers of fluvial sediment, with around 70% of the overall annual sediment loads coming from these regions (Milliman et al., 1999). In comparison with the globe's largest landmasses, including Africa, Asia, Europe, North America, and South America, the region containing Oceania and Indonesia produces the world's highest sediment yield, as defined in terms of sediment load divided by total drainage area (Syvitski et al., 2005). In fact, these two regions accounted for ca. 800 and 543 tons of sediment per square kilometer per annum, respectively. Milliman et al. (1999) pointed out the particular importance of the rivers in Sumatra, Java, Borneo, Sulawesi, Timor and the New Guinean islands (Fig. 1.2). They estimated that these six large islands alone significantly discharge about 4.2 Gt of sediment per year, despite the fact that they constitute only 2% of the world's total land area which drains into the global ocean. The rivers located on these islands are responsible for about 20-25 % of global sediment export. Sumatra alone contributes 498 Megatons through fluvial transport per annum, an amount which makes it the second largest source in this group of six islands (Milliman et al. 1999). Even though some of the larger river basins in Sumatra such as the Siak, Kampar, Rokan, Indragiri and Batanghari might not have been included in the above-mentioned calculations, further estimates suggest that the sediment volumes are substantial.
Most of rivers and estuaries in the eastern coast of Sumatra drain large area of peatland which are characterized by high humic substance of black water masses. Therefore, the environmental state of the tropical peatlands has attracted a big concern for pollutant transport and climate change. Page et al. (2002) reveal that the stability of tropical peatlands is particularly relevance for the climate change because these peatlands are one of the largest near-surface reserves of terrestrial organic carbon estimated for ca. 26 – 50 Gt. They also remain that persistence environmental change in the tropical peatlands owing to drainage and forest clearing have threaten the stability of ca. 20-m thick of peat deposit, and make the peatlands being
12 susceptible to fires. In Indonesia, peatland fires have long been an environmental problem (see section 1.3.2. for further explanation). The worst peatland fire episode occurred during severe El Niño event (1997/1998) damaging of ca. 6.8 Mha (or 34%) of Indonesia peatland, and emitted up to 2.57 Gt Carbon in 1997 (Page et al., 2002). It is relatively similar to the global net emission of CO2 from land-use change was recently estimated for 2.4 Gt y-1 (IPCC, 2000). The
environmental and health effects of the haze and smoke had been widely recognized ever since (e.g. Langmann et al., 2007; Fang et al., 1999; Parameswaran et al. 2004). Of particularly important is that the peatland fires are mostly anthropogenic as part of land clearance activities before establishing crops (Page et al., 2002).
Unfortunately, the magnitude of organic pollutants, particularly that of PAHs – the burning by products - found in Indonesia's rivers, estuaries and coastal areas has only been superficially evaluated at best. To date there have been no long-term or in-depth studies carried out, especially ones covering extremely large surface areas or multiple islands in this region. Nevertheless, the significance of organic pollution fluxes stemming from Indonesian rivers can be assumed to be directly proportional to the significance of material fluxes previously mentioned by the above literature sources. It is therefore reasonable to assume that increasing levels of "flushing out" SPM from these river systems will eventually contribute to an increase in the magnitude of pollutants in the coastal zone, given that suspended particulate matter (SPM) is a significant carrier for most organic pollutant such as PAHs (e.g. Kayal & Connell, 1989; Ollivon et al, 1995; Deng et al., 2006; Law et al., 1997; Fernandes et al., 1999; Heemken et al., 2000; Witt & Siegel 2000; Kowalewska et al., 2003; Ross & Oros, 2004 and Cao et al., 2005), polychlorinated biphenyls (PCBs) (e.g. Mai et al., 2002; Telli-Karakoç et al., 2002), polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs) (e.g. Ishaq et al., 2003). The coastal zone of Sumatra can thus be viewed as being very susceptible to organic contaminants. Conversely, dissolved organic matter such as dissolved humic substances has been shown to enhance the solubility of dissolved PAHs in aqueous solution (e.g. Liu & Amy, 1993). The implication of this additional factor is that the residence times of PAHs in the aqueous phase will be increased, which in turn promotes more far-reaching transport towards and into the open ocean. The relevance of Sumatran rivers, which often drain peat bogs and other humic soils, for the transport of dissolved organic matter into the open ocean has already been recognized (e.g. Baum et al., 2007).
13
Fig. 1.2. Sediment discharge (106 T y-1) from the six East Indies islands. Arrow widths are proportional to
the annual load. The letters S, J, B, C, T and NG refer to Sumatra, Java, Borneo, Sulawesi (Celebes), Timor and New Guinea, respectively. Shaded areas represent water depths of less than 1000 m, although most of these areas are in less than 100 m of water depth (adopted from Milliman et al., 1999).
1.3.2. Environmental settings of the study areas: An Overview
Our chosen study areas include the Siak River, its estuary, and the coastal areas Indonesia's Riau Province, extending from the Northwest to the Southwest coast in the Selat Panjang (the coastal channel running between offshore islands and the Sumatra mainland) (Fig. 1.3.). Details of the exact sampling locations are provided in each result’s chapter. The areas selected for this study provide an interesting opportunity to comprehensively study the spatial and phase distributions of PAHs in different river systems, all of which ultimately empty their waters into the Malacca Strait. The sampled areas are located in the equatorial wet climate of Sumatra Island, which is affected by a monsoon season mainly bringing rainfall between the months of September and March. However, rainfall can also occur at almost any time throughout the year. The average annual rainfall in 2004, for example, was 2088 meters with the heaviest rainfall occurring between October and December (BPDAS, 2004, unpublished seminar notes). However, even such small changes in regional weather patterns can affect the flow of terrestrial pollutants and organic materials into the marine environment through various pathways such as varying land erosion, storm water discharge, and surface runoff volumes.
The Siak River system and its catchment area are composed of ca. 300 km of waterways, including several tributary rivers (the Tapung Kanan, Tapung Kiri and Mandau) and various sub-basin streams. Including the Siak's estuary system, the entire catchment area totals roughly 1 million hectares, amounting to ca. 10% of the total surface area of Riau Province (BPDAS, 2004). The average river flow rate is normally 200 m3 s-1 (a total of 6.307 x 109 m3 y-1), reaching a peak flow rate of 1700 m3 s-1 during the monsoon season and a minimum of 45 m3 s-1 in the course of droughts (unpublished data, PEMPROV RIAU, 2005). The Siak and its estuary
14 mainly drain large areas of low-laying land characterized by widespread peat swamps, which are scattered among various landscapes, including huge palm-oil plantations, forests and secondary swamp-forests. These areas account for ~53%, ~23%, and ~11% of catchment land-use coverage, respectively (BPDAS, 2004). The main river correspondingly receives large amounts of water input stemming from smaller channels and tributaries along its course. Moreover, the main river basin is sparsely inhabited. It nevertheless drains some highly-urbanized centers, such as the Riau Province capital of Pekanbaru, the smaller city of Siak Sri Indrapura, a pulp-paper industry estate located in Perawang, and an oil refinery in the Siak river mouth, thus making it inevitable that urban and industrial discharges will enter the river's water. The Siak River and its estuary dump their waters in the coastal channel located between Bengkalis Island and the Sumatra mainland.
Adjacent to the Malacca Strait, the coastal areas are constricted and range from Dumai City in the northwest of the River down to the Selat Panjang channel in the southwest. The coastal waters are characterized by observable plumes of suspended particulate materials. These plumes stem mainly from the Siak River and some other neighboring rivers (such as the Rokan and Kampar River in the northwest and southwest, respectively) and many smaller streams emerging along the coast (Fig. 1.5.). The Siak River plume is caused by dark-colored peat materials and stands out distinctively from the plumes of other rivers. The Siak's current pushes river water through the channel and into the Malacca strait during the low tide. Its plume pulls back during high tide. The tide is semi-diurnal, meaning that two high tides and one ebb tide occur during one day, however, the high tides arrive with differing magnitudes. Unfortunately, there is no information available on the residence times of the coastal water masses.
15 Fig. 1.3. An overview map of the Siak river mouth (satellite image provided by H. Siegel (2004), Baltic Sea Research Institute (IOW) Warnemuende, Rostock, Germany). The Siak River stretches over ca. 300 km up to the estuary (the mouth). The coastal areas extend to far northwest of the mouth and to southwest of the coastal channel between the offshore islands and the Sumatra mainland.
Fig. 1.4. (a) Surface runoff of organic-rich black water from a small waterway entering the Siak River around a palm-oil plantation, (b) typical small water channel found in the estuary. These are typical water inputs feeding the Siak River (source: courtesy of SPICE).
MALAYSIA Siak River
Riau Province,
Sumatra, INDONESIA
Siak Estuary Riau Coast Bengkalis IslandA
B
16 Fig. 1.5. Plumes of SPM and dissolved humic material from several rivers: the Rokan, Siak and Kampar River (as seen from northwest to southwest) into the Malacca strait. At low tide the plume is flushed outwards by the Siak River and at high tide it is pushed backwards. Image from H. Siegel (2004), Baltic Sea Research Institute (IOW) Warnemuende, Rostock, Germany using MODIS Terra, Data: NASA-RRS.
With regards to PAH pollution, there are two potential sources which have their roots in the expanding economic development of the Province. Riau is well-known for its oil production industry. Oil-related facilities such as various plants, refineries and ridges are scattered all along the river, the estuary and coastal areas (Fig. 1.5). Oil spills might appear along the Siak River, the estuary and the coast, and most likely stem from boats, oil-transporting vessels, harbor facilities, and many other petroleum-related activities.
Fig. 1.6. Some oil-related activities observed in the studied areas. Refineries are mainly located in the coast of Dumai city (left), ridges (middle) in the channel (Panjang Strait), and typical port activities in the Siak River (source: courtesy of SPICE).
Rokan River Siak River Kampar River Sumatra, INDONESIA Siak Estuary
17 Moreover, oil industries can also be found in surrounding areas of two neighboring countries, Malaysia and Singapore (the biggest oil refinery center in Asia). Malacca Strait is likely destined to become one of the busiest straits in the world, one which is highly susceptible to oil pollution. Zakaria et al. (2000) found that oil pollution levels detected in the Strait were due mainly to spillage (through accidents or routine tanker operations such as ballast-water discharges) from the tens of thousands of vessels transporting crude oil and other petroleum products.
The region's second-largest industry, plantations for palm-oil production, covers more than 50% of the total catchment area. The studied areas were affected by high levels of organic material combustion, in particular anthropogenic, slash-and-burn agriculture and intense instances of naturally-occurring forest and swamp fires. Especially important for this region is the documented fact that uncontrolled agricultural burning for purposes such as land clearing has worsened overall air quality for over a decade.
Fig. 1.7 shows the trend of hotspots from 1996-2006 and Fig. 1.8 shows an example of a daily hotspot distribution. Data was extracted from the hotspot daily observation provided by the Indonesian Ministry of Forestry (www.dephut.go.id). Elevated numbers of hotspots were observed during the severe El Niño event of 1997-1998. It is not to extrapolate that the numbers of hotspots correlated directly with El Niño, because high numbers have also occurred during 2005 and 2006, which did not fall under the pronounced effects of El Niño. Severe burning happening during 1997/1998 resulted in a strong impact on the distribution of aerosols over Sumatra and the neighboring counties (e.g. Fang et al., 1999) and the effects even reached into the Indian Ocean. Parameswaran et al. (2004) observed a large aerosol plume which formed over the eastern equatorial (5oN to 10oS) and eventually reached 60o E latitude during September-November of 1997. This plume was directly tied with the large-scale fires occurring in Southeast Asia, particularly Indonesia, at the time of a severe drought caused by an El Niño event. Unfortunately, the observed number of annual hotspots has remained at a level of >3000 ever since then and has recently become even worse. The occurrence of agricultural burning is most likely during the dry season, the period between March and September. Nonetheless, (as can be seen in the Fig. 1.7b), large numbers of hotspot have been observed even in the assumed rainy season. This is to say that forest and swamp fires, added to uncontrolled agricultural burning, are potentially the largest source of PAH contamination in the studied areas.
