Untersuchungen zur Speziation von verschmutzungsrelevanten ... · Ost-Hainan generell höher als in Nordost-Brasilien. Das ästuarine Verhalten der Das ästuarine Verhalten der Metalle

Untersuchungen zur Speziation von verschmutzungsrelevanten Spurenmetallen in

tropischen Ästuar- und Küstensystem (Südchina und Nord-Ost-Brasilien)

Dissertation zum Erlangen des akademischen Grades eines

„Doktor der Naturwissenschaften“

(Dr. rer. nat.)

des Fachbereichs 02, Biologie/Chemie

der Universität Bremen

Vorgelegt von:

Jun Fu

Bremen, 2014

I

Untersuchungen zur Speziation von verschmutzungsrelevanten Spuren-metallen in tropischen Ästuar- und Küstensystem (Südchina und Nord-Ost-Brasilien).

Dissertation zum Erlangen des akademischen Grades eines „Doktor der

Naturwissenschaften (Dr. rer. nat.)“ des Fachbereichs 02, Biologie/Chemie der

Universität Bremen.

Vorgelegt von: Jun Fu

Gutachter: 1. Prof. Dr. Wolfgang Balzer

2. Prof. Dr. Wolfram Thiemann

Prüfer: 1. Prof. Dr. Otto Schrems

2. Dr. Uwe Schüßler

Datum des Kolloquiums: 04.02.2014

II

Selbstständigkeitserklärung:

Hiermit erkläre ich, dass ich die Doktorarbeit mit dem Titel:

„Untersuchungen zur Speziation von verschmutzungsrelevanten Spuren-metallen in

tropischen Ästuar- und Küstensystem (Südchina und Nord-Ost-Brasilien).“

Selbstständig verfasst und geschrieben habe und außer den angegebenen Quellen

keine weiteren Hilfsmittel verwendet habe.

Jun Fu

III

Danksagung

Mein besonderer Dank gilt Herrn Prof. Dr. Wolfgang Balzer von der Universität

Bremen für die Ermöglichung dieser Arbeit sowie seine wissenschaftlichen

Anregungen und Diskussionen als Doktorvater während dieser Arbeit. Ich möchte

mich bei meinem Zweitgutachter Herrn Prof. Dr. Wolfram Thiemann von der

Universität bedanken. Herrn Prof. Dr. Otto Schrems vom Alfred-Wegener-Institut für

Polar- und Meeresforschung (AWI) und Herrn Dr. Uwe Schüßler danke ich für ihre

Funktion als Prüfer.

Für die gute Zusammenarbeit und die hilfreichen Erörterungen organisatorischer

sowie analytischer und inhaltlicher Problemstellungen möchte ich mich bei meinen

Kollegen und ehemaligen Kollegen Dr. U. Schüßler, Dr. W. Barkmann, Dr. M.

Lukman, Dr. B. Bach, U. Wolpmann, I. Becker, O. Wilhelm, T. Daberkow und X.-L.

Tang bedanken. Insbesondere danke ich Herrn u.a. X.-L. Tang für die Durchführung

von Analysen eines Teils der verwendeten Proben.

Daneben wurden zu Vergleichs- und Interpretationszwecken einige Ergebnisse

(Balzer, unpublished) von Expeditionen für die beiden Projekte in die vorliegend

Arbeit mit aufgenommen, die vor Beginn dieser Dissertation durchgeführt worden

waren; dies gilt auch für die Verwendung von DOC-Daten, die sämtlich im Rahmen

eines anderen Vorhabens erhoben wurden.

Diese Arbeit wurde im Rahmen des Projekts "LANCET (Land-Sea Interactions in

Coastal Ecosystems of Tropical China: Hainan)" bzw. des Projekts „POLCAMAR

(Impact of Pollutants from Sugar Cane Monoculture on Estuaries and Coastal Waters

of NE-Brazil)“ durchgeführt. Gedankt sei den Kommissionen der jeweiligen Projekte

für die finanzielle Unterstützung.

Ich möchte mich auch bei den vielen Wissenschaftlerinnen und Wissenschaftlern, die

an den beiden wissenschaftlichen Projekten „LANCET“ und „POLCAMAR“ beteiligt

waren, für die Kollegialität bedanken und ganz besonders bei den Koordinatoren Dr.

L. Dsikowitzky und Dr. D.-R. Wang für „LANCET“ sowie O. Wilhelm und Prof. Dr. B.

Knoppers für „POLCAMAR“, die mir in China bzw. in Brasilien stets hilfreich zur Seite

standen.

Schließlich danke ich meiner Familie und meinen Freunden für ihre Ermutigung und

Unterstützung während der Arbeit.

IV

Kurzfassung

Diese Arbeit befasste sich im Wesentlichen mit der Konzentration, der Verteilung und

dem Verhalten von fünf verschmutzungsrelevanten Spurenmetallen (Nickel, Cobalt,

Kupfer, Cadmium und Blei) plus Eisen in der wässrigen und in der partikulären

Phase sowie der Abschätzung des Kontaminationsniveaus der untersuchten Metalle

in tropischen Ästuaren und Lagunen von China und Brasilien.

Die Feldarbeiten fanden im Wanquan Ästuar und im Wenchang-Wenjiao Ästuar auf

der Südchinesischen Insel Hainan sowie in der Mundaú Lagune und in der

Manguaba Lagune in Nordostbrasilien statt, die alle durch relativ niedrige DOC- und

Huminstoff-Konzentrationen gekennzeichnet sind. Die Oberflächenwasserproben

wurden durch Filtration in eine gelöste (< 0.45 μm) und eine partikuläre (> 0.45 μm)

Phase separiert. Danach erfolgte eine weitere Separation der gelösten Phase mittels

Querstrom-Ultrafiltration in zumeist zwei kolloidale Fraktionen (HMW "high molecular

weight“: 10 kDa - 0.45 μm; LMW "low molecular weight“: 5 kDa - 10 kDa) sowie eine

als "echt gelöst" angesehene Fraktion (TD "truely dissolved": < 5 kDa).

Die Konzentrationen der sechs verschmutzungsrelevanten Spurenmetalle liegen in

Ost-Hainan generell höher als in Nordost-Brasilien. Das ästuarine Verhalten der

Metalle in den beiden Studiengebieten ist jedoch kaum voneinander unterscheidbar.

Das gelöste Cadmium scheint mehr oder weniger in der gelösten Phase angereichert

zu werden, wohingegen gelöstes Eisen und Blei mit zunehmendem Salzgehalt rasch

aus der gelösten Phase entfernt werden. Nickel, Cobalt und Kupfer in Lösung sowie

alle Spurenmetalle in partikulärer Form folgen allgemein dem Muster konservativer

Mischung. In Hinblick auf die kolloidalen Fraktionen gewinnen die echt gelösten

Metalle (< 5 kDa; TD) mit steigender Salinität zunehmend an Bedeutung. Gelöstes

Eisen, Nickel und Blei sind hauptsächlich in der hochmolekularen Fraktion (10 kDa -

0.45 μm; HMW) vorhanden, wohingegen ein dominanter Anteil des gelösten

Cadmium in der echt gelösten Fraktion vorliegt. Beim gelösten Cobalt und Kupfer

sind zwar jeweils die echt gelösten Fraktionen ebenfalls vorherrschend; jedoch

können die hochmolekulare und die Fraktion mit niedrigem Molekulargewicht (LMW:

5 kDa - 10 kDa) nicht vernachlässigt werden.

Trotz der intensiven Aquakulturindustrie in Ost-Hainan und des großflächigen

Anbaus von Zuckerrohr in Monokultur in Nordost-Brasilien sowie der anthropogenen

Aktivitäten in den beiden Studiengebieten konnten keine nennenswerten

Umweltbeeinträchtigungen in Form von Spurenmetallverschmutzungen festgestellt

werden. Wenn erhöhte Metallkonzentrationen auftraten (im Wenchang-Wenjiao

V

Ästuar und in der Mundaú Lagune), wurden sie im Ästuar effektiv gedämpft. Die

Metallkonzentrationen bei hohen Salinitäten in Ost-Hainan und Nordost-Brasilien

liegen etwa auf dem gleichen Niveau. Im weltweiten Vergleich liegen sowohl die

gelösten als auch die partikulären Metallenkonzentrationen in den beiden

Studiengebieten auf einem niedrigen bis mittleren Niveau, mit Ausnahme von Eisen,

was vermutlich auf die geologische Gegebenheiten zurückzuführen ist.

VI

Abstract

This work investigated the concentrations, the distributions and the behaviors of five

trace metals (nickel, cobalt, copper, cadmium and lead) plus iron in the aqueous and

in the particulate phase in two tropical estuarine systems of China and Brazil, and

estimated their contamination levels. The field work was carried out in the Wanquan

Estuary and in the Wenchang-Wenjiao Estuary of the Southern Chinese island of

Hainan as well as in the Mundaú Lagoon and in the Manguaba Lagoon of Northeast-

Brazil, which are all characterised by relatively low contents of dissolved humics and

DOC. The surface water samples were seperated into the dissolved (< 0.45 μm) and

the particulate (> 0.45 μm) phase by conventional filtration followed by a further

separation of the dissolved phase into two colloidal fractions (HMW "high molecular

weight“: 10 kDa - 0.45 μm; LMW "low molecular weight“: 5 kDa - 10 kDa) and a

fraction (TD: < 5 kDa) considered as "truly dissolved" fraction.

The levels of these trace metals with high contamination riscs is generally higher in

East-Hainan than in Northeast-Brazil. However, the estuarine behaviors of these

metals in the two study areas are very similar. Dissolved cadmium appears more or

less to be enriched in the dissolved phase, while dissolved iron and lead are rapidly

removed from the dissolved phase when salinity increases. Nickel, cobalt and copper

in solution, as well as all trace metals in particulate forms essentially follow a

conservative mixing line. The relative significance of the truly dissolved fraction (< 5

kDa; TD) of the metals increased with increasing salinity. Dissolved iron, nickel and

lead are mainly associated with the high molecular weight fraction (10 kDa - 0.45 μm;

HMW), whereas dissolved cadmium predominantly occurs in the truly dissolved

fraction. The truly dissolved fraction is also dominant for dissolved cobalt and copper;

however, the high molecular weight fraction and the low molecular weight fraction (5

kDa - 10 kDa; LMW) cannot be neglected for these two metals.

Despite the massive aquaculture industry in East-Hainan and the extensive

cultivation of sugar cane in monocultural form in Northeast-Brazil as well as general

anthropogenic activities, no significant impact of trace metals on the local

environment could be detected in the two study areas. Although elevated metal

concentrations occurred (in the Wenchang-Wenjiao Estuary and in the Mundaú

Lagoon), they were leveled out effectively during estuarine processing. At high

salinities, the metal concentrations in East-Hainan and in Northeast-Brazil are

approximately at the same level. In comparison with averages of estuaries worldwide,

both dissolved and particulate metal concentrations in the two study areas are similar

VII

or lower, except for iron probably originating from specific geographic settings.

VIII

Inhaltsverzeichnis Abbildungsverzeichnis 1 Tabellenverzeichnis 5 1. Einführung 6

1.1. Verschmutzungsrelevante Spurenmetalle in aquatischen Systemen 8

1.1.1. Eisen 8

1.1.2. Nickel 8

1.1.3. Cobalt 9

1.1.4. Kupfer 9

1.1.5. Cadmium 10

1.1.6. Blei 10

1.2. Ästuar-Typen 11

1.3. Mastervariablen für Spurenmetalle 12

1.3.1. Salinität und chemische Zusammensetzung 12

1.3.2. Residenzzeit 13

1.3.3. Suspendiertes partikuläres Material 13

1.3.4. Größenspeziation (Kolloide) 14

1.3.5. Redoxpotential und pH-Wert 15

1.3.6. Biologische Aktivität 16

1.4. Ästuarine Mischungsprozesse für Spurenmetalle 17

2. Material und Methoden 18

2.1. Studiengebiete 18

2.1.1. Ästuare auf der südchinesischen Insel Hainan 18

2.1.2. Lagunen-System in Nordost-Brasilien 19

2.2. Maßnahmen zur Kontaminationsminderung 20

2.3. Probenahme 21

2.4. Filtration und Querstrom-Ultrafiltration 21

2.5. UV-Aufschluss 27

2.6. Flüssig-Flüssig-Extraktion 27

2.7. Mikrowellen-Aufschluss-Verfahren 28

2.8. Massenspektrometrie mit HR-ICP-MS 30

2.9. Referenzmaterialien 30

2.10. Andere Parameter 32

IX

3. Vergleichende Diskussion der Ergebnisse 33 3.1. Eisen 34

3.2. Nickel 37

3.3. Cobalt 39

3.4. Kupfer 42

3.5. Cadmium 44

3.6. Blei 47

3.7. Verteilungskoeffizienten 49

3.8. Vergleich der Studiengebiete und ihrer Kontaminationsniveaus 52

3.9. Offene Fragen und Perspektiven 57

4. Publikation 1 (erschienen) : "Estuarine modification of dissolved

and particulate trace metals in major rivers of East-Hainan" Continental Shelf Research 57 (2013): 59-72 59

Abstract 59

4.1. Introduction 61

4.2. Materials and Methods 62

4.2.1. Study area 62

4.2.2. Sampling 64

4.2.3. Pretreatment and measurement of dissolved metals 64

4.2.4. Pretreatment and analysis of particulate metals 65

4.2.5. Other parameters 65

4.3. Results and discussion 66

4.3.1. Distribution of suspended particulate matter and environmental

Conditions 66

4.3.2. Dissolved trace metals 69

4.3.3. Particulate trace metals 75

4.3.4. Distribution coefficients 79

4.3.5. Comparison of WR and WWR with other Chinese estuaries 82

4.4. Conclusions 84

Acknowledgements 85

References 86

5. Entwurf zur Publikation 2: "Transport of trace metals in colloidal

form through estuarine systems of East-Hainan" 91 Abstract 91


X



5.2.2. Sampling and filtration 95

5.2.3. Pretreatment and measurement 96

5.2.4. CFF blanks and mass balance 97

5.3. Results and discussion 98

5.3.1. Iron 98

5.3.2. Lead 101

5.3.3. Cadmium 101

5.3.4. Copper 102

5.3.5. Cobalt 103

5.3.6. Nickel 104

5.3.7. Spatial variability of colloidal trace metals in East-Hainan 104

5.4. Conclusions 105

6. Entwurf zur Publikation 3: "Cd, Co, Cu, Fe, Ni and Pb in solution,

colloidal and particulate phases of Northeast Brazilian estuaries" 107 Abstract 107




6.2.2. Sampling and filtration 111

6.2.3. Pretreatment and measurement of dissolved and colloidal metals 112

6.2.4. Cross-flow filtration Blanks and Mass balance 112

6.2.5. Pretreatment and measurement of particulate metals 114

6.2.6. Other parameters 115

6.3. Results and Discussion 115

6.3.1. Salinity, suspended particulate matter and dissolved organic carbon 115

6.3.2. Dissolved trace metals 117

6.3.3. Particulate trace metals 130

6.3.4. Distribution coefficient 131

6.3.5. Colloidal trace metals 133

6.3.6. Spatial and seasonal differences 137

6.4. Conclusions 138

Literaturverzeichnis 141 Anhang 155

Abbildungsverzeichnis

- 1 -


Kapitel 1-3

Abb. 1. Korngrößenverteilung von verschiedenen Kolloiden und Partikeln

in der Umwelt. 15

Abb. 2. Die Studiengebiete: (a) Wenchang-Wenjiao Ästuar und (b)

Wanquan Ästuar in Ost-Hainan. 18

Abb. 3. Das Studiengebiet Mundaú-Manguaba Lagunen-System in

Nordost-Brasilien. 19

Abb. 4. Schematische Darstellung der Querstrom-Ultrafiltrations-Kassette

Vivaflow® von Sartorius. 25

Abb. 5. Ablauf der Querstrom-Ultrafiltrations-Prozedur. 25

Abb. 6. Salinitäts-Abhängigkeit des gelösten Eisens. 35

Abb. 7. Salinitäts-Abhängigkeit des partikulären Eisens. 36

Abb. 8. Salinitäts-Abhängigkeit der kolloidalen Verteilung von Eisen. 36

Abb. 9. Salinitäts-Abhängigkeit des gelösten Nickels. 37

Abb. 10. Salinitäts-Abhängigkeit des partikulären Nickels. 38

Abb. 11. Salinitäts-Abhängigkeit der kolloidalen Verteilung von Nickel. 38

Abb. 12. Salinitäts-Abhängigkeit des gelösten Cobalts. 40

Abb. 13. Salinitäts-Abhängigkeit des partikulären Cobalts. 40

Abb. 14. Salinitäts-Abhängigkeit der kolloidalen Verteilung von Cobalt. 41

Abb. 15. Salinitäts-Abhängigkeit des gelösten Kupfers. 42

Abb. 16. Salinitäts-Abhängigkeit des partikulären Kupfers. 43

Abb. 17. Salinitäts-Abhängigkeit der kolloidalen Verteilung von Kupfer. 43

Abb. 18. Salinitäts-Abhängigkeit des gelösten Cadmiums. 45

Abb. 19. Salinitäts-Abhängigkeit des partikulären Cadmiums. 45

Abb. 20. Salinitäts-Abhängigkeit der kolloidalen Verteilung von Cadmium. 46

Abb. 21. Salinitäts-Abhängigkeit des gelösten Bleis. 47

Abb. 22. Salinitäts-Abhängigkeit des partikulären Bleis. 48


- 2 -

Abb. 23. Salinitäts-Abhängigkeit der kolloidalen Verteilung von Blei. 48

Abb. 24. Salinitäts-Abhängigkeit des log KD. 50

Abb. 25. Gelöstes Cadmium in der Weser an der FGG-Weser-Messstelle

Bremerhaven. 57

Kapitel 4

Fig. 1. Estuaries of East Hainan under investigation: (a)

Wenchang/Wenjiao River estuary and (b) Wanquan River estuary. 62

Fig. 2. Distribution of SPM and DOC vs. salinity. 66

Fig. 3. Dissolved trace metals vs. salinity. 68

Fig. 4. Dissolved Pb vs. dissolved Fe. 72

Fig. 5. Dissolved Co vs. dissolved Fe during the wet seasons. 72

Fig. 6. Particulate trace metals vs. salinity. 74

Fig. 7. Ratio of the percentages of the particulate metal on the sum of the

dissolved and particulate metal: Pb vs. Fe. 78

Fig. 8. Log10KD vs. salinity. 80

Kapitel 5

Fig. 1. Cross-flow filtration apparatus. 93

Fig. 2. Wenchang/Wenjiao River estuary (a) and Wanquan River estuary

(b). 94

Fig. 3. Flow diagram of the sample filtration procedure. 96

Fig. 4. Colloidal and truly dissolved concentration vs. total dissolved

concentration. 97

Fig. 5. Trace metal ratios of various size fractions vs. salinity. 99

Fig. 6. Fe concentration vs. DOC in the HMW fraction. 101

Fig. 7. Pb vs. Fe concentrations in the HMW fraction. 101

Fig. 8. Co vs. Fe concentrations in the HMW fraction and Co vs. DOC in

the UD fraction. 104


- 3 -

Kapitel 6

Fig. 1. The Mundaú-Manguaba lagoon system in Northeast Brazil. 110

Fig. 2. Calculated total dissolved concentration vs. separately measured

total dissolved trace metal concentration. 113

Fig. 3. The estuarine distribution of suspended particulate matter vs.

salinity (a) in the Mundaú Lagoon and (b) in the Manguaba

Lagoon. 115

Fig. 4. The estuarine distribution of DOC vs. salinity (a) in the Mundaú

Lagoon and (b) in the Manguaba Lagoon. 116

Fig. 5. Estuarine distribution of Cd in the Mundaú Lagoon: (a) total

dissolved Cd, (b) particulate Cd and (c) Cd percentage of colloidal

size fractions. 117

Fig. 6. Estuarine distribution of Cd in the Manguaba Lagoon: (a) total

dissolved Cd, (b) particulate Cd and (c) Cd percentage of colloidal

size fractions. 118

Fig. 7. Estuarine distribution of Cu in the Mundaú Lagoon: (a) total

dissolved Cu, (b) particulate Cu and (c) Cu percentage of colloidal

size fractions. 120

Fig. 8. Estuarine distribution of Cu in the Manguaba Lagoon: (a) total

dissolved Cu, (b) particulate Cu and (c) Cu percentage of colloidal

size fractions. 121

Fig. 9. Estuarine distribution of Fe in the Mundaú Lagoon: (a) total

dissolved Fe, (b) particulate Fe and (c) Fe percentage of colloidal

size fractions. 122

Fig. 10. Estuarine distribution of Fe in the Manguaba Lagoon: (a) total

dissolved Fe, (b) particulate Fe and (c) Fe percentage of colloidal

size fractions. 123

Fig. 11. Estuarine distribution of Pb in the Mundaú Lagoon: (a) total

dissolved Pb, (b) particulate Pb and (c) Pb percentage of colloidal

size fractions. 124

Fig. 12. Estuarine distribution of Pb in the Manguaba Lagoon: (a) total

dissolved Pb, (b) particulate Pb and (c) Pb percentage of colloidal 125


- 4 -

size fractions.

Fig. 13. Estuarine distribution of Co in the Mundaú Lagoon: (a) total

dissolved Co, (b) particulate Co and (c) Co percentage of colloidal

size fractions. 126

Fig. 14. Estuarine distribution of Co in the Manguaba Lagoon: (a) total

dissolved Co, (b) particulate Co and (c) Co percentage of colloidal

size fractions. 127

Fig. 15. Estuarine distribution of Ni in the Mundaú Lagoon: (a) total

dissolved Ni, (b) particulate Ni and (c) Ni percentage of colloidal

size fractions. 128

Fig. 16. Estuarine distribution of Ni in the Manguaba Lagoon: (a) total

dissolved Ni, (b) particulate Ni and (c) Ni percentage of colloidal

size fractions. 129

Fig. 17. Log10KD (L/kg) vs. salinity. 132

Tabellenverzeichnis

5

Tabellenverzeichnis

Kapitel 1-3

Tab. 1. Mikrowellen-Programme einschließlich der Parameter. 29

Tab. 2. Massenbilanzen der Metalle. 32

Tab. 3. Ästuarines Verhalten von gelösten Spurenmetallen. 34

Tab. 4. Durchschnittliche Konzentrationen der gelösten Spurenmetalle

und des DOC in den vier Studiengebieten. 52

Tab. 5. Durchschnittliche Konzentrationen der partikulären Spurenmetalle

und des SPM in den vier Studiengebieten. 53

Tab. 6. Konzentrationen der gelösten Spurenmetalle in Ästuaren. 55

Tab. 7. Konzentrationen der partikulären Spurenmetalle und des SPMs in

Ästuaren. 56

Kapitel 4

Tab. 1. Dissolved trace metals in freshwater, in the estuarine mixing zone

and at high salinities of the WR and the WWR during the dry

season (2006) and the wet seasons (2007 and 2008). 69

Tab. 2. Particulate trace metals in freshwater, in the estuarine mixing

zone and at high salinities of the WR and the WWR during the dry

season (2006) and the wet seasons (2007 and 2008). 75

Tab. 3. Particulate trace metals and SPM of East-Hainan estuaries in

comparison to selected Chinese rivers and estuaries. 79

Tab. 4. Dissolved trace metals of East-Hainan estuaries in comparison to

selected Chinese rivers and estuaries. 83

Kapitel1: Einführung

- 6 -

1. Einführung

Der Ausdruck „Umwelt“ wurde von Jakob Johann von Uexküll als zentraler Begriff

der Ökologie eingeführt. Mit seinem Buch „Umwelt und Innenwelt der Tiere“ (1909)

war er maßgeblich beteiligt an der Geburt des Begriffs „Umwelt“ als „die Umgebung

eines Lebewesens, die auf dieses einwirkt und seine Lebensumstände

beeinflusst“ (z.B. Mildenberger, 2009). Seit der menschlichen Zivilisation, spätestens

seit der Industrialisierung, ist das Streben der Menschen nach Energie und

Rohstoffen nicht mehr zu stoppen. Dementsprechend werden auch eine Vielzahl von

Umweltchemikalien (Stoffe, die durch menschliches Zutun in die Umwelt gebracht

werden und in Mengen oder Konzentrationen auftreten können, die geeignet sind,

Lebewesen, insbesondere den Menschen, zu gefährden; Bundesregierung, 1972)

absichtlich oder unbeabsichtigt in die Umwelt entlassen, zu den auch Spurenmetalle

gehören. Insbesondere die industrielle Anwendung ermöglicht die Ausbreitung von

Spurenmetallen, die anderenfalls nicht leicht in die Natur gelangen würden (Hem,

1970). Die dadurch entstandenen Umweltveränderungen können lokale und

regionale Ausmaße haben, teilweise jedoch auch globalen Charakter annehmen

(Koch, 1995). Bisher ist nur ein Teil der Umweltchemikalien einschließlich deren

Einflüssen und Wirkungen auf natürliche Ökosysteme sowie auf die

Lebensgrundlagen und die Gesundheit des Menschen ausreichend erforscht (z.B.

Judson et al., 2009). Umso wichtiger ist es daher, auch kleine negative

Veränderungen in der Umwelt durch gezielte Beobachtung und Überwachung

frühzeitig zu erkennen, ihre Ursachen aufzuklären und ggf. ihrem Fortgang

entgegenzuwirken.

Diese Arbeit beschäftigt sich mit der Konzentration, der Verteilung und dem

Verhalten von verschmutzungsrelevanten Spurenmetallen in der wässrigen und

partikulären Phase sowie der Abschätzung der Kontaminationsniveaus der

untersuchten Metalle in den jeweiligen Studiengebieten. Dabei ging es um zwei

ästuarbezogene wissenschaftliche Projekte, die jeweils zu den Verbund-Projekten

„LANCET (Land-Sea Interactions along Coastal Ecosystems of Tropical China:

Hainan) “ und „POLCAMAR (Impact of Pollutants from Sugar Cane Monoculture on

Estuaries and Coastal Waters of North-East-Brazil)“ gehörten. Bei dem Projekt

LANCET sollte aufgeklärt werden, inwieweit anthropogene Einflüsse zum

Niedergang der Korallenriffe an der Ost-Küste der chinesischen Insel Hainan

beitragen, und bei dem Projekt POLCAMAR stand die Untersuchung der Auswirkung

der Zuckerrohrmonokultur auf Ästuare und Küstenwässer Nordostbrasiliens im

Vordergrund. Da Ost-Hainan möglicherweise durch lokale Aquakultur verschmutzt ist


- 7 -

und das Mundaú-Manguaba Lagunen-System von Abwässern der

Zuckerrohrmonokultur und der Zuckerrohrfabriken belastet sein kann, standen die

Beeinträchtigung der jeweilige Studiengebiete durch Aquakultur bzw.

Zuckerrohrmonokultur im Fokus der beiden Projekte. Die Forschungsergebnisse

können dazu beitragen, die Wirkung von anthropogenen Aktivitäten mit dem

Schwerpunkt Spurenmetalle auf lokale Ökosysteme generell besser zu verstehen.

Die Ziele dieser Arbeit waren daher:

1. Bestimmung der gelösten, partikulären und kolloidalen Konzentrationen von

sechs verschmutzungsrelevanten Spurenmetallen im Wanquan Ästuar und im

Wenchang-Wenjiao Ästuar in Ost-Hainan sowie in der Mundaú Lagune und in der

Manguaba Lagune in Nordost-Brasilien.

2. Interpretation der ästuarinen Verteilung und des ästuarinen Verhaltens der

gelösten, partikulären und kolloidalen Spurenmetalle in den oben genannten

Studiengebieten.

3. Vergleich der Konzentrationen und des Verhaltens der sechs untersuchten

Spurenmetalle in den genannten Studiengebieten mit anderen Ästuaren weltweit

sowie Beurteilung der Kontaminationsniveaus in den Studiengebieten.


- 8 -

1.1. Verschmutzungsrelevante Spurenmetalle in aquatischen Systemen

Spurenmetalle sind in der Umwelt allgegenwärtig. Dabei sind die Hydrosphäre und

vor allem die Ozeane von besonderer Bedeutung. Denn Ozeane können als

Depositionsort aller Materialien, die aus anderen Geosphären stammen, betrachtet

werden (Chester, 1990). Spurenmetalle gelangen in die Ozeane hauptsächlich durch

Transport über Flüsse, hydrothermale Prozesse und Einträge aus der Atmosphäre.

Der erste Weg ist besonders von Interesse, da verschmutzungsrelevante

Spurenmetalle als Folge von anthropogenen Aktivitäten in großem Umfang in die

Ozeane gelangen. Während Organismen gegenüber essentiellen Metallen eine

relativ hohe Toleranz haben, wirken unbenötigte Metalle oft toxisch (z.B. Bach, 2012).

Schließlich kann eine kleine Erhöhung des Eintrags verschmutzungsrelevanter

Spurenmetalle große Beeinträchtigungen für Lebewesen im Meer mit sich bringen.

1.1.1. Eisen

Eisen ist als Bestandteil verschiedener Enzyme und Proteine von essentieller

biologischer Bedeutung. Negative Auswirkungen an Menschen infolge einer

Überdosis von Eisen wurden jedoch auch nachgewiesen (z.B. Pigeon et al., 2001).

Eisen ist an der Ästuarchemie entscheidend beteiligt (z.B. Santschi et al., 1997).

Deshalb wurde es in dieser Arbeit untersucht, obwohl es nicht zu den

verschmutzungsrelevanten Spurenmetallen im engeren Sinne gehört.

Im Meer liegt Eisen entweder als reduziertes Eisen(II) oder als oxidiertes Eisen (III)

vor (z.B. Rose and Waite, 2005). Als Spezies kommen vor allem Eisen(hydr)oxide

(z.B. Sulzberger et al., 1989) und organische Eisenkomplexe (z.B. Mill, 1980) in

Betracht sowie evtl. Eisensulfide in den Sedimenten (z.B. Davison et al., 1999; Lapp

and Balzer, 1994). Jedoch ist die Löslichkeit des Eisens durch die Gleichgewichte

zwischen gelöster Phase und festen Phasen begrenzt (z.B. Balzer, 1982). Unter

oxischen Bedingungen sollte die maximale Konzentration für gelöstes Eisen im Meer

nur etwa 0.2 nmol/L betragen (Millero, 1995), wegen organischer Komplexierung liegt

sie jedoch um den Faktor 3-5 höher (Rue and Bruland, 1995). Das fluviale Eisen wird

hauptsächlich in Form von Kolloiden und Partikeln in die Ozeanen transportiert (z.B.

Fox and Wofsy, 1983).

1.1.2. Nickel

Nickel findet vor allem Verwendung bei der Produktion von Stahl und Legierungen


- 9 -

sowie in der Galvanik- und Elektronikindustrie (Galler, 1999). Nickel wird bei

manchen Tieren als essentielles Spurenelement angesehen; auch für Pflanzen und

Plankton kann es bioaktiv wirken (z.B. Bruland et al., 1991; Galler, 1999).

Nickel existiert in natürlichen Wässern als Nickel(II) (z.B. Mouvet and Bourg, 1983).

In Meerwasser liegt Nickel zum Teil in Komplexen mit organischen Liganden vor, die

eine hohe Stabilität besitzen (z.B. Achterberg and Van den Berg, 1997).

Anorganisches Nickel ist entweder hydratisiert oder mit u.a. Chlorid- und Carbonat-

Ionen komplexiert (z.B. Donat et al., 1994).

1.1.3. Cobalt

Cobalt hat eine Vielzahl von Anwendungen z.B. in der Farbenindustrie, Glas- und

Keramikproduktion, Lebensmittelindustrie und als industrieller Katalysator (z.B.

Hamilton, 1994). In der Natur ist Cobalt essentiell für die biotische Stickstofffixierung,

kann jedoch auch schädlich für lebende Organismen sein (z.B. Kim et al., 2006).

Zwar ist Cobalt für die Bildung von Vitamin B12 (Cobalamin) im menschlichen Körper

notwendig, doch eine Überdosis kann die Schilddrüsenaktivität mindern (z.B.

Barceloux, 1999).

In Ozeanen ist Cobalt primär als Cobalt(II) vorhanden (z.B. Nolan et al., 1992). Ältere

Untersuchungen behaupteten, dass hydratisierte Ionen sowie Ionenpaare mit Chlorid

und Sulfat die überwiegenden Cobalt-Spezies im Meerwasser darstellen (z.B.

Ćosović et al., 1982). Neuere Studien weisen jedoch darauf hin, dass mindestens ein

Teil des Cobalts organisch gebunden ist (z.B. Ellwood and Van den Berg, 2001). Als

biologisches Abbauprodukt von Vitamin B12 soll inertes Cobalt(III) ebenfalls

anwesend sein (z.B. Saito and Moffett, 2001).

1.1.4. Kupfer

Neben Gold, Silber und Zink war Kupfer ein Metall, welches als eines der Ersten in

der Menschheit eine Anwendung fand. Es ist vor allem wegen seiner sehr guten

elektrischen und thermischen Leitfähigkeit vielseitig verwendbar (z.B. Wiberg, 2007),

und dementsprechend in der Industrie allgegenwärtig (z.B. Blake et al., 2004). Kupfer

ist ein für die Ernährung aller Lebewesen essentielles Element, jedoch können bei

Überschuss toxische Wirkungen bei Pflanzen und Tieren auftreten (z.B. Galler, 1999).

Obwohl unter anoxischen Bedingungen das reduziertes Kupfer(I) die dominante

Spezies sein kann (Boulègue, 1983), ist im sauerstoffhaltigen Meerwasser Kupfer(II)


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vorherrschend (z.B. Gledhill et al., 1997). Es bestehen Gleichgewichte zwischen

freien Kupfer(II)-Ionen und organisch sowie anorganisch komplexiertem Kupfer(II)

(z.B. Gledhill et al., 1997), wobei organisches Kupfer(II) über 90% der gesamten

Kupfer(II)-Spezies darstellt (z.B. Buckley and Van den Berg, 1986).

1.1.5. Cadmium

Cadmium ist ein toxisches Metall, welches auch in großem Umfang eine industrielle

Herkunft hat (Järup and Åkesson, 2009). Es kann bei Korrosionsschutz und

Akkumulatorherstellung sowie als Stabilisator für Kunststoffe eingesetzt werden (z.B.

Wurtz and Maeder, 2002). Cadmium wirkt bereits in geringer Konzentration toxisch

und ist daneben relativ mobil (Galler, 1999). Es greift vor allem die Nieren und den

Knochen des menschlichen Körper an und führt zu z.B. Nierenversagen und der so

genannten „Itai-Itai Krankheit“ (z.B. Satarug et al., 2010). Doch es kann auch die

biokatalytische Aktivität von u.a. marinen Kieselalgen ankurbeln (z.B. Lane and Morel,

2000).

Unter marinen Bedingungen wird Cadmium(II) als Hauptspezies angesehen (z.B.

Turner et al., 1981). Obwohl ein gewisser Anteil organischer Cadmiumspezies

identifiziert wurde (z.B. Ellwood, 2004), scheint die Assoziation von Cadmium mit

organischen Materien in den Ozeanen nicht sehr effektiv zu sein (z.B. Sunda and

Huntsman, 1998). Vielmehr ist Cadmium stark mit Chloriden komplexiert (z.B. Byrne

et al., 1988).

1.1.6. Blei

Blei ist eines der am längsten von Menschen verwendeten Metalle überhaupt.

Bereits in der frühen Bronzezeit wurde Blei verwendet, um Bronzen zu erzeugen.

Gegenwärtig findet es u.a. bei der Akkumulatoren-, Pigment- und

Legierungsherstellung Verwendung und bis vor kurzem als Treibstoffzusatz (z.B.

Wurtz and Maeder, 2002). Blei ist für die Menschen krebserregend und

nervensystemschädigend. Es kann einerseits eine große Zahl von Sulfhydryl-haltigen

Enzymen hemmen, Nukleinsäuren deformieren und die oxidative Phosphorylierung

stören, doch andererseits kann es mit Carboxypeptidase A gebunden z.B. die

Estersubstrate effektiv hydrolysieren (Vallee and Ulmer, 1972).

Die Oxidationsstufe +II ist die stabilste für Blei, und kommt in Ozeanen am häufigsten

vor (z.B. Town and Filella, 2002). Etwa 50% des gelösten Bleis in


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Oberflächenmeerwasser ist assoziiert mit organischen Materien (z.B. Capodaglio et

al., 1990). Die andere Hälfte besteht wesentlich aus Carbonat- und Chlorid-

Komplexen (z.B. Millero et al., 2009).

1.2. Ästuar-Typen

Flüsse gehören zu den wichtigsten Trägern für die chemischen Signale, die sich in

Ozeanen zusammenfinden (z.B. Duce and Tindale, 1991). Dabei spielt das Ästuar

eine fundamentale Rolle für die Modifikation der Signale. “An estuary is a semi-

enclosed coastal body of water which has a free connection with the open sea and

within which sea water is measurably diluted with fresh water derived from land

drainage" (Pritchard, 1967). Als komplexe aquatische Systeme sind Ästuare sowohl

fluvialen als auch marinen Einflüssen ausgesetzt (z.B. Bach, 2012). Darüber hinaus

zählen die Ästuargebiete zu den am dichtesten besiedelten Gebieten der Welt. Auf

Grund dessen sind die Ästuare auch stark von anthropogenen Aktivitäten beeinflusst.

Wegen der unterschiedlichen Zusammensetzung von Fluss- und Meerwasser spielen

sich in Ästuaren eine Vielzahl von biologischen, physikalischen und chemischen

Reaktionen bzw. Umverteilungen ab. Angesichts solcher Modifikationen kann das

Ästuar als eine Art „Filter“ angesehen werden, welcher die fluvialen chemischen

Signale in der Mischzone erheblich verändern kann (Chester, 1990). Diese Fähigkeit

des Ästuars basiert größtenteils auf der Wechselwirkung zwischen gelöster und

fester Phase, welche sowohl durch physikochemische als auch durch biologische

Prozesse gesteuert wird.

Ästuare lassen sich in unterschiedliche Typen unterteilen. Nach u.a. Davies (1964)

werden Ästuare je nach ihre Tide zu micro-tidal (< 2 m), meso-tidal (2-4 m) und

macro-tidal (> 6 m) Typ klassifiziert. Jedoch ist eine ausreichende Vermischung von

Frisch- und Salzwasser einer der wichtigen Voraussetzungen für die filternde

Wirkung eines Ästuars. So werden Ästuare nach ihrem Mischungsgrad zwischen

salt-wedge, highly-stratified, partially-stratified und well-mixed Ästuaren

unterschieden, wobei sich Frisch- und Salzwasser in den salt-wedge Ästuaren am

schlechtesten und in den well-mixed Ästuaren am besten miteinander vermischen

(z.B. Berner and Berner, 1996). Außerdem ist von einem Ästuar zum anderen die

Geomorphologie auch unterschiedlich (z.B. Kjerfve and Magill, 1989). Unter den in

dieser Arbeit behandelten Studiengebieten gehört beispielsweise das Wanquan

Ästuar zu den klassischen Ästuaren mit einer direkten Flussmündung, wohingegen in

dem Wenchang-Wenjiao Ästuar und dem Mundaú-Manguaba Lagunen-System der


- 12 -

Fluss durch eine Lagune mit dem Meer verbunden ist. Unterschiede wie solche

können die ästuarinen Mastervariablen entscheidend beeinflussen, und so auch die

ästuarine Verteilung der Spurenmetalle.

Die in dieser Arbeit untersuchten Ästuarsysteme werden durch sog.

Weißwasserflüsse (Ertel et al., 1986; Konhauser et al., 1994; Bach, 2012) gespeist,

die sich durch relativ niedrige Huminstoff- und DOC- Konzentrationen (< ca. 5 mg/L)

sowie eine blasse Färbung auszeichnen; sie sind - im Gegensatz zu den sog.

Schwarzwasserflüssen - häufig durch höhere pH-Werte, niedrigere

Lösungskonzentrationen für Spurenmetalle und eher physikalische Verwitterungspro-

zesse im mineralisch geprägten Einzugsgebiet gekennzeichnet.

1.3. Mastervariablen für Spurenmetalle

Spurenmetalle können in verschiedenen physikochemischen Formen vorkommen,

u.a. als gelöste anorganische und organische Komplexe, Metallspezies in Form von

Kolloiden und Metallspezies, die an Kolloide oder Schwebstoffe adsorbieren (Mota

and Correia Dos Santos, 1995). Dabei streben sie nach der höchst möglichen

Stabilität, welche sie durch die Bindung an eine andere Spezies erreichen können

(Bach, 2012). Das Verhalten, der Transport und die Bioverfügbarkeit von

Spurenmetallen in Ästuaren hängen eng mit ihrer chemischen Speziation zusammen

(Chester, 1990). Einige Metalle wie z.B. Eisen (z.B. Wen et al., 1999) und Cadmium

(z.B. Elbaz-Poulichet et al., 1996) existieren in Ästuaren hauptsächlich in

anorganischer Form, während andere Metalle wie Kupfer eher als organische

Komplexe vorliegen (z.B. Shank et al., 2004). Jedoch sind die meiste Metalle in

Ästuaren über ein weiteres Spektrum von freien Ionen bis hin zu makromolekularen

Formen verteilt. Eine Menge von Mastervariablen, die teilweise voneinander

abhängen, können dabei ausschlaggebend sein, u.a. die Salinität, die chemische

Zusammensetzung des Wassers, die Residenzzeit, das suspendierte partikuläre

Material, das Vorkommen von Kolloiden, das Redoxpotential, der pH-Wert sowie die

biologische Aktivität.

1.3.1. Salinität und chemische Zusammensetzung

Als Salinität wird der Salzgehalt in einer wässrigen Lösung bezeichnet (z.B. Chester,

1990). Häufig wird für die Salinität in der ästuarinen Chemie die dimensionslose

Einheit „psu“ (practical salinity units) verwendet. In den meisten Veröffentlichungen


- 13 -

wird die Spurenmetallverteilung im Ästuar als eine Funktion der Salinität diskutiert.

Dabei handelt es sich darum, dass bei einer Veränderung der Salinität die

Grenzfläche Lösung/Festkörper, die Löslichkeit der Spurenmetalle und der

Aggregatszustand sich ändern können (z.B. Stolpe and Hassellöv, 2007). Solche

Veränderungen treten vor allem bei niedriger bis mittlerer Salinität auf. Das

berühmteste Beispiel dafür ist die Salz-induzierte Aggregation des Eisens bei

niedriger Salinität (z.B. Nowostawska et al., 2008).

Außerdem sind im Flusswasser Calcium und Bicarbonat die wichtigsten

anorganischen Ionen, während das Meerwasser von Natrium und Chlorid dominiert

ist (Chester, 1990). Diese Veränderung spiegelt sich im Ästuar wider, und kann eine

Speziationsänderung der Spurenmetalle hervorrufen. Nickel z.B. konkurriert mit

Calcium und Magnesium um die Bindungsstellen an organischen Materien (z.B.

Mandal et al., 2002), und die Zunahme von Chloridionen zum Meer hin kann die

Bildung von stabilen Cadmiumchlorid-Komplexe begünstigen (z.B. Dabrin et al.,

2009).

1.3.2. Residenzzeit

Die Zeitdauer, in der das Flusswasser im Ästuar verbleibt, wird als Residenzzeit

bezeichnet (z.B. Duinker, 1986). Die Residenzzeit eines Ästuars ist einer der

entscheidenden Faktoren in der ästuarinen Chemie. Je länger diese ist, desto besser

ist das ästuarine Wasser vermischt und umso mehr Zeit besteht für

Gleichgewichtseinstellungen zwischen Lösungsbestandteilen und verschiedenen

Partikeln. Die ästuarine Residenzzeit hängt u.a. stark von den Gezeiten und der

Geomorphologie eines Ästuars ab, und variiert beispielsweise für Ästuare in Europa

und den USA zwischen 1 und 200 Tagen (Uncles et al., 2002). In geomorphologisch

fast geschlossenen Ästuaren (z.B. Lagunen) kann die Residenzzeit extrem lang sein

- im Gegensatz zu klassischen Ästuaren (Kjerfve and Magill, 1989). Darüber hinaus

kann sich die Residenzzeit in einem Ästuar stark verändern. Je nach Tidensituation

variiert die Residenzzeit beispielsweise in der San Francisco Bay von 1 Tag bis zu

60 Tagen (Walters et al., 1985).

1.3.3. Suspendiertes partikuläres Material

Suspendiertes partikuläres Material (suspended particulate matter, SPM) hat große

Bedeutung für die ästuarine Modifikation der Spurenmetalle (z.B. Turner, 1996). Je


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nach den Eigenschaften der jeweiligen Oberfläche können die im Wasser

schwebenden Partikeln sowohl adsorbierende als auch abstoßende Wirkung für die

im Wasser gelösten Stoffe haben. Spurenmetalle können sowohl an organische

Oberflächenfilme des SPM oder direkt an organische Partikel (z.B. Lion et al., 1982)

als auch an anorganische Partikel/Kolloide wie Eisen(hydr)oxide adsorbiert werden

(z.B. Turner et al., 2004). Wenn metallreiches SPM auf metallarmes Wasser trifft,

kann es auch zur Freisetzung von Metallen aus dem SPM in das Wasser kommen

(z.B. Edmond et al., 1985). Windaktivität und Flussmorphologie bewirken häufig eine

Aufwirbelung des Sediments, die beim bodennahen Zusammentreffen von Fluss-

und Meerwasser zu einem "Trübungsmaximum" - meist bei niedriger bis mittlerer

Salinität - führen kann (z.B. Turner et al., 1992; Chiffoleau et al., 1994; Kraepiel et al.,

1997; Martino et al., 2002). Dies ermöglicht einen intensiven Kontakt zwischen der

Festphase und der gelösten Phase, und kann einen besseren Austausch zwischen

den beiden Phasen hervorrufen. Zudem ist dort die Fließgeschwindigkeit des

Wassers sehr deutlich verringert. Etwa 90% des fluvialen SPM sedimentieren in den

Ästuaren (z.B. Chester, 1990), so dass die im SPM gebundenen Spurenmetalle

durch die Ästuare zum größten Teil abgefangen werden können, bevor sie das Meer

erreichen.

1.3.4. Größenspeziation (Kolloide)

Kolloide, bei denen häufig zwischen hoch- und niedermolekularen Entitäten

unterschieden wird, spielen eine zentrale Rolle für die Regulierung von Konzentration

und Speziation, den Transport und die Bioverfügbarkeit im natürlichen wässrigen

System (z.B. Guo et al., 2000a). Aufgrund ihrer vergleichsweise großen spezifischen

Oberfläche, die reaktive funktionale Gruppen enthält, können Kolloide eine Reihe von

Spurenmetallen adsorbieren (z.B. Waeles et al., 2008a). „An aquatic colloid is any

constituent that provides a molecular milieu into and onto which chemicals can

escape from the bulk aqueous solution, while its vertical movement is not significantly

affected by gravitational settling“ (Gustafsson and Gschwend, 1997). Bei Kolloiden

wird die Wechselwirkung an der Grenzfläche von den Oberflächeneigenschaften des

Partikels bestimmt (z.B. Lead and Wilkinson, 2007). Typische anorganische Kolloide

im natürlichen Gewässern beinhalten Aluminiumsilikate, Eisen- und Mangan-Oxide,

während Kohlehydrate und Fulvin-/ Huminstoffe zu den wichtigsten organischen

Kolloiden zählen (Filella, 2007). Aluminiumsilikate können z.B. Cadmium und Blei

adsorbieren (z.B. Kozar, et al., 1992) und Polysaccharide sind in der Lage, Kupfer,

Nickel und Blei aufzunehmen (z.B. Gérente, et al., 2000). Darüber hinaus ist in


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wässriger Phase eine Reihe von Spurenmetallen mit Eisen-/Mangan-Oxiden und

Fulvin-/Huminstoffen assoziiert (z.B. Tessier, et al., 1996).

Nach IUPAC-Definition sollte ein Kolloid zumindest in einer Dimension eine Länge

zwischen typischerweise 1 nm und 1 μm haben (z.B. Everett, 1988). In Abb. 1 sind

typische umweltrelevante Kolloide und deren Größen aufgelistet. Auf Grund

methodischer Probleme wurden Kolloide lange Zeit zur gelösten Phase gezählt und

Prozesse auf kolloidaler Ebene, wie das Entfernen gelöster Spurenmetalle durch die

Ausflockung von Kolloiden in Ästuaren, nicht erkannt oder falsch interpretiert (z.B.

Bach, 2012). Die Verwendung neuartiger Aufbereitungs- und Analysentechniken

ermöglicht heutzutage die Separation und Bestimmung von Kolloiden. Jedoch ist

eine exakte Separation von Kolloiden nach physikalischer Größe wegen technischer

Schwierigkeiten nicht möglich (z.B. Buesseler et al., 1996). Dies wird in einem

späteren Kapitel diskutiert.

Abbildung 1. Korngrößenverteilung von verschiedenen Kolloiden und Partikeln in der Umwelt (Lead and Wilkinson, 2007).


- 16 -

1.3.5. Redoxpotential und pH-Wert

Das Redoxpotential ist von zentraler Bedeutung bei der ästuarinen Verteilung von

Spurenmetallen. So soll z.B. im Danshuei Ästuar kolloidales Cadmium nur unter

anoxischen Bedingungen relevant sein und nur relativ wenig unter oxischen

Bedingungen (Jiann et al., 2005). Bei Anwesenheit von Sauerstoff favorisieren Eisen

und Mangan die Festphase und können andere Spurenmetalle fixieren und mit in das

Sediment nehmen (z.B. Turner et al., 2008).

Der pH-Wert ist eine weitere Mastervariable in Ästuaren. Das Flusswasser hat in der

Regel einen pH-Wert von 5 bis 8, wohingegen das Oberflächen-Meerwasser einen

durchschnittlichen pH-Wert von 8.2 besitzt (Chester, 1990). Diese pH-Unterschiede

führen zu einem Gradient des pH-Werts in Ästuaren. Für die meisten Spurenmetalle

steigt die Löslichkeit mit fallendem pH-Wert (z.B. Hatje et al., 2003a). Außer direkter

Konkurrenz (zwischen H+ und Metallkationen) um die Bindungsstellen kann eine

Veränderung des pH-Werts u.a. die Umwandlungen von Metallspezies und die

Modifikationen von metall-bindenden Liganden beeinflussen (z.B. Kola and Wilkinson,

2005). Kinetische Studien offenbarten, dass mit steigendem pH-Wert die

Adsorptionsquoten von Eisen, Cobalt und Cadmium an das SPM zunehmen (Hatje et

al., 2003b). Auch die Bildung von Metall-Hydroxiden und -Carbonaten, die häufig in

wässrigen Systemen vorkommen (z.B. Turner et al., 1981), kann dabei entscheidend

gestört werden.

1.3.6. Biologische Aktivität

Da viele Spurenmetalle biologisch aktiv sind (z.B. Norisuye et al., 2007), kann deren

Verhalten im Ästuar auch von biologischen Aktivitäten abhängen, die wiederum u.a.

vom Sauerstoffgehalt (z.B. Beck and Sañudo-Wilhelmy, 2007) und der Residenzzeit

(z.B. Crump et al., 2004) beeinflusst werden. So kann z.B. die Adsorption von Blei an

der SPM-Oberfläche durch Algen und Bakterien gefördert werden (z.B. Nelson et al.,

1999). Die Bioakkumulation von u.a. Nickel (z.B. Hong et al., 2009), Cobalt (z.B.

Nolan et al., 1992) und Kupfer (z.B. Zamuda and Sunda, 1982) ist nachgewiesen

worden. Außerdem kann eine Algenblüte den pH-Wert steigen lassen und darüber

hinaus - wie eben diskutiert - das Adsorptionsverhalten der Spurenmetalle

beeinflussen (Hatje et al., 2003a).

Allerdings ist die biologische Verfügbarkeit der Spurenmetalle durch ihre

Ionenaktivität beschränkt (z.B. Zamuda and Sunda, 1982); anders ausgedrückt, nicht

alle Spurenmetallspezies sind biologisch verfügbar. Abgesehen von elementaren


- 17 -

Unterschieden sind die unkomplexierten Metallionen generell eher biologisch

verfügbar als die organisch komplexierten (z.B. Helmers, 1994).

1.4. Ästuarine Mischungsprozesse für Spurenmetalle

Aufgrund von Änderungen bei den oben genannten ästuarinen Variablen spielen sich

unterschiedliche chemische, physikalische und biologische Prozesse im Ästuar ab.

Dadurch lässt sich der Aggregatszustand der Spurenmetalle verändern.

Eisenkolloide koagulieren beispielsweise gern mit organischen Materien und flocken

unter ästuarinen Bedingungen aus (z.B. Mylon et al., 2004). Auch Blei wird leicht aus

der Lösung entfernt als Konsequenz der Adsorption an Eisen- oder Mangan-Kolloide

sowie an das SPM (z.B. Waeles et al., 2008b). Außerdem können z.B. Nickel und

Cadmium (Luoma et al., 1998) sowie Kupfer (Sharp et al., 1982) durch biologische

Aktivitäten aus der wässrigen Phase entfernt werden. Auf der anderen Seite ist auch

eine Freisetzung von Spurenmetallen aus dem Sediment oder aus dem SPM möglich.

Nickel verliert z.B. mit steigender Präsenz von Calcium- und Magnesiumionen

zunehmend seine Bindungsstellen auf der organischen Materie (z.B. Mandal et al.,

2002). Cobalt kann reduktiv aus dem Sediment ins Wasser freigesetzt werden (z.B.

Audry et al., 2006).

Außer zu Wechseln beim Aggregatszustand kann es in Ästuaren auch zu

Modifikationen bei der Komplexbildung der Spurenmetalle geben. Kupfer und

Cadmium können beispielsweise durch Abbau der organischen Substanz ihren

kolloidalen Charakter verlieren (z.B. Masson et al., 2011; Waeles et al., 2005). Der

Anteil des organisch gebundenen Nickels kann in Ästuaren auch abnehmen (Turner

and Martino, 2006).

Kapitel 2: Material und Methoden

- 18 -

2. Material und Methoden

2.1. Studiengebiete

Die Feldarbeiten wurden im Rahmen der Projekte LANCET und POLCAMAR auf der

Südchinesischen Insel Hainan bzw. an der Küste Nordostbrasiliens durchgeführt.

2.1.1. Ästuare auf der südchinesischen Insel Hainan

Abbildung 2. Die Studiengebiete: (a) Wenchang-Wenjiao Ästuar und (b) Wanquan Ästuar in Ost-Hainan.

Die tropische Insel Hainan liegt zwischen 18°09’-20°10’N und 108°03’-111°03’E im

südchinesischen Meer. Es herrschen warme und feuchte klimatische Bedingungen.

Das Wanquan Ästuar und das Wenchang-Wenjiao Ästuar liegen beide an der

Ostküste der chinesischen Insel (Abb. 2). In diesem Studiengebiet fallen etwa 80%

der ganzjährigen Niederschläge von 1740 mm in der Regensaison zwischen Mai und

November (Ma et al., 2007; Liu et al., 2011).

Der Wanquan Fluss ist mit einer Länge von 156.6 km, einem Einzugsgebiet von

3693 km2 und einem durchschnittlichen Wassereintrag von 163.9 m3/s der drittgrößte

Fluss auf Hainan (Ge et al., 2003). Der Wenchang Fluss (Länge 37 km,

Einzugsgebiet 381 km2, durchschnittlicher Wassereintrag 9.1 m3/s) und der Wenjiao

Fluss (Länge 56 km, Einzugsgebiet 522 km2, durchschnittlicher Wassereintrag 11.6

m3/s) münden beide in die Bamen Bucht (Wasseroberfläche ca. 40 km2; Herbeck et

al., 2013) und stellen gemeinsam das Wenchang-Wenjiao Ästuar dar (Zeng and

Zeng, 1989). Je nach Gezeiten kann das Salzwasser jeweils von der Yudai


- 19 -

Sandbarriere und von der Bamen Bucht 5 km flussaufwärts in die Ästuare eindringen.

Die Lage der Mischzone für Frisch- und Salzwasser im Wanquan Ästuar ist

angesichts des Zusammenwirkens von Frischwassereintrag sowie Tiden- und

Windeinfluss äußerst variabel. Hingegen bleibt im seichten Wenchang-Wenjiao

Ästuar (< 3 m, die Schiffroute ausgenommen; Fu et al., 2013) der Salinitätsgradient

relativ stabil. Die Residenzzeit des Wassers beträgt 0.2-4.7 Tage im Wanquan

Ästuar (Li et al., 2013) und 7.8 Tage im Wenchang-Wenjiao Ästuar (Liu et al., 2011).

Im Vergleich zu dem tourismusorientierten und umweltgeschützten Wanquan Ästuar

(z.B. Gao et al., 2004) ist das Wenchang-Wenjiao Ästuar von der Aquakultur geprägt.

Die gesamte von der Aquakultur bewirtschaftete Fläche in der Region wird auf ca.

21.6 km2 mit einem jährlichen Abwassereintrag ins Ästuar von ca. 210·106 m3

geschätzt (Herbeck et al., 2013; Unger et al., 2013). Dadurch ist das Wenchang-

Wenjiao Ästuar mit einiger Wahrscheinlichkeit umweltbelastet.

2.1.2. Lagunen-System in Nordost-Brasilien

Abbildung 3. Das Studiengebiet Mundaú-Manguaba Lagunen-System in Nordost-Brasilien.

Das Mundaú-Manguaba Lagunen-System liegt im nordostbrasilianischen

Bundesstaat Alagoas (Abb. 3). Seine tropische Lage deutet darauf hin, dass das


- 20 -

Klima dort warm und feucht ist. Es herrscht eine klar definierte Regensaison von

April bis August und eine Trockensaison von September bis März (Oliveira and

Kjerfve, 1993; Brockmeyer and Spitzy, 2011). Die durchschnittliche Temperatur

beträgt etwa 24°C (Diegues, 1994).

Das Mundaú-Manguaba Lagunen-System besteht aus zwei Lagunen. Die nördlich

liegende Mundaú Lagune hat den Rio Mundaú (Einzugsgebiet 2135 km2) als

Hauptfrischwasserzufuhr, eine Wasseroberfläche von 24 km2 und eine

durchschnittliche Wassertiefe von 1.5 m (Melo-Magalhães et al., 2009). Der Rio

Paraíba do Meio mit einem Einzugsgebiet von 3299 km2 ist der Hauptzufluss für die

südlich liegende Manguaba Lagune, welche eine Wasseroberfläche von 43 km2 und

eine durchschnittliche Wassertiefe von 2 m hat (z.B. De Souza et al, 2002; Melo-

Magalhães et al., 2009; Costa et al, 2011). Die beiden Lagunen sind miteinander

durch mehrere schmale Kanäle verbunden. Durch eine einzige Öffnung fließt das

Wasser aus den Kanälen in den Atlantik. Durch das Kanalsystem wird das

Eindringen von Salzwasser weitgehend gedämpft. Solche sogenannten

„choked“ Lagunen sind durch lange Residenzzeiten und die Anfälligkeit gegenüber

anthropogenen Beeinträchtigungen charakterisiert (Knoppers et al., 1991). Vor allem

in der Regensaison sind die beiden Lagunen vom Eintrag aus den Flüssen dominiert

(Spörl, 2011). Die durchschnittliche Residenzzeit beträgt 1-2 Wochen für die Mundaú

Lagune und 5-7 Wochen für die Manguaba Lagune (Oliveira and Kjerfve, 1993).

In den letzten Jahrzehnten ist das Mundaú-Manguaba Lagunen-System von der

Zuckerrohrmonokultur in den Einzugsgebieten seiner Zuflüsse stark geprägt worden.

Dies führt zu erhöhten Nährstoffeinträgen in die Lagunen und kann saisonale

Unterschiede in den Nährstoffkonzentrationen verursachen (Spörl, 2011). Darüber

hinaus liegt die Millionenstadt Maceío direkt am Ufer der Mundaú Lagune, was das

System zusätzlich belastet.

2.2. Maßnahmen zur Kontaminationsminderung

Die Spurenmetallkonzentration im Ästuar, insbesondere im Meerwasser-Endglied, ist

meistens sehr gering (z.B. Bruland, 1983). Während die analytischen Methoden

immer empfindlicher geworden sind, müssen mögliche Kontaminationen während der

Aufarbeitung als ein einschränkender Faktor gesehen werden (z.B. Kosta, 1982).

Aufgrund dessen fanden außer Probenahme und Filtration/Ultrafiltration alle Labor-

Aktivitäten in einem Reinraumlabor (Class 1000) statt; die Probenbehandlungen

wurden unter einer zusätzlichen Clean-Bench (Class 100) durchgeführt. Zum Einsatz


- 21 -

kamen ausschließlich supra-pur®-Chemikalien, außer Salpetersäure und "Freon"

(1,1,2-Trichlor-trifluorethan), die unter Reinraumbedingung destilliert wurden (sub-

boiled HNO3). Alle verwendeten Materialien (u.a. Gefäße) wurden zur Entfettung für

drei Tage in ein 2%-iges Mucasol-Bad (anionisches Tensid) eingelegt, danach mit

warmem Leitungswasser und dann mit Milli-Q®-Wasser gründlich gespült. Um die

Metalle an der Oberfläche der Materialien zu entfernen, wurden die Materialien

anschließend in Salzsäure (3 mol/L) und später in Salpetersäure (3 mol/L) für jeweils

drei Tage eingelegt. Dazwischen und am Ende wurden die Materialien mit Milli-Q®-

Wasser gründlich gespült und dann unter einer Clean-Bench getrocknet ("supra-

sauber"). Die Pipettenspitzen und Autosamplercups für das Massenspektrometer mit

induktiv gekoppeltem Plasma wurden für ca. eine Woche in Salpetersäure (1 mol/L)

eingelegt, mit Milli-Q®-Wasser gründlich gespült und unter einer Clean-Bench

getrocknet (Salpetersäure-sauber). Für einige Materialien, wie z.B. die Gefäße für

Mikrowellen-Aufschlüsse und Querstrom-Ultrafiltrations-Kassetten waren

methodenspezifische Reinigungsverfahren gefordert, welche in den jeweiligen

Kapiteln behandelt werden.

2.3. Probenahme

Die Probenahme fand in den jeweiligen Studiengebieten der südchinesischen Insel

Hainan bzw. in Nordost-Brasilien statt.

Die Wanquan und Wenchang-Wenjiao Ästuare wurden in der Trockensaison 2006

(Dezember), in der Regensaison 2007 (August) und in der Regensaison 2008

(Juli/August) beprobt.

Die Probenahme im Mundaú-Manguaba Lagunen-System erfolgte in der

Regensaison 2007 (September/Oktober), der Trockensaison 2008 (Februar) und der

Trockensaison 2009 (März).

Oberflächenwasserproben aus einer Tiefe von ca. 0.3 m wurden flussaufwärts von

einem Boot aus direkt in eine Polyethylenflasche (LDPE, Nalgene®) gefüllt. Vor der

Behandlung wurden die Proben gekühlt gelagert.

2.4. Filtration und Querstrom-Ultrafiltration

Im Labor vor Ort wurden die Oberflächenwasserproben innerhalb von 24 Stunden

filtriert und anschließend ultra-filtriert. Polycarbonatfilter (Nuclepore®, Porengröße

0.45 μm; vorher mit Salpetersäure gereinigt) wurden benutzt, um die Proben zu


- 22 -

filtrieren. Die mit Filterkuchen beladenen Filter wurden in einer Petrischale und das

Filtrat in einer Polyethylenflasche (LDPE, Nalgene®) aufbewahrt. Bevor ein Filtrat mit

konzentrierter Salpetersäure (sub-boiled, Verhältnis zur Probe 1:1000) angesäuert

wurde, wurden 10 mL davon für die Bestimmung von gelöstem organischem

Kohlenstoff (dissolved organic carbon; DOC) für ein anderes Projekt in eine

Glassampulle abgefüllt, mit Phosphorsäure (Verhältnis zur Probe 1:100) angesäuert

und durch Abschmelzen verschlossen. Etwa 1 L des Filtrats wurde nicht angesäuert

und stand für die Querstrom-Ultrafiltration zur Verfügung.

Die Querstrom-Ultrafiltration (Cross-Flow-Filtration, Tangential-Flow-Filtration) wird

seit den 90-er Jahren des letzten Jahrhunderts vermehrt in der Ästuarforschung

eingesetzt, um die Kolloide nach ihrer physikalischen Größe zu trennen, bzw. um die

kolloidalen Spezies in Ästuaren zu studieren (z.B. Whitehouse et al., 1990; Buesseler

et al., 1996; Guo and Santschi, 1996; Gustafsson et al., 1996; Wen et al., 1996; Guo

and Santschi, 2007). Diese Technik beruht auf einen Membran-Trennverfahren, bei

dem durch die Anströmung quer zur Filtrationsrichtung an der Membranoberfläche

eine hohe Strömungsgeschwindigkeit erzeugt wird, um die Ablagerung von Kolloiden

bzw. die Bildung von Deckschichten möglichst weitgehend zu verhindern. Da die

Eigenschaften der dennoch auftretenden Ablagerungen sich im Verlauf der Filtration

ändern können (z.B. Gustafsson et al., 1996), spielt der Konzentrationsfaktor CF eine

gewisse Rolle, welcher gegeben ist durch:

VVV

VVCF

R

PR

R

+== 0

Dabei ist VR das Volumen des Retentats, VP das Volumen des Permeats und V0 das

Ausgangsvolumen. Zur Überprüfung der Effizienz von Querstrom-Ultrafiltrationen

wird der Begriff der „Wiederfindung“ verwendet, welche dem Prozentsatz der Summe

der Gehalte aller erhaltenen Fraktionen bezogen auf die Ausgangsmenge entspricht

(z.B. Whitehouse et al., 1990):

100(%)00

××

×+×=

VCVCVCungWiederfind RRPP

CP, CR und V0 stehen jeweils für die Konzentration des Permeats, die des Retentats

und die Ausgangskonzentration. Bei einer Wiederfindung größer als 100% kommt

Kontamination in Betracht (z.B. Martin et al., 1995), eine kleinere Wiederfindung als

100% deutet auf einen Verlust während der Querstrom-Ultrafiltration hin. Um einen

hohen Konzentrationsfaktor zu erreichen, braucht die Querstrom-Ultrafiltration zwar

viel Probe (Waeles et al., 2008a); vorteilhaft ist jedoch, dass dadurch die späteren


- 23 -

Analyten im Retentat angereichert werden und so die Anforderungen an die

Messinstrumente reduziert werden.

Um die Bedeutung des Konzentrationsfaktors gibt es erhebliche Uneinigkeiten. Für

Meerwasser wurde z.B. ein Konzentrationsfaktor von über 40 empfohlen (Guo and

Santschi, 1996). Auf der anderen Seite kann ein hoher Konzentrationsfaktor jedoch

zum Durchbruch des Retentats durch die Membran führen (z.B. Dai et al., 1998).

Auch Matrix-Effekte (z.B. Whitehouse et al., 1990) und Aggregation von Kolloiden

(z.B. Waeles et al., 2008a) können vermehrt auftreten. In manchen Literaturstellen

wird ein Konzentrationsfaktor von 5-10 als optimaler Konzentrationsfaktor angesehen

(z.B. Wen et al., 1996). Außerdem verändert sich die Performance der

Ultrafiltrations-Membran je nach Konzentration und Natur der Probe (z.B. Baker and

Strathmann, 1970). Eine Übereinstimmung über einen optimalen

Konzentrationsfaktor ist noch nicht gegeben (Guo et al., 2000b).

Außer der Querstrom-Ultrafiltration ist die Frontal-Ultrafiltration (z.B. Jiann et al.,

2005; Guo et al., 2011) eine gängige Methode, um die Kolloide nach ihrer

physikalischen Größe zu trennen. Diese Technik ist ausgezeichnet vor allem durch

einfache Bedienung (Waeles et al., 2008a), führt jedoch leicht zu Überladung des

Filters, welche die effektive Porengröße des Filters vermindern und eine einheitliche

Fraktionierung von Kolloiden verhindern können (z.B. Benoit and Rozan, 1999).

Deswegen hat man sich in meisten Kolloidstudien für die Querstrom-Ultrafiltration

entschieden.

Verschiedene Querstrom-Ultrafiltrations-Systeme wie z.B. Amicon® (z.B. Benoit et al.,

1994; Tang et al., 2001), Millipore Pellicon® (z.B. Whitehouse et al., 1990; Dai et al.,

1995) und Vivaflow® (z.B. Gaffney et al., 2008; Schlosser and Croot, 2008) wurden

entwickelt. Im Wesentlichen unterscheiden sich diese Systeme hinsichtlich des

Designs, des Membranmaterials und der Membranoberfläche (z.B. Guo and Santschi,

2007). Ein System aus Edelstahl ist gut geeignet für Fraktionierung organischer

Materien, während bei der Spurenmetallanalyse Kunststoffmaterialien einzusetzen

sind. Eine Tiefenfilter-Membran kann beispielsweise nur 60% Poren haben, die der

effektiven Filtrationsrate entsprechen, während die anderen 40% entweder größer

oder kleiner sind (Guo and Santschi, 2007). Zudem ist die Sorptionskapazität der

Membran auch von Material zu Material unterschiedlich. Eine große aktive

Membranoberfläche ermöglicht zwar die schnelle Ultrafiltration, damit vergrößert sich

jedoch gleichzeitig das Totvolumen des Systems.

Da die wichtigen kolloidalen Spurenmetallenträger, wie Huminstoffe und


- 24 -

Eisen(hydr)oxide, unterschiedliche Größen und -bereiche aufweisen (z.B. Lead and

Wilkinson, 2007), wird manchmal eine Mehrfachfraktionierung von Kolloiden

gewünscht. Dafür wird der Begriff NMWCO (nominal molecular weight cut-off) als

Maßstab eingesetzt, um die Membranen zu charakterisieren. Bei einem definierten

NMWCO werden rund 90% der spezifischen globularen Moleküle, die größer als die

Poren der Membran sind, zurückgehalten (z.B. Guo and Santschi, 2007). Da

natürliche Kolloide selten kugelförmig sind, kann die reale Filtrationsrate jedoch von

der idealen abweichen.

Konzeptionell kommen Kaskaden- und Parallel-Prozeduren in Betracht. Bei der

Kaskaden-Prozedur werden die Membranen nach Porengrößen bzw. NMWCOs

sukzessiv zugeschaltet. Je nach der Anforderung werden aus einer Probe mehrere

Subproben mit unterschiedlichen Kolloidgrößen. Das Permeat der jeweils

vorausgegangenen Ultrafiltration wird als Ausgangsprobe des nachfolgenden

Ultrafiltrationsvorgangs eingesetzt. Diese Prozedur hat den Vorteil, dass die Kolloide

nach Größe nacheinander entfernt werden. Dadurch wird die Konzentration des

Retentats möglichst klein gehalten und die Aggregation von Kolloiden an der

Membranoberfläche minimiert (Waeles et al., 2008a). Da bei der Querstrom-

Ultrafiltration nicht nur größere Kolloide in konzentrierter Form, sondern auch

kleinere Kolloide im Retentat vorhanden ist, wird die Konzentration der größeren

Kolloide CCol wie folgt berechnet:

CFCCC PR

Col−

=

Dabei stellt die Fortpflanzung von analytischen Fehlern das größte Problem dar.

Besonders bei der kleinsten "Kolloid"-Fraktion, die das Permeat (sog. "echt gelöste"

Fraktion) des letzten Querstrom-Ultrafiltrationsvorgangs darstellt und nicht

aufkonzentriert ist, kann eine Kontamination schwerwiegende Folgen für die

Berechnung der Konzentrationen größerer Kolloid-Fraktionen haben. Vorteilhafter in

Hinblick darauf ist die Parallel-Prozedur (z.B. Laxen and Chandler, 1983), bei der

mehrere Aliquote einer einzigen Probe durch Filter unterschiedlicher Porengröße

parallel ultrafiltriert werden, um Fraktionen unterschiedlicher Kolloidgröße zu erhalten.

Dadurch wird die Fehlerfortpflanzung zwar verhindert, aber besonders bei der

Fraktionierung kleiner Kolloidgrößen kommt es leicht zur Aggregation an der

Membranoberfläche und die Modifikation des Retentats ist schwer zu vermeiden.


- 25 -

Abbildung 4. Schematische Darstellung der Querstrom-Ultrafiltrations-Kassette Vivaflow® von Sartorius (www.sartorius-stedim.com).

Abbildung 5. Ablauf der Querstrom-Ultrafiltrations-Prozedur.

In dieser Arbeit wurden Querstrom-Ultrafiltrations-Kassetten aus Polycarbonat

(Vivaflow® 200, Sartorius; Abb. 4) mit einer Polyethersulfon-Membran und einer


- 26 -

aktiven Oberfläche von 200 cm2 eingesetzt; sie wurden mit Porengrößen von 30 kDa,

10 kDa und 5 kDa kaskadisch betrieben. Das System ist vor allem ausgezeichnet

durch hohe Fluss- und gute Wiederfindungsraten. Die Querstrom-Ultrafiltrations-

Kassetten wurden vor Gebrauch jeweils 30 Minuten lang erst mit Salzsäure (1%) und

dann mit EDTA (10 mmol/L) durchspült. Am Anfang, zwischen den beiden

Reinigungsprozeduren und am Ende wurden die Querstrom-Ultrafiltrations-Kassetten

jeweils mit großen Mengen an Milli-Q®-Wasser gespült. Ein schematischer Ablauf der

Querstrom-Ultrafiltrations-Prozedur einschließlich der Filtrationsprozedur ist in Abb. 5

dargestellt. Nach der konventionellen Filtration (0.45 μm Filter) wurden die Filtrate

zuerst durch eine Membran mit einem NMWCO von 30 kDa ultrafiltriert. Die

Fliessgeschwindigkeit der Probe durch die Querstrom-Ultrafiltrations-Kassette betrug

dabei etwa 500 mL pro Minute bei einem Druck von ca. 2 bar. Das Permeat wurde

als Ausgangsprobe für nachfolgende Fraktionierung von Kolloiden mit einem

NMWCO von 10 kDa verwendet. Schließlich wurde das daraus entstandene Permeat

weiter als Ausgangsprobe für die Fraktionierung der 5 kDa Kolloide eingesetzt. Nach

der Querstrom-Ultrafiltration wurden die drei Retentate und das Permeat (< 5 kDa)

jeweils in eine Polyethylenflasche (Nalgene®) abgefüllt und mit konzentrierter

Salpetersäure auf einen pH-Wert von ca. 2 angesäuert. Zur Volumenbestimmung

wurden die drei Retentate und das Permeat sowie die Ausgangsprobe vorher

gewogen. Jeweils 10 mL von den Retentaten und dem Permeat wurden ebenfalls vor

der Ansäuerung für die DOC-Bestimmung in eine Glassampulle abgefüllt, mit

Phosphorsäure angesäuert und abgeschmolzen. Die durchschnittlichen

Konzentrationsfaktoren betrugen ca. 7.5 für die Fraktion mit einer Größe zwischen 30

kDa und 0.45 μm, ca. 6 für die Fraktion mit einer Größe zwischen 10 kDa und 30 kDa,

und ca. 4.5 für die Fraktion mit einer Größe zwischen 5 kDa und 10 kDa. Um die

Membranen zu konditionieren, wurden die Kassetten vor jeder Ultrafiltration mit der

zu filtrierenden Probe gespült.

Bei der Untersuchung hat sich herausgestellt, dass die Kolloidfraktion mit einer

Größe zwischen 10 kDa und 30 kDa im Hinblick auf die ästuarine Modifikation relativ

unrelevant ist. In der folgenden Diskussion wurde sie daher mit der Kolloidfraktion mit

einer Größe zwischen 30 kDa und 0.45 μm zusammen betrachtet und HMW-Fraktion

(high molecular weight) genannt. Die Kolloidfraktion mit einer Größe zwischen 5 kDa

und 10 kDa und die echt gelöste Fraktion (< 5 kDa) wurden als LMW- (low molecular

weight) bzw. TD-Fraktion (truly dissolved) bezeichnet.

Die Querstrom-Ultrafiltrations-Technik - und die gegenwärtigen Techniken zur

Kolloidfraktionierung überhaupt - basieren hauptsächlich auf der physikalischen


- 27 -

Abtrennung von Kolloiden (z.B. Lyvén et al., 2003). Dabei sind die Struktur und das

Verhalten von Kolloiden nicht unbedingt von ihren physikalischen Größen abhängig

(Lead and Wilkinson, 2007). Neben der chemischen Zusammensetzung sind

Faktoren wie Beladung, funktionelle Gruppen und spezifische Oberfläche ebenfalls

von Bedeutung. Wie oben schon erwähnt, können die Ablagerung und der

Durchbruch von Kolloiden die Separation beeinträchtigen. Eine Ablagerung kann

jedoch unter Umständen effektiv bekämpft werden (z.B. Schlosser and Croot, 2008),

während der Durchbruch durch Verringerung des Konzentrationsfaktors weitgehend

zu dämpfen ist (Guo et al., 2000b). Gleichzeitig wurden auch neue Techniken wie z.B.

Flow-Field-Flow-Fraktionierung (z.B. Stolpe et al., 2010) entwickelt. Solche Verfahren

können zum besseren Verstehen von ästuarinen Kolloiden beitragen.

2.5. UV-Aufschluss

Organische Moleküle und organische Partikel können bekanntlich mit Metallionen

Komplexe bilden, die im Wasser sehr stabil und zum Teil resistent gegen Säure sind

(z.B. Golimowski and Golimowska, 1996). Infolgedessen wird traditionell ein UV-

Aufschluss-Verfahren unter Anwesenheit von Oxidationsmitteln bei der Analyse

natürlicher Wässer angewendet, um die organischen Matrix zu zerstören bzw. die zu

untersuchenden Metallionen zu befreien (z.B. Achterberg et al., 2001). Im Reinraum

wurden alle wässrigen Proben (im Wesentlichen nach: Danielsson et al., 1978) unter

Einsatz von Wasserstoffperoxid (30%) als Oxidationsmittel (Verhältnis zur Probe

1:100) in Quarzglasgefäßen für 2 Stunden mit UV-Licht bestrahlt. Nach der

Behandlung wurden die Proben wieder in eine vorgereinigte LDPE-Flasche gefüllt

und waren bereit für den nächsten Verarbeitungsschritt.

Zur Reinigung wurden die Quarzgefäße zuerst in Mucasol (5%) und anschließend in

Salpetersäure (3 mol/L) für jeweils zwei Stunden mit Ultraschall behandelt. Nach

dem Mucasol-Bad wurden sie mit warmem Leitungswasser und dann mit Milli-Q®-

Wasser gründlich gespült; nach dem Salpetersäure-Bad erfolgte ebenfalls ein

Spülvorgang mit Milli-Q®-Wasser. Schließlich wurden sie unter einer Clean-Bench

getrocknet.

2.6. Flüssig-Flüssig-Extraktion

Die Flüssig-Flüssig-Extraktion ist eine gebräuchliche Methode zur Untersuchung von

Spurenmetallen in Meerwasser bzw. in ästuarinen Wässern (z.B. Komjarova and


- 28 -

Blust, 2006), da der hohe Salzanteil im Meerwasser und Ästuarwasser die

Auswertung der analytischen Signale stören und den hochempfindlichen Detektor

des Messinstruments beschädigen kann. Darüber hinaus treten in solchen Medien

häufig Matrixeffekte auf, was zu Verfälschungen des Messergebnisses führen kann

(z.B. Rodushkin and Ruth, 1997). Durch die Flüssig-Flüssig-Extraktion werden die

störenden Alkali- und Erdalkali-Ionen ausgeschlossen und Matrixeffekte weitgehend

minimiert. Die zu untersuchenden Spurenmetalle werden zudem aufkonzentriert.

Trotz aller Vorteile ist das Verfahren jedoch relativ aufwendig und benötigt eine

Reihe von Instrumenten und Chemikalien. Eine Kontamination während der

Verarbeitung gilt es zu verhindern.

Im Rahmen dieser Arbeit wurde eine Flüssig-Flüssig-Extraktions-Methode im

Wesentlichen nach Bruland et al. (1985) entwickelt. Zuerst wurden je nach

Einschätzung der Probekonzentration 40-100 g Probe in einen fluorierten Ethylen-

Propylen Scheidetrichter (FEP, Nalgene®) eingewogen. Die Lösung wurde mit

Ammoniumacetat-Puffer (selbst hergestellt aus Ammoniak und Essigsäure) auf einen

pH-Wert von 4.0-4.5 gebracht. Als Chelatbildner wurden 500 μL einer Mischung von

Natriumdiethyldithiocarbamat und Ammoniumpyrrolidindithiocarbamat (Verhältnis 1:1,

jeweils 0.05 mol/L) zugesetzt und durchmischt. Nach Zugabe von 10 mL Freon

wurde der Scheidetrichter für 15 Minuten geschüttelt. Nach der Phasenseparation

wurde die untere organische Phase in ein Polypropylenröhrchen abgelassen und mit

100 μL konzentrierter Salpetersäure versetzt. Das Polypropylenröhrchen wurde

einschließend kräftig geschüttelt und für kurze Zeit ruhig gestellt, bevor 1900 μL Milli-

Q®-Wasser zugegeben und das Polypropylenröhrchen erneut 3 Minuten lang

geschüttelt wurde. Die obere wässrige Phase wurde schließlich mit einer Pipette

sorgfältig abgezogen und in einem Polypropylenröhrchen aufbewahrt.

Um die Effizienz des Verfahrens zu überprüfen, wurden einige Süßwasserproben

direkt vermessen. Ein Vergleich der Ergebnisse mit flüssig-flüssig-extrahierten

Proben wies nur auf geringfügige Abweichungen hin (< 5%).

2.7. Mikrowellen-Aufschluss-Verfahren

Das Mikrowellen-Aufschluss-Verfahren kommt in der analytischen Chemie und

Umweltchemie häufig zum Einsatz (z.B. Zlotorzynski, 1995), da es in der

Spurenmetallanalyse eine schnelle und effiziente Methode zur Überführung der

Festproben in die lösliche Phase darstellt (z.B. Sandroni et al., 2003). Dieses

Verfahren wird in die Gruppe der „Nassaufschlussmethoden unter Verwendung


- 29 -

lösender und/oder oxidierender Aufschlussreagenzien im geschlossenen

System“ einordnet, die dadurch die Anwendung höherer Temperaturen erlauben

(Stoeppler, 1994).

Tabelle 1. Mikrowellen-Programme einschließlich der Parameter.

Mikrowellen-Programm Zeit maximale Temperatur

maximale Leistung

Filterzersetzen mit Zusatz von 5 mL HNO3 (65%) 30 Min 60 °C 250 W

Druckaufschluss mit Zusatz von 5 mL HNO3 (65%), 1 mL HF (40%), 0.5 mL HCl (30%) und

0.5 mL HClO4 (70%)

5 Min 60 °C 250 W

1 Min 60 °C 0 W

5 Min 90 °C 300 W

5 Min 90 °C 400 W

4 Min 90 °C 600 W

3 Min 90 °C 400 W

10 Min 90 °C 350 W

erstes Abrauchen 30 Min 100 °C 300 W

zweites Abrauchen mit Zusatz von 3 mL HNO3 (65%) 30 Min 100 °C 300 W

Lösen mit Zusatz von 5.00 mL HNO3 (0.5 mol/L) 15 Min 65 °C 320 W

Reinigung mit Zusatz von 5 mL HNO3 (65%), 1 mL HF (40%) und 1 mL HClO4 (70%)

5 Min 80 °C 300 W

1 Min 80 °C 0 W

2 Min 80 °C 400 W

1 Min 100 °C 0 W

2 Min 80 °C 500 W

Im dieser Arbeit wurde die SPM-Probe einschließlich Filter nach Schüßler et al.

(2005) zunächst gewogen und dann in einem Teflon-Gefäß mit entsprechendem

Mikrowellen-Programm aufgeschlossen. Die Parameter des Mikrowellen-Programms

sind in Tab. 1 aufgelistet. Sobald die eingestellte maximale Temperatur im Gefäß

erreicht wurde, wurde die Mikrowellen-Leistung heruntergefahren, um die

Temperatur konstant zu halten. Am Ende wurde die Probe in genau 5.00 mL

Salpetersäure (0.5 mol/L) aufgelöst und war damit bereit für die Analyse


- 30 -

2.8. Massenspektrometrie mit HR-ICP-MS

Die hochauflösende Massenspektrometrie mit induktiv gekoppeltem Plasma (high-

resolution inductively-coupled-plasma mass-spectrometry; HR-ICP-MS) ist eine

relativ neue instrumentelle Technik zur Untersuchung von Spurenelementen in

Umweltmatrices (z.B. Liang et al., 2000). Dabei wird die Probe unter Einsatz von

Argon durch einen hochfrequenten Strom auf 5000 °C bis 10000 °C erhitzt bzw.

ionisiert. Die dabei entstandenen Ionen werden massenspektrometrisch getrennt und

gemessen. Mit dieser Technik ist man in der Lage, Spurenmetallanalysen schnell

und präzise durchzuführen (z.B. Jenner et al., 1990). Im Rahmen dieser Arbeit wurde

ein HR-ICP-MS-Gerät (Element 2, Thermo Scientific) verwendet. Rhodium wurde als

interner Standard eingesetzt. Vor jeder Messung wurde zuerst ein Tuning des ICP-

MS durchgeführt. Zum Einsatz gekommen sind dabei eine Gasflussrate von ca. 16

L/min für cool gas, ca. 1 L/min für sample gas und ca. 1 L/min für auxiliary gas sowie

eine Generatorleistung von 1200 W. Anschließend wurde das Mass-Offset bestimmt.

Die Konzentrationen von Eisen, Nickel, Cobalt, Kupfer, Cadmium und Blei wurden

mittels externer Kalibrierung quantifiziert. Für die externe Kalibrierung wurden 7

Kalibrationspunkte verwendet. Mit einem Bestimmtheitsmaß R2 ≥ 0.999995 wurden

die Probekonzentrationen anhand der Kalibrationsgleichung berechnet. Überschritt

eine gemessene Konzentration den Kalibrationsbereich, erfolgte eine erneute

Bestimmung nach einer Verdünnung der Probe. Innerhalb einer Messung wurden

routinemäßig mehrfach die Konzentrationen der Blanks (0.5 mol/L HNO3) vermessen.

Bei der Berechnung wurden die Mittelwerte von den Blanks für die jeweiligen Metalle

ermittelt und von den Probekonzentrationen abgezogen.

2.9. Referenzmaterialien

Spurenmetalle sind in der Hydrosphäre nur in geringen Mengen vorhanden (z.B.

Chester, 1990). Bezogen auf diese Arbeit schwankten die Konzentrationen in Lösung

meist im Bereich von nmol/L. Entsprechend musste großer Wert auf saubere und

präzise Arbeit gelegt werden. Mögliche Kontaminationsquellen wurden bereits

mehrfach erwähnt. Gegenmaßnahmen, wie z.B. die sorgfältige Probenahme, die

Aufarbeitung, die gründliche Reinigung von Chemikalien und Gefäßen sowie das

Arbeiten im Reinraum und unter einer Clean-Bench, sollten kontaminationsarme

Ergebnisse ermöglichen.

Wenn die Konzentration eines bestimmten Analyts sich der Konzentration Null nähert,

droht das Signal im Grundrauschen zu verschwinden. Man ist ständig konfrontiert mit


- 31 -

der Frage, ob solche Signale auswertbar sind. In der analytischen Chemie wurde der

Begriff „Nachweisgrenze (detection limit)“ eingeführt, um die Aussagekraft solcher

Signale quantitativ zu bewerten (z.B. Willard et al., 1988). Häufig ist die

Nachweisgrenze NG gegeben durch:

σ3+= XNG

wobei X für den durchschnittlichen Blank-Wert und σ für die relative

Standardabweichung steht (z.B. Schwedt, 2008). Die Messwerte, die unterhalb der

Nachweisgrenze liegen, weisen statistisch betrachtet eine große Ungenauigkeit auf,

und werden als „nicht nachweisbar“ bezeichnet.

Für jede Messung durch ICP-MS wurde eine instrumentelle Nachweisgrenze für alle

sechs untersuchten Metalle ermittelt. Im Vergleich zu den Probekonzentrationen

waren die Nachweisgrenzen so klein, dass bis auf einige wenige Kolloid-

Konzentrationen für Cadmium und Eisen alle Metallkonzentrationen deutlich darüber

lagen. Für die Berechnungen der kolloidalen Konzentrationen wurden die Werte

unterhalb der Nachweisgrenze auf Null gesetzt.

Referenzmaterialien finden in der analytischen Chemie regelmäßig Verwendung, z.B.

um eine Messmethode zu beurteilen oder um Apparaturen zu kalibrieren. Mit Hilfe

solcher Referenzmaterialien können die Einflüsse der Matrix auf das analytische

Signal eingeschätzt werden. Zum Einsatz gekommen sind in dieser Arbeit die von

NRCC (National Research Council of Canada) zertifizierten Referenzmaterialien.

SLRS-5 (Flusswasser), SLEW-3 (ästuarines Wasser) sowie CASS-4 und NASS-5

(Meerwasser) wurden für den UV-Aufschluss bzw. für die Flüssig-Flüssig-Extraktion

verwendet. Für den Mikrowellen-Aufschluss kamen MESS-1 und MAG-1 zum Einsatz.

Die Metallkonzentrationen für alle Referenzmaterialien wurden mindestens dreifach

gemessen. Alle Messergebnisse lagen im zertifizierten Wertebereich.

Für die Ultrafiltration gibt es keine Referenzmaterialien. Die Berechnung von

Massenbilanzen (Vergleich der direkt gemessenen Lösungs-Gesamtkonzentration in

einer Probe mit der Summe der Konzentrationen der einzelnen Kolloidfraktionen)

kann jedoch Aufschluss geben über die Effizienz der Prozeduren. Bei Annahme einer

fehlerfreien Gesamtmessung ist eine Massenbilanz kleiner als 100% wahrscheinlich

auf einen Verlust der entsprechenden Metalle durch Adsorption an den Membranfilter

oder an die Oberfläche des Probengefäßes zurückzuführen, während bei einer

Massenbilanz > 100% eine Kontamination oder analytische Schwankungen bei den

Kolloid-Messungen zu vermuten sind (z.B. Martin et al., 1995; Powell et al., 1996).


- 32 -

Tabelle 2. Massenbilanzen der Metalle (N: die gesamte Anzahl der Proben; n: die verwendete Anzahl der Proben).

Studiengebiet Fe Ni Co Cu Cd Pb

Ost-Hainan 75% ± 30%

80% ± 20%

100% ± 30%

80% ± 20%

75% ± 30%

75% ± 30%

N=20 n=12 n=12 n=17 n=18 n=15 n=11

Nordost-Brasilien (Regensaison 2007)

80% ± 40%

110% ± 40%

100% ± 30%

110% ± 30%

90% ± 40%

70% ± 40%

N=15 n=7 n=5 n=9 n=10 n=9 n=8

Nordost-Brasilien (Trockensaison

2009)

100% ± 50%

100% ± 30%

100% ± 20%

100% ± 30%

100% ± 40%

100% ± 50%

N=16 n=7 n=15 n=12 n=14 n=11 n=7

Da die ermittelten Massenbilanzen stark variierten, wurden Niveaus und

Bereichsgrenzen festgelegt (Tab. 2). Nur die Proben, deren Massenbilanz innerhalb

der Bereichsgrenzen lag, wurden weiter verwendet. Obwohl bei den beiden

Projekten nach gleichen Kriterien und mit identischen Materialien gearbeitet wurde,

sind unterschiedliche Ergebnisse herausgekommen. Bei dem LANCET Projekt waren

die Massenbilanzen meist kleiner als 100%. Gleichzeitig wiesen die Massenbilanzen

aus dem POLCAMAR Projekt hohe Wiederfindungsraten auf. Die dafür eingesetzten

Filtrationskartuschen waren zwar von der gleichen Firma, doch wurden sie für die

jeweilige Feldarbeit separat bestellt. Qualitätsunterschiede der Versuchsmaterialien

könnten die unterschiedlichen Ergebnisse verursacht haben. Allerdings traten beim

POLCAMAR Projekt während der Regensaison 2007 teilweise erhöhte

Massenbilanzen auf. Neben der Fortpflanzung analytischer Fehler, die bei der

Berechnung kolloidaler Fraktionen eine Rolle spielen (siehe oben), sind

Kontaminationen trotz aller getroffenen Maßnahmen nicht auszuschließen.

2.10. Andere Parameter

Die Parameter wie z.B. die Salinität, der pH-Wert und die Wassertemperatur wurden

bei den Probenahmen mit einer Multisensor-Einheit (MultiLine® P3, WTW) vor Ort

vermessen. Die Bestimmung von DOC erfolgte durch eine katalytische

Hochtemperatur-Oxidation (Total Organic Carbon Analyzer; Apollo 9000; Tecmar)

mit einer Nachweisgrenze von 0.1 mg/L.

Kapitel 3: Vergleichende Diskussion der Ergebnisse

- 33 -

3. Vergleichende Diskussion der Ergebnisse

Das physikalische Zusammenmischen von Süßwasser und Salzwasser ist einer der

grundlegenden Prozesse in Ästuaren, die die fluvialen chemischen Signale

modifizieren (Chester, 1990). Bei Abwesenheit jeglicher biologischen, geologischen,

chemischen und anthropogenen Prozesse würde die Konzentration einer bestimmten

chemischen Komponente in Ästuaren entlang des Salinitätsgradients linear verlaufen.

In der Realität treten solche Prozesse zumeist auf; Auswirkungen können nicht

ausgeschlossen werden. Wird Linearität beobachtet, sind Einwirkungen gering oder

nicht vorhanden oder sie gleichen sich aus; man spricht von einem konservativen

Mischen. Eine Abweichung der Konzentrationsverteilung von der idealen

Verdünnungslinie wird als non-konservatives Verhalten bezeichnet. Dabei kann eine

Verteilungskurve sowohl über (positive Abweichung) als auch unter (negative

Abweichung) der idealen Verdünnungslinie verlaufen, was einen Eintrag in die

wässrige Phase bzw. eine Entfernung aus der wässrigen Phase bedeutet. Ein

Eintrag kommt zustande, wenn eine Komponente aus der partikulären Phase (SPM

und/oder Sediment) sich zum Teil löst und ins Wasser übertritt, während bei einer

Entfernung die gelöste Komponente an das SPM adsorbiert wird oder Kolloide zu

größeren Partikeln koagulieren. Wegen der Bedeutung von Ästuaren für die

Materialzufuhr zum Ozean hat sich eine Reihe von wissenschaftlichen

Veröffentlichungen mit dem ästuarinen Verhalten von Spurenmetallen

auseinandergesetzt. In Tab. 3 sind einige wichtige Ergebnisse aufgelistet. Sowohl

Eintrag in die wässrige Phase als auch Entfernung aus der wässrigen Phase können

bei allen genannten Metallen festgestellt werden. Außer bei Cobalt wurde

konservatives Verhalten bei allen untersuchten Metallen beobachtet. In manchen

Ästuaren wie z.B. dem Changjiang Ästuar (Wang and Liu, 2003) und der Galveston

Bucht (Wen et al., 1999) sind sogar saisonale Unterschiede beim

Spurenmetallverhalten erkennbar. Diese Vielfalt im Verhalten der Metalle ist das

Resultat einer Verflechtung von ästuarinen Variablen, und deutet auf die Komplexität

der ästuarinen Chemie hin, die sich im Wanquan Ästuar (im Folgenden: WR) und im

Wenchang-Wenjiao Ästuar (im Folgenden: WWR) an der Ost-Küste Hainans bzw. in

der Mundaú-Lagune (im Folgenden: MD) und in der Manguaba-Lagune (im

Folgenden: MG) Nordostbrasiliens widerspiegelt. In den nachfolgenden Kapiteln wird

das Verhalten der sechs verschmutzungsrelevanten Metalle Eisen, Nickel, Cobalt,

Kupfer, Cadmium und Blei in den tropischen Studiengebieten WR, WWR, MD und

MG diskutiert.


- 34 -

Tabelle 3. Ästuarines Verhalten von gelösten Spurenmetallen (P: Positive Abweichung von der idealen Verdünnungslinie; N: Negative Abweichung von der idealen Verdünnungslinie; K: Konservatives Verhalten; --: nicht untersucht).

Studiengebiet Fe Ni Co Cu Cd Pb Literatur

Amazonas Ästuar -- K -- K P -- Boyle et al., 1982

Bristol Kanal und outer Severn Ästuar -- -- -- K K -- Harper, 1991

Changjiang Ästuar N P N P P N Wang and Liu, 2003

Changjiang Ästuar -- -- -- K P K Wang et al., 2009

Danshuei Ästuar -- P P P P -- Fang and Lin, 2002

Delaware Ästuar N N N N N -- Sharp et al., 1982

Donau Ästuar K K P K P -- Guieu et al., 1998

Forth Ästuar N P -- P K N Balls et al., 1994

Galveston Bucht Juli 1993 -- K -- K -- N Wen et al., 1999

Galveston Bucht Mai 1994 -- P -- P -- N Wen et al., 1999

Galveston Bucht Juli 1995 N P P P P N Wen et al., 1999

Gironde Ästuar N P -- P P K Kraepiel et al., 1997

Göta Ästuar N K -- K N N Danielsson et al., 1983

Humber Ästuar -- P -- P P -- Comber et al., 1995

Loire Ästuar -- -- -- P P -- Waeles et al., 2004

Mersey Ästuar -- P P K P P Martino et al., 2002

Mississippi Ästuar K K -- K P -- Shiller and Boyle, 1991

Nerbioi-Ibaizabal Ästuar P K -- K -- -- Fernández et al., 2008

Ob Ästuar N K -- K P N Dai and Martin, 1995

Ochlockonee Ästuar N K -- K K -- Powell et al, 1996

Penzé Ästuar -- -- -- P P N Waeles et al., 2008a

Rhône Ästuar -- K -- K P K Elbaz-Poulichet et al., 1996

Scheldt Ästuar -- P P P P -- Chaudry and Zwolsman, 2008

Seine Ästuar -- P P P P N Chiffoleau et al., 1994

Venedig Lagune -- K -- N P K Martin et al., 1995

Weser Ästuar N P -- P -- N/P Turner et al., 1992

Yenisey Ästuar N P -- K P N Dai and Martin, 1995

3.1. Eisen

Dass Eisen in Ästuaren bei niedrigen Salinitäten ausflockt, ist seit langem etabliert

(z.B. Boyle et al., 1977; Sharp et al., 1982; Danielsson et al., 1983; Turner et al.,

1992; Balls et al., 1994; Dai and Martin, 1995; Powell et al, 1996; Kraepiel et al.,


- 35 -

1997; Wen et al., 1999; Wang and Liu, 2003). Ein signifikanter Anteil (> 80%) des in

den Flüssen transportierten gelösten Eisens wird dadurch aus der wässrigen Phase

entfernt und erreicht nicht den Ozean (z.B. Sholkovitz et al., 1978). Lediglich im

Mississippi Ästuar (Shiller and Boyle, 1991) und im Donau Ästuar (Guieu et al., 1998),

wo gelöstes Eisen im Fluss-Endglied extrem niedrig ist (20-40 nmol/L), ist dieser

Prozess nicht deutlich zu erkennen. Interessanterweise wurde im Nerbioi-Ibaizabal

Ästuar in Spanien eine Erhöhung des gelösten Eisens mit steigender Salinität

festgestellt, wo gelöstes Eisen vergleichsweise sehr hoch ist (1.5-7.6 μmol/L;

Fernández et al., 2008). Die anoxischen Bedingungen ermöglichen vermutlich die

Freisetzung aus partikulärem Eisen (z.B. Pakhomova et al., 2007).

WR

0

5

10

15

0 10 20 30Salinity

Fe (μ

mol

/L)

200620072008

WWR

0

5

10

15

0 10 20 30Salinity

Fe (μ

mol

/L)

MD

0

5

10

15

0 10 20 30Salinity

Fe (μ

mol

/L)

200720082009

MG

0

5

10

15

0 10 20 30Salinity

Fe (μ

mol

/L)

Abbildung 6. Salinitäts-Abhängigkeit des gelösten Eisens.

Mit bis zu 14 μmol/L an der Ost-Küste Hainans bzw. bis zu 6 μmol/L in Nordost-

Brasilien ist die Konzentration des gelösten Eisens in den Studiengebieten ebenfalls

sehr hoch. Eine Anreicherung gelösten Eisens ist jedoch nicht sichtbar. Wie in Abb. 6

dargestellt, sinkt die Konzentration gelösten Eisens bei niedriger Salinität schnell auf

ein Niveau von 1-2 μmol/L, und weiterhin auf unterhalb von 0.5 μmol/L bei hohen

Salinitäten. Gleichzeitig sinkt die partikuläre Konzentration von Eisen von

typischerweise 1.5 mmol/g (schnell v. a. von S = 0 bis S = 10) auf 0.5 mmol/g mit

steigender Salinität (Abb. 7). Die Analyse der Kolloidfraktionen weist darauf hin, dass

bei niedrigen bis mittleren Salinitäten über 80% des gelösten Eisens aus HMW-

Kolloiden besteht. Bei hohen Salinitäten nimmt der HMW-Anteil mehr oder weniger

ab, wobei der Anteil der Eisenkolloide in der LMW-Fraktion und der TD-Fraktion

zunimmt (Abb. 8). Es liegt wahrscheinlich zum einen an der Salz-induzierten


- 36 -

Aggregation (z.B. Nowostawska et al., 2008), durch die die HMW-Fraktion bevorzugt

aus der wässrigen Phase entfernt wird, zum anderen an der möglichen Dissoziation

vom HMW-Eisen wie z.B. Eisenhuminstoffen (z.B. Batchelli et al., 2010).

WR

0.0

0.5

1.0

1.5

2.0

2.5

0 10 20 30Salinity

Fe (m

mol

/g)

200620072008

WWR

0.0

0.5

1.0

1.5

2.0

2.5

0 10 20 30Salinity

Fe (m

mol

/g)

MD

0.0

0.5

1.0

1.5

2.0

2.5

0 10 20 30Salinity

Fe (m

mol

/g)

200720082009

MG

0.0

0.5

1.0

1.5

2.0

2.5

0 10 20 30Salinity

Fe (m

mol

/g)

Abbildung 7. Salinitäts-Abhängigkeit des partikulären Eisens.

WR 2008

0

25

50

75

100

0 10 20 30Salinity

Fe (%

) HMWLMWTD

WWR 2008

0

25

50

75

100

0 10 20 30Salinity

Fe (%

)

MD 2007

0

25

50

75

100

0 10 20 30Salinity

Fe (%

)

HMWLMWTD

MG 2007

0

25

50

75

100

0 10 20 30Salinity

Fe (%

)

MD 2009

0

25

50

75

100

0 10 20 30Salinity

Fe (%

)

HMWLMWTD

MG 2009

0

25

50

75

100

0 10 20 30Salinity

Fe (%

)

Abbildung 8. Salinitäts-Abhängigkeit der kolloidalen Verteilung von Eisen.


- 37 -

3.2. Nickel

Nickel ist relativ unreaktiv und verhält sich in Ästuaren zumeist konservativ (z.B.

Boyle et al., 1982; Danielsson et al., 1983; Shiller and Boyle, 1991; Dai and Martin,

1995; Martin et al., 1995; Elbaz-Poulichet et al., 1996; Powell et al, 1996; Guieu et al.,

1998; Wen et al., 1999; Fernández et al., 2008). Dennoch konnten sowohl positive

(z.B. Campbell et al., 1988; Turner et al., 1992; Balls et al., 1994; Comber et al.,

1995; Dai and Martin, 1995; Kraepiel et al., 1997; Wen et al., 1999; Fang and Lin,

2002; Martino et al., 2002; Wang and Liu, 2003; Chaudry and Zwolsman, 2008) als

auch negative (z.B. Sharp et al., 1982; Turner et al., 1998) Abweichungen von der

theoretischen Verdünnungslinie beobachtet werden. Ein Überschuss des gelösten

Nickels kann sowohl aus dem Sediment (z.B. Martino et al., 2004) als auch aus dem

SPM (z.B. Edmond et al., 1985) stammen. Kinetikstudien bestätigen auch, dass die

Dissoziation des Organonickels durch Konkurrenz zwischen Calcium-/Magnesium-

und Nickelionen um die Bindungsstellen an organischen Materialien begünstigt wird

(z.B. Mandal et al., 2002). Andererseits kann das in die Lösung freigesetzte Nickel

durch die Präsenz von organischen Molekülen wie z.B. Ethylendiamintetraacetat

(Bedsworth and Sedlak, 1999) stabilisiert und die Adsorption von Nickel dadurch

effektiv reduziert werden (Turner et al., 1998; Martino et al., 2003). Außerdem kann

Nickel auch mit Huminstoffen und Eisenoxiden koagulieren (z.B. Sholkovitz, 1978),

oder beispielsweise durch Bioakkumulation aus der wässrigen Phase entfernen

werden (z.B. Hong et al., 2009).

WR

0

5

10

15

20

0 10 20 30Salinity

Ni (

nmol

/L)

200620072008

WWR

0

5

10

15

20

0 10 20 30Salinity

Ni (

nmol

/L)

MD

0

5

10

15

20

0 10 20 30Salinity

Ni (

nmol

/L)

200720082009

MG

0

5

10

15

20

0 10 20 30Salinity

Ni (

nmol

/L)

Abbildung 9. Salinitäts-Abhängigkeit des gelösten Nickels.


- 38 -

WR

0.0

0.5

1.0

1.5

2.0

0 10 20 30Salinity

Ni (μm

ol/g

)

200620072008

WWR

0.0

0.5

1.0

1.5

2.0

0 10 20 30Salinity

Ni (μm

ol/g

)

MD

0.0

0.5

1.0

1.5

2.0

0 10 20 30Salinity

Ni (μm

ol/g

)

200720082009

MG

0.0

0.5

1.0

1.5

2.0

0 10 20 30Salinity

Ni (μm

ol/g

)

Abbildung 10. Salinitäts-Abhängigkeit des partikulären Nickels.

WR 2008

0

25

50

75

100

0 10 20 30Salinity

Ni (

%)

HMWLMWTD

WWR 2008

0

25

50

75

100

0 10 20 30Salinity

Ni (

%)

MD 2007

0

25

50

75

100

0 10 20 30Salinity

Ni (

%)

HMWLMWTD

MG 2007

0

25

50

75

100

0 10 20 30Salinity

Ni (

%)

MD 2009

0

25

50

75

100

0 10 20 30Salinity

Ni (

%)

HMWLMWTD

MG 2009

0

25

50

75

100

0 10 20 30Salinity

Ni (

%)

Abbildung 11. Salinitäts-Abhängigkeit der kolloidalen Verteilung von Nickel.

Trotz Fluktuationen bei niedriger Salinität verhält sich gelöstes Nickel in allen vier

Studiengebieten nahezu konservativ (Abb. 9). Die Konzentrationen bei niedrigen

Salinitäten variieren zwischen 18 nmol/L in WWR und kleiner als 1 nmol/L in

Nordostbrasilien. Bei einer Salinität von S = 10 bis S = 15 vereinigen sich die

Konzentrationen des gelösten Nickels typischerweise auf ein Niveau von etwa 5


- 39 -

nmol/L, und sinken seewärts weiter ab. Die Modifikationen bei niedrigen Salinitäten

können aufgrund der Calcium-/Magnesium-induzierten Auslösung aus dem

partikulären Nickel (z.B. Mandal et al., 2002) oder der Koagulation mit u.a.

Eisenoxiden (z.B. Sholkovitz, 1978) zustande gekommen sein. Das partikuläre Nickel

verhält sich im Wesentlichen ebenfalls konservativ, wobei der Gehalt in Ost-Hainan

mit steigender Salinität abnimmt und in Nordost-Brasilien konstant bleibt (Abb. 10).

Die Unterschiede in den mittleren Niveaus des partikulären Nickels zwischen Hainan

und Nordost-Brasilien sind wohl auf die jeweiligen geologischen Randbedingungen

zurückzuführen. Angesichts der Beobachtung, dass der TD-Anteil mit zunehmender

Salinität steigt, während die HMW- und LMW-Anteile sinken (Abb. 11), und

gleichzeitig die Konzentration des gelösten Nickels sich nahezu konservativ verhält,

kann man vermuten, dass entweder eine interne Konversion von HMW- und LMW-

zur TD-Fraktion aufgetreten ist oder die Dissoziation von Koagulation begleitet wird.

3.3. Cobalt

Auch wenn bei Feldarbeiten (Sharp et al., 1982; Wang and Liu, 2003) und

Mischungsexperimenten (Sholkovitz, 1978) auch negative Abweichung festgestellt

werden konnten, weicht die Verteilungskurve von Cobalt im Ästuar überwiegend

positiv von der theoretischen Verdünnungslinie ab (z.B. Chiffoleau et al., 1994; Guieu

et al., 1998; Wen et al., 1999; Fang and Lin, 2002; Martino et al., 2002; Tovar-

Sánchez et al., 2004; Chaudry and Zwolsman, 2008). Das Verhalten von Cobalt im

Ästuar wird häufig mit Mangan in Verbindung gebracht (z.B. Takara et al., 2010).

Ästuarinen Untersuchungen von z.B. Chaudry and Zwolsman (2008) zufolge findet

unter oxischen Bedingungen eine gemeinsame Sedimentation von Cobalt und

Mangan statt. Cobalt kann in gelöster Form aus dem Sediment freigesetzt werden,

wenn partikuläres Mangan (IV) während der Mineralisierung von organischer

Substanz reduktiv gelöst wird (z.B. Moffett and Ho, 1996; Audry et al., 2006).

Allerdings ist ein gewisser Anteil von gelöstem Cobalt im Ästuar organisch gebunden,

und diese Bindung scheint sehr stabil zu sein (z.B. Zhang et al., 1990).


- 40 -

WR

0

3

6

9

0 10 20 30Salinity

Co

(nm

ol/L

)

200620072008

WWR

0

3

6

9

0 10 20 30Salinity

Co

(nm

ol/L

)

MD

0

3

6

9

0 10 20 30Salinity

Co

(nm

ol/L

)

200720082009

MG

0

3

6

9

0 10 20 30Salinity

Co

(nm

ol/L

)

Abbildung 12. Salinitäts-Abhängigkeit des gelösten Cobalts.

WR

0.00

0.25

0.50

0.75

1.00

0 10 20 30Salinity

Co

(μm

ol/g

)

200620072008

WWR

0.00

0.25

0.50

0.75

1.00

0 10 20 30Salinity

Co

(μm

ol/g

)

MD

0.00

0.25

0.50

0.75

1.00

0 10 20 30Salinity

Co

(μm

ol/g

)

200720082009

MG

0.00

0.25

0.50

0.75

1.00

0 10 20 30Salinity

Co

(μm

ol/g

)

Abbildung 13. Salinitäts-Abhängigkeit des partikulären Cobalts.


- 41 -

WR 2008

0

25

50

75

100

0 10 20 30Salinity

Co

(%) HMW

LMWTD

WWR 2008

0

25

50

75

100

0 10 20 30Salinity

Co

(%)

MD 2007

0

25

50

75

100

0 10 20 30Salinity

Co

(%)

HMWLMWTD

MG 2007

0

25

50

75

100

0 10 20 30Salinity

Co

(%)

MD 2009

0

25

50

75

100

0 10 20 30Salinity

Co

(%)

HMWLMWTD

MG 2009

0

25

50

75

100

0 10 20 30Salinity

Co

(%)

Abbildung 14. Salinitäts-Abhängigkeit der kolloidalen Verteilung von Cobalt.

Außer in WWR ist das Signal von gelöstem Cobalt im Flusswasser selten höher als 3

nmol/L. Nach anfänglichen Fluktuationen sinkt es kontinuierlich auf ein Niveau von

weniger als 1 nmol/L bei hohen Salinitäten (Abb. 12). Das gilt auch für das partikuläre

Cobalt, dessen Konzentration typischerweise von 0.5 μmol/g auf 0.2 μmol/g in Ost-

Hainan bzw. von 0.25 μmol/g auf 0.05 μmol/g in Nordost-Brasilien abnimmt (Abb. 13).

Besonderes in der Trockensaison von 2006 im WWR und in Nordost-Brasilien, ist

teilweise eine starke Erhöhung von gelöstem Cobalt bei niedriger Salinität (S < 5)

gegenüber dem Flusswasser zu beobachten. Da der Sauerstoffgehalt nicht

gemessen wurde, kann ein Zusammenhang mit anoxischen Bedingungen an der

Sediment-Wasser-Grenzfläche nicht ermittelt werden. Eventuell wurden die obersten

Sedimentschichten im Süßwasser-Zufuhrbereich aufgewirbelt, wodurch die

Freisetzung von Cobalt aus dem Sediment begünstigt würde. Steigende Anteile von

TD-Fraktion mit steigender Salinität (Abb. 14) liefern Hinweise dafür, dass das

freigesetzte Cobalt hauptsächlich hydratisierte Ionen und kleine lösliche Komplexe

mit Chlorid und Sulfat (z.B. Ćosović et al., 1982) und/oder mit kleineren organischen

Liganden (z.B. Zhang et al., 1990) bildet.


- 42 -

3.4. Kupfer

Zahlreiche Studien haben sich mit dem ästuarinen Verhalten des Kupfers beschäftigt.

Demnach kann sich Kupfer sowohl konservativ (z.B. Boyle et al., 1982; Danielsson et

al., 1983; Harper, 1991; Shiller and Boyle, 1991; Dai and Martin, 1995; Elbaz-

Poulichet et al., 1996; Powell et al, 1996; Guieu et al., 1998; Wen et al., 1999;

Martino et al., 2002; Fernández et al., 2008; Wang et al., 2009) als auch non-

konservativ verhalten (z.B. Sharp et al., 1982; Turner et al., 1992; Balls et al., 1994;

Chiffoleau et al., 1994; Comber et al., 1995; Martin et al., 1995; Kraepiel et al., 1997;

Wen et al., 1999; Fang and Lin, 2002; Wang and Liu, 2003; Waeles et al., 2004;

Chaudry and Zwolsman, 2008; Waeles et al., 2008a). Dabei kommt eine Zufuhr zum

gelösten Kupfer meist durch Freisetzung aus dem SPM und/oder dem Sediment

zustande (z.B. Santos-Echeandia et al., 2008), möglicherweise jedoch auch infolge

des Abbaus organischen Substanz (z.B. Masson et al., 2011). Verluste hingegen

sind immer mit Adsorption (z.B. Sholkovitz, 1978) oder biochemischen Aktivitäten

(z.B. Sharp et al., 1982) verbunden. Kolloidales Kupfer besteht überwiegend aus Cu-

Komplexen mit organischen Liganden wie z.B. Huminstoffen (z.B. Kogut and Voelker,

2003). In organisch-reichen Ästuaren ist gelöstes Kupfer sogar fast ausschließlich

gebunden an organische Entitäten (z.B. Shank et al., 2004). Dies liegt zum einen an

der starken Neigung von Kupfer zu organischen Funktionsgruppen (z.B. Newell and

Sanders, 1986), zum anderen kann es auch eine Konsequenz der Primärproduktion

sein (z.B. Van den Berg and De Luca Rebello, 1986).

WR

0

5

10

15

20

0 10 20 30Salinity

Cu

(nm

ol/L

)

200620072008

WWR

0

5

10

15

20

0 10 20 30Salinity

Cu

(nm

ol/L

)

MD

0

5

10

15

20

0 10 20 30Salinity

Cu

(nm

ol/L

)

200720082009

MG

0

5

10

15

20

0 10 20 30Salinity

Cu

(nm

ol/L

)

Abbildung 15. Salinitäts-Abhängigkeit des gelösten Kupfers.


- 43 -

WR

0.0

0.5

1.0

1.5

0 10 20 30Salinity

Cu

(μm

ol/g

)

200620072008

WWR

0.0

0.5

1.0

1.5

0 10 20 30Salinity

Cu

(μm

ol/g

)

MD

0.0

0.5

1.0

1.5

0 10 20 30Salinity

Cu

(μm

ol/g

)

200720082009

MG

0.0

0.5

1.0

1.5

0 10 20 30Salinity

Cu

(μm

ol/g

)

Abbildung 16. Salinitäts-Abhängigkeit des partikulären Kupfers.

WR 2008

0

25

50

75

100

0 10 20 30Salinity

Cu

(%) HMW

LMWTD

WWR 2008

0

25

50

75

100

0 10 20 30Salinity

Cu

(%)

MD 2007

0

25

50

75

100

0 10 20 30Salinity

Cu

(%)

HMWLMWTD

MG 2007

0

25

50

75

100

0 10 20 30Salinity

Cu

(%)

MD 2009

0

25

50

75

100

0 10 20 30Salinity

Cu

(%)

HMWLMWTD

MG 2009

0

25

50

75

100

0 10 20 30Salinity

Cu

(%)

Abbildung 17. Salinitäts-Abhängigkeit der kolloidalen Verteilung von Kupfer.

In den vier Studiengebieten nimmt die Konzentration von gelöstem Kupfer

typischerweise von 10-15 nmol/L im Flusswasser auf weniger als 5 nmol/L im

Meerwasser ab (Abb. 15). Ein konservatives Verhalten ist deutlich zu erkennen. Auch

das partikuläre Kupfer verhält sich konservativ mit einer seewärts zumeist ebenfalls

fallenden Konzentration, typischerweise von 0.6 μmol/g auf 0.3 μmol/g in Ost-Hainan


- 44 -

bzw. mit relativ konstanter Konzentration von ca. 0.3 μmol/g in Nordost-Brasilien

(Abb. 16). Dabei fällt der Kupfergehalt auf Hainan in 2008 deutlich höher aus als

sonst üblich. Aus den Kolloid-Verteilungen ist jedoch ersichtlich, dass eine ästuarine

Modifikation trotzdem stattfindet: während die HMW- und LMW-Fraktionen

schrumpfen, steigt der Anteil der TD-Fraktion (schnell bei niedrigen Salinitäten und

langsam bei mittleren bis hohen Salinitäten) auf ein Niveau von 65% im

Küstenwasser an - unabhängig von der Natur des Flusswassers (Abb. 17).

Möglicherweise werden große kolloidale organische Kupfermoleküle zu kleineren

organischen Kupfer-Komplexen abgebaut als Konsequenz der Veränderung der

Ionenstärke (Waeles et al., 2004).

3.5. Cadmium

Manchmal wird konservatives ästuarines Verhalten von Cadmium behauptet; jedoch

scheint dabei entweder die Meerwasser-Endglied-Konzentration nicht berücksichtigt

worden zu sein (z.B. Balls et al., 1994) oder es wurde nur ein Ausschnitt von

mittleren bis zu hohen Salinitäten dargestellt (z.B. Powell et al, 1996). Die meisten

ästuarinen Untersuchungen zeigen eine Anreicherung von gelöstem Cadmium im

Verlauf der ästuarinen Mischung (z.B. Boyle et al., 1982; Elbaz-Poulichet et al., 1987;

Shiller and Boyle, 1991; Chiffoleau et al., 1994; Comber et al., 1995; Dai and Martin,

1995; Martin et al., 1995; Elbaz-Poulichet et al., 1996; Kraepiel et al., 1997; Guieu et

al., 1998; Wen et al., 1999; Fang and Lin, 2002; Martino et al., 2002; Wang and Liu,

2003; Waeles et al., 2004; Chaudry and Zwolsman, 2008; Waeles et al., 2008a;

Wang et al., 2009). Dabei ist die Mobilisierung von SPM- und/oder Sediment-

gebundenem Cadmium von großer Bedeutung (z.B. Comans and van Dijk, 1988).

Der Transfer von Cadmium in die Lösungsphase wird vor allem durch

Komplexbildung mit Chloridionen begünstigt (z.B. Dabrin et al., 2009). Darüber

hinaus zeigt die Speziesanalyse, dass Cadmium sowohl mit Chloridionen als auch

mit organischen Liganden stabile Komplexe bildet (z.B. Waeles et al., 2005). In

einigen Fällen ist aber auch Sedimentation (d.h. Entfernung aus der Lösungsphase)

bei niedriger Salinität nachgewiesen worden, wie z.B. auf der subtropischen Insel

Taiwan (Jiann et al., 2005), im kalten Schweden (Danielsson et al., 1983) und in den

moderat klimatischen Ästuaren des Mersey (Comber et al., 1995) und des Delaware

(Sharp et al., 1982). Obwohl sich ein solches Verhalten in den meisten Ästuaren

nicht deutlich bemerkbar macht, scheint es doch ein globales Phänomen zu sein. In

der Tat deuten Mischungsexperimente daraufhin, dass Cadmium durchaus

sedimentieren kann (Sholkovitz, 1978), und dass dies vor allem durch Adsorbieren


- 45 -

an die Oberfläche von Mangan- und/oder Eisenoxiden geschieht (z.B. Yang and

Sañudo-Wilhelmy, 1998; Turner et al., 2008).

WR

0.00

0.05

0.10

0.15

0 10 20 30Salinity

Cd

(nm

ol/L

)

200620072008

WWR

0.00

0.05

0.10

0.15

0 10 20 30Salinity

Cd

(nm

ol/L

)

MD

0.00

0.05

0.10

0.15

0 10 20 30Salinity

Cd

(nm

ol/L

)

200720082009

MG

0.00

0.05

0.10

0.15

0 10 20 30Salinity

Cd

(nm

ol/L

)

Abbildung 18. Salinitäts-Abhängigkeit des gelösten Cadmiums.

WR

0

4

8

12

0 10 20 30Salinity

Cd

(nm

ol/g

)

200620072008

WWR

0

4

8

12

0 10 20 30Salinity

Cd

(nm

ol/g

)

MD

0

4

8

12

0 10 20 30Salinity

Cd

(nm

ol/g

)

200720082009

MG

0

4

8

12

0 10 20 30Salinity

Cd

(nm

ol/g

)

Abbildung 19. Salinitäts-Abhängigkeit des partikulären Cadmiums.


- 46 -

WR 2008

0

25

50

75

100

0 10 20 30Salinity

Cd

(%) HMW

LMWTD

WWR 2008

0

25

50

75

100

0 10 20 30Salinity

Cd

(%)

MD 2007

0

25

50

75

100

0 10 20 30Salinity

Cd

(%)

HMWLMWTD

MG 2007

0

25

50

75

100

0 10 20 30Salinity

Cd

(%)

MD 2009

0

25

50

75

100

0 10 20 30Salinity

Cd

(%)

HMWLMWTD

MG 2009

0

25

50

75

100

0 10 20 30Salinity

Cd

(%)

Abbildung 20. Salinitäts-Abhängigkeit der kolloidalen Verteilung von Cadmium.

Die ästuarine Verteilung des gelösten Cadmiums in allen vier Studiengebieten ist in

Abb. 18 aufgetragen. In den meisten Fällen steigt die Konzentration mit steigender

Salinität von 0.01-0.05 nmol/L auf typischerweise 0.1 nmol/L im Küstenbereich. Eine

Mobilisierung aus dem partikulären Cadmium scheint aufzutreten, wobei dieser Effekt

in Ost-Hainan ausgeprägter ist als in Nordost-Brasilien. In der MG in 2007 z.B. bleibt

die Konzentration bei etwa 0.1 nmol/L stabil. In der MD vermindert sich die

Konzentration von gelöstem Cadmium von 0.1 nmol/L im Flusswasser sogar um zwei

Drittel auf 0.03 nmol/L bei niedrigen und mittleren Salinitäten. Auch in 2008 und 2009

ist auf ersten Blick nur eine mit der Salinität steigende konservative

Konzentrationskurve zu sehen. Wenn die weitaus kleinere Konzentration an der

Küste (in 2009; 0.03 nmol/L) im Betrag gezogen wird, ist eine Mobilisierung aus dem

partikulären Cadmium jedoch durchaus möglich, wie es im Huanghe (gelben Fluss)

Ästuar auch beobachtet worden ist (Elbaz-Poulichet et al., 1987). Das partikuläre

Cadmium verhält sich in allen vier Studiengebieten im Wesentlichen konservativ,

wobei zum Teil eine starke Abnahme bei niedriger Salinität vor allem in Ost-Hainan

zu beobachten ist (Abb. 19). Die Adsorption u.a. an Eisenoxide gefolgt von

gemeinsamer Ausfällung wird hier vermutlich eine wichtige Rolle gespielt haben. Die

Analysen der Kolloidfraktionen ergeben, dass bei niedriger Salinität die Anteile der


- 47 -

Kolloide (HMW und LMW) deutlich abnehmen (Abb. 20), was eventuell auf

zunehmende Chloro-Komplexierung des Cadmium zurückzuführen ist.

3.6. Blei

Im Allgemeinen wird Blei im Verlauf der ästuarinen Vermischung aus der wässrigen

Phase entfernt (z.B. Danielsson et al., 1983; Turner et al., 1992; Balls et al., 1994;

Chiffoleau et al., 1994; Dai and Martin, 1995; Wen et al., 1999; Wang and Liu, 2003;

Monbet, 2006; Waeles et al., 2008a). Allerdings ist konservatives Verhalten auch zu

beobachten (z.B. Harper, 1991; Martin et al., 1995; Elbaz-Poulichet et al., 1996;

Kraepiel et al., 1997; Wang et al., 2009). Zudem wird über Konzentrationsmaxima

von gelöstem Blei in der Turbiditäts-Maximum-Zone infolge einer Freisetzung von

Blei aus resuspendierten Sedimentpartikeln berichtet (Turner et al., 1992; Martino et

al., 2002); auch Stabilisierung durch Komplexbildung mit reduzierten Schwefel-

Spezies soll eine Rolle spielen (Tang et al., 2002). Dennoch kontrolliert in den

meisten Fällen die Adsorption an das SPM oder die Ausflockung mit Eisenkolloiden

die Ästuarchemie von Blei (z.B. Elbaz-Poulichet et al., 1984; Waeles et al., 2008b).

WR

0

1

2

3

0 10 20 30Salinity

Pb (n

mol

/L)

200620072008

WWR

0

1

2

3

0 10 20 30Salinity

Pb (n

mol

/L)

MD

0

1

2

3

0 10 20 30Salinity

Pb (n

mol

/L)

200720082009

MG

0

1

2

3

0 10 20 30Salinity

Pb (n

mol

/L)

Abbildung 21. Salinitäts-Abhängigkeit des gelösten Bleis.


- 48 -

WR

0.0

0.1

0.2

0.3

0.4

0 10 20 30Salinity

Pb (μ

mol

/g)

200620072008

WWR

0.0

0.1

0.2

0.3

0.4

0 10 20 30Salinity

Pb (μ

mol

/g)

MD

0.0

0.1

0.2

0.3

0.4

0 10 20 30Salinity

Pb (μ

mol

/g)

200720082009

MG

0.0

0.1

0.2

0.3

0.4

0 10 20 30Salinity

Pb (μ

mol

/g)

Abbildung 22. Salinitäts-Abhängigkeit des partikulären Bleis.

WR 2008

0

25

50

75

100

0 10 20 30Salinity

Pb (%

) HMWLMWTD

WWR 2008

0

25

50

75

100

0 10 20 30Salinity

Pb (%

)

MD 2007

0

25

50

75

100

0 10 20 30Salinity

Pb (%

)

HMWLMWTD

MG 2007

0

25

50

75

100

0 10 20 30Salinity

Pb (%

)

MD 2009

0

25

50

75

100

0 10 20 30Salinity

Pb (%

)

HMWLMWTD

MG 2009

0

25

50

75

100

0 10 20 30Salinity

Pb (%

)

Abbildung 23. Salinitäts-Abhängigkeit der kolloidalen Verteilung von Blei.

Trotz einiger Ausreißer sinkt generell der Gehalt von gelöstem Blei in den beiden

vergleichsweise bleireichen Ästuaren Hainans (bis zu 2.2 nmol/L), und zwar um mehr

als 50% bereits bei niedriger Salinität (S < 10); eine solche Abnahme ist im

bleiarmen ästuarinen Wasser in Nordost-Brasilien (< 1 nmol/L) nicht klar erkennbar


- 49 -

(Abb. 21). Ab S > 10 sinkt die gelöste Bleikonzentration nur noch wenig bis auf ein

Niveau von ca. 0.5 nmol/L im Küstenwasser. Partikuläres Blei folgt im Allgemeinen

der konservativen Mischungslinie (Abb. 22). Die Erhöhungen bei höherer Salinität in

WWR könnten durch die Aufwirbelung von Blei-haltigen Sedimenten hervorgerufen

worden sein. Dass die kolloidale Verteilung von gelöstem Blei der von Eisen ähnelt

(Abb. 23), legt eine Affinität von Blei zu Eisenkolloiden nahe. Offensichtlich ist Blei im

Salinitätsbereich S < 20 primär mit der HMW-Fraktion assoziiert. Jedoch steigt auch

beim Blei die relative Bedeutung des "echt gelösten" Pb (TD-Fraktion) mit

zunehmendem Salzgehalt.

3.7. Verteilungskoeffizienten

Der Verteilungskoeffizient (distribution coefficient; partition coefficient), der die

relative Affinität eines Stoffes zur Festphase widerspiegelt (z.B. Chiffoleau et al.,

1994), ist ein Schlüsselparameter zur Interpretation des chemischen Transports in

Ästuaren (Turner et al., 1992). Der Verteilungskoeffizient ist gegeben durch:

CC

Ks

pD =

wobei Cp (in: mol/kg) die partikuläre und CS (in: mol/L) die gelöste Konzentration

darstellen (z.B. Benoit et al., 1994). Cadmium und Kupfer z.B. neigen zum

bevorzugten Auftreten in gelöster Form und haben daher einen kleineren

Verteilungskoeffizient, während Eisen und Blei wegen ihrer geringen Löslichkeit im

Wasser größere Verteilungskoeffizienten aufweisen.

Häufig wird der Verteilungskoeffizient in Abhängigkeit von der Salinität diskutiert (z.B.

Balls et al., 1994). Demnach ist eine Zunahme des Verteilungskoeffizienten mit

steigender Salinität ein Anzeichen für die Dominanz von Metalladsorption und eine

Abnahme für das Überwiegen der Desorption. Eine steigende Kurve in Diagrammen

KD vs. Salinität deutet auf eine zunehmende Entfernung aus der wässrigen Phase im

Ästuar hin und eine fallende Kurve offenbart Eintrag in die wässrige Phase.


- 50 -

Fe in East-Hainan

3

4

5

6

7

8

0 10 20 30Salinity

LogK

D (L

/kg)

WRWWR

Fe in Northeast-Brazil

3

4

5

6

7

8

0 10 20 30Salinity

LogK

D (L

/kg)

MDMG

Ni in East-Hainan

3

4

5

6

7

8

0 10 20 30Salinity

LogK

D (L

/kg)

Ni in Northeast-Brazil

3

4

5

6

7

8

0 10 20 30Salinity

LogK

D (L

/kg)

Co in East-Hainan

3

4

5

6

7

8

0 10 20 30Salinity

LogK

D (L

/kg)

Co in Northeast-Brazil

3

4

5

6

7

8

0 10 20 30Salinity

LogK

D (L

/kg)

Cu in East-Hainan

3

4

5

6

7

8

0 10 20 30Salinity

LogK

D (L

/kg)

Cu in Northeast-Brazil

3

4

5

6

7

8

0 10 20 30Salinity

LogK

D (L

/kg)

Cd in East-Hainan

3

4

5

6

7

8

0 10 20 30Salinity

LogK

D (L

/kg)

Cd in Northeast-Brazil

3

4

5

6

7

8

0 10 20 30Salinity

LogK

D (L

/kg)

Pb in East-Hainan

3

4

5

6

7

8

0 10 20 30Salinity

LogK

D (L

/kg)

Pb in Northeast-Brazil

3

4

5

6

7

8

0 10 20 30Salinity

LogK

D (L

/kg)

Abbildung 24. Salinitäts-Abhängigkeit des log KD.


- 51 -

In Abb. 24 ist ersichtlich, dass Cadmium das einzige Element ist, dessen

Verteilungskoeffizient negativ mit der Salinität korreliert ist. Die Bildung von stabilen

Cadmium-Chloro-Komplexen und die erfolgreiche Konkurrenz von Meerwasser-

kationen mit Cadmiumionen um Sorptionsstellen am SPM ist vermutlich dafür

verantwortlich (z.B. Turner, 1996; Dabrin et al., 2009). Die KD-Werte von Nickel,

Cobalt und Kupfer können annähernd als konstant betrachtet werden, während die

von Eisen und Blei mit steigender Salinität teilweise deutlich (logarithmische Skala)

zunehmen, was als eine Folge von Ausfällungen der beiden Metalle aus der gelösten

Phase gewertet werden kann (z.B. Dai and Martin, 1995).

Vergleiche der Höhe der Verteilungskoeffizienten unterschiedlicher Metalle ergeben

zumeist die folgende Reihung: Fe > Pb > Co > Ni > Cd ≥ Cu (z.B. Balls, 1989; Turner

et al., 1992; Chiffoleau et al., 1994; Munksgaard and Parry, 2001; Tang et al., 2002).

Mit durchschnittlichen log KD-Werten von

Fe : Pb : Co : Ni : Cd : Cu = 6.0 : 5.7 : 5.3 : 5.1 : 4.8 : 4.8 auf Hainan bzw.

Fe : Pb : Co : Ni : Cd : Cu = 6.1 : 5.8 : 5.0 : 4.8 : 4.6 : 4.5 in Nordost-Brasilien

stimmen die Ergebnisse dieser Arbeit im Wesentlichen mit den Literatur-Reihungen

überein (z.B. Balls, 1989; Turner et al., 1992; Balls et al., 1994; Munksgaard and

Parry, 2001; Fang and Lin, 2002; Tang et al., 2002; Guo et al., 2000b; Jiann et al.,

2005). Allerdings muss man sich dabei im Klaren sein, dass je nach den

angewendeten analytischen Methoden die Verteilungskoeffizienten sehr stark

variieren können (z.B. Turner et al., 1992): beispielsweise werden in Hinblick auf die

SPM-Bestimmung sowohl - wie in der vorliegenden Arbeit - die totalen

Metallkonzentrationen (z.B. Zhang, 1999) als auch die - deutlich niedrigeren -

säureverfügbaren Metallkonzentrationen (acid leachable; z.B. Fang and Lin, 2002)

verwendet, um den Verteilungskoeffizient zu berechnen.


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3.8. Vergleich der Studiengebiete und ihrer Kontaminationsniveaus

Die durchschnittlichen Konzentrationen in den vier Studiengebieten von gelösten und

partikulären Metallen sowie des DOC und des SPM werden in Tab. 4 und Tab. 5

dargestellt.

Tabelle 4. Durchschnittliche Konzentrationen der gelösten Spurenmetalle und des DOC in den vier Studiengebieten.

Studiengebiet Fe Ni Co Cu Cd Pb DOC

μmol/L nmol/L nmol/L nmol/L nmol/L nmol/L mg/L

WR

Trockensaison 2006 0.1 4.8 3.2 17.6 0.06 0.27 1.4

Regensaison 2007 2.6 5.6 1.2 7.5 0.06 0.92 1.7

Regensaison 2008 2.6 4.2 1.1 6.3 0.06 0.71 2.1

gesamt 2.1 4.9 1.6 9.3 0.06 0.70 1.8

WWR

Trockensaison 2006 0.3 8.4 3.3 8.4 0.11 0.09 1.9

Regensaison 2007 1.5 7.7 2.4 7.2 0.07 0.27 2.1

Regensaison 2008 5.2 8.8 3.0 11.5 0.06 0.99 5.6

gesamt 2.8 8.4 2.9 9.4 0.07 0.53 3.8

MD

Regensaison 2007 1.3 5.6 1.0 7.5 0.04 0.16 3.0

Trockensaison 2008 1.3 2.0 2.0 5.6 0.02 0.23 4.3

Trockensaison 2009 1.2 7.7 2.0 7.7 0.03 0.34 5.4

gesamt 1.3 4.9 1.8 6.8 0.03 0.25 4.4

MG

Regensaison 2007 1.7 6.8 0.9 10.7 0.01 0.45 3.8

Trockensaison 2008 1.3 3.7 2.0 6.3 0.04 0.21 4.5

Trockensaison 2009 0.9 7.7 1.9 6.7 0.04 0.45 6.6

gesamt 1.3 5.8 1.6 7.9 0.03 0.35 4.8


- 53 -

Tabelle 5. Durchschnittliche Konzentrationen der partikulären Spurenmetalle und des SPM in den vier Studiengebieten.

Studiengebiet Fe Ni Co Cu Cd Pb SPM

mmol/g μmol/g μmol/g μmol/g nmol/g μmol/g mg/L

WR

Trockensaison 2006 1.02 0.52 0.29 0.30 4.2 0.23 10.9

Regensaison 2007 0.91 0.58 0.29 0.46 2.8 0.21 2.0

Regensaison 2008 0.92 0.71 0.30 0.71 4.4 0.28 3.0

gesamt 0.94 0.61 0.29 0.52 3.7 0.24 4.3

WWR

Trockensaison 2006 1.12 0.91 0.29 0.46 3.2 0.13 9.1

Regensaison 2007 0.86 0.94 0.36 0.35 3.5 0.17 2.7

Regensaison 2008 0.73 0.94 0.33 0.86 3.5 0.17 1.4

gesamt 0.87 0.93 0.33 0.60 3.4 0.16 3.8

MD

Regensaison 2007 0.89 0.27 0.16 0.22 2.3 0.17 20.8

Trockensaison 2008 0.72 0.25 0.16 0.21 0.7 0.12 6.5

Trockensaison 2009 0.67 0.28 0.16 0.21 0.9 0.13 9.1

gesamt 0.75 0.27 0.16 0.21 1.2 0.13 11.1

MG

Regensaison 2007 0.79 0.29 0.15 0.21 0.6 0.13 24.1

Trockensaison 2008 0.57 0.21 0.10 0.18 0.6 0.11 9.0

Trockensaison 2009 0.87 0.22 0.14 0.17 0.8 0.14 20.4

gesamt 0.73 0.24 0.13 0.19 0.7 0.13 17.1

Der DOC-Gehalt in Nordost-Brasilien mit einer durchschnittlichen Konzentration von

4.6 mg/L ist zwar mehr als 50% höher als in Ost-Hainan (2.9 mg/L), unterscheidet

sich jedoch nicht wesentlich von anderen Ästuaren, wie z.B. dem europäischen

Mersey Ästuar (1.2-9.6 mg/L; Martino et al., 2004), dem asiatischen Danshuei Ästuar

(2.6 mg/L; Jiann et al., 2005), dem nordamerikanischen Ochlockonee Ästuar (2.3-

18.8 mg/L; Powell et al, 1996) oder dem südamerikanischen Amazonas Ästuar (0.8-

3.7 mg/L; Sholkovitz et al., 1978).


- 54 -

Die SPM-Beladung in Nordost-Brasilien mit einer durchschnittlichen Konzentration

von 14.0 mg/L ist ebenfalls deutlich größer als in Ost-Hainan (4.0 mg/L). Dennoch

gehören die vier Studiengebiete zu den am wenigsten SPM-beladenen Ästuaren der

Welt (vgl.: Sholkovitz et al., 1978; Munksgaard and Parry, 2001; Martino et al., 2002;

Wang and Liu, 2003; Tovar-Sánchez et al., 2004).

In Ost-Hainan ist der Metallgehalt sowohl in der gelösten als auch in der partikulären

Phase generell höher als in Nordost-Brasilien. Dies könnte zum Teil auf die stärkeren

anthropogenen Aktivitäten auf Hainan zurückgeführt werden, liegt aber vermutlich

primär an der unterschiedlichen geologischen Natur der beiden Regionen. Der

höhere SPM-Gehalt kann in geringem Umfang zu stärkerer Adsorption der Metalle

beitragen (z.B. Turner, 1996). Das WWR weist zumeist die höchsten

Metallkonzentrationen von den vier Studiengebieten auf. Mit Ausnahme von Kupfer

und Cadmium, die zur gelösten Phase neigen, nehmen die Metallkonzentrationen

jedoch durch Prozesse im Ästuar mit steigender Salinität schnell ab. Ansonsten

liegen die Metallkonzentrationen bei hohen Salinitäten in Ost-Hainan und Nordost-

Brasilien etwa auf dem gleichen Niveau.

Erwähnenswert ist noch, dass in der Trockensaison von 2009 ein

Konzentrationsmaximum (S ≈ 10) in der MD bei allen gelösten Metallen mehr oder

weniger deutlich auftritt, was vielleicht auf eine Verschmutzungsquelle aus der

naheliegenden Stadt Maceío hindeutet. Wie zuvor ausgeführt, werden die erhöhten

Konzentrationen durch ästuarine Einwirkungen bald gemildert; die filternde Wirkung

des Ästuars scheint einwandfrei zu funktionieren.

In Tab. 6 und Tab. 7 sind typische Konzentrationen für gelöste bzw. partikuläre

Spurenmetalle in einigen Ästuaren weltweit aufgelistet. Ein Vergleich macht deutlich,

dass die gelösten und die partikulären Spurenmetallenkonzentration auf Hainan und

in Nordost-Brasilien sich nicht viel von denen anderer, nicht-industriell geprägter

Ästuare in der Welt unterscheiden. Vielmehr gehören die gelösten und die

partikulären Konzentrationen aller Spurenmetalle außer Eisen auf Hainan und in

Nordost-Brasilien zu den niedrigsten, die in Tab. 6 und Tab. 7 gelistet sind. Auch im

Vergleich zu den weniger industriell belasteten Ästuaren, wie z.B. den Ästuaren auf

Sumatra (Bach, 2012), fallen die beiden Studiengebiete nicht durch höhere Werte auf.

Was die verschmutzungsrelevanten Spurenmetalle anbelangt, scheint die Aquakultur

auf Hainan und die Zuckerrohrmonokultur in Nordost-Brasilien keine bzw. nur

begrenzte Auswirkungen auf die Umweltkonzentrationen zu haben.


- 55 -

Tabelle 6. Konzentrationen der gelösten Spurenmetalle in Ästuaren.

Studiengebiet Fe Ni Co Cu Cd Pb

Literatur μmol/L nmol/L nmol/L nmol/L nmol/L nmol/L

WR 2.1 4.9 1.6 9.3 0.06 0.70

diese Arbeit WWR 2.8 8.4 2.9 9.4 0.07 0.53

MD 1.3 4.9 1.8 6.8 0.03 0.25

MG 1.3 5.8 1.6 7.9 0.03 0.35

Bristol Kanal und outer Severn Ästuar -- -- -- 26.8-

74.0 0.01-1.25

0.01-48.3 Harper, 1991

Changjiang Ästuar 0.1-0.6 18-28 1.4-2.9 19-26 0.02-0.36 2.3-3.4 Wang and

Liu, 2003

Danshuei Ästuar Oktober 2001

0.01-1.8 138 -- 60 0.2 0.9 Jiann et al.,

2005

Forth Ästuar 1.4 13.6 -- 25 1.1 1.0 Balls et al., 1994

Galveston Bucht 0.01-1.1

4.7-32.3 0.4-9.6 3.5-

18.6 0.03-0.17 0.1-0.8 Wen et al.,

1999

Huanghe Ästuar -- -- -- 1.6-70 0.9-29 1.1-6.4 Tang et al., 2010

Mersey Ästuar -- 10-150 0.1-40 20-90 0.1-4.5 0.1-6 Martino et al., 2002

Mississippi Ästuar 0.03 1-30 -- 1-28 0.01-0.3 -- Shiller and

Boyle, 1991

Nordaustralische Küste und ästuarines

Meerwasser

0.001-0.6 2.0-9.4 0.1-1.6 2.3-16 0.01-

0.31 0.01-0.3

Munksgaard and Parry,

2001

Ob Ästuar 0.4-0.7 20-24 -- 29-38 0.005-0.008

0.06-0.08

Dai and Martin, 1995

Ochlockonee Ästuar 0.04-7.0 2.6-5.5 -- 2.0-5.4 0.04-

0.05 -- Powell et al, 1996

Penzé Ästuar -- -- -- 1.0-4.5 0.1-0.4 0.04-0.30

Waeles et al., 2008a

Rhône Ästuar -- 4-18 -- 3-23 0.1-0.3 -- Dai et al., 1995

Venedig Lagune 0.001-1.1

6.7-20.9 -- 5.6-

15.8 0.01-0.08 0.1-1.1 Martin et al.,

1995

Weser Ästuar 0.03 64 -- 47 0.22-0.62 0.3 Turner et al.,

1992

Yenisey Ästuar 0.3 8.8-9.4 -- 22-29 0.01-0.02 0.03 Dai and

Martin, 1995

Gironde Ästuar 0.05 7 -- 15 0.5 0.2 Kraepiel et al., 1997


- 56 -

Tabelle 7. Konzentrationen der partikulären Spurenmetalle und des SPMs in Ästuaren.

Studiengebiet Fe Ni Co Cu Cd Pb SPM

Literatur mmol/g μmol/g μmol/g μmol/g nmol/g μmol/g mg/L

WR 0.94 0.61 0.29 0.52 3.7 0.24 4.3

diese Arbeit

WWR 0.87 0.93 0.33 0.60 3.4 0.16 3.8

MD 0.75 0.27 0.16 0.21 1.2 0.13 11.1

MG 0.73 0.24 0.13 0.19 0.7 0.13 17.1

Changjiang Ästuar 0.18-1.01

0.22-9.58

0.12-0.88

0.32-2.48

0.18-9.79

0.08-0.58 70 Wang and

Liu, 2003

Conwy Ästuar 0.57 2.98 -- 2.58 370 2.64 < 140 Zhou et al., 2003

Danshuei Ästuar 0.19-1.2 0.1-1.8 0.03-

0.32 0.35-7.9 -- -- -- Fang and

Lin, 2002

Elbe Ästuar 0.09-0.15

0.36-0.68

0.25-0.32

0.16-0.49 -- 0.26-

0.31 13-289 Turner et al., 1991

Galveston Bucht 0.09 -- -- 0.31 -- 0.97 10 Benoit et al., 1994

Gironde Ästuar 0.89 0.88 -- 0.58 4.8 0.27 -- Kraepiel et al., 1997

Humber Ästuar 0.09-0.52

0.32-0.53

0.19-0.49 0.2-3.8 -- 0.14-

0.33 35-693 Turner et al., 1991

Mersey Ästuar 0.30 0.38 0.02 0.41 82 0.45 < 700 Martino et al., 2002

Patos Bucht < 0.2 -- -- < 0.22 -- < 0.07 < 261

Niencheski and

Baumgarten, 2000

Rhône Ästuar 9.8 -- -- 0.72 -- 0.21 -- Elbaz-

Poulichet et al., 1996

Scheldt Ästuar 0.01-0.07

0.20-0.58

0.14-0.29

0.05-0.69 -- 0.11-

0.29 8-239 Turner et al., 1991

Seine Ästuar -- 0.22-1.02

0.07-0.26 0.7-4.0 14-59 0.24-

0.82 < 300 Chiffoleau et al., 1994

Sepetiba Bucht -- -- -- 0.74 2.4 0.07 160 Lacerda et al., 1987

Thames Ästuar 0.07-0.15

0.26-0.75

0.10-0.25

0.25-0.71 -- 0.22-

0.55 19-424 Turner et al., 1991

Venedig Lagune 0.14-0.30

0.26-0.64 -- 0.29-

3.8 3.2-11.0

0.038-0.075

3.5-23.9

Martin et al., 1995

Weser Ästuar 0.73 1.18 -- 0.65 5.2-64.1 0.55 < 400 Turner et

al., 1992


- 57 -

Bemerkenswert ist in Tabelle 6, dass die gelöste Konzentration von Cadmium im

Weser Ästuar deutlich höher ausgefallen ist als in den meisten anderen Ästuaren

weltweit. Die Daten von Turner (1992) in Tab. 6 stammen aus einer Expedition im

August 1989. Infolge des seit dem 10. Jahrhundert nach unserer Zeitrechnung

bestehenden Bergbaus im Harz sind die Harzer Oberböden zum Teil stark mit

Cadmium angereichert (Umweltbericht Niedersachsen, 2013). Durch Erosion gelangt

das Cadmium in die Aller, die einen Nebenfluss der Weser darstellt. Über die

Schwermetallfracht der Aller wird schließlich die Wasserqualität der Weser

entscheidend u.a. mit Cadmium belastet (FGG Weser, 2012). Jedoch deuten die

veröffentlichten Messdaten an der FGG-Weser-Messstelle Bremerhaven zwischen

den Jahren 1979 und 2000 auf einen Rückgang des gelösten Cadmiums in der

Weser hin (Abb. 25); allerdings sind keine Daten nach dem Jahre 2000 vorhanden.

Die durchschnittliche Konzentration des gelösten Cadmiums an dieser Messstelle ist

von ca. 19 nmol/L in 1980 auf ca. 7 nmol/L in 1998, d.h. um knapp zwei Drittel,

gesunken.

Abbildung 25. Gelöstes Cadmium in der Weser an der FGG-Weser-Messstelle Bremerhaven (Datenquelle: Weser-Datenbank, in: FGG Weser, 2012).

3.9. Offene Fragen und Perspektive

In dieser Arbeit wurden die gelösten, partikulären und kolloidalen Konzentrationen

von sechs Spurenmetallen im Wanquan Ästuar und Wenchang-Wenjiao Ästuar in

Ost-Hainan sowie in der Mundaú Lagune und der Manguaba Lagune in Nordost-

Brasilien bestimmt. Die ästuarinen Verteilungs- und Verhaltensmuster der gelösten,

partikulären und kolloidalen Spurenmetalle wurden diskutiert. Daneben wurden die

Konzentrationen und das Verhalten der sechs untersuchten Spurenmetalle in den

genannten Studiengebieten mit anderen Ästuaren weltweit verglichen und die


- 58 -

Kontaminationsniveaus in Bezug auf die untersuchten Spurenmetalle beurteilt. Einige

Fragen bedürfen jedoch weiterer Bearbeitung:

1. Welche Rolle spielt das Auftreten von oxischen/anoxischen Bedingungen in den

Oberflächensedimenten für das Spurenmetallverhalten in den genannten

Studiengebieten?

2. Wie groß ist der Anteil an den partikulären Spurenmetallen, der sich aktiv an den

ästuarinen Prozessen beteiligt, d.h. welches Verhältnis besteht zwischen den

Spurenmetallen an den Partikeloberflächen zu denen in Gitterpositionen?

3. In welchem Umfang sind die "echt gelösten" bzw. die kolloidalen Spurenmetalle

mit organischen Liganden bzw. organischen Partikeln assoziiert?

Um diese Fragen zu beantworten wäre es sinnvoll:

1. bei einer evtl. erneuten Beprobung der Studiengebiete den gelösten

Sauerstoffgehalt nahe der Sedimentoberfläche zu vermessen. Anhand der

Beziehung zwischen dem gelösten Sauerstoffgehalt und den Spurenmetall-

Konzentrationen könnte das Verhalten mancher Sauerstoff-empfindlicher

Spurenmetalle wie Cobalt und Cadmium besser verstanden werden.

2. zusätzlich andere Methoden zur Untersuchung der partikulären Spurenmetalle

wie z.B. die Festphasen-Extraktion (z.B. Danielsson et al., 1983; Fang and Lin,

2002) zu verwenden. Damit könnten Hinweise auf die Menge der tatsächlich

beteiligten partikulären Spurenmetalle gewonnen werden.

3. zusätzlich die physikalische Größe der Kolloide durch alternative Methoden, wie

z.B. mittels Field-Flow-Fraktionierung (z.B. Stolpe et al., 2010) zu bestimmen, um

operationelle Artefakte zu mindern.

4. die chemische Zusammensetzung unterschiedlicher Kolloidfraktionen durch

alternative Methoden, wie z.B. Ionenaustausch-Chromatographie (z.B. Wen et al.,

2011) oder "hyphenated methods" wie HPLC-ICP-MS zu untersuchen.

Kapitel 4: Publikation 1

- 59 -

4. Publikation 1:

erschienen als: Continental Shelf Research 57 (2013): 59-72.

Estuarine modification of dissolved and particulate trace metals in major rivers of East-Hainan

Jun Fu a, Xiao-Liang Tang a, Jing Zhang b, Wolfgang Balzer a, * a University of Bremen, FB2, Marine Chemistry, D- 28334 Bremen, Germany

b State Key Laboratory of Estuarine and Coastal Research, East China Normal University,

3663 Zhongshan Road North, Shanghai 200062, People’s Republic of China

* Corresponding author: e-mail address: [email protected]

Abstract

Dissolved and particulate cadmium, copper, iron, lead, cobalt and nickel were

analyzed in surface waters of the Wanquan River estuary and the

Wenchang/Wenjiao River estuary in East-Hainan Island during the dry season

(December 2006) and two wet seasons (August 2007 and July/August 2008). A major

difference to other Chinese rivers was the very low concentration of suspended

particles in these tropical Hainan estuaries. In the dissolved phase, a positive

deviation from the theoretical dilution line was observed for Cd during different

expeditions. Dissolved Cu and Ni essentially behaved conservatively, while Fe, Pb

and partly also Co correlated in their negative deviation from simple mixing. Strong

seasonal variability was observed only for dissolved Fe, Pb and Cd: sorption by the

much higher loading with suspended particles during the dry season lead to a strong

lowering of dissolved Fe and Pb, while the opposite was observed for dissolved Cd.

In both estuaries all six metals in particulate form showed almost constant values

with a tendency for slight decreases along the salinity profile. The normalization to

particulate Al revealed some specific particle properties during the different

expeditions. The dynamics of Fe chemistry dominated the distribution of Pb in all

forms. The distribution coefficients KD showed a general decrease in the order Fe >

Pb > Co > Ni > Cu ≈ Cd. There was no ‘‘particle concentration effect’’; rather the KD’s

of Fe and Pb exhibited slightly positive correlations with the suspended particle

loadings. Elevated concentrations levels in the Wenchang/Wenjiao River estuary,

especially during the wet season 2008, were ascribed to diffuse inputs from


- 60 -

aquaculture ponds which girdle the upper estuary. In comparison to major Chinese

rivers, the tropical Hainan estuaries (S > 0) showed similar levels for Cd, Cu, Pb, Co

and Ni in particles and solution, while Fe was enriched in both matrices. On a global

scale, neither in the Wanquan River estuary nor in the Wenchang/Wenjiao River

estuary significant trace metal contamination was observed.

Keywords: Trace metals; Particulate; Dissolved; Estuary; Distribution coefficient;

Hainan


- 61 -

4.1. Introduction

River-transported signals are subjected to a variety of physical, chemical and

biological processes in the estuarine mixing zone, in which the boundary conditions

are extremely variable in both space and time; estuaries can be thought of as acting

as filters of the river-transported chemical signals, which can often emerge from the

mixing zone in a form that is considerably modified with respect to that which entered

the system (Chester, 1990). This concept of the estuarine filter is based on the fact

that the mixing of the two very different end-member waters will result in the setting

up of strong physico-chemical gradients in an environment. It is these gradients

which are the driving force behind the filter. In addition, also organic matter

production (Louis et al., 2009), the oxygenation state (Khalid et al., 1978; Zwolsman

and van Eck, 1999), possible inputs from anoxic sediments (Salomons et al., 1987)

and/or other processes may have influence.

Trace metals are a particularly interesting aspect of estuarine chemistry because

their differing physical chemistries lead to a variety of geochemical behaviours

(Shiller and Boyle, 1991). For example, previous studies have indicated an affinity of

cobalt to manganese oxide phases and a removal of lead through adsorption onto

the suspended particulate matter (Chiffoleau et al., 1994), the desorption of cadmium

from suspended particles and the flocculation of iron colloids (Roux et al., 1998), the

removal of copper from the dissolved phase at low salinities (Comber et al., 1995)

and non-conservative behaviour of nickel (Wang and Liu, 2003). The trace metal

behaviour in the large rivers and major estuaries of China at mid latitudes such as

the Changjiang (Yangtze River) and the Huanghe (Yellow River) has been well

studied (e.g., Zhang, 1995; Zhang and Liu, 2002; Wang et al., 2009; Tang et al.,

2010), but about the trace metal chemistry in estuaries of tropical China, especially of

Hainan Island, very little is known.

Adsorption/desorption is one of the most significant factors that affect the solid-

solution interaction (O’Connor and Connolly, 1980). Under the assumption that

equilibrium conditions prevail, the solid-solution-interaction can be described in terms

of the conditional distribution coefficient (or partition coefficient) KD, which is defined

as the ratio of the adsorbed or the total particulate concentration (CP, w/w) to the

dissolved concentration (CS, w/v) of a chemical constituent: KD = CP / CS. KD’s are of

fundamental significance to geochemical modelling and pollution impact assessment

(e.g., Wood et al., 1995). The distribution coefficients provide empirical information

on the combined effect of heterogeneous reactions of an element at the solid-solution

interface. An elevated KD value may indicate that an element is associated and


- 62 -

transported with the solid phase, which eventually may become part of the sediment

and may never reach the ocean. From the dependence of the distribution coefficients

on salinity it is possible to identify whether or not there is a tendency for the release

of the respective metal from the particulate matter, when proceeding from fresh water

through brackish to seawater conditions. Estuarine cycling of trace metals covering

all these aspects and including also changes over decades has been studied

extensively in the Scheldt estuary; it is characterised as a system with strong

physico-chemical gradients, large anthropogenic inputs, a very long mixing zone of

fresh- and saltwater (up to 100 km) and a long residence time of about 3 months of

water in the upper estuary (e.g., Paucot and Wollast, 1997; Zwolsman et al., 1997;

Zwolsman and van Eck, 1999).

In this paper we investigate the environmentally relevant trace metals cadmium,

copper, iron, lead, cobalt and nickel in dissolved and particulate form. In addition to

contributing to the general knowledge about the trace metal behaviour in tropical

estuaries, a specific objective was to estimate the contamination state of the

Wanquan River and the Wenchang/Wenjiao River estuaries of Hainan Island (~

20°N).

4.2. Materials and Methods

4.2.1. Study area

Fig. 1. Estuaries of East Hainan under investigation: (a) Wenchang/Wenjiao River estuary (WWR) and (b) Wanquan River estuary (WR).

The Wanquan River (in the following: WR) and the Wenchang/Wenjiao River system


- 63 -

(in the following: WWR) in East-Hainan are both located at the northern part of the

tropical zone with a humid warm climate (Fig. 1). The Wanquan River, with a

drainage area of 3693 km2, a total length of 156.6 km and a mean annual discharge

of 163.9 m3/s, is the third largest river in Hainan Island (Ge et al., 2003). The

Wenchang River (drainage area 381 km2, total length 37 km, mean annual discharge

9.1 m3/s) and the Wenjiao River (drainage area 522 km2, total length 56 km, mean

annual discharge 11.6 m3/s) empty into the Bamen Bay (Zeng and Zeng, 1989) and

form the Wenchang/Wenjiao River estuary (areal extent: ~ 40 km2; Herbeck et al.,

2013). Both estuaries have a micro-tidal, irregular diurnal tidal regime with a mean

range of about 0.7-0.8 m (Zhu et al., 2005). The salinity intrusion may extend 5 km

from the Yudai Sand Barrier into the WR and from the Bamen Bay end into the WWR,

respectively. The WWR is a shallow estuary ( < 3 m; except for the shipping channel)

with a low flow rate, while in the WR the position of the mixing zone is highly variable

due to changes of the river discharge rate, tides and winds. About 80% of the annual

rainfall (1740 mm/yr; Liu et al., 2011) and of the water flow in this region occurs in the

wet period from May to November (Ma et al., 2007). Typhoons with heavy rainfall and

elevated input of soil erosion products, nutrients and suspended particles (Herbeck et

al., 2011) did not occur during the sampling for this study. The water residence time

in the estuaries was estimated to be 7.8 day in WWR (Liu et al., 2011) and 0.2-4.7

day in WR (Li et al., 2013). Comprehensive nutrient investigations of the WWR

system (Liu et al., 2011) revealed that the tributaries are enriched in dissolved

inorganic nitrogen and depleted in dissolved inorganic phosphorus and have major

contributions of dissolved organic forms to the total N and P dissolved concentrations;

these elements - in contrast to Si - show non-conservative behaviour. The WWR

system is characterized by lower silicate than average in tropical systems (Liu et al.,

2011), while the Wanquan River estuary (WR) is enriched in silicate (Li et al., 2013).

Because of the generally very low depths of about 3 m except for a few more meters

in the shipping channels, oxygen depletion or anoxic conditions with their far

reaching effects on redox-dependent trace metals (e.g., Balzer, 1982) were not

observed and could not be expected. Ma et al. (2007) investigated major and trace

elements in laterites of Northern Hainan which are in-situ weathering products

developed from basalts. Typical soils of the study area are Oxisols which are

especially rich in Fe2O3 (Li et al., 2012) which explains the comparatively high Fe

values in the estuary found during the present study. The Shilu Iron Mine as China’s

largest iron mine is located in the western part of the Hainan Island (Hsieh and

Zhong, 1990), but is outside the drainage area of either estuary studied.

Nevertheless, the high Fe content of the soil and possibly also some atmospheric


- 64 -

transport from this mine might affect aquatic processes of East-Hainan. In contrast to

the WR, the upper WWR estuarine system is girdled by shrimp and fish ponds with

an estimated area of ~ 21.6 km2 and estimated effluents of ~ 210 • 106 m3/yr

(Herbeck et al., 2013; Unger et al., 2013). Minor contamination of both systems may

also originate from municipal discharge of small towns. Hainan has a population

density of 255 inhabitants/km2.

4.2.2. Sampling

Surface water samples were collected during the dry season (Dec. 2006) and the wet

seasons (Aug. 2007 and Jul./Aug. 2008) in the WR and the WWR (exact positions

may be obtained from the corresponding author upon request). Upstream of the ship,

low-density polyethylene bottles (LDPE, 1 L, Nalgene®) were filled by hand with

surface water from a depth of about 0.3 m. Within 20 h the water samples were

drawn through acid-cleaned 0.4 mm polycarbonate filters (pre-weighed, Nuclepore®,

Whatman) into LDPE bottles (Nalgene®). A 10 mL aliquot of each sample was filled

into a glass ampoule for the determination of the dissolved organic carbon (DOC)

followed by the addition of phosphoric acid to remove the inorganic carbon and for

sterilization. After acidification to pH 2 with sub-boiled concentrated nitric acid,

sample bottles for trace elements were wrapped in polyethylene bags and stored at

room temperature. The filter cakes were transferred to Petri dishes and stored at -

10 °C.

4.2.3. Pretreatment and measurement of dissolved metals

Sample work-up was performed in a clean-air room (class 1000) with additional clean

benches of class 100. The filtrates were first irradiated in a UV-digester at 90 °C for 2

h with addition of hydrogen peroxide in order to destroy the organic complexes. Then

the trace metals were preconcentrated by an extraction/backextraction procedure

developed by our workgroup based on the work of Danielsson et al. (1978). A 40-100

g sample (depending on the metal concentration) was weighed into a fluorinated

ethylene propylene separatory funnel (FEP, Nalgene®) and after adding (i)

ammonium acetate buffer to adjust the pH value of the sample to 4.0-4.5, (ii) 500 μL

APDC/NaDEDTC solution (0.06 mol/L) and (iii) 10 mL Freon, the sample had been

well shaken (15 min). After phase separation, the lower organic phase was drawn off

into an acid-cleaned polypropylene tube and 100 μL concentrated nitric acid were

added, followed by a further addition of 1900 μL Milli-Q® water. After phase


- 65 -

separation, the upper aqueous extract was removed into an acid-cleaned sample

tube by using a pipette. The extract was analyzed by high-resolution inductively-

coupled-plasma mass-spectrometry (HR-ICP-MS; Element 2, Thermo Scientific).

Four certified reference materials from the National Research Council of Canada

(NRCC) were investigated: SLRS-5 (river water), SLEW-3 (estuarine water), CASS-4

and NASS-5 (seawater); the results for all metals in all reference materials were

inside the certified ranges. The detection limits (3σ; n = 13) for Milli-Q® water being

subjected to UV-digestion and liquid/liquid-extraction were 0.004, 0.016, 2.63, 0.067,

0.004, 0.127 for Cd, Pb, Fe, Ni, Co, Cu, respectively. The mean precisions of

replicates during the analysis of the certified reference materials were 0.7%, 2.6%,

10.2%, 5.7%, 2.6%, 5.3% for Cd, Pb, Fe, Ni, Co and Cu, respectively.

4.2.4. Pretreatment and analysis of particulate metals

The filter cakes were weighed and completely decomposed using a microwave-

assisted pressure digestion technique after adding a mixture of concentrated nitric,

hydrochloric and hydrofluoric acid (for details, see: Balzer et al., 2013). At regular

intervals the certified reference material MESS-1 (marine sediment) from the NRCC

was analyzed and all values were found to be in the certified ranges. The protocol

implies a total destruction of the particles and includes trace metals in surface layers

and in lattice positions. Therefore, results cannot be directly compared to many other

studies of estuarine particles, during which the metals are leached from the surface

layers by employing e.g., acetic acid or dilute hydrochloric acid. The metal

concentrations were determined by HR-ICP-MS.

4.2.5. Other parameters

Environmental parameters such as salinity and pH were measured on the spot using

a multi-sensor-device (MultiLine P3, WTW). Dissolved organic carbon (DOC) in the

sealed glass ampoules was determined by high temperature catalytic oxidation (Total

Carbon Analyzer, Apollo 9000; Tecmar) with a detection limit of 0.1 mg/L. Replicates

from 19 samples gave an average relative standard deviations of 2.5% (n = 5).

Before use, the glass ampoules were pre-combusted at 450 °C for 4 h.


- 66 -

4.3. Results and discussion

4.3.1. Distribution of suspended particulate matter and environmental conditions

WR

0

5

10

15

20

0 10 20 30

SPM

(mg/

L)

WWR

0

5

10

15

20

0 10 20 30

SPM

(mg/

L)

WR

0

2

4

6

8

0 10 20 30Salinity (psu)

DO

C (p

pm)

WWR

0

2

4

6

8


DO

C (p

pm)

Fig. 2. Distribution of SPM and DOC vs. salinity (open symbols for the dry season and solid symbols for the wet seasons; : WR 2006; : WWR 2006; : WR 2007; : WWR 2007; ▲: WR 2008; : WWR 2008).

The distributions of the suspended particulate matter (SPM) and the dissolved

organic carbon (DOC), which were obtained during the three expeditions in the two

estuaries, are shown in Fig.2.

Except for a slight minimum at mid salinities, the SPM concentration was essentially

constant in both estuaries during the wet seasons (Aug. 2007: on average 2.3 mg/L

and Jul./Aug. 2008: on average 2.0 mg/L). The SPM values during the dry season

(Dec. 2006: on average 9.8 mg/L) showed more variability and were distinctly higher

than during both wet seasons, possibly originating from higher resuspension loadings

at lower water levels and/or from enhanced dust deposition. Also Li et al., 2013 report

higher SPM values for the dry season (Dec. 2006, Mar./Apr. 2009) in comparison to

the wet seasons (Aug. 2007, Jul/Aug. 2008) for the Wanquan estuary (WR). The

SPM concentrations were much lower than in other Chinese estuaries e.g., in the

Daliao River (8-1001 mg/L; Guo et al., 2009), in the Changjiang estuary (70-350


- 67 -

mg/L; Koshikawa et al., 2007), or in the Zhujiang (Pearl River) Delta (70-247 mg/L; Ni

et al., 2008). Summarizing the studies in several Chinese estuaries, Zhang and Liu

(2002) reported a typical SPM range from 10 mg/L to 20000 mg/L. The SPM content

in both study areas is obviously at the lower limit of this range.

The DOC content generally decreased with increasing salinity with concentration

ranges of 1-3 ppm during the dry season and 1-8 ppm during the two wet seasons.

Comparable data was determined in the Amazon estuary (1-4 ppm; Sholkovitz et al.,

1978) and in the Zhujiang estuary (1-4 ppm; Dai et al., 2000). The DOC content in

the WWR was considerably higher than in the WR and its distribution showed much

more scatter especially during the wet season of 2008. In accordance with the fact

that the WWR is an aquacultural region with some biological activity, anthropogenic

sources from the aquacultural activities and overflow during storms were suspected

to be responsible for the higher DOC concentration in the WWR. In contrast to WR,

both the river and the upper estuary (Bamen Bay) of the WWR system are

surrounded by aquaculture ponds as shown in detail by Unger et al., 2013. Also the

different DOC levels between the two wet seasons in WWR are probably related to

different aquaculture inputs during the specific sampling campaigns.

The pH increased with increasing salinity (not shown). It ranged between 7.0 and 8.3

in the WR and between 6.8 and 8.5 in the WWR which is comparable with other

Chinese estuaries: e.g., in the Zhujiang estuary (7.0-8.0; Dai et al., 2000) and in the

Changjiang estuary (7.5-8.3; Zhang and Zhang, 2003).


- 68 -

WR

0.00

0.05

0.10

0.15

0 10 20 30

Cd

(nm

ol/L

)

WWR

0.00

0.05

0.10

0.15

0 10 20 30

0

5

10

15

20

0 10 20 30

Cu

(nm

ol/L

)

0

5

10

15

20

0 10 20 30

0

5

10

15

0 10 20 30

Fe (μ

mol

/L)

0

5

10

15

0 10 20 30

0

1

2

3

0 10 20 30

Pb (n

mol

/L)

0

1

2

3

0 10 20 30

0

3

6

9

0 10 20 30

Co

(nm

ol/L

)

0

3

6

9

0 10 20 30

0

5

10

15

20


Ni (

nmol

/L)

0

5

10

15

20


Fig. 3. Dissolved trace metals (left column: WR, right column: WWR) vs. salinity (open symbols: dry season, solid symbols: wet seasons; : WR 2006; : WWR 2006; : WR 2007; : WWR 2007; ▲: WR 2008; : WWR 2008); for further details: see text.


- 69 -

4.3.2. Dissolved trace metals

Tab. 1. Dissolved trace metals (average values ± 1 standard deviation; n = number of values) in freshwater (S = 0), in the estuarine mixing zone (0 < S < 28) and at high salinities (S > 28) of the WR and the WWR during the dry season (2006) and the wet seasons (2007 and 2008).

Expedition Cd Cu Fe Pb Co Ni

nmol/L nmol/L μmol/L nmol/L nmol/L nmol/L

WR 2006

S=0 (n=3) 0.02±0.01 17±16 0.23±0.09 0.25±0.26 0.27±0.08 3.2±1.0

0<S<28 (n=3) 0.11±0.02 19±19 0.06±0.05 0.30±0.13 6.1±9.0 6.5±2.2

S>28 (n=0) -- -- -- -- -- --

WWR 2006

S=0 (n=2) 0.07±0.09 8.2±0.5 0.44±0.39 0.06±0.06 2.6±3.0 8.1±0.5

0<S<28 (n=3) 0.10±0.01 8.7±2.1 0.59±0.79 0.12±0.13 6.8±1.7 12.1±2.6

S>28 (n=4) 0.14±0.01 8.4±2.9 0.04±0.03 0.08±0.04 0.98±0.34 5.9±0.7

WR 2007

S = 0 (n=2) 0.06±0.02 9.7±0.7 4.2±0.2 1.4±0.2 1.6±0.2 6.6±0.8

0<S<28 (n=8) 0.06±0.01 7.6±0.9 2.5±0.8 0.81±0.18 1.2±0.4 5.6±1.2

S>28 (n=1) 0.06 2.3 0.28 0.15 0.41 3.8

WWR 2007

S = 0 (n=2) 0.02±0.01 8.3±1.1 3.7±2.6 0.51±0.51 2.4±0.2 11.9±8.3

0<S<28 (n=8) 0.07±0.03 7.1±1.0 1.1±1.3 0.23±0.17 2.6±1.4 7.0±2.3

S>28 (n=1) 0.09 6.2 0.16 0.06 0.91 4.6

WR 2008

S = 0 (n=4) 0.03±0.00 6.6±1.8 3.7±0.7 1.0±0.2 1.6±0.4 2.8±1.3

0<S<28 (n=5) 0.08±0.02 7.0±3.0 2.2±2.4 0.60±0.58 0.90±0.59 5.3±2.6

S>28 (n=1) 0.10 1.7 0.16 0.01 0.09 4.2

WWR 2008

S = 0 (n=4) 0.02±0.01 11.4±1.1 9.6±1.6 1.6±0.6 4.2±0.6 6.8±2.8

0<S<28 (n=10) 0.07±0.03 12.0±0.9 4.0±4.1 0.85±0.61 2.7±1.3 10.2±3.0

S>28 (n=1) 0.09 7.6 0.35 0.25 0.67 2.8

The dissolved trace metal concentrations obtained during the three expeditions in the

two estuaries are plotted as a function of salinity in Fig. 3. The dissolved metal

concentrations depicted for salinity zero, which show considerable variability for most

metals, are not always representative for the geographical freshwater entrance to the

estuaries but also include samples which were taken upstream the rivers at different

dates. The average concentrations for freshwater, in the estuarine mixing zone (0 < S

< 28) and at high salinities of the WR and the WWR during the three expeditions are

shown in Table 1.


- 70 -

Cadmium: Dissolved Cd concentrations generally increased with increasing salinity

from 0.01 to 0.05 nmol/L in the freshwater to about 0.14 nmol/L in the saline waters

during the dry season (only WWR) and to about 0.10 nmol/L at salinities above S =

28 during the wet seasons. During the dry season of 2006 and the wet season of

Jul./Aug. 2008, the dissolved Cd showed a clear positive deviation from the

theoretical dilution line in both estuaries. A similar non-conservative behaviour of Cd

was also observed in many other studies, including the Changjiang estuary (Edmond

et al., 1985; Elbaz-Poulichet et al., 1987), the Hudson River estuary (Yang and

Sañudo-Wilhelmy, 1998), the Loire estuary (Waeles et al., 2004) and during

laboratory experiments (e.g., Roux et al., 1998). This behaviour indicates a

mobilization of dissolved Cd from the SPM into the solution phase due to the

formation of highly stable Cd-chloro-complexes (Van der Weijden et al., 1977;

Kraepiel, et al., 1997). Although showing some scatter, the dissolved Cd distribution

during the wet season of August 2007 seems to be conservative. Such a behaviour,

which was also observed e.g., in the outer Severn estuary (Harper, 1991) and in the

Rhône estuary in July 1987 (Elbaz-Poulichet et al., 1996), might be related to low

particulate Cd per volume and a short residence time of water and particles in the

estuary as suggested by Kraepiel et al. (1997). While the residence times in the

Hainan estuaries are indeed short (see above), the actual concentrations of

particulate Cd per volume in 2007 and 2008 do not explain such a different behaviour

between the two years. Thus, the exact reason for it remains open. Dilution by

enhanced rainfall and the slightly lower pH during the dry season might have

contributed to the seasonally lower concentrations of dissolved Cd during the wet

seasons (Hatje et al., 2003). Because the dry season is also associated with

comparably high levels of SPM, very low concentrations of dissolved Fe and Pb (see

below), and essentially equal values for dissolved Cu and Ni, unknown properties of

the additional particles might have played a role for the removal of dissolved Fe and

Pb and the simultaneous release of Cd.

Copper: With increasing salinity dissolved Cu exhibited a slight but continuous and

essentially conservative decline, which is clearly evident in the WR during all

investigated seasons, but can be identified in the WWR estuary only as a tendency.

Elevated Cu concentrations were observed during the wet season 2008 in WWR,

coinciding with higher levels of DOC (Fig. 2). In combination with more scatter this

may be explained by irregular inputs from adjacent aquaculture ponds (Balzer et al.,

2013). Dissolved copper decreased from typically about 8-12 nmol/L in the

freshwater to about 3-7 nmol/L at higher salinities. A similar conservative behavior of


- 71 -

dissolved Cu was also observed in the Amazon plume (Boyle et al., 1982), in the

Changjiang estuary (Edmond et al., 1985), in the Mississippi River Delta (Shiller and

Boyle, 1991) and in the tropical Bang Pakong estuary (Windom et al., 1988), and

seems to be related to the Cu speciation in form of stable complexes with organic

ligands (Shiller and Boyle, 1991; Roux et al., 1998). Strong positive deviations from

simple dilution of dissolved Cu as observed e.g., in the Scheldt estuary were

explained by sulfide precipitation under anoxic conditions at low salinities (Paucot

and Wollast, 1997; Zwolsman et al., 1997); however, oxygen depletion was not likely

to occur in the estuaries under investigation.

Iron: Dissolved Fe in both study areas and during both wet seasons decreased from

typically 5-10 μmol/L in the riverine water to 0.1-0.2 μmol/L at higher salinities.

Although the concentration of dissolved Fe in both estuaries during the dry season

was about 10 times lower than during both wet seasons, there was also a slight

decrease with salinity in both estuaries. Dissolved Fe showed a non-conservative

behavior and appeared to be removed from solution in both estuaries and during

both dry and wet seasons. Two elevated data points for Fe (and Pb) at low salinities

are considered outliers being due to either contamination or irregular input from

aquaculture ponds. Field and laboratory studies indicate that the main process

affecting Fe during estuarine mixing is flocculation of colloidal Fe forms and

subsequent sedimentation, globally resulting in a reduction of more than 90% of the

Fe river input during its way to the ocean (Boyle et al., 1977; Sholkovitz, 1978). In the

wet season of 2008, the dissolved Fe of the freshwater and at low salinities in the

WWR was two times higher than during all other expeditions and in the WR,

suggesting lateral inputs from aquaculture ponds as discussed before. The relatively

low concentrations of Fe (and Pb) in solution during the dry season 2006 coincide

with similar particulate Fe levels and similar Fe/Al ratios but an about six times higher

SPM concentration. This suggests enhanced adsorption onto unsaturated sites of the

additional particle load during 2006. However, the suspended particles form only a

fraction of the total particulate load with a major contribution from bed sediments in

addition to riverine particles (Kraepiel et al., 1997). Thus, the striking difference

between dissolved Fe and Pb during the dry and the wet seasons might also arise

from different ratios of the two particle types.


- 72 -

0

1

2

0 5 10 15Fe (μmol/L)Pb

(nm

ol/L

)

Fig. 4. Dissolved Pb vs. dissolved Fe (open symbols: dry season, solid symbols: wet seasons; : WR 2006; : WWR 2006; : WR 2007; : WWR 2007; ▲: WR 2008; : WWR 2008). Broken and solid line: correlation for all WWR data (r2 = 0.849) and for all WR data (r2 = 0.923), respectively.

Lead: In the estuaries, dissolved Pb during both wet seasons exhibited (similar to Fe)

a very strong decline from riverine waters (1-2 nmol/L) to marine waters (0.02-0.3

nmol/L). Such a distribution was also observed in the Gironde estuary (Elbaz-

Poulichet et al., 1984), in the Bristol Channel (Harper, 1991), in the Bang Pakong

estuary (Windom et al., 1988) and in the Cochin estuary (Ouseph, 1992) and is often

explained by rapid sorption of dissolved Pb onto re-suspended particles. In contrast,

dissolved Pb during the dry season showed only a small decline with increasing

salinity and was 40-80% lower than during the wet seasons. Considering the good

correlation between dissolved Fe and dissolved Pb (r2 = 0.849 for all WR data, r2 =

0.923 for all WWR data; Fig. 4), a removal of dissolved Pb with freshly precipitated

Fe-oxy-hydroxides may be suspected (e.g., Balls, 1990; Wen et al., 2008).

0

2

4

6

0 5 10 15Fe (μmol/L)

Co

(nm

ol/L

)

Fig. 5. Dissolved Co vs. dissolved Fe during the wet seasons (: WR 2007; : WWR 2007; ▲: WR 2008; : WWR 2008). Broken and solid line: correlation for all WWR data (r2 = 0.887) and for all WR data (r2 = 0.515), respectively.

Cobalt: During both wet seasons, dissolved Co in WR decreased continuously with


- 73 -

increasing salinity from 2 to 5 nmol/L to less than 1 nmol/L and behaved similar as Fe

and Pb. Also the differences between WR and WWR were similar to Fe and suggest

that the distribution of Co might also be controlled by co-precipitation with Fe-oxy-

hydroxides (Fig. 5). The good correlation of dissolved Co vs. Fe in WR (r2 = 0.887)

during the wet seasons supports the predominance of Fe-related processes. On the

other hand, the weak correlation for the WWR system (r2 = 0.515) provides evidence

for contribution from other processes; most probable are lateral inputs from

aquaculture ponds which girdle the WWR rivers and upper estuary in contrast to the

WR estuary (Herbeck et al., 2013; Unger et al., 2013). This behavior of Co during the

wet seasons contrasts with reports of dissolved Co maxima at low or mid salinities

(Windom et al., 1988; Martino et al., 2002; Takara et al., 2010). While the few

observations during the dry season in WR and WWR might be interpreted as low

salinity maxima or conservative behaviour (at S > 1), it is clear that additional work is

needed to understand the estuarine behavior of cobalt.

Nickel: During both wet seasons dissolved Ni concentration was essentially constant

at about 5 nmol/L over the whole salinity profile in the WR, while in the WWR Ni

decreased from 10 to 15 nmol/L to also about 5 nmol/L at higher salinities. When

excluding two higher values, dissolved Ni might be considered conservative in both

estuaries during the wet seasons, as also observed e.g., in the Göta River estuary

(Danielsson et al., 1983). A part of the Ni variability at low salinities may be explained

by fluctuations in the incoming river water (Chaudry and Zwolsman, 2008) and (in the

WWR) may also contain an anthropogenic contribution from the Wenchang town

being located at the Wenchang tributary. In addition, at salinities of about S = 20

diffuse inputs in WWR from aquaculture may be suspected. During the dry season,

dissolved Ni had a tendency for a positive deviation from conservative behavior,

which might be attributed to anthropogenic discharges (e.g., Campbell et al., 1988;

Martino et al., 2004).


- 74 -

WR

0

4

8

12

0 10 20 30

Cd

(nm

ol/g

)

WWR

0

4

8

12

0 10 20 30

0

1

2

3

0 10 20 30

Cu

(μm

ol/g

)

0

1

2

3

0 10 20 30

0

1

2

3

0 10 20 30

Fe (m

mol

/g)

0

1

2

3

0 10 20 30

0.0

0.3

0.6

0.9

0 10 20 30

Pb

(μm

ol/g

)

0.0

0.3

0.6

0.9

0 10 20 30

0.0

0.3

0.6

0.9

0 10 20 30

Co

(μm

ol/g

)

0.0

0.3

0.6

0.9

0 10 20 30

0

1

2

3


Ni (μm

ol/g

)

0

1

2

3


Fig. 6. Particulate trace metals (left column: WR, right column: WWR) vs. salinity (open symbols: dry season, solid symbols: wet seasons; : WR 2006; : WWR 2006; : WR 2007; : WWR 2007; ▲: WR 2008; : WWR 2008); for further details: see text.


- 75 -

4.3.3. Particulate trace metals

Table 2. Particulate trace metals (average values ± 1 standard deviation; n = number of values) in freshwater (S = 0), in the estuarine mixing zone (0 < S < 28) and at high salinities (S > 28) of the WR and the WWR during the dry season (2006) and the wet seasons (2007 and 2008).

Expedition Cd Cu Fe Pb Co Ni

nmol/g μmol/g mmol/g μmol/g μmol/g μmol/g

WR 2006

S=0 (n=3) 5.3±0.1 0.33±0.02 1.2±0.0 0.26±0.00 0.30±0.03 0.57±0.04

0<S<28 (n=3) 3.1±1.1 0.28±0.02 0.88±0.24 0.20±0.05 0.28±0.12 0.46±0.06

S>28 (n=0) -- -- -- -- -- --

WWR 2006

S=0 (n=2) 6.1±4.6 0.57±0.32 1.8±0.5 0.18±0.01 0.34±0.20 1.3±0.7

0<S<28 (n=3) 3.3±1.0 0.55±0.20 1.3±0.6 0.11±0.02 0.34±0.07 1.1±0.5

S>28 (n=4) 1.7±0.3 0.35±0.04 0.61±0.09 0.13±0.02 0.23±0.04 0.60±0.08

WR 2007

S = 0 (n=2) 4.6±0.7 0.46±0.02 1.1±0.0 0.26±0.01 0.44±0.01 0.81±0.06

0<S<28 (n=8) 2.2±0.6 0.48±0.25 0.91±0.19 0.21±0.04 0.27±0.05 0.54±0.10

S>28 (n=1) 3.2 0.25 0.57 0.17 0.22 0.45

WWR 2007

S = 0 (n=2) 3.2±1.7 0.52±0.30 1.9±0.3 0.23±0.06 0.91±0.69 1.2±0.3

0<S<28 (n=8) 3.5±1.1 0.32±0.20 0.68±0.55 0.17±0.09 0.25±0.20 0.9±1.2

S>28 (n=1) 4.2 0.26 0.26 0.09 0.11 0.32

WR 2008

S = 0 (n=4) 6.7±1.6 0.81±0.29 1.0±0.4 0.32±0.15 0.40±0.15 0.73±0.28

0<S<28 (n=5) 2.9±0.4 0.68±0.11 0.94±0.12 0.27±0.02 0.25±0.07 0.76±0.13

S>28 (n=1) 2.8 0.50 0.42 0.16 0.15 0.37

WWR 2008

S = 0 (n=4) 3.2±1.3 0.81±0.32 0.82±0.28 0.17±0.02 0.39±0.15 1.0±0.4

0<S<28 (n=10) 3.5±0.7 0.88±0.21 0.70±0.25 0.170.03 0.32±0.14 0.88±0.28

S>28 (n=1) 3.8 0.89 0.58 0.22 0.21 1.3

The concentrations of Cd, Cu, Fe, Pb, Co and Ni in the particles obtained during the

three expeditions in the two estuaries are plotted as a function of salinity in Fig. 6.

The average values are shown in Table 2. In the WR system, all particulate metals

show a very homogeneous distribution with constant or slightly decreasing levels

over the whole estuarine salinity profile. Except for some variability among the

freshwater samples, no difference occurred between the seasons. In contrast, the

salinity profiles of the particulate trace metals in the WWR system show much more

scatter in the data, but mainly at the same concentration levels as in WR. Some


- 76 -

slightly positive deviations from simple endmember mixing suggest that in WWR

particles with different properties might have contributed to the composite signals.

Likely candidates with slightly differing properties are riverine particles entering the

estuary, resuspended bottom sediments and possibly also some aquacultural

particles.

In order to identify particles of potentially differing origin among the expeditions and

to obtain hints for biogenic and/or anthropogenic contributions (Zhang and Liu, 2002),

the trace metal contents of all SPM samples were normalized to aluminium as a

proxy for the mineral particle matrix (not shown). The particulate Al itself decreased

linearly along the salinity profile from 3.0 ± 1 mmol/g in the freshwaters and in the low

salinity region to 2.2 ± 0.3 mmol/g near the coastal endmember (S > 28). The only

exception were some low Al values during the wet season 2007 in WWR at S = 20-22;

concomitant high ratios to Al for Cd and Pb but not for Cu, Co, Ni suggests an

anthropogenic contribution; alternatively, it may originate from an unusually high

fraction of biogenic particles with elevated contents of particulate carbonates and/or

organic matter (e.g., Koshikawa et al., 2007).

From high riverine values in the WR particulate Cd rapidly decreased and remained

constant over the whole salinity range during both wet seasons, which was also

found in the Gironde estuary (Kraepiel et al., 1997). In the WWR system, the

particulate Cd appeared to slightly increase, but at a similar level as in the WR. The

Cd/Al ratio was constant along the salinity profile except for elevated ratios of Cd/Al

(and also of Pb/Al) at S = 20-28 during the wet season 2007 in WWR. Exponential

decreases of particulate Cd (and also Cu) over the whole salinity profile (i.e.,

negative deviations from simple dilution) being attributed to desorption into the

dissolved phase (e.g., in the highly polluted Scheldt estuary; Zwolsman et al., 1997;

Paucot and Wollast, 1997) cannot be detected in the Hainan estuaries. With few

other exceptions (e.g., enrichment of particulate Cd at mid salinities in the Danshuei

River estuary; Jiann et al., 2005), simple estuarine mixing seems to be the dominant

feature of particulate Cd worldwide (e.g., Balls, 1990; Paalman and Van der Weijden,

1992; Wen et al., 2008; Zwolsman and van Eck, 1999).

In both estuaries and during dry and wet seasons the particulate Cu showed no

considerable change except for a slight decrease towards the coast. Such

distributions are often found in estuaries, e.g., in several Texas estuaries (Benoit et

al., 1994) and in the Gironde estuary (Kraepiel et al., 1997). However, also a positive

relation of particulate Cu (and Pb) to salinity as a result of pH dependent sorption to

Fe-Mn-oxides has been reported (Hatje et al., 2001). In contrast to any other of the


- 77 -

six particulate trace metals, elevated values of particulate Cu/Al were observed

during the wet season 2008 in WWR over the whole salinity range. In accordance

with comparatively high values for dissolved and particulate arsenic during the same

season in WWR (Balzer et al., 2013), a relatively high contribution of copper-rich

inputs from adjacent aquaculture ponds may be suspected. This presumption is

corroborated by a few measurements of aquaculture effluents which showed some

very high dissolved Cu values.

Particulate Fe, Ni and Co during all seasons decreased with increasing salinity,

generally with a larger slope in the WWR than in the WR. Such a distribution pattern

was often found, e.g., in the Clyde estuary (Balls, 1990), in the Rhine/Meuse estuary

(Paalman and Van der Weijden, 1992) and in the Tanshui River estuary (Fang and

Lin, 2002). Also particulate Pb showed a slight decrease with salinity but no temporal

variability. This essentially conservative salinity profile agrees with observations in the

Gironde estuary (Kraepiel et al., 1997) and - combined with a negative relation to

salinity - it was also observed in the Weser estuary (Turner et al., 1992). The ratios to

Al of particulate Fe, Pb, Co and Ni during all expeditions (and seasons) were rather

constant over the whole salinity range except for some high values of Fe/Al in WWR

at low to mid salinities in 2006 and 2007, and of Pb during the wet season 2008. It

cannot be ruled out that the high Fe/Al during the dry season 2006 partly correspond

to the extremely low dissolved Fe during that expedition.

A striking feature of the ratios to Al of all trace metals under investigation was their

rather constant lower limit over the whole salinity range. The ratios (in: mol/mol) of

Cd/Al = 6.0•10-7, Cu/Al = 1.0•10-4, Fe/Al = 2.4•10-1, Pb/Al = 6.0•10-5, Co/Al = 8.0•10-5

and Ni/Al = 1.8•10-4 might be representative for the lattice composition of the particles

from the drainage area. However, these ratios to Al for Cu, Fe, Co and Ni are lower

by 18-70% than the top layer of the East-Hainan laterites which are the products of

extreme tropical weathering (Ma et al., 2007; no data for Cd). In contrast, the Pb/Al of

the estuarine particles was considerably higher than in the weathered laterites. If the

above metal/Al ratios represent the lattice composition of the estuarine particles and

if these particles are the weathering products of basalts from which the laterites

formed, then the four metals Cu, Fe, Co and Ni were more refractory than Al in the

parent material.

On a volume basis, the percentages of particulate metal on the total metal

concentration (dissolved plus particulate) were essentially constant for Cd (~ 10 %),

Cu (~ 14 %), Co (~ 30 %) and Ni (~ 22 %) in the estuarine waters (S > 0), but with

generally higher values during the dry season. In contrast to this behavior of


- 78 -

relatively constant particulate concentrations over the whole salinity range, the

particulate concentrations per volume of Fe, Al, Cu, Ni, Pb and Cd in Galveston Bay

all showed a general decrease (Wen et al., 2008).

0

20

40

60

80

100

0 20 40 60 80 100Fe (%)

Pb (%

)

Fig. 7. Ratio of the percentages of the particulate metal on the sum of the dissolved and particulate metal: Pb vs. Fe (open symbols: dry season, solid symbols: wet seasons; : WR 2006; : WWR 2006; : WR 2007; : WWR 2007; ▲: WR 2008; : WWR 2008).

The percentage of particulate Fe and Pb on the total concentration increased from

about 10 % to more than 80% along the salinity profile, which reflects the intense

conversion of Fe and Pb from solution to particles in the estuaries and the strong

preference of these two metals for particulate forms (Fig. 7). This also confirms the

assessment made before, that flocculation of Fe colloids and co-precipitation of Pb

were the dominant processes of particle-solution interaction for these metals in the

two study areas.

The freshwater concentrations of particulate metals as one endmember for the

mixing processes in the two estuaries were generally much lower than in typical

European and North-American estuaries with a more industrialized character (Balls,

1989, 1990), but they agreed with other Chinese estuaries (Table 3). For the marine

endmember with salinities S ≈ 30, neither considerable temporal nor spatial variability

was observed in the two estuaries, and the concentrations were comparable with

typical levels of particulate Fe, Cu, Co and Ni in the South China Sea (Ho et al.,

2007).


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Table 3. Particulate trace metals and SPM of East-Hainan estuaries (average values ± 1 standard deviation) in comparison to selected Chinese rivers (S = 0) and estuaries (S > 0) (averages or typical ranges; *: estimated from the figures; **: extract or leach).

Study area SPM Cd Cu Fe Pb Co Ni

Source mg/L nmol/g μmol/g mmol/g μmol/g μmol/g μmol/g

East Hainan

S=0 5.3±4.4 4.9±2.1 0.62±0.29 1.21±0.46 0.24±0.09 0.44±0.27 0.90±0.40

a 0<S<28 3.0±3.3 3.1±0.9 0.57±0.29 0.85±0.38 0.19±0.06 0.29±0.12 0.79±0.61

S>28 6.1±4.6 2.7±1.0 0.44±0.21 0.54±0.14 0.15±0.04 0.20±0.05 0.59±0.3

Daliaohe (S=0) 851 -- 0.80 0.93 0.69 1.42 2.91 b

Yalujiang (S=0) 127 5.6 0.62 0.69 0.27 0.23 0.63 c

Shuangtaizihe (S=0) 4651 6.9 0.60 0.78 0.40 0.35 0.86 c

Luanhe (S=0) 4633 7.4 0.86 0.90 0.33 0.29 0.88 c

Haihe (S=0) 3927 7.7 0.61 -- 0.37 -- 0.88 b

Huanghe (S=0) 25581 1.6 0.44 0.63 0.08 -- -- d

Huanghe (S=0) 2000-31500 1.5 0.42 0.67 0.08 0.24 0.69 e

Huanghe (S=0) 26829 1.5 0.42 0.67 0.08 0.24 0.69 b

Changjiang (S=0) 494 2.7 0.76 0.84 0.16 -- -- f

Changjiang (S=0) 539 2.6 0.98 0.93 0.24 0.32 2.11 b

Changjiang (S=0) 10-20000 3.6 0.98 0.93 0.19 0.32 1.09 c

Changjiang estuary (S>0) 100-1000 -- 0.8* -- 0.14* -- 1.0* g

Changjiang estuary (S>0)** 70 2* 1.1* 0.5* 0.2* 0.4* 2.6* h

Qiantangjiang (S=0) 187 2.5 1.41 -- 0.37 -- 1.58 b

Jiaojiang (S=0) 1254 6.7 0.57 0.65 0.26 0.29 0.79 c

Minjiang (S=0) 128 2.8 0.82 -- 0.30 -- -- b

Jiulongjiang (S=0) 211 4.4 0.62 1.03 0.29 0.39 1.38 b

Zhujiang (S=0) 348 5.3 0.77 -- 0.29 0.32 -- d

Zhujiang (S=0) 227 6.7 0.80 0.88 0.36 0.30 1.05 c

Tanshui River estuary (S>0)** 15* -- 0.35-7.9 0.19-1.2 -- 0.03-0.32 0.1-1.8 i

Tanshui River estuary (S>0) 10* 2-20* 1-11* -- 0.2-0.7* -- 1-6* j

Open South China Sea 0.4 -- 0.68 0.40 -- 0.25 0.46 k

(a): This study; (b): Zhang, 1995; (c): Zhang and Liu, 2002; (d): Qu and Yan, 1990; (e): Huang et al., 1992; (f): Martin and Meybeck, 1979; (g): Zhang, 1999; (h): Wang and Liu, 2003; (i): Fang and Lin, 2002; (j): Jiann et al., 2005; (k): Ho et al., 2007.

4.3.4. Distribution coefficients

The use of partition coefficients KD = CP / CS is a common approach for describing

solid-solution interactions; the KD, however, is not a true equilibrium coefficient, but

an empirical term which depends on various factors such as solution and particle

composition, speciation, analytical approach to particulate forms, ratio of colloidal


- 80 -

forms to truly dissolved ions, DOC and SPM concentration, pH, temperature, and

others (Bourg, 1987; Comans and van Dijk, 1988; Balls, 1989; Turner, 1996; Hatje et

al, 2003). The situation is further complicated by the fact that suspended particles

represent mixtures of particles originating from the rivers entering the estuary and

from resuspended bottom sediments. For the samples of this study, the logarithms of

the distribution coefficient were plotted versus salinity in Fig. 8.

Cd

3

4

5

6

7

8

0 5 10 15 20 25 30 35Salinity

logK

D

Cu

3

4

5

6

7

8

0 5 10 15 20 25 30 35Salinity

Fe

3

4

5

6

7

8

0 5 10 15 20 25 30 35Salinity

logK

D

Pb

3

4

5

6

7

8

0 5 10 15 20 25 30 35Salinity

Co

3

4

5

6

7

8

0 5 10 15 20 25 30 35Salinity

logK

D

Ni

3

4

5

6

7

8

0 5 10 15 20 25 30 35Salinity

Fig. 8. Log10KD (L/kg) vs. salinity (open symbol: dry season, solid symbols: wet seasons; : WR 2006; : WWR 2006; : WR 2007; : WWR 2007; ▲: WR 2008; : WWR 2008).

Among the six elements under investigation, only the KD value of Cd clearly

decreased along the salinity profile in the estuary. This appears to result from

complexation of dissolved Cd by seawater anions forming highly stable Cd-chloro-


- 81 -

complexes and from successful competition of seawater cations for sorption sites on

the SPM (e.g., Balls et al., 1994; Turner, 1996). In contrast, the KD’s of Cu, Ni and Co

showed neither temporal nor spatial variation nor any significant change with salinity.

Fe and Pb have the greatest affinity for the particulate phase with a log10KD increase

along the salinity profile from 5.4 to 8.0 L/kg for Fe and from 4.9 to 7.2 L/kg for Pb

during the wet seasons. Such a positive correlation of KD’s for Fe and Pb with salinity

and no relation to salinity for Cu and Ni was often observed in estuaries (e.g., Balls et

al., 1994). The higher KD’s for Fe and Pb in both estuaries during the dry season

were a consequence of its very low dissolved Fe and Pb and seemed to be related to

the specific properties of its much higher SPM loading. Several processes may

contribute to these peculiarities of the dry season, such as the additional input of

particles with differing surface properties from the drainage basin and/or from bottom

sediments, and the enhanced flocculation of colloidal Fe. An explanation is difficult,

because the dissolved Fe during the dry season is - per volume - more than one

order of magnitude lower than the Fe content in the particles which have a similar

composition at all seasons (Fig.6).

The general order of the log10KD for all samples was on average: Fe (6.0) > Pb

(5.7) > Co (5.3) > Ni (5.1) > Cu (4.8) ≈ Cd (4.8), and essentially agreed with the

results presented for the Seine estuary (Pb > Co > Cu > Zn > Ni > Cd; total digestion;

Chiffoleau et al., 1994) and for North Australian estuaries (Fe > Pb > Zn > Ni ≈ Cd ≈

Cu; extract at pH 2; Munksgaard and Parry, 2001). Although sorption processes of

the trace metals are sensitive for the source and the composition of the particles, and

although the KD is influenced by the method adopted for the particulate metal

determination (total destruction vs. partial extraction; Turner et al., 1992), the KD’s

obtained in this study differ from published data (e.g., Chiffoleau et al., 1994; Zhang,

1995; Kraepiel et al., 1997; Munksgaard and Parry, 2001; Tang et al., 2002) by less

than one order of magnitude.

In estuarine studies a so-called “particle concentration effect” is often observed,

which means that for a particular element the KD decreases with increasing SPM

loadings (e.g., Balls, 1989; Benoit and Rozan, 1999; Jiann et al., 2005). When the

KD’s of this study are plotted vs. SPM (not shown), the values for all investigated

metals form one cluster for the wet seasons at SPM loadings of 1-4 mg/L and a

second cluster for the dry season at SPM loadings of 8-12 mg/L (cf. Fig. 2). Possibly

owing to the overall relatively low SPM concentrations and the low particulate

concentrations in WR and WWR, no relationship at all was found between KD’s and

SPM for Cd, Cu, Co and Ni, while Fe and Pb even showed a slight increase. Also in


- 82 -

the Changjiang KD’s of Cd, Cu and Ni did not change with increasing SPM

concentration between 40 and 350 mg/L, while the KD’s of Fe, Mn and Al showed

drastic declines as a consequence of their high ratio of particulate metal to dissolved

metal (Koshikawa et al., 2007). A stable concentration of particulate Cu, Ni and Pb in

the Changjiang relative to the SPM, which varied over several orders of magnitude,

was reported earlier by Zhang (1999).

4.3.5. Comparison of WR and WWR with other Chinese estuaries

This study asserted that all dissolved and particulate metals were similar or

somewhat lower (Co, Ni) in the WR with its low industrial and agricultural run-off than

in the WWR estuary, which is probably affected by aquacultural contamination plus

some municipal sewage along the river (Table1, Table 2). This is partly a

consequence of the exceptional characteristics during the wet season of 2008 in the

WWR, when not only DOC, but also dissolved Cu, Fe, Pb and Co at low salinities as

well as the particulate Cu, Co and Ni were pronouncedly higher than during the other

two expeditions. As also reported for dissolved arsenic (Balzer et al., 2013), this

points to extraordinary lateral inputs probably originating from adjacent aquaculture

ponds as shown for nutrients, dissolved organic matter and suspended particles by

Herbeck et al., 2013.

A comparison of the particulate metal concentrations with large Chinese rivers

demonstrates that the two Hainan estuaries contained similar levels of Cd, Cu, Pb,

Co and Ni (Table 3). The lower concentrations especially in the Huanghe and the

Changjiang probably originate from lower trace metal contents in soils of the northern

drainage areas (Zhang 1995). It is noteworthy that the particulate and dissolved Fe

concentrations in both study areas mostly exceeded those of the other estuaries

listed in Table 4. This is probably related to the generally high Fe content in the soils

of Hainan Island, although a small contamination - via atmospheric transport - by the

Shilu Iron Mine cannot be excluded,

The dissolved concentrations of the other trace metals in the freshwater part of the

WR and the WWR were similar (Cd, Ni, Cu) or higher (Fe, Pb, Co) than in the large

Chinese rivers (Table 4). The most striking difference between the Hainan estuaries

and the large rivers of mainland China is the orders of magnitude higher SPM loading

(Zhang, 1995; Zhang and Liu, 2002). Higher SPM, however, possibly fostered by

longer residence times only serve for a faster attainment of sorption equilibria, but

gives no general direction for concentration changes in the solution phase.


- 83 -

Table 4. Dissolved trace metals of East-Hainan estuaries (average values ± 1 standard deviation) in comparison to selected Chinese rivers (S = 0) and estuaries (S > 0) (averages or typical ranges; *: estimated from the figures).

Study area SPM Cd Cu Fe Pb Co Ni

Source mg/L nmol/L nmol/L μmol/L nmol/L nmol/L nmol/L

East Hainan

S=0 5.3±4.4 0.03±0.03 10±6.8 4.1±3.6 0.88±0.65 2.2±1.6 5.9±4.0

a 0<S<28 3.0±3.3 0.07±0.03 9.6±5.7 2.2±2.7 0.59±0.49 2.7±3.0 7.7±3.2

S>28 6.1±4.6 0.11±0.03 6.4±3.4 0.14±0.13 0.10±0.08 0.75±0.40 4.9±1.3

Huanghe (S=0) 25600 0.02 22 0.35 0.14 0.21 6.9 b

Huanghe (S=0) 27000 0.03 20 0.24 0.13 0.30 7.5 c

Changjiang (S=0) 20-5000 0.03 26 0.55 0.26 -- 2.5 d

Changjiang estuary(S>0) 70 0.02-0.36 19-26 0.1-0.6 2.3-3.4 1.4-2.9 18-28 e

Minjiang (S=0) 128 0.04 14 -- 0.63 -- -- c

Jiulongjiang (S=0) 211 0.03 13 16 0.07 0.14 1.4 c

Tanshui River estuary (S>0) 15* -- 5-53 0.39-3.4 -- 0.3-6.1 7-310 f

Tanshui River estuary (S>0) 10* 0.20 60 -- 0.88 -- 138.0 g

Open South China Sea -- 0.01 0.8 0.19 -- -- 2.3 h

(a): This study; (b): Zhang et al., 1994; (c): Zhang, 1995; (d): Edmond et al., 1985; (e): Wang and Liu, 2003; (f): Fang and Lin, 2002; (g): Jiann et al., 2005; (h): Wen, et al., 2006.

With respect to the estuarine chemistry of the Chinese aquatic systems, detailed

information is only available for the Changjiang with its rather stable estuarine trace

metal distribution (Zhang, 1999; Wang and Liu, 2003) and about the subtropical

Tanshui (Fang and Lin, 2002; Jiann et al., 2005). Compared to these estuarine

systems, the tropical East-Hainan estuaries (S > 0) generally contained similar or

lower trace metal concentrations in the suspended particles (Table 3), and also

similar or lower concentrations for Cd, Cu, Pb and Ni, but higher levels of Fe in the

solution phase (Table 4). However, the concentration differences, the number of

regions for comparison and the amount of necessary subsidiary data are not large

enough to derive specific features for the estuarine chemistry of tropical East-Hainan.

The information compiled in Table 3 and Table 4 demonstrate, that on a global scale

the Chinese aquatic systems including the two Hainan estuaries have lower levels of

dissolved and particulate metals than the estuaries of countries with a longer history

of industrialization and with stronger signals of anthropogenic activities (Balls, 1989;

Zhang, 1995). The concentrations of the investigated trace metals in solution and in


- 84 -

the particles suggest a low level of contamination especially in these two Hainan

estuaries.

4.4. Conclusions

This study provides data for dissolved and particulate cadmium, copper, iron, lead,

cobalt and nickel concentrations plus particulate aluminium along the salinity

gradients of two estuaries of the tropical East-Hainan Island during both the wet and

the dry season. The estuarine chemistry of the Wanquan River (WR) with its low

industrial and agricultural run-off and of the Wenchang/Wenjiao River (WWR), which

probably is affected by aquacultural contamination plus some municipal sewage, is

similar but not identical.

It was found that the dissolved trace metals mostly behaved conservatively. A positive

deviation from the theoretical dilution line was observed for dissolved cadmium, while

the converse situation applied for iron and lead. Seasonal variability was observed

during the dry season for dissolved iron and lead with a strong lowering and for

dissolved Cd with higher values. In both estuaries all six metals in particulate form

showed almost constant values with a tendency for slight decreases along the

salinity profile. The normalization to particulate Al, which itself decreased with salinity,

revealed specific particle properties during the different expeditions, e.g., Cu-rich

particles during the wet season 2008 in WWR, which coincided with very high DOC

levels and generally low percentages of particulate metals on the total metal

concentration (i.e., sum of dissolved and particulate) per unit volume. In general, the

dynamics of Fe chemistry dominated the distribution of Pb in all forms and partly also

of Co as revealed by correlations in the dissolved state and in their ratio of particulate

metal to the total metal per unit volume.

The distribution coefficients of the trace metals KD = CP / CS decreased in the order

Fe > Pb > Co > Ni > Cu ≈ Cd. Except for the expected KD decrease of Cd with

increasing salinity, the KD´s of Cu, Co and Ni were constant, and Fe and Pb exhibited

slightly positive correlations with salinity. There was no "particle concentration effect",

i.e., a negative dependence of the KD on the SPM level.

The most striking difference between the large Chinese rivers especially from

temperate latitudes and the two Hainan water systems is the extremely low SPM in

the latter ones. Together with its low levels of particulate trace metals, this entails the

possibility that equilibrium desorption might be so small in absolute amounts that the

transfer is hardly detectable in the solution phase.


- 85 -

One of the aims of this study was to use the distributions of environmentally relevant

trace metals during estuarine mixing for assessing the pollution state of the East-

Hainan estuaries. In most cases, both particulate and dissolved metal concentrations

in the Wenchang/Wenjiao River estuary were higher than in the Wanquan River

estuary, which is probably due to diffuse inputs from aquaculture ponds to the WWR

system. The trace metal levels of these East-Hainan aquatic systems were similar to

large rivers in mainland China and showed no significant contamination when

compared to rivers and estuaries in Europe and North America. This is especially

important with respect to the recently observed health decline of the coral reefs at the

east coast of Hainan.

Acknowledgements

The authors wish to thank Dr. Dao-Ru Wang and Dr. Larissa Dsikowitzky for their

coordinating and managing efforts during the study. Appreciation is extended to the

colleagues from the Centre for Tropical Marine Ecology (ZMT) of Germany, the

Hainan Provincial Marine Development and Design Institute (HNMDDI) of China, the

East China Normal University (ECNU) and the Ocean University of China (OUC) for

their help in the field and during laboratory works. We thank Dr. Uwe Schüßler for his

help with the ICP-MS analyses. We are especially thankful for the constructive

criticism of two anonymous reviewers. We gratefully acknowledge the funding by the

"Bundesministerium für Bildung und Forschung" (BMBF, German Federal Ministry of

Education and Research) at Berlin under the contract number 03F0457 C and the

financial support from the Ministry of Science and Technology (MOST) through the

contract No. 2007DFB20380 and under the agreement between the State Oceanic

Administration of China (SOA) and the BMBF.


- 86 -

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Kapitel 5: Entwurf zur Publikation 2

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5. Entwurf zur Publikation 2:

Transport of trace metals in colloidal form through estuarine systems of East-Hainan

Abstract

Cross flow ultrafiltration in conjunction with high resolution inductively coupled

plasma mass spectrometry was used to estimate the colloidal distribution of iron,

lead, cadmium, copper, cobalt and nickel in two estuarine systems of East Hainan.

Metal concentrations along the salinity gradient were determined in three operatically

defined size fractions: 0.45 μm - 10 kDa (large colloids), 5 kDa - 10 kDa (small

colloids) and < 5 kDa (assumed to be in true solution) in order to investigate the

changes of trace metals. Dissolved iron and lead were found to exist mainly as large

colloids, while dissolved cadmium, copper and nickel were present in low molecular

size fractions. Both the larger and the smaller size fractions were important for

dissolved cobalt. Modifications in size distribution during the estuarine mixing are

observed for all examined metals with the involvement of estuarine processes such

as coagulation, degradation and dissociation. For all investigated metals the truly

dissolved fraction becomes more important with increasing salinity. Furthermore, the

difference of colloidal size fraction between the two estuarine systems was discussed.

Keywords: Trace metals; Estuaries; Colloids; East Hainan


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5.1. Introduction

An estuarine environment may be simply defined as one in which sea water is

substantially diluted with fresh water entering from land drainage. These two types of

water have different compositions, and as a result estuaries are very complex

environments in which the boundary conditions are extremely variable in both space

and time (Chester, 1990). It is now widely accepted that colloids are important in

controlling the cycling of trace metals that may either be scavenged e.g., onto Fe-

oxyhydroxides or complexed with large aggregates such as organic matter in the

estuarine environment (Powell et al., 1996), which can be scavenged in the estuary.

As a consequence, such colloidally mediated scavenging results in removal of both

bulk matter and associated trace components (Gustafsson et al., 1996).

The general pattern of colloidal partitioning of trace metals in the water column

depends primarily on the metal (Benoit et al., 1994; Farag, et al., 2007). Various size

fractionation studies indicate different proportions of colloidal forms for different trace

metals. For example, particle reactive elements such as Fe and Pb probably have

very small truly dissolved fractions (Ren et al., 2010). In contradistinction, cadmium

should have larger fractions in the truly dissolved pool via formation of strong chloride

complexes (Wen et al., 1999). The complexation of cobalt with dissolved organic

carbon was reported (Collins and Kinsela, 2010). Competition between different size

ligands may control the colloidal distribution of copper and nickel, which form strong

organic complexes (Wang et al., 2003).

Little is known about the trace metals in colloidal form in the estuarine systems of

East Hainan. The aims of this study were therefore to investigate the size distribution

of colloidal trace metals as well as their behavior in the Wanquan River and

Wenchang/Wenjiao River during estuarine mixing.

Although dissolved, colloidal and particulate phases are purely operational and are

not necessarily related to real differences in structure or to environmental or chemical

behavior (Lead and Wilkinson, 2007), present knowledge of natural water colloids is

largely based on the physical separation techniques, of which cross flow filtration

(tangential flow filtration; CFF) is currently widely used to provide concentrates of

colloidal material for further analyses. This approach has made it possible to

separate colloidal size material from true solutes (Buesseler et al., 1996; Guo et al.,

2000a). An exemplary representation of the CFF apparatus which was also used

during this study is shown in Fig. 1. The colloid concentration is given by: CCol = (CR-

CP)/CF, where CR is the retentate concentration, CP is the permeate concentration


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and CF is the concentration factor. The CF is defined as: CF = (VR+VP)/ VR, where VR

is the volume of retentate and VP is the volume of permeate. A high CF can minimize

the retention at the membrane surface (Guo and Santschi, 1996), but a breakthrough

becomes more significant (Dai et al., 1998). On the other side, a low CF reduces the

coagulation of colloids at the membrane surface (Waeles et al., 2008), but the

recovery will be more sensitive for the system performance (Whitehouse et al., 1990).

There is still no general consensus as to what the optimal CF should be for

ultrafiltration (e.g., Guo et al., 2000b; Kottelat et al., 2008; Batchelli et al., 2009).

Fig. 1. Cross-flow filtration apparatus (source: www.sartorius-stedim.com).

Colloid size is defined in a range of 1 nm to 1 μm (Buffle and Leppard, 1995) with 1

nm ≈ 1000 Daltons (1 KDa; e.g., Wen et al., 1999). But the exact molecular weight or

size cut-off of the ultrafilter is not very sharp (Buesseler et al., 1996). To clarify the

notation for the present work, the size fraction between 10 kDa and 0.45 μm is

operationally defined as the high molecular weight colloid fraction (HMW), between 5

kDa and 10 kDa as the low molecular weight colloid fraction (LMW), and the

permeate from the ultrafiltration (< 5 kDa) as the truly dissolved fraction (UD).


5.2.1. Study area

The Wanquan River and the Wenchang/Wenjiao River estuaries (Fig. 2) are

important estuaries of East-Hainan which are both located at the northern part of the

tropical zone with a humid warm climate (Zhu et al., 2005). About 80% of the annual

rainfall in this region occurs in the period from May to November (Ma et al., 2007).

The Wanquan River (in the following: WR) is the third largest river in Hainan Island


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with a drainage area of 3693 km2, a total length of 156.6 km and a mean annual

discharge of 163.9 m3/s (Ge, et al., 2003). The Wenchang River (drainage area 381

km2, total length 37 km, mean annual discharge 9.1 m3/s) and the Wenjiao River

(drainage area 522 km2, total length 56 km, mean annual discharge 11.6 m3/s) empty

into the Bamen Bay (Zeng and Zeng, 1989) and form the Wenchang/Wenjiao River

estuary (in the following: WWR). Both estuaries have a micro-tidal, irregular diurnal

tidal regime with a mean range of about 0.7-0.8 m (Zhu et al., 2005). The salinity

intrusion may extend 5 kilometers from the Yudai Sand Barrier into the WR and from

the Bamen Bay end into the WWR, respectively. The WWR is a shallow lagoon

region (< 2 m; except for the ship channel) with a low flow rate, as compared with the

WR, in which the position of the mix zone is highly variable due to changes of river

discharge rate, tides and wind.

Fig. 2. Wenchang/Wenjiao River estuary (a) and Wanquan River estuary (b).

These two estuarine systems of East Hainan are distinguished by their extremely low

suspended particulate matter and their low levels of particulate trace metals

compared to large Chinese rivers (< 15 mg/L; Fu et al., 2013). Also dissolved trace

metal levels of these two aquatic systems are comparable to the rivers and estuaries

in Europe and North America (< 15 μmol/L for Fe, < 2.5 nmol/L for Pb, < 0.15 nmol/L

for Cd, < 15 nmol/L for Cu except 2 abnormalities, < 9 nmol/L for Co except 1

abnormality and < 20 nmol/L for Ni; Fu et al., 2013). But the nutrient level in these

two estuaries, especially in the WWR where is dominated by aquaculture

contamination, is higher than the average global level (Herbeck et al., 2013; Li et al.,

2013). Also dissolved organic carbon (DOC) and dissolved trace metals in the WWR


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are higher than in the WR (Fu et al., 2013).

5.2.2. Sampling and filtration

Surface water samples in the WR and the WWR were collected in July/August 2008.

Low-density polyethylene (LDPE) bottles (1 L, Nalgene®) were filled by hand with

surface water from a depth of 0.3 m upstream of the ship. In the laboratory, the water

samples were drawn through acid-cleaned 0.45 μm polycarbonate filters (pre-

weighed) within 24 hours before they were further separated into 30 kDa - 0.45 μm,

10-30 kDa, 5-10 kDa and < 5 kDa fractions using CFF with a flow rate of 500 mL/min

(2 bar). The cascade way was carried out to separate the colloid size fractions. This

method, in contrast to the parallel one, minimizes aggregation of colloids at the

membrane surface (Waeles et al., 2008). About 1 liter of the 0.45 μm filtrates were

first ultrafiltrated through a 30 kDa polyethersulfone membrane using a cross flow

filtration device (Vivaflow® 200, Sartorius; with an active membrane area of 200 cm2).

The permeate was used for the second ultrafiltration step employing a 10 kDa

polyethersulfone membrane. Eventually, the permeate from the second ultrafiltration

step was ultrafiltrated using a 5 kDa polyethersulfone membrane. A schematic

description of the filtration/ultrafiltration procedure is shown in Fig. 3. The average

CF’s ranged from 4.5 to 7.5. The filtrates, retentates and permeates were weighted

before the next step began. After 10 mL of each retentate/permeate was filled into a

pre-combusted glass ampoule with addition of phosphoric acid for dissolved organic

carbon measurements (Balzer et al., 2012), the sub-samples were filled into LDPE

bottles (Nalgene®) and acidified to pH 2 with concentrated nitric acid. The sample

bottles were wrapped in polyethylene bags and stored at room temperature. Before

use the CFF devices were washed with HCl (0.1 %, 30 min) and EDTA (0.02 mol/L,

30 min) and subsequently rinsed with Milli-Q® water. The environmental parameters

such as salinity and pH value were measured on the spot using a multi-sensor-

device (MultiLine P3, WTW).


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Fig. 3. Flow diagram of the sample filtration procedure (SPM: suspended particular matter; Perm: permeat; Ret: retentat).

5.2.3. Pretreatment and measurement

Sample work-up was performed in a clean-air room (class 1000) with additional clean

benches of class 100. The sub-samples were first irradiated in a UV-digester for 2

hours with addition of hydrogen peroxide (1 mL per 100 mL sample) in order to

destroy the organic complexes, before the trace metals were extracted. The

extraction/backextraction procedure was developed by our workgroup based on the

work of e.g., Danielsson et al. (1978). At first, 40-100 g sample (depending on the

metal concentration) was weighed into a fluorinated ethylene propylene (FEP)

separatory funnel (Nalgene®). Ammonium acetate buffer (pH ~ 9) was added to

adjust the pH value of the sample to 4.0-4.5. The sample was well shaken (15 min)

after 500 μL APDC/NaDETC solution (0.06 mol/L) and 10 mL Freon were added.

Following the phase separation the lower organic phase was drawn off into an acid-

cleaned polypropylene tube. After addition of 100 μL concentrated nitric acid the tube

was shaken by hand for 3 min. Finally, 1900 μL Milli-Q® water was added and the

tube was again shaken by hand for 2 min for back-extraction. After the phase


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separation, the upper aqueous extract was removed into an acid-cleaned sample

tube using a pipette. The metal concentrations were determined by high resolution

inductively coupled plasma mass spectrometry (HR-ICP-MS; Element 2, Thermo

Scientific).

5.2.4. CFF blanks and mass balance

Fe: y = 0.6422xR2 = 0.92

0

4

8

12

0 4 8 12TD

HM

W+L

MW

+UD

WRWWR

Pb: y = 0.5845xR2 = 0.8629

0.0

0.5

1.0

1.5

2.0

0.0 0.5 1.0 1.5 2.0TD

Cd: y = 0.6506xR2 = 0.678

0.00

0.04

0.08

0.12

0.00 0.04 0.08 0.12TD

HM

W+L

MW

+UD

Cu: y = 0.7884xR2 = 0.9025

0

5

10

15

0 5 10 15TD

Co: y = 0.8801xR2 = 0.944

0

1

2

3

4

5

0 1 2 3 4 5TD

HM

W+L

MW

+UD

Ni: y = 0.7924xR2 = 0.8184

0

5

10

15

0 5 10 15TD

Fig. 4. Colloidal and truly dissolved concentration vs. total dissolved concentration (nmol/L; for Fe μmol/L).

To produce CFF blanks, new CFF devices were pre-washed with HCl (0.1 %, 30 min)

and EDTA (0.02 mol/L, 30 min) and subsequently rinsed with Milli-Q® water. Milli-Q®

water was ultrafiltrated into 4 colloidal fractions (30 kDa - 0.45 μm; 10-30 kDa; 5-10

kDa and < 5 kDa). All sub-samples were analyzed using ICP-MS for trace metal

concentrations. Since elevated Cu concentration occurred, the devices were pre-


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washed once again with HCl and EDTA and additional Milli-Q® samples were

ultrafiltrated and analyzed. All trace metal concentrations of the blanks were less than

1 % of the sample concentrations.

Since there is no standard reference material to quantify the accuracy of the

ultrafiltration analysis, a concentration mass balance between the sum of all size

fractions (HMW + LMW + UD) and the 0.45 μm filtrated sample (TD) was calculated.

For most samples, the mass balance was < 100% indicating a loss of material

probably due to scavenging onto the filter membrane or sample reservoir (Martin et

al., 1995; Powell et al., 1996). Also the propagation of sampling and analytical errors

of the sub-samples, of the UD samples in particular, may cause a deviation of the

mass balance from 100%. Especially for lower concentrations, a small deviation may

create a remarkable aberrance in the mass balance. The available data sets were

therefore operationally defined with a mass balance within a range of 75% ± 30%,

55% ± 20%, 65% ± 20%, 80% ± 20%, 100% ± 30% and 80% ± 20% for Fe, Pb, Cd,

Cu, Co and Ni, respectively. In total 20 analyzed samples, 12 for Fe, 11 for Pb, 15 for

Cd, 18 for Cu, 17 for Co and 12 for Ni have been selected, whose mass balances

were in the respective defined ranges. The calculated coefficients of determination

showed generally good correlation between the measured and the calculated total

dissolved concentrations (Fig. 4). Theoretically, the curves must have a slope of 1,

which indicates a mass balance of 100%. But indeed, all of the slopes resulted

smaller, which could be attributed to the sorption to the membrane as discussed.

5.3. Results and discussion

5.3.1. Iron

The dissolved Fe in the WR and the WWR was found mainly in the HMW fraction

with over 95% on average of the calculated TD in both study areas (Fig. 5). The LMW

and UD fractions with less than 5% in total of the calculated TD appear not to be

relevant in both study areas. The affinity of Fe to large colloids in the fresh water is

well known (Pham and Garnier, 1998; Ren et al., 2010). Also the loss of the colloidal

Fe through flocculation during the estuarine mixing was widely reported (e.g.,

Kraepiel et al., 1997; Nowostawska et al., 2008). This phenomenon bases on the fact,

that riverine colloidal Fe-oxyhydroxides closely associated with organic matter

aggregate to form larger particles due to interactions between the negatively charged

particles and Mg2+ and Ca2+ ions introduced by the seawater end member (e.g.,

Boyle et al., 1977; Nowostawska et al., 2008). As a consequence, the Fe ratios of the


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LMW and the UD fraction increased with increasing salinity in both study areas. Iron

is also able to form stable organic complexes (Powell et al., 1996). Since the

solubility of inorganic iron in seawater is extremely low (e.g., Balzer, 1982), there

must be a significant part of dissolved Fe, which is associated with organic

compounds (Liu and Millero, 1999). Humic material in particular degrades only slowly,

and can protect Fe against coagulation by expanding the effective distance between

approaching colloids as well as by altering their net surface electrostatic charge

(Batchelli et al., 2010). For instance, humic ligands were identified to be a stabilizer

for the transport of riverine Fe to the ocean (Laglera and van den Berg, 2009). The

Fe-organic complexation may also be inferred by the correlation between DOC and

Fe concentrations in the HMW fraction (Fig. 6). The higher DOC content and

probably higher concentration of Fe-humic-substance in the WWR enhanced the

solubility of Fe and caused the difference in the slope between the WR and the WWR

in Fig. 6.

WR

0%

25%

50%

75%

100%

0 0 5 8 11 17 23 34Salinity

Fe ra

tio

WWR

0%

25%

50%

75%

100%

0 0 3 6 8 10 11 13 16 18 20 28Salinity

HMWLMWUD

0%

25%

50%

75%

100%

0 0 5 8 11 17 23 34Salinity

Pb r

atio

0%

25%

50%

75%

100%

0 0 3 6 8 10 11 13 16 18 20 28Salinity

Fig. 5. Trace metal ratios of various size fractions vs. salinity (left side: WR; right side: WWR).


- 100 -

0%

25%

50%

75%

100%

0 0 5 8 11 17 23 34Salinity

Cd

ratio

0%

25%

50%

75%

100%

0 0 3 6 8 10 11 13 16 18 20 28Salinity

0%

25%

50%

75%

100%

0 0 5 8 11 17 23 34Salinity

Cu

ratio

0%

25%

50%

75%

100%

0 0 3 6 8 10 11 13 16 18 20 28Salinity

0%

25%

50%

75%

100%

0 0 5 8 11 17 23 34Salinity

Co

ratio

0%

25%

50%

75%

100%

0 0 3 6 8 10 11 13 16 18 20 28Salinity

0%

25%

50%

75%

100%

0 0 5 8 11 17 23 34Salinity

Ni r

atio

0%

25%

50%

75%

100%

0 0 3 6 8 10 11 13 16 18 20 28Salinity

Fig. 5. (continued).


- 101 -

0

2

4

6

8

0 1 2 3DOC (ppm)

Fe (μ

mol

/L)

WR HMWWWR HMW

Fig. 6. Fe concentration vs. DOC in the HMW fraction.

5.3.2. Lead

The colloidal size distribution of dissolved Pb was comparable with dissolved Fe in

both study areas with Pb in the HMW fraction as the most important fraction (92%

and 95% on average in both estuaries, respectively). The ratios of the other size

fractions were less than 10% overall in both study areas and increased with

increasing salinity (Fig. 5). Dissolved Pb was widely found to associate with macro-

size Fe-rich colloids e.g., in the estuaries in Texas, USA (Benoit et al., 1994), in the

Penzé estuary (Waeles et al., 2008) and in the Mississippi Delta (Stolpe et al., 2010).

This is confirmed by the good correlations between HMW-Fe and HMW-Pb (Fig. 7).

Although labor experiment pointed out that Pb was taken up by both organic and Fe

colloids (Lyvén et al., 2003), the co-precipitation of Pb with Fe colloids appears to be

a typical characteristic in the estuarine chemistry of East Hainan (Fu et al., 2013).

0.0

0.5

1.0

1.5

0 2 4 6 8Fe (μmol/L)

Pb (n

mol

/L)

WR HMWWWR HMW

Fig. 7. Pb vs. Fe concentrations in the HMW fraction.

5.3.3. Cadmium

Fig. 5 shows the increase of both the concentration and the ratio of Cd in the UD

fraction with increasing salinity probably as a consequence of chloride induced


- 102 -

mobilization of particulate Cd (e.g., Comans and van Dijk, 1988). The Cd ratio in the

HMW and the LMW fraction decreases with increasing salinity, whereas their ratio

appears to be steady, perhaps with a small reduction of Cd in the HMW fraction at

low salinity due to adsorption of dissolved Cd on Mn oxide (Turner et al., 2008). It is

not surprising that Cd in the UD fraction was the greatest fraction with around 75% in

both estuaries due to formation of highly stable Cd-chloro-complexes (e.g., Van der

Weijden, 1977; Salomons, 1980; Comans and van Dijk, 1988). Such high

contribution of the Cd in the UD fraction was generally observed e.g., in the Rhône

delta (more than 90%; Dai et al., 1995), in the Venice Lagoon (65%; Martin et al.,

1995), in the Danube River estuary (60-70%; Guieu et al., 1998) and in the Seine

estuary (90-95%; Chiffoleau et al., 2001). Since the Cd ratio in the UD fraction in the

WR (81%) was higher than in the WWR (72%), the difference of the Cd ratio in the

UD fraction indicates a more complicated background than simple formation of highly

stable Cd-chloro-complexes. Organic complexes can also exert a significant control

of the Cd speciation in certain systems (Waeles et al., 2008). For instance, Cd in the

Penzé estuary was highlighted in winter and autumn to be constituted over 50% as

organic complexes due to its release from particle or sediment subsequent to

degradation of high molecular organic Cd (Waeles et al., 2005). Such degradation

could be inhibited by the relatively higher DOC content in the WWR (Fu et al., 2013)

and caused the difference in Cd speciation between the two study areas. On the

other hand, it also could be an evidence for the formation of organic Cd species

which aggregated in the HMW fraction. In conclusion, higher Cd ratio in the UD

fraction probably presents the dominance of Cd-chloro complexes. In systems with

relatively high colloidal Cd, both Cd-chloro and Cd-organic species may influence the

Cd behavior.

5.3.4. Copper

The Cu concentration in the UD fraction, with an average percentage of 40% and

52%, was the most significant size fraction, followed by the LMW fraction with 38%

and 25% and the HMW fraction with about 14% and 18% in the WR and the WWR,

respectively. Since dissolved Cu behaves generally conservatively (Fu et al., 2013), a

colloidal analysis suggested a conversion from the HMW fraction to the UD fraction

at low salinities with stronger tendency in the WR than in the WWR (Fig. 5). Copper

in the LMW fraction had a constant percentage along the whole salinity gradient. It is

believed that the competition between high and low molecular weight ligands may

control the Cu size distribution (Wen et al., 1999). For example, dissolved Cu was


- 103 -

mainly found in colloidal forms (5 kDa - 0.45 μm) which presented 94% of TD-Cu in

the Penzè estuary (Waeles et al., 2008). But the dominance of UD-Cu (< 10 kDa)

with more than 50% was observed overall in the Mississippi estuary (Wen et al.,

2011). These differences should be related to the nature of organic matter within the

systems of concern (Waeles et al., 2008). In this case, over 80% of dissolved Cu was

identified as low molecular weight ligands (< 10 kDa) in the saline water. In the

freshwater, however, most Cu occurred in the colloidal fraction (HMW and LMW),

especially in the WR with almost 100% of colloidal Cu. In association with the organic

nature of Cu in estuaries (e.g., Shank et al., 2004), the degradation of macro

molecular organic matter might cause such a drift of Cu from the HMW to the UD

fraction during the estuarine mixing (Masson et al., 2011). Such degradation was less

distinctive in the WWR probably because of the higher DOC content as discussed

before.

5.3.5. Cobalt

The investigation in the Galveston Bay demonstrated a permanently strong UD

fraction of dissolved Co (about 80%) with the argument of metabolical interaction with

Zn (Wen et al., 1999), while over 50% of dissolved Co was reported to be controlled

by Fe colloids (HMW) in Northwest Russia (Pokrovsky and Schott, 2002). In this

study, the HMW fraction as the largest size fraction of dissolved Co decreased

seawards from 75-80% to typically 35% in the seawater in both study areas (Fig. 11).

Conversely, the content of Co in the UD fraction increased from 10% to 30-40% with

increasing salinity. The Co in the LMW fraction with typically 20% varied

insignificantly with salinity. Previous study suggested a co-precipitation with Fe

colloids (Sharp et al., 1982; Fu et al., 2013), which could be confirmed by the good

correlation between Fe and Co in the HMW fraction (Fig. 8). Furthermore, previous

study suggested that lateral inputs from aquaculture ponds could influence the

behavior of dissolved Co and the elevated DOC content in the WWR (Fu et al., 2013).

Cobalt is also well known to be able to form strong organic complexes in nature

waters (e.g., Zhang et al., 1990; Qian et al., 1998; Saito and Moffett, 2001). The

correlation between organic matter and Co in the UD fraction (Fig. 8) might indicate

such an association of Co in the UD fraction with low molecular organic matter.


- 104 -

0

1

2

3

0 2 4 6 8Fe (μmol/L)

Co

(nm

ol/L

)

WR HMWWWR HMW

0

1

2

3

0 1 2 3 4DOC (ppm)

Co

(nm

ol/L

)

WR UDWWR UD

Fig. 8. Co vs. Fe concentrations in the HMW fraction and Co vs. DOC in the UD fraction.

5.3.6. Nickel

The UD fraction of Ni was the dominant size fraction in the both estuaries, (Fig. 5)

which was also observed e.g., in the Venice Lagoon (Martin et al., 1995), in the San

Francisco Bay (Sañudo-Wilhelmy et al., 1996) and in the Ochlockonee estuary

(Powell et al., 1996). At the very beginning of the estuarine mixing, the LMW-Ni and

the HMW-Ni behaved conservatively, which indicates stable Ni complexes in both

size fractions, while UD-Ni shows fluctuations compared to the theoretical dilution line.

Biological regeneration processes may affect the colloidal distribution of Ni and may

cause such an excess (Dai and Martin, 1995). But no relationship between Ni and

DOC was observed in this study. It is more likely that riverine particle-bound Ni was

replaced by Ca2+ and Mg2+ during the estuarine mixing (e.g., Powell et al., 1996;

Mandal et al., 2002). However, Ni is largely unreactive in estuaries. Both small

suspended particular matter fluxes in the estuary and the relatively low particle

affinity of Ni which is due to its strong association with dissolved organic matter can

inhibit the interactions between sediment, suspended particular matter, the colloid

and the UD fractions (Turner et al., 1998). All size fractions behaved therefore

conservatively at higher salinities in both study areas. However, the more significant

HMW fraction of dissolved Ni in the WWR than in the WR probably indicated stronger

involvement of organic matter in the colloidal distribution of dissolved Ni in the WWR

due to its higher DOC content.

5.3.7. Spatial variability of colloidal trace metals in East-Hainan

As mentioned before, for all studied dissolved trace metals the ratio of their HMW

fraction decreased with increasing salinity, whereas the ratio of their UD fraction

became more important, especially at low salinity (Fig. 5). Such variability was widely

observed e.g., in the Ochlockonee estuary (Powell et al., 1996), in the Narragansett


- 105 -

Bay (Wells et al., 1998) and during laboratory experiments (Stolpe and Hassellöv,

2007). Since a negative deviation from the theoretical dilution line was observed for

dissolved Fe in the WR and the WWR (Fu et al., 2013), the more importance of the

UD fraction with increasing salinity could be related to the loss of large size colloidal

Fe due to its flocculation during the estuarine mixing (e.g., Kraepiel et al., 1997).

Similarly, a removal of large size colloidal Pb and Co with freshly precipitated Fe-

oxyhydroxides (e.g., Sharp et al., 1982; Balls, 1990) was suspected to be responsible

for the increasing significance of Pb and Co in the UD fraction. Conversely, the

dissolved Cd showed a positive deviation from the theoretical dilution line in both

estuaries (Fu et al., 2013). The mobilization of Cd due to formation of highly stable

Cd-chloro-complexes (e.g., Comans and van Dijk, 1988) enhanced the Cd content in

the UD fraction. Despite the conservative behavior of dissolved Cu and Ni in both

estuaries (Fu et al., 2013), a conversion from the HMW to the UD fraction appeared

to occur with stronger tendency for Cu than for Ni. In accordance with the affinity of

Cu to organic material (e.g., Wang et al., 2003; Shank et al., 2004) such conversion

could be attributed to the degradation of macro molecular organic matter (Masson et

al., 2011). During the estuarine mixing Ni in the HMW fraction was replaced by Ca2+

and Mg2+ although such modification was not pronounced (e.g., Powell et al., 1996;

Mandal et al., 2002).

5.4. Conclusions

In this study, a cross flow ultrafiltration technique was used to separate various

colloidal fractions of trace metals in estuarine water. The colloidal distribution and its

modification along the salinity gradient were investigated in the Wanquan River and

Wenchang/Wenjiao River estuaries of East Hainan. Iron and lead were mainly

associated with high molecular weight colloids in the size range 10 kDa - 0.45 μm (>

90%), whereas cadmium (76%) and nickel (65%) appeared in the truly dissolved

fractions (< 5 kDa). Copper and cobalt favor in the high molecular weight colloids at

low salinities and are preferentially tied to the truly dissolved fractions at high

salinities. Some processes affecting trace metals during their transport from river to

ocean could be identified. The removal of Fe and Pb is controlled by flocculation of

Fe-oxyhydroxides. Also cobalt might be scavenged during co-precipitation with Fe

colloids at low salinity. The formation of cadmium chloride complexes controlled the

estuarine distribution of dissolved cadmium. The competition between nickel and

calcium/magnesium for the binding site at the surface of macro size material

influenced colloidal distribution of dissolved nickel during the estuarine mixing. The


- 106 -

degradation of organic copper was assumed to be responsible for the modification of

dissolved copper distribution. For all of the investigated metals the truly dissolved

fraction becomes more important with increasing salinity whereas the ratio of high

molecular weight fraction decreased. For iron, cadmium, copper and nickel the high

molecular weight fraction was more present in the WWR than in the WR.

Acknowledgements

The authors are thankful to Dr. Dao-Ru Wang and Dr. Larissa Dsikowitzky for their

coordinating and managing efforts during the study. Appreciation is extended to the

colleagues from the Centre for Tropical Marine Ecology (ZMT) of Germany, the

Hainan Provincial Marine Development and Design Institute (HNMDDI) of China, the

East China Normal University (ECNU) and the Ocean University of China (OUC) for

their help in the field and during laboratory works. Our special thanks will be given to

Dr. Uwe Schüßler for his help with the HR-ICP-MS analyses. We gratefully

acknowledge the funding by the "Bundesministerium für Bildung und Forschung"

(BMBF, German Federal Ministry of Education and Research) at Berlin under the

contract number 03F0457 C and the financial support from the Ministry of Science

and Technology (MOST) through the contract No. 2007DFB20380 and under the

agreement between the State Oceanic Administration of China (SOA) and the BMBF.


- 107 -

6. Entwurf zur Publikation 3:

Cd, Co, Cu, Fe, Ni and Pb in solution, colloidal and particulate phases of Northeast Brazilian estuaries

Abstract

Dissolved and particulate cadmium, copper, iron, lead, cobalt and nickel were

analyzed in surface waters of the Mundaú-Manguaba lagoon system in Northeast

Brazil during the wet season (September/October 2007) and two dry seasons

(February 2008 and March 2009). During the wet season, dissolved Cadmium

showed precipitation in the Mundaú lagoon and conservative behavior in the

Manguaba lagoon, while enrichment in the solution phase seemingly occurred,

especially at high salinities, during both dry seasons. In general, the behavior of

dissolved copper and nickel could be identified as conservative. Dissolved iron

showed a negative deviation from the theoretical dilution line, whereas the opposite

was observed for dissolved cobalt. Dissolved Pb co-precipitated with Fe in the

Mundaú lagoon but was partially enriched in the Manguaba lagoon. Essentially,

particulate metals behaved nearly conservatively and showed no seasonal variability

except for particulate Fe and Pb, which are scavenged during estuarine mixing. The

distribution coefficients were calculated with an order Fe > Pb > Co > Ni > Cu > Cd.

The colloidal partition and distribution of trace metals was investigated during the wet

season of 2007 and the dry season of 2009. Most metals had a significant presence

in the colloidal fraction (5 kDa - 0.45 μm). The truly dissolved fraction (< 5 kDa) for all

the metals was more important during the dry season than during the wet season.

Keywords: Trace metals; Estuaries; Colloids; Northeast-Brazil


- 108 -

6.1. Introduction

A coastal lagoon is an inland water body which is separated from the ocean by a

barrier and connected to the ocean by one or more restricted inlets (Kjerfve, 1994).

The geomorphology often creates an estuarine system with low tidal influence.

Estuaries often suffer from strong anthropogenic impacts. Estuarine systems can

also be thought of as acting as filters of the river-transported chemical signals, which

often emerge from the mixing zone in a form that is considerably modified with

respect to that which entered the system (Chester, 1990). In view of the fact that

rivers are one of the most important pathways for the transport of anthropogenic

materials to the ocean, it is necessary to understand how the estuarine filter operates

and especially, which part is retained before reaching the ocean.

The estuarine filter is selective in the manner in which it acts on different elements,

and the effects of the filter can vary widely from one estuary to another (Chester,

1990). Especially trace metals are a particularly interesting aspect of estuarine

chemistry because their differing physical chemistries lead to a variety of

geochemical behaviors (Shiller and Boyle, 1991) which are strongly influenced by the

chemical form in which the elements occur (Sigleo and Helz, 1981). The behavior of

trace metals in estuaries differs from one to another. In a given region, estuaries may

have comparable environmental conditions. However, the trace metal distribution in

different estuaries may differ from each other (e.g., Benoit et al., 1994; Dai and

Martin, 1995; Fu et al., 2013). Even in a certain estuary, seasonal variability of trace

metal behavior is not infrequent (e.g., Zwolsman et al., 1997; Jiann et al., 2005;

Waeles et al., 2005). Many estuarine processes have been investigated in the last

decades: for instance, chloride-induced desorption of Cd is characteristic for most

estuaries (e.g., Elbaz-Poulichet et al., 1996; Waeles et al., 2004; Dabrin et al., 2009).

Copper behaves conservatively in many estuaries (e.g., Shiller and Boyle, 1991;

Guieu et al., 1998; Martino et al., 2002), but both positive (e.g., Santos-Echeandia et

al., 2008; Masson et al., 2011) and negative deviation from the theoretical dilution

line (e.g., Sholkovitz, 1978; Sharp et al., 1982) have also been noted. The rapid

flocculation of Fe and Pb are reported for most estuaries (e.g., Turner et al., 1992;

Wen et al., 1999; Wang and Liu, 2003), whereas Ni is relatively unreactive during the

estuarine mixing and behaves mostly conservatively (e.g., Shiller and Boyle, 1991;

Turner et al., 1998; Fernández et al., 2008). Manganese involved desorption of Co is

widely observed in estuaries (e.g., Chiffoleau et al., 1994; Wen et al., 1999; Takata et

al., 2010).

The trace metal chemistry has been studied in various Brazilian estuarine and lagoon


- 109 -

systems along the Atlantic coast, e.g., in the Amazon estuary (Boyle et al., 1982), in

the Guanabara Bay (e.g., De Luca Rebello et al., 1986), in the Sepetiba Bay (e.g., de

Lacerda et al., 1987), in the Patos lagoon (e.g., Windom et al., 1999) and in the

Paraíba do Sul estuary (e.g., Carvalho et al., 2002). But there are no publications

which investigate the behavior of trace metals in the Mundaú and Manguaba lagoons.

Thus, the first aim of this study is to determine the estuarine distribution of selected

metals along the salinity gradients in these two study areas.

Colloids are important ligands affecting the speciation, fate, transport,

biogeochemistry, bioavailability and toxicity of trace metals in aquatic systems

(Doucet et al., 2007). For many trace metals, the colloidal size fraction is

predominant in the estuarine waters. For instance, most of total dissolved Fe and Pb

was found to be associated with colloidal material, e.g., in the Venice Lagoon (Martin

et al., 1995), in the Penzé estuary (Waeles et al., 2008a) and in the Mississippi River

delta (Shim et al., 2012). Also Cd and Cu may be strongly associated with colloidal

materials (e.g., Martin et al., 1995; Waeles et al., 2005; Wen et al., 2011). It is

possible that colloidal trace metals also undergo estuarine modification and

consequently change their distribution and state. Hence, the second aim of this study

is to investigate the colloidal trace metal behavior in these two study areas. Because

of close links between dissolved trace metals and dissolved organic carbon (DOC),

especially in colloidal fractions, analyses of DOC (Balzer, unpublished) were

evaluated and used for interpretation of the trace metal results.

Colloidal material is defined to have at least one dimension between 1 nm and 1 μm

(e.g., Everett, 1988; Buffle and Leppard, 1995) with 1 nm ≈ 1000 Daltons (1 kDa; e.g.,

Wen et al. 1999). Colloids are mostly separated from larger fractions by using

ultrafiltration. However, the exact molecular weight or size cut-off of the ultra-filter is

not very sharp (Buesseler et al., 1996). For the sequential separation protocol of this

study, the size fraction between 10 kDa and 0.45 μm is operationally defined as high

molecular weight colloid fraction (HMW) and the size fraction between 5 kDa and 10

kDa as low molecular weight colloid fraction (LMW), while the permeate from the

ultrafiltration (< 5 kDa) is considered to be the truly dissolved fraction (UD) in the

present work.


- 110 -


6.2.1. Study area

Fig. 1. The Mundaú-Manguaba lagoon system in Northeast Brazil.

The Mundaú-Manguaba lagoon system is located in the state of Alagoas, Northeast

Brazil (Fig. 1). Its climate is tropical, semi-humid with well-defined dry (September-

March) and wet (April-August) seasons (Oliveira and Kjerfve, 1993; Brockmeyer and

Spitzy, 2011), with mean annual temperatures around 24°C (Diegues, 1994). The

Mundaú-Manguaba estuarine system is composed of two large lagoons (Melo-

Magalhães et al., 2009): the northern Mundaú lagoon (in the following: MD), with the

Rio Mundaú as its major tributary (catchment area 2135 km2), has a total surface

water area of 24 km2 and an average depth of 1.5 m; the southern Manguaba lagoon

(in the following: MG), which receives fresh water mainly from the Rio Paraíba do

Meio (catchment area 3299 km2) and the Rio Sumaúma (catchment area 372 km2),

has a total surface water area of 43 km2 and an average depth of 2 m (de Souza et

al., 2002; Costa et al., 2011). Both lagoons are interconnected by mainly mangrove-

lined narrow channels with a single inlet connecting them to the Atlantic. Such so-

called “choked lagoons” are characterized by one or more long and narrow entrance

channels, long residence times of water in the estuary and dominant wind forcing


- 111 -

(Kjerfve and Magill, 1989), and are particularly vulnerable to natural and

anthropogenic impacts (Knoppers et al., 1991). In both lagoons, the flux of salt water

from the coastal ocean into the lagoons is strongly dampened by the channel system,

where tidal cycles determine the direction and magnitude of currents (Brockmeyer

and Spitzy, 2011). During the wet season, the lagoons are strongly influenced by

freshwater and sediment transport in the rivers, while the river flow is low during the

dry season which means that the system is mainly controlled by tidal variation (Spörl,

2011). The average water residence time ranges from around 1-2 weeks for MD to 5-

7 weeks for MG (Oliveira and Kjerfve, 1993). In the river catchment areas, sugar

cane is the dominant land cover for both lagoons with a further impact by urban

effluents for the MD from the city of Maceío (population ca. 1 million; e.g.,

Brockmeyer and Spitzy, 2011).

The Mundaú-Manguaba lagoon system is strongly influenced by monocultural sugar

cane cultivation during the last decades, which led to raised nutrient inputs and may

result in spatial and seasonal differences of nutrient concentrations in the system

(Spörl, 2011). This may be the reason why dissolved organic carbon concentration in

the study area (4-11 mg/L) is considerably higher than in other estuaries in Brazil,

such as in the Amazon estuary (1-4 mg/L; Sholkovitz et al., 1978), in the Sepetiba

Bay (1.7-5.8 mg/L; de Lacerda et al., 2001) and in the Paraíba do Sul River estuary

(1.5-5.6 mg/L; Krüger et al., 2006).

6.2.2. Sampling and filtration

The sampling was carried out during September/October 2007 (nominal wet season)

as well as during February 2008 and March 2009 (nominal dry season) with the aid

of small boats. Surface waters from a depth of 0.3 m were filled by hand into pre-

cleaned low-density polyethylene (LDPE) bottles (Nalgene®, 1 L) upstream of the

ship. Within 24 hours the water samples were drawn through acid-cleaned 0.45 μm

polycarbonate filters (Nuclepore®; pre-weighed) in the laboratory. The filter cakes

were transferred to Petri dishes and stored at -10 °C. Subsequently, the filtrates were

separated into fractions of 30 kDa-0.45 μm, 10-30 kDa, 5-10 kDa and < 5 kDa using

cross flow filtration devices (CFF) with a flow rate of 500 mL/min (2 bar). The

fractions of 30 kDa-0.45 μm and 10-30 kDa have been put together, because these

two fractions contained no additional information. The filtrate, the retentate and

permeate were weighted before the next step began. After weighing, 10 mL of each

retentate/permeate was filled into an ampoule with addition of phosphoric acid for the


- 112 -

dissolved organic carbon measurement. The rest of the sample was filled into LDPE

bottles (Nalgene®) and acidified to pH 2 with sub-boiled concentrated nitric acid for

trace metal analysis. The sample bottles were wrapped in polyethylene bags and

stored at room temperature. The used cross-flow filtration devices (Vivaflow® 200,

Sartorius) constructed of polyethersulfone with an active membrane area of 200 cm2

were pre-washed with HCl (0.1%, 30 min) and EDTA (0.02 M, 30 min) and

subsequently rinsed with Milli-Q® water.

6.2.3. Pretreatment and measurement of dissolved and colloidal metals

Sample work-up was performed in a clean-air laboratory (class 1000) with additional

clean benches of class 100. The samples were first irradiated in a UV-digester (500W;

UV-digester 705, Metrohm) for 2 hours with addition of hydrogen peroxide in order to

destroy the organic complexes. Afterwards they were extracted by an

extraction/backextraction procedure based on the work of Danielsson et al. (1978): at

first, 40-100 g sample (depending on the metal concentration) was weighed into a

fluorinated ethylene propylene (FEP) separatory funnel (Nalgene®). After pH

adjustment to 4.0-4.5 with ammonium acetate buffer (pH ~ 9), the sample was well

shaken (15 min) after addition of 500 μL APDC/NaDETC solution (0.06 M) and 10 mL

Freon. The lower organic phase was drawn off into a polypropylene tube following

the phase separation. After addition of 100 μL concentrated nitric acid the tube was

shaken by hand for 3 min. Finally, 1900 μL Milli-Q® water was added and the tube

was again shaken by hand for 2 min for back-extraction. After the phase separation,

the upper aqueous extract was removed into an acid-cleaned sample tube using a

pipette. The extract was measured by high-resolution-inductively-coupled-plasma

mass-spectrometry (HR-ICP-MS; Element 2, Thermo Scientific).

6.2.4. Cross-flow filtration Blanks and Mass balance

To obtain CFF blanks by using a new pre-washed CFF device, a Milli-Q® water

sample was separated by ultrafiltration into 4 colloidal fractions (30 kDa - 0.45 μm;

10-30 kDa; 5-10 kDa and < 5 kDa). All acidified sub-samples were analyzed for trace

metals using HR-ICP-MS. All trace metal concentrations of the blanks were less than

1% of the minimum sample concentrations or not detectable.


- 113 -

Cd2007: y = 0.6777x

2009: y = 0.9382x0.00

0.04

0.08

0.12

0.00 0.04 0.08 0.12total dissolved (nmol/L)

HM

W+L

MW

+UD

(n

mol

/L)

Cu2007: y = 0.9862x

2009: y = 1.0632x0

10

20

30

0 10 20 30total dissolved (nmol/L)

HM

W+L

MW

+UD

(n

mol

/L)

Fe2007: y = 0.5416x

2009: y = 0.5724x0

2

4

6

0 2 4 6total dissolved (μmol/L)

HM

W+L

MW

+UD

(μ

mol

/L)

Pb2007: y = 0.6393x

2009: y = 1.0114x0.00

0.25

0.50

0.75

1.00

0.00 0.25 0.50 0.75 1.00total dissolved (nmol/L)

HM

W+L

MW

+UD

(n

mol

/L)

Co2007: y = 1.1526x

2009: y = 1.0386x0

1

2

3

4

0 1 2 3 4total dissolved (nmol/L)

HM

W+L

MW

+UD

(n

mol

/L)

Ni2007: y = 1.1725x

2009: y = 1.0003x0

5

10

15

20

0 5 10 15 20total dissolved (nmol/L)

HM

W+L

MW

+UD

(n

mol

/L)

Fig. 2. Calculated total dissolved concentration (HMW + LMW + UD) vs. separately measured total dissolved trace metal concentration : MD 2007; : MG 2007; : MD 2009; : MG 2009. Solid lines: Correlation between calculated and separately measured total dissolved concentration for Cd, Cu, Fe, Pb, Co and Ni during the wet season of 2007; Broken line: Correlation between calculated and separately measured total dissolved concentration for Cd, Cu, Fe, Pb, Co and Ni during the dry season of 2009).

Since there is no standard reference material to quantify the accuracy of the

ultrafiltration analysis, a concentration mass balance between the sum of all size

fractions (HMW + LMW + UD) and the 0.45 μm filtrated sample (total dissolved

fraction) was calculated. A mass balance < 100% indicates a loss of material

probably due to scavenging onto the filter membrane or sample reservoir, while a

mass balance > 100% may be a evidence for a contamination during the sample

pretreatment (e.g., Martin et al., 1995; Powell et al., 1996). Also the propagation of

sampling and analytical errors for the colloidal sub-samples and for the UD sample in

particular may cause a deviation of the mass balance from 100%. Especially for


- 114 -

lower concentrations, a small absolute error may create a remarkable aberrance in

the mass balance. Most calculations showed a mass balance around 100%. Only

those data sets were evaluated, which had a mass balance within a range of 90% ±

40%, 110% ± 30%, 80% ± 40%, 70% ± 40%, 100% ± 30% and 110% ± 40% during

the wet season of 2007, and 100% ± 40%, 100% ± 30%, 100% ± 50%, 100% ± 50%,

100% ± 20% and 100% ± 30% during the dry season of 2009 for Cd, Cu, Fe, Pb, Co

and Ni, respectively. The calculated coefficients of determination generally showed a

good correlation between the separately measured total dissolved trace metal

concentrations and the sum of the colloidal and the UD fractions (Fig. 2). The slopes

were mainly within a range of 0.9-1.1. However, the slopes for Cd, Fe and Pb during

the wet season of 2007 and the slope for Fe during the dry season of 2009 were

lower probably indicating a loss of material, but have not influenced the operative

mass balance. During the wet season of 2007, elevated slopes were found for Co

and Ni possibly evidencing a contamination. Errors during the sampling or the

calculation could also cause the fluctuation of the mass balance. From a total of 31

analyzed samples, 20 for Cd, 24 for Cu, 14 for Fe, 15 for Pb, 21 for Co and 20 for Ni

were selected for further evaluation, which had a mass balance in the respective

defined range.

6.2.5. Pretreatment and measurement of particulate metals

The pretreatment of particulate samples was also performed in a clean-air laboratory

(class 1000) with additional clean benches of class 100. The filter cakes were

weighed at least threefold with one day storage over conc. H2SO4 in between, before

the filter cakes have been dried over conc. H2SO4 for several days. After weighting,

the filter cakes were decomposed with the filter in sealed PTFE vessels using a

microwave-assisted pressure digestion technique (MWS-1200 mega microwave

digestion system with HPR 1000-6 rotor; MLS, Leutkirch, Germany) as Balzer et al.

(2013) did: The samples were destructed under a radiation power of up to 600 W

after adding a mixture of concentrated HNO3, HCl, HClO4 and HF (sub-boiled or

suprapure®), and subsequently under the same power after only adding HNO3 (for

further details, see Schüßler et al., 2005). The metal concentrations were also

measured by HR-ICP-MS. The certified reference material MESS-1 (marine sediment)

from the NRCC was analyzed in order to ensure the validity of the method. The

results were found to be in the certified ranges.


- 115 -

6.2.6. Other parameters

Environmental parameters such as salinity and water temperature were measured on

the spot using a multi-sensor-device (MultiLine P3, WTW). Dissolved organic carbon

was determined by high temperature catalytic oxidation (Total Carbon Analyzer,

Apollo 9000; Tecmar).

6.3. Results and Discussion

6.3.1. Salinity, suspended particulate matter and dissolved organic carbon

wet season 2007

0

20

40

60

80

0 10 20 30

SPM

(mg/

L)

dry season 2008

0

20

40

60

80

0 10 20 30

SPM

(mg/

L)

dry season 2009

0

20

40

60

80

0 10 20 30

SPM

(mg/

L)

(b)

wet season 2007

0

20

40

60

80

0 10 20 30

SPM

(mg/

L)

dry season 2008

0

20

40

60

80

0 10 20 30

SPM

(mg/

L)

dry season 2009

0

20

40

60

80

0 10 20 30

SPM

(mg/

L)

Fig. 3. The estuarine distribution of suspended particulate matter (SPM) vs. salinity (a) in the Mundaú Lagoon and (b) in the Manguaba Lagoon : MD; : MG; : Sumaúma; : coastal water; left side of dotted line: lagoon water; right side of dotted line: channel and coastal water).

The concentration of suspended particulate matter (SPM) decreased with increasing

salinity from typically 20 mg/L in the MD and 30 mg/L in the MG to less than 5 mg/L

during both the wet season and the dry seasons (Fig. 3). The extremely high SPM

concentration (76 mg/L) at salinity of 3 psu in the MD during the wet season of 2007

(Fig. 3a) probably originated from higher resuspension loadings. The SPM level in

the MG was generally higher than in the MD. The SPM concentration decreased with

increasing salinity from typically 30-40 mg/L to less than 5 mg/L during the wet

season of 2007 and the dry season of 2009 in the MG despite some scatter (Fig. 3b).

In the MG during the dry season of 2008, the SPM concentration in the riverine water

(S = 0) was relatively low (10 mg/L), and a mid turbidity maximum (20 mg/L) was

observed at a salinity between 20 and 25 psu. The freshwater end-member SPM in

the Rio Sumaúma was approximately at the same level as in the Rio Paraíba do

Meio, and could not influence the SPM distribution in the MG significantly, because of


- 116 -

its low discharge (Oliveira and Kjerfve, 1993). In comparison with estuaries in Europe

and North America (e.g., Uncles et al., 2002), the SPM concentration in the MD and

MG estuaries was at the mid level, and lower than most estuaries in South America

such as the Paraíba do Sul estuary (5-193 mg/L; Carvalho et al., 2002), the Río de la

Plata estuary (31-70 mg/L; Calliari et al., 2005) and the Bahía Blanca Estuary (78

mg/L; Guinder et al., 2009).

(a) wet season 2007

0

4

8

12

0 10 20 30

DOC

(mg/

L)

dry season 2008

0

4

8

12

0 10 20 30

DOC

(mg/

L)

dry season 2009

0

4

8

12

0 10 20 30

DOC

(mg/

L)

(b)

wet season 2007

0

4

8

12

0 10 20 30

DOC

(mg/

L)

dry season 2008

0

4

8

12

0 10 20 30

DO

C (m

g/L)

dry season 2009

0

4

8

12

0 10 20 30

DOC

(mg/

L)

Fig. 4. The estuarine distribution of DOC vs. salinity (a) in the Mundaú Lagoon and (b) in the Manguaba Lagoon : MD; : MG; : Sumaúma; : coastal water).

The distribution of DOC as determined by Balzer (unpublished) is depicted in Fig. 4.

During the wet season of 2007 and the dry season of 2008, the DOC concentration

generally decreased with increasing salinity from about 4 mg/L in the riverine water to

1 mg/L in coastal waters of both study areas, except for the MG during the dry

season of 2008 and the dry season of 2009. During the latter season, the DOC

concentration decreased from 9 mg/L in the MD and 12 mg/L in the MG to 4 mg/L

with increasing salinity. Cyanobacteria blooms were frequently observed in both

study areas (Porfirio et al., 1999; Spörl, 2011) which may promote the increase of

DOC (e.g., Norrman et al., 1995). The mid salinity maximum of DOC in the MG

during the dry season of 2008 and the DOC peak at a salinity of 15 psu in the MD

during the dry season of 2009 may therefore be a consequence of local

cyanobacteria blooms. The observed DOC data was comparable with earlier studies

in this study area (Wolf et al., 2010; Brockmeyer and Spitzy, 2011), but was higher

than in other Brazilian estuaries e.g., Sepetiba Bay (Luciani et al., 2007). It was

obviously at the higher limit when compared with the range of the European estuaries

(Abril et al., 2002).


- 117 -

6.3.2. Dissolved trace metals

(a) wet season 2007

0.00

0.04

0.08

0.12

0 10 20 30

Cd

(nm

ol/L

) dry season 2008

0.00

0.04

0.08

0.12

0 10 20 30

Cd

(nm

ol/L

) dry season 2009

0.00

0.04

0.08

0.12

0 10 20 30

Cd

(nm

ol/L

)

(b)

wet season 2007

0

1

2

3

4

0 10 20 30

Cd

(nm

ol/g

) dry season 2008

0

1

2

3

4

0 10 20 30

Cd

(nm

ol/g

) dry season 2009

0

1

2

3

4

0 10 20 30

Cd

(nm

ol/g

)

(c)

wet season 2007

0%

25%

50%

75%

100%

0 0 3 5 21 36Salinity

Cd

ratio

HMWLMWUD

dry season 2009

0%

25%

50%

75%

100%

0 3 6 10 12 14 21 25 28Salinity

Cd

ratio

HMWLMWUD

Fig. 5. Estuarine distribution of Cd in the Mundaú Lagoon : MD; : coastal water): (a) total dissolved Cd, (b) particulate Cd and (c) Cd percentage of colloidal size fractions; dotted: high molecular weight fraction; hatched: low molecular weight fraction; solid: truly dissolved fraction.


- 118 -

(a) wet season 2007

0.00

0.04

0.08

0.12

0 10 20 30

Cd

(nm

ol/L

) dry season 2008

0.00

0.04

0.08

0.12

0 10 20 30

Cd

(nm

ol/L

) dry season 2009

0.00

0.04

0.08

0.12

0 10 20 30

Cd

(nm

ol/L

)

(b)

wet season 2007

0

1

2

3

4

0 10 20 30

Cd

(nm

ol/g

) dry season 2008

0

1

2

3

4

0 10 20 30

Cd

(nm

ol/g

) dry season 2009

0

1

2

3

4

0 10 20 30

Cd

(nm

ol/g

)

(c)

wet season 2007

0%

25%

50%

75%

100%

0 0 0 1 3 5 9 10 19Salinity

Cd

ratio

HMWLMWUD

dry season 2009

0%

25%

50%

75%

100%

0 0 2 4 8 17 34Salinity

Cd

ratio

HMWLMWUD

Fig. 6. Estuarine distribution of Cd in the Manguaba Lagoon : MG; : Sumaúma; : coastal water): (a) total dissolved Cd, (b) particulate Cd and (c) Cd percentage of colloidal size fractions; dotted: high molecular weight fraction; hatched: low molecular weight fraction; solid: truly dissolved fraction.

Cadmium: Numerous estuarine investigations and laboratory experiments suggest

that an enrichment of dissolved Cd due to chloride-induced desorption from particles

is a typical feature in estuaries (e.g., Chiffoleau et al., 2001; Waeles et al., 2004;

Dabrin et al., 2009). But this effect was not clear in this study. In the MD during the

wet season of 2007, dissolved Cd decreased rapidly from about 0.08 nmol/L in the

freshwater to 0.03 nmol/L at high salinities (Fig. 5a). Such elimination (at low salinity)

was also observed e.g., in the Delaware estuary (Sharp et al., 1982), in the Göta

River estuary (Danielsson et al., 1983) and in the Danshuei River estuary (Jiann et

al., 2005) and is often attributed to adsorption onto Fe/Mn oxides (e.g., Turner et al.,

2008). In the MG, however, the freshwater dissolved Cd concentration was quite low

(< 0.02 nmol/L) and varied insignificantly with salinity (Fig. 6a). During both dry

seasons (2008 and 2009), dissolved Cd increased from typically 0.02 nmol/L in the

riverine water to about 0.1 nmol/L at high salinities. Such a more or less conservative

behavior was also observed e.g., in the Rhône estuary (Elbaz-Poulichet et al., 1996)

and in estuaries of East Hainan (Fu et al., 2013). However, during the dry season of


- 119 -

2009, the dissolved Cd concentration in the coastal water outside the lagoon system

was quite low (0.03 nmol/L) probably representing waters from remote oceanic sites.

Dissolved Cd seemingly shows a positive deviation from the theoretical dilution line.

A previous study reported a dissolved Cd concentration of 0.02 nmol/L in Brazilian

coastal water (Kremling, 1985). The calculation of mass balance showed also a

positive Cd net flux into the ocean (particulate + dissolved concentration in nmol/L;

not shown). Thus, in addition to possible anthropogenic contributions, desorption

from the sediment and/or suspended particles inside the lagoon system is assumed.

While at lower salinities (0-20 psu), such desorption was not pronounced, the

increase was significant at higher salinities as also observed in the Huanghe estuary

(Elbaz-Poulichet et al., 1987). The elevated Cd concentrations occurred mostly in the

channel (Fig. 3). This may be due to higher flow rate in the channel section and

subsequently stronger water-sediment exchange. Dissolved oxygen was neither

measured nor published for the dry seasons. But during the wet season of 2008,

dissolved oxygen in the lagoons was quite low in the surface water (about 5 mg/L;

Spörl, 2011), and presumably much lower at the water-sediment-interface during the

dry season. Such suboxic conditions may have inhibited the release of Cd in the

lagoons (e.g., Jiann et al., 2005).


- 120 -

(a) wet season 2007

0

5

10

15

20

0 10 20 30

Cu

(nm

ol/L

) dry season 2008

0

5

10

15

20

0 10 20 30

Cu

(nm

ol/L

) dry season 2009

0

5

10

15

20

0 10 20 30

Cu

(nm

ol/L

)

(b)

wet season 2007

0.00

0.15

0.30

0.45

0 10 20 30

Cu (μ

mol

/g) dry season 2008

0.00

0.15

0.30

0.45

0 10 20 30

Cu (μ

mol

/g) dry season 2009

0.00

0.15

0.30

0.45

0 10 20 30

Cu (μ

mol

/g)

(c)

wet season 2007

0%

25%

50%

75%

100%

0 0 3 5 21 36Salinity

Cu

ratio

HMWLMWUD

dry season 2009

0%

25%

50%

75%

100%

0 3 6 10 12 14 21 25 28Salinity

Cu

ratio

HMWLMWUD

Fig. 7. Estuarine distribution of Cu in the Mundaú Lagoon ( : MD; : coastal water): (a) total dissolved Cu, (b) particulate Cu and (c) Cu percentage of colloidal size fractions; dotted: high molecular weight fraction; hatched: low molecular weight fraction; solid: truly dissolved fraction.


- 121 -

(a) wet season 2007

0

5

10

15

20

0 10 20 30

Cu

(nm

ol/L

) dry season 2008

0

5

10

15

20

0 10 20 30

Cu

(nm

ol/L

) dry season 2009

0

5

10

15

20

0 10 20 30

Cu

(nm

ol/L

)

(b)

wet season 2007

0.00

0.15

0.30

0.45

0 10 20 30

Cu

(μm

ol/g

) dry season 2008

0.00

0.15

0.30

0.45

0 10 20 30

Cu

(μm

ol/g

) dry season 2009

0.00

0.15

0.30

0.45

0 10 20 30

Cu

(μm

ol/g

)

(c)

wet season 2007

0%

25%

50%

75%

100%

0 0 0 1 3 5 9 10 19Salinity

Cu

ratio

HMWLMWUD

dry season 2009

0%

25%

50%

75%

100%

0 0 2 4 8 17 34Salinity

Cu

ratio

HMWLMWUD

Fig. 8. Estuarine distribution of Cu in the Manguaba Lagoon : MG; : Sumaúma; : coastal water): (a) total dissolved Cu, (b) particulate Cu and (c) Cu percentage of colloidal size fractions; dotted: high molecular weight fraction; hatched: low molecular weight fraction; solid: truly dissolved fraction.

Copper: A supply of Cu to the solution phase during the estuarine mixing may be a

result of external sources such as desorption from particulate Cu or anthropogenic

activities (e.g., Zwolsman et al., 1997; Guo, et al., 2000b; Santos-Echeandia et al.,

2008). On the other side, a removal of Cu from the dissolved phase may occur due to

the formation of copper sulfide (Jiann and Wen, 2009). In this study, dissolved Cu

concentration decreased from typically 10 nmol/L to 1-2 nmol/L with increasing

salinity during all seasons in both study areas (Fig. 7a and Fig. 8a). Dissolved Cu

appeared to behave conservatively during both wet and dry seasons, which was

widely observed e.g., in the Mississippi River estuary (Wen et al., 2011) and in the

Nerbioi-Ibaizabal River estuary (Fernández et al., 2008). The fluctuations at low

salinities, especially in the MG, may be attributed to irregular anthropogenic inputs

into the riverine water.


- 122 -

(a) wet season 2007

0

2

4

6

0 10 20 30

Fe (μ

mol

/L)

dry season 2008

0

2

4

6

0 10 20 30

Fe (μ

mol

/L)

dry season 2009

0

2

4

6

0 10 20 30

Fe (μ

mol

/L)

(b)

wet season 2007

0.0

0.5

1.0

1.5

2.0

0 10 20 30

Fe (m

mol

/g) dry season 2008

0.0

0.5

1.0

1.5

2.0

0 10 20 30

Fe (m

mol

/g) dry season 2009

0.0

0.5

1.0

1.5

2.0

0 10 20 30

Fe (m

mol

/g)

(c)

wet season 2007

0%

25%

50%

75%

100%

0 0 3 5 21 36Salinity

Fe ra

tio

HMWLMWUD

dry season 2009

0%

25%

50%

75%

100%

0 3 6 10 12 14 21 25 28Salinity

Fe ra

tio

HMWLMWUD

Fig. 9. Estuarine distribution of Fe in the Mundaú Lagoon ( : MD; : coastal water): (a) total dissolved Fe, (b) particulate Fe and (c) Fe percentage of colloidal size fractions; dotted: high molecular weight fraction; hatched: low molecular weight fraction; solid: truly dissolved fraction.


- 123 -

(a) wet season 2007

0

2

4

6

0 10 20 30

Fe (μ

mol

/L)

dry season 2008

0

2

4

6

0 10 20 30

Fe (μ

mol

/L)

dry season 2009

0

2

4

6

0 10 20 30

Fe (μ

mol

/L)

(b)

wet season 2007

0.0

0.5

1.0

1.5

2.0

0 10 20 30

Fe (m

mol

/g) dry season 2008

0.0

0.5

1.0

1.5

2.0

0 10 20 30

Fe (m

mol

/g) dry season 2009

0.0

0.5

1.0

1.5

2.0

0 10 20 30

Fe (m

mol

/g)

(c)

wet season 2007

0%

25%

50%

75%

100%

0 0 0 1 3 5 9 10 19Salinity

Fe ra

tio

HMWLMWUD

dry season 2009

0%

25%

50%

75%

100%

0 0 2 4 8 17 34Salinity

Fe ra

tio

HMWLMWUD

Fig. 10. Estuarine distribution of Fe in the Manguaba Lagoon : MG; : Sumaúma;

: coastal water): (a) total dissolved Fe, (b) particulate Fe and (c) Fe percentage of colloidal size fractions; dotted: high molecular weight fraction; hatched: low molecular weight fraction; solid: truly dissolved fraction.

Iron: In the MD (Fig. 9a), dissolved Fe decreased rapidly from 2-6 μmol/L during the

wet season and 4 μmol/L during both dry seasons to typically 0.2 μmol/L at low

salinities (S < 10) and varied insignificantly at higher salinities (S = 10-33). In the MG,

dissolved Fe also decreased with increasing salinity from typically 1-6 μmol/L to 0.2

μmol/L at S > 31 (Fig. 10a). The considerable fluctuations in the freshwater and some

elevated outliers at low salinities are probably due to irregular input from sewage and

agricultural activities (especially in the MD during the dry season of 2009 and in the

Rio Sumaúma during the wet season of 2007). During various field works and

laboratory experiments, the estuarine removal of dissolved Fe (e.g., Aston and

Chester, 1973; Sholkovitz et al., 1978) was estimated to result in a loss of more than

90% (e.g., Boyle et al., 1977; Sholkovitz, 1978) as a consequence of flocculation of

colloidal Fe and subsequent sedimentation (e.g., Mayer, 1982; Fox and Wofsy, 1983).

In both lagoons, dissolved Fe behaved non-conservatively and also underwent

estuarine flocculation. The riverine dissolved Fe in the MD was mostly higher than in


- 124 -

the MG suggesting geochemical differences of the catchment areas between the Rio

Mundaú and the Rio Paraíba do Meio.

(a) wet season 2007

0.0

0.5

1.0

1.5

2.0

0 10 20 30

Pb (n

mol

/L) dry season 2008

0.0

0.5

1.0

1.5

2.0

0 10 20 30Pb

(nm

ol/L

) dry season 2009

0.0

0.5

1.0

1.5

2.0

0 10 20 30

Pb (n

mol

/L)

(b)

wet season 2007

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Pb (μ

mol

/g) dry season 2008

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Pb (μ

mol

/g) dry season 2009

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Pb (μ

mol

/g)

(c)

wet season 2007

0%

25%

50%

75%

100%

0 0 3 5 21 36Salinity

Pb ra

tio

HMWLMWUD

dry season 2009

0%

25%

50%

75%

100%

0 3 6 10 12 14 21 25 28Salinity

Pb

ratio

HMWLMWUD

Fig. 11. Estuarine distribution of Pb in the Mundaú Lagoon ( : MD; : coastal water): (a) total dissolved Pb, (b) particulate Pb and (c) Pb percentage of colloidal size fractions; dotted: high molecular weight fraction; hatched: low molecular weight fraction; solid: truly dissolved fraction.


- 125 -

(a) wet season 2007

0.0

0.5

1.0

1.5

2.0

0 10 20 30

Pb (n

mol

/L) dry season 2008

0.0

0.5

1.0

1.5

2.0

0 10 20 30

Pb (n

mol

/L) dry season 2009

0.0

0.5

1.0

1.5

2.0

0 10 20 30

Pb (n

mol

/L)

(b)

wet season 2007

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Pb (μ

mol

/g) dry season 2008

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Pb (μ

mol

/g) dry season 2009

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Pb (μ

mol

/g)

(c)

wet season 2007

0%

25%

50%

75%

100%

0 0 0 1 3 5 9 10 19Salinity

Pb ra

tio

HMWLMWUD

dry season 2009

0%

25%

50%

75%

100%

0 0 2 4 8 17 34Salinity

Pb

ratio

HMWLMWUD

Fig. 12. Estuarine distribution of Pb in the Manguaba Lagoon : MG; : Sumaúma;

: coastal water): (a) total dissolved Pb, (b) particulate Pb and (c) Pb percentage of colloidal size fractions; dotted: high molecular weight fraction; hatched: low molecular weight fraction; solid: truly dissolved fraction.

Lead: Dissolved Pb decreased with increasing salinity from typically 0.4 nmol/L to

0.1-0.2 nmol/L in the MD during both the wet season and two dry seasons (Fig. 11a).

The estuarine removal of dissolved Pb was established by various estuarine

investigations (e.g., Dai and Martin, 1995; Wen et al., 1999; Monbet, 2006), which

was attributed to co-precipitation with Fe and Mn oxides (e.g., Wen et al., 2008). In

the MG, however, such a decrease was not pronounced (Fig. 12a). During the dry

season of 2008, dissolved Pb was constant at about 0.2-0.3 nmol/L, such as the

conservative behavior e.g., in the Rhône Estuary (Elbaz-Poulichet et al., 1996) and in

the Gironde Estuary (Kraepiel et al., 1997). During the wet season of 2007 and the

dry season of 2009, dissolved Pb concentration in the river and the sea end-

members differed from the dry season of 2008 not significantly, but the estuarine

distribution showed a positive deviation from the theoretical dilution line. A similar

distribution was also noted e.g., in the Mersey estuary due to anthropogenic activities

(Comber et al., 1995) and in the Penzé estuary due to resuspension of historically


- 126 -

contaminated sediment (Waeles et al., 2007). Since the SPM load was higher during

both surveys (Fig. 3), a mobilization from labile particulates might have taken place.

The elevation of dissolved Pb in the MD at s ≈ 10 during the dry season of 2009 and

in the Rio Sumaúma during the wet season of 2007 is quite similar to dissolved Fe

and Ni. Thus, a contamination during the sample pretreatment might be excluded. It

is more likely, as discussed before, that irregular municipal input was responsible for

the excesses of dissolved Fe and Pb.

(a) wet season 2007

0

1

2

3

4

0 10 20 30

Co (n

mol

/L) dry season 2008

0

1

2

3

4

0 10 20 30

Co (n

mol

/L) dry season 2009

0

1

2

3

4

0 10 20 30

Co (n

mol

/L)

(b)

wet season 2007

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Co

(μm

ol/g

) dry season 2008

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Co

(μm

ol/g

) dry season 2009

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Co

(μm

ol/g

)

(c)

wet season 2007

0%

25%

50%

75%

100%

0 0 3 5 21 36Salinity

Co

ratio

HMWLMWUD

dry season 2009

0%

25%

50%

75%

100%

0 3 6 10 12 14 21 25 28Salinity

Co

ratio

HMWLMWUD

Fig. 13. Estuarine distribution of Co in the Mundaú Lagoon ( : MD; : coastal water): (a) total dissolved Co, (b) particulate Co and (c) Co percentage of colloidal size fractions; dotted: high molecular weight fraction; hatched: low molecular weight fraction; solid: truly dissolved fraction.


- 127 -

(a) wet season 2007

0

1

2

3

4

0 10 20 30

Co

(nm

ol/L

) dry season 2008

0

1

2

3

4

0 10 20 30

Co

(nm

ol/L

) dry season 2009

0

1

2

3

4

0 10 20 30

Co

(nm

ol/L

)

(b)

wet season 2007

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Co

(μm

ol/g

) dry season 2008

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Co

(μm

ol/g

) dry season 2009

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Co

(μm

ol/g

)

(c)

wet season 2007

0%

25%

50%

75%

100%

0 0 0 1 3 5 9 10 19Salinity

Co

ratio

HMWLMWUD

dry season 2009

0%

25%

50%

75%

100%

0 0 2 4 8 17 34Salinity

Co

ratio

HMWLMWUD

Fig. 14. Estuarine distribution of Co in the Manguaba Lagoon : MG; : Sumaúma;

: coastal water): (a) total dissolved Co, (b) particulate Co and (c) Co percentage of colloidal size fractions; dotted: high molecular weight fraction; hatched: low molecular weight fraction; solid: truly dissolved fraction.

Cobalt: Dissolved Co in the MD and MG was plotted as a function of salinity in Fig.

13a and Fig. 14a. During the wet season in both study areas, dissolved Co

decreased from about 0.6 nmol/L in the riverine water to 0.4 nmol/L in coastal waters.

During both dry seasons, dissolved Co also decreased from typically 2 nmol/L to less

than 0.4 nmol/L. The estuarine distribution behaved non-conservatively, and was

characterized by a positive deviation from the theoretical dilution line with a maximum

of about 2 nmol/L at low salinity during the wet season and maxima of about 3 nmol/L

during both dry seasons. The enrichment of dissolved Co appeared to take place in

particular at low salinities. Although removal of dissolved Co was observed e.g., in

the Changjiang estuary (Wang and Liu, 2003) and in East Hainan (Fu et al., 2013),

most field works and labor experiments indicated the inverse situation (e.g., Li et al.,

1984; Martino et al., 2002; Tovar-Sánchez et al., 2004; Takata et al., 2010). Release

of Co from the resuspended sediment may have contributed to the enrichment of

dissolved Co (e.g., Chiffoleau et al., 1994; Chaudry and Zwolsman, 2008).


- 128 -

(a) wet season 2007

0

5

10

15

0 10 20 30

Ni (

nmol

/L)

dry season 2008

0

5

10

15

0 10 20 30

Ni (

nmol

/L)

dry season 2009

0

5

10

15

0 10 20 30

Ni (

nmol

/L)

(b)

wet season 2007

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Ni (μm

ol/g

)

dry season 2008

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Ni (μm

ol/g

)

dry season 2009

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Ni (μm

ol/g

)

(c)

wet season 2007

0%

25%

50%

75%

100%

0 0 3 5 21 36Salinity

Ni r

atio

HMWLMWUD

dry season 2009

0%

25%

50%

75%

100%

0 3 6 10 12 14 21 25 28Salinity

Ni r

atio

HMWLMWUD

Fig. 15. Estuarine distribution of Ni in the Mundaú Lagoon ( : MD; : coastal water): (a) total dissolved Ni, (b) particulate Ni and (c) Ni percentage of colloidal size fractions; dotted: high molecular weight fraction; hatched: low molecular weight fraction; solid: truly dissolved fraction.


- 129 -

(a) wet season 2007

0

5

10

15

0 10 20 30

Ni (

nmol

/L)

dry season 2008

0

5

10

15

0 10 20 30

Ni (

nmol

/L)

dry season 2009

0

5

10

15

0 10 20 30

Ni (

nmol

/L)

(b)

wet season 2007

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Ni (μm

ol/g

)

dry season 2008

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Ni (μm

ol/g

)

dry season 2009

0.0

0.1

0.2

0.3

0.4

0 10 20 30

Ni (μm

ol/g

)

(c)

wet season 2007

0%

25%

50%

75%

100%

0 0 0 1 3 5 9 10 19Salinity

Ni r

atio

HMWLMWUD

dry season 2009

0%

25%

50%

75%

100%

0 0 2 4 8 17 34Salinity

Ni r

atio

HMWLMWUD

Fig. 16. Estuarine distribution of Ni in the Manguaba Lagoon : MG; : Sumaúma;

: coastal water): (a) total dissolved Ni, (b) particulate Ni and (c) Ni percentage of colloidal size fractions; dotted: high molecular weight fraction; hatched: low molecular weight fraction; solid: truly dissolved fraction.

Nickel: Despite some variabilities at low salinities, which may be explained by

fluctuations in the incoming river water (Chaudry and Zwolsman, 2008), dissolved Ni

decreased generally with increasing salinity from typically 10 nmol/L to less than 3

nmol/L in the coastal ocean. This was observed during the wet season and both dry

seasons and in both study areas except in the MD during the dry season of 2008,

during which dissolved Ni concentration was essentially constant at about 2 nmol/L

(Fig. 15a and Fig. 16a). When the high dissolved Ni concentration (15 nmol/L) at the

salinity of 34 psu in the MG during the dry season of 2009 could be identified as

irregular discharge or contamination during the sample pretreatment, dissolved Ni

essentially has a conservative behavior in both study areas and during all three

sample periods. Although also non-conservative estuarine behavior of dissolved Ni

has been reported (e.g., Sholkovitz 1978; Campbell et al., 1988; Wang and Liu,

2003), Ni is largely unreactive and mainly behaves conservatively in estuaries (e.g.,

Danielsson et al., 1983; Dai and Martin, 1995; Turner and Martino, 2006).


- 130 -

6.3.3. Particulate trace metals

Only during the wet season in the MD (Fig. 5b), the distribution pattern of particulate

Cd resembled the dissolved one. Its concentration decreased rapidly from about 5

nmol/g in the freshwater to about 0.5 nmol/g at high salinities. Such a decrease of

particulate Cd has been observed before, but always combined with an enrichment of

dissolved Cd (e.g., Paucot and Wollast, 1997). Such a “Fe-like” estuarine distribution

of both particulate and dissolved Cd was not found in any other estuary. Rapid co-

precipitation with Fe/Mn oxides (e.g., Turner et al., 2008) might be an explanation;

alternatively, riverine Cd might be removed by formation of Cd sulfides under anoxic

conditions as discussed before (e.g., Wen et al., 2011). During both dry seasons in

the MD, particulate Cd varied between 0.5 and 1.2 nmol/g and showed no

considerable change with salinity, which was widely observed e.g., in the Huanghe

estuary (Elbaz-Poulichet et al., 1987), in several other Chinese estuaries (Zhang and

Liu, 2002) and in the Galveston Bay (Wen et al., 2008). Similar levels for particulate

Cd (about 0.6 nmol/g) were also found in the MG during all three seasons (Fig. 6b)

and appear to be the predominant feature for both lagoons. Although seasonal

variability occurred in both study areas with respect to dissolved Cd, particulate Cd

does not seem to be influenced by seasonality. Simple dilution of particulate Cd is not

only the dominant feature in the MD and MG, but also in other estuaries in the world

(e.g., Paalman and van der Weijden, 1992; Wen et al., 2008).

Similar to the dissolved phase, particulate Cu generally followed the theoretical

dilution line in both study areas and during both wet and dry seasons despite some

scatter. Its value decreased from typically 0.2-0.3 μmol/g in the riverine water to

about 0.1-0.15 μmol/g at high salinities (Fig. 7b and Fig. 8b). Such a conservative

distribution of particulate Cu was often observed, e.g., in some Texas estuaries

(Benoit et al., 1994) and in the Gironde estuary (Kraepiel et al., 1997).

Particulate Fe rapidly decreased from typically 1.5 mmol/g to about 0.5 mmol/g at low

salinities and varied only slightly at higher salinities in both study areas and during

both wet and dry seasons (Fig. 9b and Fig. 10b). A similar distribution was observed

for particulate Pb, which decreased from typically 0.2 μmol/g to about 0.1 μmol/g and

stabilized at this level (Fig. 11b and Fig. 12b). Both particulate Fe and particulate Pb

were removed during the estuarine mixing as it occurs worldwide (e.g., Duinker and

Nolting, 1977; Benoit et al., 1994, Wen et al., 2008) as a consequence of flocculation

or co-precipitation as discussed before.

Particulate Co slightly decreased from typically 0.2 μmol/g at low salinities to about


- 131 -

0.1 μmol/g at high salinities in both study areas and during all seasons (Fig. 13b and

Fig. 14b). Particulate Ni varied between 0.2 and 0.3 μmol/g in both estuaries and

during all seasons (Fig. 15b and Fig. 16b), when a few low values are neglected.

Despite some fluctuations in the incoming river water, the estuarine distribution of

particulate Co and Ni may be considered as conservative mixing, which was widely

observed e.g., in the Rhine/Meuse estuary (Paalman and van der Weijden, 1992), in

the Hudson River estuary (Tovar-Sánchez et al., 2004) and in the estuaries of East

Hainan (Fu et al., 2013). This may be taken as a confirmation for the assumptions

made before, that the excess of dissolved Co has a sediment origin, and that Ni is

unreactive in the estuaries.

6.3.4. Distribution coefficient

The relative affinity of trace metals for dissolved and particulate phases can be

evaluated from the distribution coefficient KD as the ratio of the particulate (in mol/kg)

to the dissolved (in mol/L) concentration (e.g., Chiffoleau et al., 1994). It is a crucial

parameter for the interpretation of chemical transport in estuaries (Turner et al.,

1992). The logarithms of KD for the individual metals in both the MD and the MG were

plotted versus salinity in Fig. 17.

Among the six metals under investigation, Cd was the only metal, for which the

logarithm of KD clearly decreased from about 5.1 to 4.1 L/kg with increasing salinity

probably due to formation of highly stable Cd-chloro-complexes during the estuarine

mixing (e.g., Chiffoleau et al., 2001). The KD´s of Ni (~ 4.8 L/kg) was constant over

the whole salinity range. The logarithm of KD showed a slight increase with increasing

salinity (from 4.4 to 4.8 L/kg) for Cu, in contrast to the most estuaries worldwide (e.g.,

Balls et al., 1994; Fu et al., 2013), where no significant changes of KD´s with salinity

could be identified. Although both dissolved and particulate Cu appears to behave

conservatively, Cu seemed to favor the particulate phase with increasing salinity in

both study areas. Fe and Pb have the greatest affinity for the particulate phase (e.g.,

Fu et al., 2013). The logarithm of KD increased with increasing salinity from 5.5 to 6.5

L/kg for Fe. Despite the fluctuation at low salinities, the KD´s of Pb varied between 4.9

and 6.5 L/kg, and seem not to be dependent on salinity, which contrasts to the

positive correlation of KD vs. salinity in the estuaries of East Hainan (Fu et al., 2013).

The partial increase of dissolved Pb with salinity (Fig. 11a and Fig. 12a) probably

inhibited the increase of the KD´s. The enrichment of Co in solution and the slight

conservative decrease of particulate Co in both study areas resulted in a slight


- 132 -

increase of the KD with salinity from 4.9 to 5.2 L/kg. The higher KD´s for Co during the

wet season is a consequence of the lower dissolved concentration during that

season.

Cd

3

4

5

6

7

8

0 10 20 30Salinity

Log 1

0KD (L

/kg)

200720082009

Cu

3

4

5

6

7

8

0 10 20 30Salinity

Log 1

0KD (L

/kg)

Fe

3

4

5

6

7

8

0 10 20 30Salinity

Log 1

0KD (L

/kg)

Pb

3

4

5

6

7

8

0 10 20 30Salinity

Log 1

0KD (L

/kg)

Co

3

4

5

6

7

8

0 10 20 30Salinity

Log 1

0KD (L

/kg)

Ni

3

4

5

6

7

8

0 10 20 30Salinity

Log 1

0KD (L

/kg)

Fig. 17. Log10KD (L/kg) vs. salinity : Wet season of 2007; : Dry season of 2008; : Dry season of 2009).

A general order of log10KD for Fe (6.1 L/kg) > Pb (5.8 L/kg) > Co (5.0 L/kg) > Ni (4.8

L/kg) > Cd (4.6 L/kg) > Cu (4.5 L/kg) was calculated on the basis of the average

values over the whole estuary. This order generally agrees with the results presented

e.g., for the Weser estuary (Turner et al., 1992), for the Seine estuary (Chiffoleau et

al., 1994), for North Australian estuaries (Munksgaard and Parry, 2001) and for the

estuaries in East-Hainan (Fu et al., 2013). The relative magnitudes of the KD´s reflect

the affinity of Fe and Pb for the particulate phase and the favor of Cu and Cd for the


- 133 -

dissolved phase. Despite the differences in the source and the composition of the

suspended particles, the KD´s obtained in this study differ from the data in other

investigations (e.g., Balls et al., 1994; Kraepiel et al., 1997; Munksgaard and Parry,

2001; Fu et al., 2013) by less than one order of magnitude.

6.3.5. Colloidal trace metals

The colloidal size fraction analysis indicates for all dissolved metals under

consideration an increased importance of the truly dissolved (UD) fraction with

increasing salinity during both investigated seasons and in both study areas. Similar

features were also observed in the Ochlockonee estuary for Cd, Cu and Ni (Powell et

al., 1996) and in the San Francisco Bay for Cu and Fe (Hurst and Bruland, 2008).

During the wet season of 2007, the percentage of the UD-Cd in the total dissolved

fraction decreased from about 50% to about 25% at low salinity (S < 5), and

increased to more than 50% at middle and high salinities in both study areas, while

the HMW-Cd showed a converse trend (Fig. 5c and Fig. 6c). The percentage of the

LMW-Cd seems also to decrease at low salinity, but was constant at middle and high

salinities (20% in the MD and 15% in the MG). During the dry season of 2009, the

UD-Cd increased from 30% in the MD and from 50% in the MG to over 90% in the

salinity range S = 0 to S = 10 and remained constant at higher salinities. The HMW-

Cd decreased from 60% in the MD and 40% in the MG to less than 10% above S ≈

10, while the occurrence of the LMW with 5% on average was not significant. The

general increase of the UD fraction percentage with increasing salinity in both study

areas probably indicated the formation of highly stable Cd-chloro-complexes (e.g.,

Dabrin et al., 2009). During the dry season of 2009, the increase of the UD fraction

percentage could also be related to release of organo-bound-Cd from resuspended

sediment subsequent to its degradation (Waeles et al., 2005). During the wet season

of 2007, the decrease of the UD fraction percentage at low salinity in the MD was

probably attributed to adsorption onto Fe/Mn oxides as discussed. But in the MG, the

percentage of the UD fraction also decreased at low salinity, while total dissolved Cd

was essentially constant (Fig. 6a). An internal conversion between the UD and the

HMW fractions may have taken place. Furthermore, the lower percentage of the UD

in the incoming river water in both study areas during the wet season of 2007 is

combined with lower percentage in the lagoons (50% on average in both lagoons,

respectively), in contrast to the higher UD percentage in the freshwater and in the

lagoons (71% in the MD and 78% in the MG on average, respectively) during the dry


- 134 -

season of 2009. The presence of colloidal Cd appears to be associated with the

riverine input as observed e.g., in the Hudson River estuary (Yang and Sañudo-

Wilhelmy, 1998). Also the level of colloidal Cd involved in the estuarine chemistry

differs from one estuary to another. For instance, more than 90% of dissolved Cd was

found to be associated with molecules smaller than 10 kDa in the Rhône delta (Dai et

al., 1995) and in the Seine estuary (Chiffoleau et al., 2001), but only 60-70% in the

Venice Lagoon (Martin et al., 1995) and in the Danube River estuary (Guieu et al.,

1998). The partitioning of Cd is influenced by many variables such as pH (e.g., Hatje

et al., 2003b), temperature (e.g., Warren and Zimmerman, 1994), Ca2+ (e.g., Waeles

et al., 2005), Cl- (e.g., Dabrin et al., 2009) and DOC (e.g., Wen et al., 1999). It is

therefore difficult to forecast the colloidal composition of Cd in the estuary. As

discussed in the methodological section, a major contribution to these differences is

probably due to the different operational conditions.

In the MD and MG, the percentage of the UD fraction in total dissolved Cu increased

with increasing salinity from typically 20% to more than 50% during both wet and dry

seasons, while the percentage of the HMW-Cu decreased from typically 40% to less

than 10% (Fig. 7c and Fig. 8c). The LMW fraction with a percentage of 37% on

average during the wet season and 38% during the dry season varied insignificantly

with salinity. The increases of the UD-Cu content at the expense of the HMW-Cu

content were widely observed e.g., in the Rhône delta (Dai et al., 1995), in the San

Francisco Bay (Sañudo-Wilhelmy et al., 1996) and in the Galveston Bay (Wen et al.,

1999). With regard to the conservative distribution of dissolved and particulate Cu

(Fig. 7a, Fig. 7b, Fig. 8a and Fig. 8b), a conversion from the HMW to the UD fraction

might be suggested. Copper is well known to have a strong affinity to organic

functional groups (e.g., Shank et al., 2004). Dissolved and colloidal Cu is controlled

by dominant organic ligands in estuaries (e.g., Wen et al., 1999). Thus, such

conversion could be a consequence of the degradation of high molecular weight

dissolved organic materials (e.g., Masson et al., 2011) and/or the processing by

phytoplankton and bacteria (e.g., Wen et al., 2011) because of the high bio-

productivity in both study areas (Melo-Magalhães et al., 2009). It is also possible that

the removal of dissolved Cu (Fig. 7a and Fig. 8a) originates from a partial flocculation

of large organic colloids, while particulate Cu may associate with strong low

molecular weight ligands (Waeles et al., 2008a). The level of colloidal Cu in estuaries

worldwide varied significantly from 46% of the total dissolved fraction in the Venice

Lagoon (10 kDa - 0.4 μm; Martin et al., 1995) to 94% in the Penzè estuary (5 kDa -

0.45 μm; Waeles et al., 2008a). In both study areas, 57% of total dissolved Cu during


- 135 -

the wet season and 53% during the dry season could be identified in the colloidal

fraction (5 kDa - 0.45 μm). The highly varying results may be related - at least partly -

to the nature of organic matter within the system of concern (Waeles et al., 2008a).

The HMW of Fe with over 80% of the total dissolved fraction was the most important

fraction in both study areas and during both wet and dry seasons, while the HMW

and the LMW fractions ranged generally less than 10%, respectively (Fig. 9c and Fig.

10c). Despite the small extent of available data, a generally decrease of HMW

fraction and an associated increase of UD fraction tends to occur also for Fe. Both

bulk colloidal Fe-hydroxides and small size colloidal organo-bound-Fe is destabilized

by sea water ions such as Mg2+ and Ca2+ and consequently flocculate (e.g., Boyle et

al., 1977; Nowostawska et al., 2008; Shim et al., 2012). It is remarkable that the UD

and the LMW fraction of Fe during the dry season of 2009 (12% and 11% on average,

respectively) were more important than during the wet season of 2007 (1% and 4%

on average, respectively). With respect to the higher DOC during the dry season of

2009, Fe in the UD and the LMW fractions might be stabilized by humic material

against coagulation (e.g., Batchelli et al., 2010).

Similar to Fe, a complete estimation of the colloidal Pb distribution was difficult

because of the scarcity of usable data. But a decreasing tendency of the HMW

fraction and an increasing tendency of the UD fraction is observable (Fig. 11c and Fig.

12c). Dissolved Pb was mostly found in the HMW fraction (65% in the MD and 88%

in the MG) during the wet season, while the UD and the LMW fraction together

comprised less than one third of total dissolved Pb. Such Fe-like distribution was also

observed e.g., in the Galveston Bay (Benoit et al., 1994) and in the estuaries of East

Hainan (Fu, unpublished; see chapter 5), and was mainly attributed to the removal of

dissolved Pb with freshly precipitated Fe-hydroxides (e.g., Waeles et al., 2008a; Wen

et al., 2008). During the dry season, around 58% of total dissolved Pb was

associated with the UD fraction and only 37% with the HMW fraction in both study

areas. This may be explained by the absence of low salinity samples in the MD. But

the low salinity samples in the MG did contain a significant portion of UD-Pb. As

mentioned before, dissolved Pb was enriched due to release from SPM. It is possible

that the introduced Pb could not be effective scavenged because of the saturation of

the binding sites on suspended particles, or dissolved Pb was bound by strong

organic ligands (Kozelka et al., 1997).

The colloidal partition of Co in estuaries is highly variable. For instance, 81% of total

dissolved Co was reported to reside in the UD fraction (< 1 kDa) throughout the

Galveston Bay (Wen et al., 1999). On the other hand, over 50% of dissolved Co was


- 136 -

found in colloidal forms (1-100 kDa) in the estuaries in NW Russia (Pokrovsky and

Schott, 2002). Around 40% of UD-Co and 40% of HMW-Co dominated the estuaries

in East Hainan (Fu, unpublished; see chapter 5). In this study, total dissolved Co

contained on average about 37% UD fraction, 36% LMW fraction and 28% HMW

fraction during the wet season, and about 59% UD fraction, 31% LMW fraction and

10% HMW fraction during the dry season (Fig. 13c and Fig. 14c). The percentage of

the UD-Co clearly increased with increasing salinity from typically 25% to 60% during

the wet season and to more than 70% during the dry season, whereas the

percentages of LMW and HMW decreased correspondingly. A formation of stable low

molecular weight (chloro-, sulphato-, carbonato-and organo-) complexes might occur

(e.g., Saito and Moffett, 2001; Takara et al., 2010).

The UD fraction of Ni with a percentage of about 60% on average of total dissolved

fraction during both wet and dry seasons was the dominant size fraction in both study

areas, and clearly increased from typically 30% in the riverine water to more than

70% in the seawater except during the wet season of 2007 because of the scarcity of

usable data (Fig. 15c and Fig. 16c). The LMW-Ni, which represented 30% of total

dissolved Ni during both wet and dry seasons, decreased with increasing salinity, but

less than the HMW-Ni. The dominance of low molecular Ni (UD + LMW; < 10 kDa) in

this study is consistent with the observations e.g., in the Venice Lagoon (Martin et al.,

1995) and in the San Francisco Bay (Sañudo-Wilhelmy et al., 1996) and in the

Mississippi River estuary (Wen et al., 2011). Also an increasing significance of the

UD-Ni fraction with salinity was reported (Powell et al., 1996). Since dissolved and

particulate Ni was considered to behave conservatively as discussed before, the

HMW and the LMW fraction appear to be converted directly into the UD fraction. Ni is

able to complex with organic materials (e.g., Martin et al., 1995; Turner et al., 1998;

Wen et al., 2011). Thus, bulk organic-bound-Ni may also dissociate due to the

replacement by Ca2+ and Mg2+ during the estuarine mixing (e.g., Powell et al., 1996).

Also biological regeneration processes may cause the high proportion of dissolved Ni

in the UD fraction (Dai and Martin, 1995).

Summarizing the seasonal differences in the colloidal distribution, the UD fraction of

all metals was generally more important during the dry season of 2009 (74%, 47%,

12%, 55%, 59% and 58% on average in the whole study area for Cd, Cu, Fe, Pb, Co

and Ni, respectively) than during the wet season of 2007 (50%, 43%, 1%, 12%, 37%

and 59% on average in the whole study areas for Cd, Cu, Fe, Pb, Co and Ni,

respectively).

The dry season of 2009 was characterized by higher water temperatures, higher


- 137 -

DOC contents and lower water discharges in comparison with the wet season of

2007. The higher water temperature probably enhanced the bioactivity. For some

metals such as Cu (e.g., Wen et al., 2011), Fe (e.g., Hurst and Bruland, 2008) and Ni

(e.g., Dai and Martin, 1995) this may induce their colloidal redistribution. The higher

DOC content provides more binding sites for the metals and may stabilize the metals

in the dissolved phase. For instance, the enrichment of dissolved Cd was found not

only to be due to the formation of Cd-chloro-complexes but also to be due to the

binding of Cd with low-molecular-weight organic ligands (Waeles et al., 2008a).

Particulate Cu is known to be able to associate with low molecular weight organic

ligands and to form strong complexes (Waeles et al., 2008a). The lower water

discharge during the dry seasons led to a longer water residence time in the lagoons.

Some metals (Cd, Co, Fe and Pb; e.g., Takara et al., 2010; Braungardt et al., 2011)

can be mobilized from the particulate phase as either free ions or low molecular

weight organic complexes due to e.g., the more complete particle-water-exchange,

although some metals are rapidly re-scavenged (Caetano et al., 2003). However,

other parameters such as pH and the nature of the incoming colloids can also

influence the seasonal variability of the colloidal partition (e.g., Braungardt et al.,

2011). Since only a few comparable studies have been carried out and because of

the diversity of the environmental variables, it is difficult to estimate the reactions,

their mechanisms and the fate for the estuarine trace metal colloids with regard to

their seasonal variability.

6.3.6. Spatial and seasonal differences

In the concentration and estuarine distribution of the six trace metals, some

differences were found between the MD and the MG. The SPM loading in the MG

was a factor of 2 higher than in the MD. This fact did not cause a remarkable

difference of the trace metal distribution in the two estuaries. Although a release of

Cd during the dry seasons, of Pb during the surveys of 2007 and 2009 and of Co

during all seasons occurred, particulate trace metals behaved conservatively or even

remained constant. The transfer of the studied metals to the solution phase seems to

have a sediment-origin rather than release from the SPM. It might be partly attributed

to the different nature of the river inputs. The higher dissolved Cu and Ni in the MG at

S = 0 could also originals from the incoming river water. The same holds for the

different distribution of Cd in the two study areas during the wet season.

Seasonal differences of the concentration and the estuarine distribution of most


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investigated metals were observed in this study. The dissolved Cd was enriched

more strongly in the channel system during both dry seasons. The lower residence

time and/or the stronger dilution of saline water as a consequence of more

freshwater input during the wet season may inhibit the resolution of Cd from the

resuspended sediment. The dissolved Co level during both dry seasons was a factor

of 2 higher than during the wet season. The unexplainably higher concentration in the

freshwater input may originate such occurrence.

For all the investigated metals, a peak in the diagram of DOC and dissolved metals

vs. salinity was observed at s ≈ 10 in the MD during the dry season of 2009 with

more intensity for Cd, Fe and Pb than for the other metals. This may reflect the

influence of urban outputs from the city of Maceío for the MD (Brockmeyer and Spitzy,

2011). This addition of dissolved trace metals, however, was efficiently leveled out

during estuarine processing.

In the Mundaú-Manguaba Lagoon system, this work estimated an averaged

concentration of 0.03 nmol/L, 7.0 nmol/L, 1.5 μmol/L, 0.3 nmol/L, 1.6 nmol/L and 5.0

nmol/L for dissolved Cd, Cu, Fe, Pb, Co and Ni, respectively, and 0.4 nmol/g, 1.6

nmol/g, 5.7 μmol/g, 1.1 nmol/g, 1.1 nmol/g and 1.8 nmol/g for particulate Cd, Cu, Fe,

Pb, Co and Ni, respectively. Compared to other estuarine systems in South America

such as the Sepetiba Bay (e.g., de Lacerda et al., 1987), the Patos Lagoon (e.g.,

Windom et al., 1999) and the Paraíba do Sul Estuary (e.g., Carvalho et al., 2002) and

to estuaries worldwide such as the Mississippi River Estuary (Shiller and Boyle,

1991), the Gironde Estuary (Kraepiel et al., 1997) and the Changjiang Estuary (Wang

and Liu, 2003), both dissolved and particulate metal concentrations are on the same

level or even lower except Fe probably indicating the different geographic feature in

both study areas. The dissolved and particulate concentrations of investigated trace

metals suggest no significant contamination in the Mundaú-Manguaba Lagoon

system.

6.4. Conclusions

This study provides data along salinity gradients for dissolved and particulate

cadmium, copper, iron, lead, cobalt and nickel concentrations in the surface water of

the Mundaú-Manguaba lagoon system in Northeast Brazil during the wet season

(September/October 2007) and two dry seasons (February 2008 and March 2009).

Some universal estuarine behaviors for metals such as the formation of chloro-

cadmium-complexes, the co-precipitation of iron and lead, the conservative


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distribution of copper due to the strong association with organic materials, the

remobilization of Cobalt and the weak activity of Nickel were also observed in this

investigation. Cadmium underwent the formation of highly stable soluble chloro-

complexes, particularly at high salinities during both dry seasons. The suboxic

conditions in both lagoons possibly inhibited the release of dissolved cadmium from

the sediment. In the channel system, where dissolved oxygen was abundant, the

enrichment of dissolved cadmium occurred as in the most estuaries worldwide. But

during the wet season, both a precipitation in the Mundaú lagoon and a conservative

mixing in the Manguaba lagoon happened. In the MG, lead was likely released from

suspended particulate matter during the surveys of 2007 and 2009 due to higher

particulate matter loads despite its affinity to the particulate phase. Dissolved cobalt

was enriched at low salinity suggesting a release of cobalt from resuspended

sediment.

Distribution coefficients were calculated with the following order of the average

logarithms of KD´s: Fe (6.1 L/kg) > Pb (5.9 L/kg) > Co (4.9 L/kg) > Ni (4.8 L/kg) > Cu

(4.6 L/kg) > Cd (4.5 L/kg). Both the order and the values of the distribution

coefficients essentially agree with other estuarine studies. The distribution coefficient

of cadmium decreased with increasing salinity, while for lead and nickel it showed no

variations with salinity. For copper, iron and cobalt, the distribution coefficient was

positively correlated with salinity.

In addition, the colloidal partitioning of the trace metals and their relation to salinity

was investigated during the wet season of 2007 and the dry season of 2009. Most

metals had a significant presence in the colloidal fraction (5 kDa - 0.45 μm). The

significance of the truly dissolved fraction for almost all the metals increased with

salinity and was more important during the dry season than during the wet season. In

the tropical Mundaú-Manguaba lagoon system, the higher bioactivity and the longer

water residence time during the dry season may favor the degradation of high

molecular weight organo-bound metals and the release of metals from the sediment.

The released metals are probably associated with small organic ligands caused by

higher DOC contents during the dry season, thus enhancing their appearance in the

true dissolved fraction.

Acknowledgments

The authors wish to thank the “Bundesministerium für Bildung und Forschung (BMBF,

German Federal Ministry of Education and Research)” at Berlin and “the Brazilian


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Research Council (CNPq)” who funded our sub-project from the Project POLCAMAR

(POllution from sugar CAne in MARine systems) within the frame of the Brazilian-

German cooperation in science and technology [Marine Sciences Program,

BMBF/CNPq research project number.: 03F0455C (BMBF) and project nr.

590002/2005-8 (CNPq)]. We would like to thank all colleagues from the POLCAMAR

team, in special Professor Dr. Bastiaan Knoppers and Olaf Wilhelm for their

coordinating and managing efforts during the study in Brazil. We thank Dr. Uwe

Schüßler for his help with the ICP-MS analyses.

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Anhang

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2.63

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0.

22

5.9

4.9

5.3

4.9

4.6

5.8

W3s

18

.3

1.8

4.09

0.

95

13.0

7 1.

44

12.2

1 0.

07

0.31

0.

43

0.63

0.

22

0.79

3.

83

0.12

5.

7 4.

7 5.

2 4.

8 4.

8 5.

6 W

4s

15.5

1.

4 4.

18

0.80

6.

71

1.43

12

.65

0.06

0.

28

0.46

0.

72

0.26

0.

89

4.72

0.

15

5.8

5.0

5.3

4.8

4.9

5.7

W5

0.0

1.2

2.30

7.

26

10.9

6 4.

08

12.9

9 0.

02

0.72

1.

24

1.65

0.

61

1.28

4.

95

0.14

5.

2 5.

2 5.

2 5.

0 5.

4 5.

3 W

6 3.

4 1.

3 5.

56

5.56

15

.32

4.59

11

.48

0.04

0.

94

1.04

1.

38

0.50

1.

11

3.33

0.

18

5.3

5.0

5.0

5.0

4.9

5.3

W7

5.9

1.4

6.68

4.

89

13.1

6 3.

75

13.9

4 0.

05

0.95

0.

92

1.37

0.

61

1.18

3.

58

0.19

5.

3 5.

0 5.

2 4.

9 4.

9 5.

3 W

8 9.

7 0.

9 5.

34

2.31

9.

47

2.36

11

.51

0.11

1.

08

0.67

0.

88

0.31

0.

74

2.49

0.

15

5.5

5.0

5.1

4.8

4.4

5.1

W9

10.5

0.

9 6.

37

1.80

9.

48

1.81

12

.25

0.07

0.

49

0.41

0.

60

0.17

0.

53

2.87

0.

13

5.4

4.8

5.0

4.6

4.7

5.4

W10

12

.5

0.9

4.91

1.

81

7.48

1.

99

11.2

6 0.

08

0.46

0.

56

0.80

0.

38

0.84

3.

98

0.16

5.

5 5.

0 5.

3 4.

9 4.

8 5.

5 W

11

0.0

4.2

7.85

9.

71

5.58

5.

04

11.2

0 0.

04

1.53

0.

64

0.85

0.

34

0.69

3.

20

0.17

4.

8 5.

2 4.

8 4.

8 4.

9 5.

0 W

12

0.0

1.7

7.79

10

.81

5.18

3.

81

10.5

1 0.

02

2.04

0.

67

0.68

0.

28

0.55

2.

06

0.17

4.

8 5.

1 4.

9 4.

7 5.

1 4.

9 W

13

0.0

1.9

6.17

10

.64

5.43

3.

82

10.7

2 0.

02

1.92

0.

74

0.80

0.

32

0.73

2.

52

0.19

4.

8 5.

2 4.

9 4.

8 5.

3 5.

0 W

14

0.2

1.4

6.59

13

.72

6.56

4.

62

11.6

8 0.

03

2.24

0.

98

0.74

0.

29

0.63

2.

50

0.20

4.

9 5.

1 4.

8 4.

7 5.

0 4.

9 W

15

7.6

0.8

5.89

7.

25

11.3

2 3.

79

10.6

4 0.

08

1.35

0.

94

0.85

0.

24

1.07

3.

90

0.22

5.

1 4.

9 4.

8 5.

0 4.

7 5.

2

Anhang

- 158 -

Anha

ng 4

. Um

wel

tpar

amet

er, d

ie g

elös

te u

nd d

ie p

artik

ulär

e M

etal

lkon

zent

ratio

nen

sow

ie d

ie L

og K

D W

erte

in d

er R

egen

sais

on 2

007

in N

ordw

est-

Bra

silie

n (n

.d.:

nich

t det

ektie

rbar

).

U

mw

eltp

aram

eter

ge

löst

e M

etal

le

part

ikul

äre

Met

alle

Lo

gKD

Stat

ion

Salin

ität

SPM

D

OC

Fe

Ni

C

o C

u C

d Pb

Fe

Ni

C

o C

u C

d Pb

Fe

Ni

C

o C

u C

d Pb

psu

mg/

L m

g/L

μmol

/L

nmol

/L

nmol

/L

nmol

/L

nmol

/L

nmol

/L

mm

ol/g

μm

ol/g

μm

ol/g

μm

ol/g

nm

ol/g

μm

ol/g

L/

kg

L/kg

L/

kg

L/kg

L/

kg

L/kg

MD

285

0.0

5.3

0.73

5.

73

3.29

0.

79

3.72

0.

09

0.37

1.

58

0.23

0.

22

0.23

3.

17

0.34

5.

4 4.

8 5.

4 4.

8 4.

5 6.

0 28

7 0.

0 9.

1 2.

55

1.36

4.

85

0.34

7.

61

0.05

0.

16

1.36

0.

24

0.19

0.

22

7.27

0.

18

6.0

4.7

5.8

4.5

5.1

6.1

288

2.9

12.3

4.

20

0.11

9.

19

1.90

14

.00

0.06

0.

04

0.83

0.

34

0.21

0.

31

1.43

0.

15

6.9

4.6

5.0

4.4

4.4

6.6

289

4.9

75.5

3.

88

0.24

5.

89

1.36

10

.86

0.01

0.

06

0.62

0.

29

0.13

0.

19

1.27

0.

11

6.4

4.7

5.0

4.3

5.1

6.3

290

36.2

16

.2

3.77

0.

40

3.50

0.

24

2.12

0.

03

0.14

0.

41

0.23

0.

07

0.13

0.

41

0.08

6.

0 4.

8 5.

5 4.

8 4.

2 5.

8 29

1 20

.8

6.2

--

0.04

6.

99

1.23

6.

87

0.03

0.

18

0.57

0.

32

0.11

0.

22

0.43

0.

12

7.1

4.7

5.0

4.5

4.2

5.8

MG

264

19.4

22

.7

2.37

1.

22

4.77

0.

95

6.45

0.

02

0.47

0.

56

0.27

0.

12

0.19

0.

43

0.11

5.

7 4.

8 5.

1 4.

5 4.

3 5.

4 26

5 10

.0

11.3

3.

46

2.06

3.

55

1.03

5.

66

0.02

0.

53

0.68

0.

30

0.12

0.

20

0.54

0.

14

5.5

4.9

5.1

4.5

4.5

5.4

266

8.6

17.4

4.

55

1.41

6.

80

1.18

13

.12

0.01

0.

66

0.55

0.

26

0.11

0.

17

0.52

0.

11

5.6

4.6

5.0

4.1

4.6

5.2

267

4.8

26.8

4.

78

1.03

6.

34

0.93

11

.87

0.01

0.

40

0.60

0.

31

0.13

0.

20

0.67

0.

14

5.8

4.7

5.1

4.2

4.8

5.5

268

3.0

-- 2.

89

4.84

6.

93

3.09

10

.47

0.02

1.

04

0.62

0.

28

0.12

0.

23

0.63

0.

12

5.1

4.6

4.6

4.3

4.5

5.1

270

1.1

27.6

4.

41

2.34

4.

37

1.21

7.

37

0.01

0.

54

0.58

0.

26

0.13

0.

18

0.47

0.

11

5.4

4.8

5.0

4.4

4.5

5.3

271

0.0

32.7

--

0.

32

11.0

6 0.

68

20.6

1 0.

01

0.12

0.

57

0.31

0.

12

0.20

0.

52

0.13

6.

3 4.

4 5.

2 4.

0 4.

8 6.

0 27

2 0.

0 --

3.63

0.

46

8.93

1.

48

10.0

2 0.

01

0.17

1.

84

0.32

0.

29

0.26

0.

97

0.18

6.

6 4.

6 5.

3 4.

4 5.

1 6.

0 27

3 0.

0 30

.0

4.00

0.

85

8.07

0.

26

11.8

8 0.

01

0.18

1.

08

0.33

0.

23

0.26

0.

91

0.17

6.

1 4.

6 6.

0 4.

3 5.

2 6.

0

Sum

aúm

a

405

0.0

12.1

8.

22

21.5

0 9.

50

3.80

13

.86

0.02

2.

02

0.76

0.

20

0.11

0.

24

1.50

0.

32

4.5

4.3

4.4

4.2

4.9

5.2

Mar

ine

33

5 31

.7

2.9

2.88

0.

01

2.83

0.

41

3.25

0.

02

0.11

0.

54

0.29

0.

09

0.19

1.

00

0.17

7.

8 5.

0 5.

4 4.

8 4.

7 6.

2 33

6 29

.2

4.3

1.17

0.

06

3.50

0.

69

4.13

0.

02

0.02

0.

46

0.27

0.

09

0.16

0.

53

0.14

6.

9 4.

9 5.

1 4.

6 4.

4 6.

8

Anhang

- 159 -

Anha

ng 5

. U

mw

eltp

aram

eter

, di

e ge

löst

e un

d di

e pa

rtik

ulär

e M

etal

lkon

zent

ratio

nen

sow

ie d

ie L

og K

D W

erte

in

der

Troc

kens

aiso

n 20

08 i

n N

ordw

est-B

rasi

lien

(n.d

.: ni

cht d

etek

tierb

ar).

ge

löst

e M

etal

le

part

ikul

äre

Met

alle

Lo

gKD

Stat

ion

Salin

ität

SPM

D

OC

Fe

Ni

C

o C

u C

d Pb

Fe

Ni

C

o C

u C

d Pb

Fe

Ni

C

o C

u C

d Pb

psu

mg/

L m

g/L

μmol

/L

nmol

/L

nmol

/L

nmol

/L

nmol

/L

nmol

/L

mm

ol/g

μm

ol/g

μm

ol/g

μm

ol/g

nm

ol/g

μm

ol/g

L/

kg

L/kg

L/

kg

L/kg

L/

kg

L/kg

MD

590

36.2

2.

8 3.

61

0.36

2.

54

0.37

1.

83

0.06

0.

12

0.13

0.

06

0.01

0.

05

0.53

0.

04

5.5

4.4

4.6

4.4

4.0

5.5

591

35.5

0.

8 2.

39

0.23

1.

36

0.48

1.

42

0.04

0.

08

0.45

0.

09

0.12

0.

09

1.12

0.

09

6.3

4.8

5.4

4.8

4.4

6.0

592

26.3

1.

0 4.

07

0.82

3.

53

2.06

3.

68

0.06

0.

39

0.56

0.

25

0.20

0.

21

1.06

0.

13

5.8

4.8

5.0

4.7

4.3

5.5

593

18.6

10

.8

2.63

0.

05

0.08

1.

58

6.30

0.

00

0.33

0.

53

0.33

0.

17

0.22

0.

55

0.11

7.

0 6.

6 5.

0 4.

5 5.

4 5.

5 59

8 16

.8

5.0

4.01

0.

08

n.d.

2.

27

5.07

0.

00

n.d.

0.

32

0.31

0.

17

0.15

0.

55

0.08

6.

6 --

4.9

4.5

5.6

-- 60

1 10

.6

11.2

3.

54

0.14

3.

26

2.73

5.

35

0.01

0.

17

0.91

0.

30

0.21

0.

24

0.65

0.

13

6.8

5.0

4.9

4.7

4.8

5.9

602

1.2

3.8

10.8

3 4.

08

2.77

3.

44

8.74

0.

00

0.27

1.

24

0.31

0.

18

0.30

0.

52

0.17

5.

5 5.

1 4.

7 4.

5 5.

2 5.

8 60

3 4.

3 5.

3 4.

45

2.81

2.

25

3.60

5.

51

0.01

0.

24

1.07

0.

27

0.17

0.

27

0.51

0.

14

5.6

5.1

4.7

4.7

5.0

5.8

680

0.3

18.1

2.

24

0.13

0.

73

1.83

7.

80

0.00

0.

01

1.28

0.

37

0.20

0.

36

0.94

0.

17

7.0

5.7

5.0

4.7

5.6

7.5

692

0.0

-- 4.

88

4.37

3.

85

2.09

9.

92

0.02

0.

67

-- --

-- --

-- --

-- --

-- --

-- --

MG

623

22.2

20

.1

6.19

0.

42

3.07

2.

27

5.06

0.

05

0.16

0.

60

0.27

0.

12

0.22

0.

50

0.13

6.

2 4.

9 4.

7 4.

6 4.

0 5.

9 62

5 24

.0

18.2

3.

44

0.03

2.

84

1.95

5.

06

0.05

0.

13

0.48

0.

27

0.12

0.

18

0.54

0.

14

7.2

5.0

4.8

4.6

4.1

6.0

626

31.0

1.

2 3.

24

0.96

3.

91

1.32

3.

87

0.07

0.

29

0.40

0.

20

0.09

0.

14

0.93

0.

10

5.6

4.7

4.8

4.6

4.1

5.5

627

35.7

--

1.67

0.

73

1.30

0.

49

2.65

0.

07

0.16

0.

32

0.18

0.

07

0.11

0.

48

0.08

5.

6 5.

1 5.

1 4.

6 3.

9 5.

7 62

8 31

.7

4.9

2.81

0.

06

1.28

0.

91

2.14

0.

04

0.06

0.

31

0.22

0.

06

0.11

0.

28

0.08

6.

7 5.

2 4.

8 4.

7 3.

9 6.

1 63

0 13

.8

4.1

6.63

0.

43

2.99

2.

33

4.96

0.

06

0.14

0.

42

0.22

0.

10

0.17

0.

92

0.11

6.

0 4.

9 4.

6 4.

5 4.

2 5.

9 63

3 0.

0 10

.1

2.47

6.

03

2.66

1.

60

3.99

n.

d.

0.21

1.

52

0.16

0.

08

0.20

0.

70

0.26

5.

4 4.

8 4.

7 4.

7 --

6.1

637

6.1

7.5

7.33

0.

99

1.17

2.

44

6.36

0.

01

0.31

0.

08

0.05

0.

04

0.05

0.

29

0.02

4.

9 4.

6 4.

2 3.

9 4.

4 4.

8 63

9 2.

6 5.

2 5.

01

1.82

7.

10

2.75

3.

88

0.04

0.

20

1.12

0.

20

0.20

0.

34

0.93

0.

08

5.8

4.5

4.9

4.9

4.3

5.6

641

4.5

10.0

5.

46

1.46

5.

47

3.27

11

.17

0.01

0.

30

0.48

0.

30

0.11

0.

23

0.49

0.

11

5.5

4.7

4.5

4.3

4.5

5.6

691

0.0

-- 4.

88

0.96

8.

39

2.38

20

.52

0.02

0.

32

-- --

-- --

-- --

-- --

-- --

-- --

Sum

aúm

a

542

0.0

11.3

4.

04

0.87

0.

64

0.96

6.

94

0.00

0.

03

0.88

0.

16

0.15

0.

23

1.46

0.

25

6.0

5.4

5.2

4.5

5.9

7.0

690

0.0

7.9

2.79

0.

12

4.24

1.

08

7.27

0.

01

0.04

--

0.36

0.

25

0.42

2.

43

0.25

--

4.9

5.4

4.8

5.4

6.8

Mar

ine

66

5 37

.1

2.0

1.16

0.

01

1.71

0.

18

2.10

0.

08

0.07

0.

26

0.18

0.

04

0.11

1.

32

0.07

7.

3 5.

0 5.

3 4.

7 4.

2 5.

9 66

6 35

.4

2.3

4.41

0.

02

2.37

0.

42

2.09

0.

09

0.07

0.

29

0.22

0.

05

0.12

0.

67

0.09

7.

3 5.

0 5.

0 4.

8 3.

9 6.

1 66

9 37

.3

2.0

1.58

0.

56

2.90

0.

29

4.07

0.

10

0.34

0.

37

0.23

0.

04

0.32

0.

98

0.13

5.

8 4.

9 5.

2 4.

9 4.

0 5.

6

Anhang

- 160 -

Anha

ng 6

. U

mw

eltp

aram

eter

, di

e ge

löst

e un

d di

e pa

rtik

ulär

e M

etal

lkon

zent

ratio

nen

sow

ie d

ie L

og K

D W

erte

in

der

Troc

kens

aiso

n 20

09 i

n N

ordw

est-B

rasi

lien

(n.d

.: ni

cht d

etek

tierb

ar).

ge

löst

e M

etal

le

part

ikul

äre

Met

alle

Lo

gKD

Stat

ion

Salin

ität

SPM

D

OC

Fe

Ni

C

o C

u C

d Pb

Fe

Ni

C

o C

u C

d Pb

Fe

Ni

C

o C

u C

d Pb

psu

mg/

L m

g/L

μmol

/L

nmol

/L

nmol

/L

nmol

/L

nmol

/L

nmol

/L

mm

ol/g

μm

ol/g

μm

ol/g

μm

ol/g

nm

ol/g

μm

ol/g

L/

kg

L/kg

L/

kg

L/kg

L/

kg

L/kg

MD

940

27.5

7.

5 2.

69

0.11

5.

42

1.03

4.

60

0.08

0.

25

0.51

0.

25

0.07

0.

20

0.57

0.

12

6.6

4.7

4.8

4.6

3.9

5.7

941

24.5

9.

5 2.

84

0.16

5.

49

1.40

5.

65

0.04

0.

17

0.51

0.

28

0.09

0.

22

0.52

0.

13

6.5

4.7

4.8

4.6

4.1

5.9

942

20.6

9.

6 3.

30

0.15

6.

17

1.65

5.

98

0.03

0.

12

0.47

0.

30

0.09

0.

22

0.51

0.

14

6.5

4.7

4.7

4.6

4.3

6.1

943

14.2

6.

8 7.

72

0.05

5.

56

1.58

6.

61

0.02

0.

10

0.42

0.

27

0.12

0.

18

0.75

0.

12

6.9

4.7

4.9

4.4

4.7

6.1

944

11.8

3.

7 5.

97

1.54

9.

25

1.78

7.

16

0.04

0.

77

0.39

0.

31

0.15

0.

19

0.71

0.

11

5.4

4.5

4.9

4.4

4.3

5.2

946

10.0

3.

7 5.

62

4.18

7.

93

3.03

9.

05

0.04

1.

20

0.43

0.

28

0.19

0.

20

1.10

0.

12

5.0

4.5

4.8

4.3

4.4

5.0

948

5.6

7.1

5.29

0.

16

7.46

2.

60

10.2

5 0.

02

0.08

0.

58

0.31

0.

16

0.20

1.

16

0.13

6.

6 4.

6 4.

8 4.

3 4.

7 6.

2 95

0 3.

0 13

.9

5.82

0.

53

8.66

2.

51

10.1

9 0.

02

0.09

0.

96

0.22

0.

19

0.19

1.

30

0.11

6.

3 4.

4 4.

9 4.

3 4.

8 6.

1 95

1 0.

1 19

.9

9.02

3.

58

13.4

9 2.

36

9.45

0.

02

0.32

1.

78

0.32

0.

39

0.27

1.

06

0.19

5.

7 4.

4 5.

2 4.

5 4.

8 5.

8

MG

900

34.0

12

.8

6.11

0.

33

15.0

4 0.

49

2.61

0.

10

0.55

0.

35

0.24

0.

08

0.13

0.

67

0.07

6.

0 4.

2 5.

2 4.

7 3.

8 5.

1 90

2 17

.3

14.5

4.

57

0.10

5.

77

2.01

6.

95

0.04

1.

35

0.59

0.

24

0.09

0.

16

0.59

0.

13

6.8

4.6

4.6

4.4

4.1

5.0

903

8.1

13.2

5.

11

1.96

5.

06

2.21

6.

47

0.03

0.

47

0.59

0.

26

0.11

0.

19

0.78

0.

15

5.5

4.7

4.7

4.5

4.4

5.5

904

2.2

20.2

7 6.

16

0.95

4.

17

2.05

4.

02

0.02

0.

14

0.94

0.

20

0.17

0.

17

0.83

0.

18

6.0

4.7

4.9

4.6

4.7

6.1

905

0.0

19.0

12

.10

1.26

4.

52

1.65

3.

20

0.01

0.

13

1.61

0.

15

0.20

0.

17

0.93

0.

24

6.1

4.5

5.1

4.7

4.9

6.3

907

3.8

36.6

4.

69

0.66

6.

84

3.00

9.

68

0.02

0.

24

0.72

0.

17

0.05

0.

11

0.71

0.

05

6.0

4.4

4.2

4.1

4.6

5.4

908

0.0

26.7

7.

79

1.16

12

.66

1.65

13

.93

0.02

0.

31

1.29

0.

28

0.33

0.

27

0.77

0.

16

6.0

4.3

5.3

4.3

4.6

5.7

Sum

aúm

a

983

0.0

36.0

2.

85

0.93

1.

85

0.22

4.

87

n.d.

0.

08

-- 0.

16

0.13

0.

23

1.75

0.

27

-- 4.

9 5.

8 4.

7 --

6.5

Mar

ine

99

5 36

.8

4.1

3.69

n.

d.

n.d.

n.

d.

0.16

0.

03

0.05

0.

26

0.22

0.

06

0.09

0.

73

0.06

--

-- --

5.8

4.4

6.1

Anhang

- 161 -

Anha

ng 7

. Met

allk

onze

ntra

tione

n in

der

HM

W-,

der L

MW

- und

der

TD

-Fra

ktio

n so

wie

die

Mas

senb

ilanz

en in

der

Tro

cken

sais

on 2

006

in O

st-H

aina

n (n

.d.:

nich

t det

ektie

rbar

).

HM

W-F

rakt

ion

LMW

-Fra

ktio

n

TD-F

rakt

ion

Mas

senb

ilanz

St

atio

n Fe

Ni

C

o C

u C

d Pb

Fe

Ni

C

o C

u C

d Pb

Fe

Ni

C

o C

u C

d Pb

Fe

Ni

C

o C

u C

d Pb

μmol

/L

nmol

/L

nmol

/L

nmol

/L

nmol

/L

nmol

/L μm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

μm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

WR

A1

1.

67

-- --

1.50

0.

01

0.38

0.

00

-- --

1.63

0.

00

0.01

n.

d.

-- --

n.d.

0.

01

0.02

0.

48

-- --

0.72

0.

64

0.56

B1

--

-- --

-- --

-- --

-- --

-- --

-- --

-- --

-- --

-- --

-- --

-- --

-- B2

0.

09

-- --

-- 0.

00

-- 0.

00

-- --

-- 0.

00

-- 0.

00

-- --

-- 0.

05

-- 0.

57

-- --

-- 0.

55

-- B3

0.

62

0.31

0.

17

0.21

0.

00

0.28

0.

00

1.56

0.

11

2.19

0.

00

0.01

0.

00

2.29

0.

21

2.31

0.

04

0.01

0.

74

0.72

1.

00

0.90

0.

46

0.65

B4

1.

11

0.51

0.

29

0.62

0.

00

-- 0.

01

0.60

0.

18

1.85

0.

00

-- 0.

00

2.48

0.

35

3.30

0.

03

-- 1.

02

0.93

1.

05

0.78

0.

56

-- B5

--

-- 0.

10

0.23

0.

00

-- --

-- 0.

06

0.83

0.

00

-- --

-- 0.

20

1.98

0.

06

-- --

-- 1.

11

0.76

0.

63

-- B6

--

0.34

0.

27

0.47

0.

00

0.26

--

0.40

0.

13

1.49

0.

00

0.01

--

3.14

0.

39

3.33

0.

04

0.02

--

0.76

0.

71

0.81

0.

62

0.46

B9

3.

84

-- 1.

53

3.18

--

-- 0.

00

-- 0.

17

2.21

--

-- n.

d.

-- 0.

22

0.18

--

-- 1.

03

-- 1.

30

0.81

--

--

WW

R

W1

0.29

--

0.18

0.

70

0.00

0.

10

0.00

--

0.09

1.

46

0.00

0.

00

0.00

--

0.33

3.

92

0.06

0.

02

0.84

--

0.90

0.

80

0.78

0.

47

W2s

--

1.12

0.

54

2.01

0.

01

-- --

0.43

0.

20

1.86

0.

00

-- --

5.48

0.

45

5.42

0.

07

-- --

0.76

0.

93

0.77

0.

68

-- W

3s

-- --

0.38

1.

48

0.01

--

-- --

0.24

1.

69

0.00

--

-- --

0.42

5.

52

0.03

--

-- --

0.72

0.

71

0.64

--

W4s

0.

38

0.88

0.

37

2.11

0.

02

0.12

0.

00

1.78

0.

49

1.71

0.

00

0.00

0.

00

3.67

0.

48

5.82

0.

03

0.01

0.

48

0.94

0.

93

0.76

0.

73

0.45

W

5 3.

67

2.60

2.

30

4.40

--

0.45

0.

00

3.18

0.

42

3.51

--

0.00

n.

d.

4.82

0.

46

2.07

--

0.00

0.

51

0.97

0.

78

0.77

--

0.63

W

6 --

2.80

1.

64

1.82

0.

01

0.35

--

0.08

0.

00

2.46

0.

00

0.00

--

6.92

2.

08

4.06

0.

02

0.01

--

0.64

0.

81

0.73

0.

71

0.39

W

7 3.

12

3.30

2.

42

2.59

0.

02

0.55

0.

00

1.69

0.

71

1.76

0.

00

0.00

0.

00

6.25

0.

64

4.91

0.

02

0.01

0.

64

0.86

1.

00

0.66

0.

73

0.58

W

8 1.

73

1.88

1.

08

1.51

--

-- 0.

00

1.36

0.

54

2.68

--

-- 0.

00

5.32

0.

71

6.22

--

-- 0.

75

0.90

0.

99

0.90

--

-- W

9 1.

32

1.34

0.

62

1.23

0.

02

0.30

0.

00

1.12

0.

43

2.47

0.

00

0.00

0.

00

5.12

0.

73

7.23

0.

02

0.01

0.

73

0.80

0.

98

0.89

0.

58

0.64

W

10

-- 1.

25

0.53

2.

19

0.01

--

-- 1.

24

0.42

2.

45

0.00

--

-- 3.

44

0.65

5.

07

0.06

--

-- 0.

79

0.81

0.

86

0.95

--

W12

7.

16

-- 2.

72

4.32

--

1.44

0.

00

-- 0.

57

3.04

--

0.00

n.

d.

-- 0.

37

2.09

--

0.02

0.

66

-- 0.

96

0.90

--

0.72

W

15

-- 1.

91

0.64

1.

17

0.01

0.

48

-- 1.

93

0.63

2.

67

0.00

0.

00

-- 4.

33

1.78

4.

31

0.04

0.

05

-- 0.

72

0.80

0.

77

0.66

0.

39

Anhang

- 162 -

Anha

ng 8

. Met

allk

onze

ntra

tione

n in

der

HM

W-,

der

LMW

- un

d de

r TD

-Fra

ktio

n so

wie

die

Mas

senb

ilanz

en in

der

Reg

ensa

ison

200

7 in

Nor

dwes

t-B

rasi

lien

(n.d

.: ni

cht d

etek

tierb

ar).

HMW

-Fra

ktio

n LM

W-F

rakt

ion

TD

-Fra

ktio

n M

asse

nbila

nz

Stat

ion

Fe

Ni

Co

Cu

Cd

Pb

Fe

Ni

Co

Cu

Cd

Pb

Fe

Ni

Co

Cu

Cd

Pb

Fe

Ni

Co

Cu

Cd

Pb

μm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

μm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

μm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

MD

28

5 2.

68

-- 0.

72

-- --

0.23

0.

02

-- 0.

13

-- --

0.05

0.

01

-- 0.

06

-- --

0.05

0.

47

-- 1.

15

-- --

0.90

28

7 --

1.04

0.

09

2.10

0.

00

0.08

--

2.65

0.

26

3.87

0.

01

0.05

--

2.63

0.

10

2.97

0.

02

0.03

--

1.30

1.

30

1.18

0.

50

1.03

28

8 --

0.64

0.

39

1.81

--

-- --

2.25

0.

79

4.29

--

-- --

4.13

0.

71

5.19

--

-- --

0.76

1.

00

0.81

--

-- 28

9 --

0.44

0.

31

2.62

0.

01

-- --

2.19

0.

68

3.69

0.

00

-- --

4.45

0.

75

4.72

0.

00

-- --

1.20

1.

28

1.02

1.

05

-- 29

0 --

-- --

-- --

-- --

-- --

-- --

-- --

-- --

-- --

-- --

-- --

-- --

-- 29

1 --

0.00

0.

03

0.73

0.

01

0.04

--

1.51

0.

53

3.00

0.

01

0.00

--

8.93

1.

00

5.15

0.

02

0.01

--

1.49

1.

27

1.29

1.

05

0.33

MG

26

4 0.

47

-- 0.

20

1.00

0.

00

0.12

0.

02

-- 0.

25

2.28

0.

00

0.01

0.

00

-- 0.

68

5.42

0.

01

0.01

0.

40

-- 1.

19

1.35

0.

93

0.30

26

5 2.

31

-- 0.

86

-- --

0.39

0.

00

-- 0.

35

-- --

0.00

0.

01

-- 0.

11

-- --

0.16

1.

12

-- 1.

28

-- --

1.03

26

6 --

-- 0.

18

2.40

0.

00

-- --

-- 0.

37

2.66

0.

00

-- --

-- 0.

68

6.68

0.

00

-- --

-- 1.

05

0.89

0.

66

-- 26

7 0.

76

-- --

-- 0.

00

-- 0.

01

-- --

-- 0.

00

-- 0.

00

-- --

-- 0.

01

-- 0.

75

-- --

-- 0.

90

-- 26

8 2.

47

-- --

2.55

0.

01

0.57

0.

00

-- --

4.28

0.

00

0.01

0.

06

-- --

7.37

0.

00

0.01

0.

52

-- --

1.36

0.

67

0.57

27

0 --

-- 0.

19

-- --

-- --

-- 0.

79

-- --

-- --

-- 0.

49

-- --

-- --

-- 1.

21

-- --

-- 27

1 --

-- --

2.42

0.

00

0.11

--

-- --

10.5

7 0.

00

0.00

--

-- --

3.35

0.

00

n.d.

--

-- --

0.79

1.

26

1.00

27

2 0.

24

-- --

3.87

0.

00

-- 0.

05

-- --

5.08

0.

00

-- 0.

00

-- --

4.09

0.

00

-- 0.

64

-- --

1.30

0.

89

-- 27

3 0.

66

2.08

--

5.31

--

0.14

0.

03

3.54

--

4.63

--

0.02

0.

03

5.31

--

3.19

--

0.00

0.

84

1.36

--

1.10

--

0.84

Anhang

- 163 -

Anha

ng 9

. Met

allk

onze

ntra

tione

n in

der

HM

W-,

der L

MW

- und

der

TD

-Fra

ktio

n so

wie

die

Mas

senb

ilanz

en in

der

Tro

cken

sais

on 2

009

in N

ordw

est-

Bra

silie

n (n

.d.:

nich

t det

ektie

rbar

).

H

MW

Fra

ktio

n

LMW

Fra

ktio

n

TD F

rakt

ion

M

asse

nbila

nz

Stat

ion

Fe

Ni

Co

Cu

Cd

Pb

Fe

Ni

Co

Cu

Cd

Pb

Fe

Ni

Co

Cu

Cd

Pb

Fe

Ni

Co

Cu

Cd

Pb

μm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

μm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

μm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

nm

ol/L

M

D

940

-- 0.

44

0.04

0.

42

0.01

0.

18

-- 1.

01

0.22

1.

45

0.00

0.

03

-- 3.

72

0.88

3.

20

0.05

0.

12

-- 0.

95

1.10

1.

10

0.74

1.

33

941

0.08

0.

34

0.01

0.

25

0.00

0.

02

0.02

1.

22

0.31

1.

84

0.00

0.

02

0.02

3.

38

1.00

2.

52

0.03

0.

07

0.76

0.

90

0.94

0.

82

0.89

0.

60

942

0.11

0.

19

0.02

0.

34

-- 0.

05

0.05

1.

59

0.46

2.

15

-- 0.

02

0.01

4.

32

1.19

3.

62

-- 0.

06

1.12

0.

99

1.01

1.

02

-- 1.

13

943

0.05

0.

00

0.04

0.

28

-- 0.

04

0.01

1.

38

0.55

2.

47

-- 0.

01

0.05

5.

76

1.15

3.

95

-- 0.

07

1.49

1.

29

1.10

1.

01

-- 1.

08

944

-- 0.

40

0.25

0.

65

-- --

-- 1.

80

0.66

2.

79

-- --

-- 3.

94

0.90

3.

35

-- --

-- 0.

66

1.01

0.

95

-- --

946

2.33

1.

61

0.96

--

0.00

--

0.02

2.

21

0.77

--

0.00

--

0.01

3.

08

0.86

--

0.05

--

0.56

0.

87

0.86

--

1.19

--

948

-- 0.

65

0.16

0.

88

0.01

--

-- 2.

51

1.16

4.

35

0.00

--

-- 3.

77

1.10

4.

09

0.01

--

-- 0.

93

0.93

0.

91

0.84

--

950

-- 1.

71

1.03

1.

05

0.01

--

-- 3.

15

1.02

6.

94

0.00

--

-- 3.

88

0.96

4.

82

0.01

--

-- 1.

01

1.20

1.

26

1.14

--

951

-- 6.

24

-- 5.

16

0.01

--

-- 5.

15

-- 4.

79

0.00

--

-- 3.

57

-- 2.

79

0.01

--

-- 1.

11

-- 1.

35

1.40

--

MG

90

0 --

-- --

-- --

0.00

--

-- --

-- --

0.00

--

-- --

-- --

0.56

--

-- --

-- --

1.02

90

2 --

0.27

0.

06

0.79

0.

00

-- --

1.28

0.

47

2.10

0.

00

-- --

4.69

1.

54

3.53

0.

04

-- --

1.08

1.

04

0.92

1.

05

-- 90

3 1.

12

0.36

0.

28

0.52

0.

00

-- 0.

03

0.91

0.

56

1.92

0.

00

-- 0.

02

4.97

1.

62

4.35

0.

04

-- 0.

59

1.23

1.

11

1.05

1.

22

-- 90

4 0.

49

0.62

0.

08

0.43

0.

00

0.08

0.

09

1.65

0.

66

1.75

0.

00

0.01

0.

02

2.56

1.

68

2.14

0.

02

0.07

0.

64

1.16

1.

18

1.07

1.

08

1.15

90

5 --

1.13

--

1.34

0.

01

-- --

1.23

--

1.24

0.

00

-- --

3.35

--

1.47

0.

01

-- --

1.26

--

1.26

1.

29

-- 90

7 0.

32

1.18

0.

23

1.12

0.

01

0.12

0.

01

1.67

1.

26

3.53

0.

00

0.01

0.

01

2.92

1.

25

3.85

0.

01

0.06

0.

52

0.84

0.

92

0.88

0.

88

0.76

90

8 --

4.34

--

6.61

--

-- --

4.39

--

5.47

--

-- --

4.49

--

3.44

--

-- --

1.04

--

1.11

--

--

Documents

Untersuchungen zur Speziation von verschmutzungsrelevanten ... · Ost-Hainan generell höher als in Nordost-Brasilien. Das ästuarine Verhalten der Das ästuarine Verhalten der Metalle