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ECOLOGICAL RISK ASSESSMENT OF
TRACE METALS IN SEDIMENTS ALONG
THE EGYPTIAN MEDITERRANEAN COAST
A Thesis
Submitted to
Institute of Graduate Studies and Research
Alexandria University
In Partial Fulfillment of the
Requirements for the Degree of
Doctor of Philosophy
In
Environmental Studies
By
Naglaa Farag Elsayed Soliman
B.Sc. Chemistry & Oceanography, 1997
Faculty of Science
M.Sc. in Environmental Studies, 2008
Alexandria University
Department of Environmental Studies
Institute of Graduate Studies and Research
Alexandria University
2012
ECOLOGICAL RISK ASSESSMENT OF
TRACE METALS IN SEDIMENTS ALONG
THE EGYPTIAN MEDITERRANEAN COAST
Presented by
Naglaa Farag Elsayed Soliman
Examiners Committee:
Approved
Prof. Dr. Mamdouh Saad Masoud
Professor of Chemistry
Department of Chemistry
Faculty of Science
…………………..
Alexandria University
Prof. Dr. Mohamed Mohamed Attia Shreadah
Professor of Marine Chemistry
President of the National Institute of Oceanography and Fisheries
Alexandria
…………………..
Prof. Dr. Samir Mahmoud Nasr
Professor of Marine Geochemistry
Department of Environmental Studies
Institute of Graduate Studies and Research
Alexandria University
Date
/
/ 2012
…………………..
Advisors’ Committee:
Prof. Dr. Samir Mahmoud Nasr
…………
Professor of Marine Geochemistry,
Department of Environmental Studies,
Institute of Graduate Studies and Research,
Alexandria University
Dr. Mohammed Abd El-Hamid Khairy
Lecturer,
Department of Environmental Science,
Faculty of Science,
Alexandria University
……...…
ACKNOWLEDGEMENT
First of all, I thank God for helping me to accomplish this research and for
providing me with such very encouraging and supporting supervisors.
I wish to express my deepest appreciation and sincere gratitude to Professor Dr.
Samir Mahmoud Nasr, Professor of Marine Geochemistry, Department of
Environmental Studies, Institute of Graduate Studies and Research, University of
Alexandria for the suggestion of the research subject and the ideas, his keen interest,
guidance and valuable discussions and his cooperation and continuous encouragement he
has given to me to accomplish the work. It is honor working under his supervision.
I would like to express my sincere thanks to Dr. Mohammed Abdel Hamid
Khairy, Lecturer, Department of Environmental Science, Faculty of Science, University
of Alexandria, for his great support, guidance, supervision and advice throughout the
study and his encouragement during the preparation of the manuscript.
No words can express my sincere gratitude to Prof. Dr. Mohammed Abdel Aziz
Okbah, Professor of Marine Chemistry, National Institute of Oceanography and
Fisheries, for his detailed and constructive comments, and for his important support
throughout this work and for his detailed review, constructive criticism and excellent
advice during the preparation of this work.
I would like also to express my deepest thanks to Mr. Mohammed Abdel Khaleq
for his help, support and encouragement at the beginning of my practical work. I would
like to thank him very much for his assistance on the field and in the laboratory during
the project period.
Last of all, I am greatly indebted to my family. Thanks to my parents, for giving
me life in the first place, for educating me for unconditional support and encouragement
to pursue my interests.
Table of Contents
Page
ACKNOWLEDGMENT ....................................................................... i
TABLE OF CONTENT ........................................................................ ii
LIST OF TABLES ................................................................................. iii
LIST OF FIGURES ............................................................................... iv
LIST OF ABBREVIATIONS ............................................................... v
I- INTRODUCTION ............................................................................. 1
II- REVIEW OF LITERATURE ......................................................... 4
2.1. Contaminants in the environment ............................................................ 4
2.2. Trace metals ................................................................................................ 4
2.2.1. Sources, uses and toxicity of trace metals .................................................. 6
2.2.2. Pollution of the aquatic environment with trace metals ........................... 18
2.2.2.1. Trace metals in the water column ........................................................................... 18
2.2.2.2. Metals in seawater .................................................................................................. 20
2.2.2.3. Metals in marine biota ............................................................................................ 21
2.2.2.4. Metals in marine sediments .................................................................................... 22
2.2.3. The importance of studying sediments ....................................................... 23
2.2.4. Effects of trace metals contamination in sediments .................................. 26
2.2.4.1. Metals polluted sediments ...................................................................................... 26
2.2.4.2. Ecology ................................................................................................................... 26
2.2.4.3. Human health.......................................................................................................... 27
2.2.5. Factors affecting trace metal distribution, concentration and fate in
sediments ................................................................................................................. 27
2.2.5.1. Trace metals input into sediment ............................................................................ 28
2.2.5.2. The influence of grain size ..................................................................................... 28
2.2.5.3. Sediment composition ............................................................................................ 29
2.2.5.4. The influence of pH values..................................................................................... 30
2.2.5.5. Influence of organic matter .................................................................................... 30
2.2.5.6. The influence of oxidation reduction potential ...................................................... 31
2.2.5.7. Adsorption/desorption reactions ............................................................................. 32
2.2.5.8 Physical transport .................................................................................................... 33
2.2.5.9. Bioturbation ............................................................................................................ 34
2.2.5.10. The influence of other species .............................................................................. 34
2.2.6. Analysis of trace metals in sediments ......................................................... 35
2.2.6.1 Acid volatile sulphide ................................................................................. 36
2.2.6.2. Sequential extraction procedure .............................................................. 37
2.3. Environmental problems and previous studies on the coastal area along
the Mediterranean Sea ..................................................................................... 41
2.4. Ecological risk assessment ......................................................................... 53
2.4.1. Ecological risk terms .................................................................................... 53
2.4.2. Phases of Ecological risk assessment .......................................................... 53
2.4.3. Ecological risk assessment in a management context ............................... 54
2.4.4. Contributions of Ecological Risk Assessment to Environmental Decision
making .................................................................................................................... 56
2.4.5. Factors Affecting the Value of Ecological Risk Assessment for
Environmental decision making ............................................................................ 56
III- SOCIOECONOMIC AND ENVIRONMENTAL ASPECTS ... 57
3.1. Description of the Egyptian Mediterranean environment coastal
environment ....................................................................................................... 57
3.1.1. History and culture....................................................................................... 57
3.1.2. Climate........................................................................................................... 57
3.1.3. Topography ................................................................................................... 58
3.1.4. Subsidence and sea level rise ....................................................................... 59
3.1.5. Current .......................................................................................................... 59
3.1.6. Wind............................................................................................................... 59
3.1.7. Waves ............................................................................................................. 61
3.1.8. Tides and storm surges ................................................................................ 61
3.1.9. Hydrology ...................................................................................................... 61
3.1.10. Biological diversity ..................................................................................... 62
3.1.11. Ecosystem features ..................................................................................... 62
3.12. Water quality ................................................................................................ 62
3.13. Sediment characteristics ............................................................................... 63
3.2. Economic activities of the Egyptian Mediterranean coastal area ......... 64
3.2.1 Population ...................................................................................................... 65
3.2.2. Main activities sectors ............................................................................... .. 66
3.2.3. Land use ..................................................................................................... .. 69
3.2.4. The economic importance of the coastal lakes ........................................ ..70
3.2.5. The economic importance of the Egyptian Mediterranean ports ......... ..71
3.3. Environmental aspects ............................................................................. ..73
IV- MATERIALS & METHODS
.................................................................................................................. 7
5
4.1. Study area ................................................................................................. ..75
4.2. Sample collection, preservation and storage ......................................... ..75
4.3. Sample pre-treatment .............................................................................. ..75
4.4. Physicochemical characterization of sediment ..................................... ..78
4.4.1. Determination of water content 9%WC) ................................................ ..78
4.4.2. Determination of pH.................................................................................. ..78
4.4.3. Determination of grain size analysis ........................................................ ..78
4.4.4. Determination of total carbonate ............................................................. ..78
4.4.5. Determination of total organic carbon (TOC) ........................................ ..79
4.5. Determination of total metals ................................................................. ..79
4.5.1. Sample preparation ................................................................................... ..79
4.5.2. Measuring system ...................................................................................... ..79
4.5.3 Analysis of certified reference materials .................................................. ..80
4.6. Determination of leachable metals ......................................................... ..80
4.7. Sequential extraction procedure for trace metals ............................... ..81
4.8. SEM/AVS analysis procedure ................................................................. ..83
4.8.1. Reagents...................................................................................................... ..83
4.8.2. Generation of H2S ...................................................................................... ..83
4.8.3. Procedure ................................................................................................... ..83
4.8.4. Analysis of sulfide ...................................................................................... ..84
4.8.5. Calculation of AVS .................................................................................... ..84
4.8.6. Determination of simultaneously extractable metals (SEM)................. ..84
V- RESULTS ....................................................................................... ..86
5.1. Physico-chemical parameters ................................................................ ..86
5.1.1. Seawater temperature ..................................................................................86
5.2.1. pH value .........................................................................................................86
5.3.1. Salinity ...........................................................................................................86
5.1.4. Dissolved oxygen ...........................................................................................86
5.2. Geochemical Analysis ............................................................................. ..88
5.2.1. pH value .........................................................................................................88
5.2.2. Calcium carbonate (CaCO3)........................................................................88
5.2.3. Total organic matter (TOM%) ...................................................................88
5.2.4. Water content (%WC) .................................................................................88
5.2.5. Grain size .......................................................................................................88
5.3. Trace metals distribution ....................................................................... ...91
5.3.1. Total metals ................................................................................................... 91
5.3.1.1. Total Iron (TFe) ...................................................................................................... 91
5.3.1.2. Total Manganese (TMn) ..................................................................................... …91
5.3.1.3. Total Zinc (TZn) ..................................................................................................... 92
5.3.1.4. Total Copper (Cu)................................................................................................... 92
5.3.1.5. Total Nickel (Ni) .................................................................................................... 93
5.3.1.6. Total Chromium (Cr).............................................................................................. 93
5.3.1.7. Total Cobalt (Co) .................................................................................................... 93
5.3.1.8. Total Lead (Pb) ....................................................................................................... 94
5.3.1.9. Total Cadmium (Cd)............................................................................................... 94
5.3.2. Leachable Metals ......................................................................................100
5.3.2.1. Leachable Iron (Fe) ............................................................................................ 100
5.3.2.2 Leachable Mnganese (Mn) .................................................................................. 100
5.2.2.3. Leachable Zinc (Zn) ........................................................................................... 100
5.2.2.4. Leachable Cupper (Cu)....................................................................................... 101
5.2.2.5 Leachable Nickel (Ni) ......................................................................................... 101
5.2.2.6 Leachable Cr (Cr) ................................................................................................ 101
5.2.2.7 Leachable Co (Co) ............................................................................................... 101
5.2.2.8 Leachable Lead (Pb) ............................................................................................ 102
5.2.2.9. Leachable Cadmium (Cd)................................................................................... 102
5.3.3. Partitioning of Metals ..............................................................................107
5.3.3.1. Fractionation of Iron (Fe) ................................................................................... 107
5.3.3.2. Fractionation of Manganese (Mn) ...................................................................... 108
5.3.3.3. Fractionation of Zinc (Zn) .................................................................................. 108
5.3.3.4. Fractionation of Nickel (Ni) ............................................................................... 109
5.3.3.5. Fractionation of Chromium (Cr) ........................................................................ 110
5.3.3.6. Fractionation of Lead (Pb).................................................................................. 110
5.3.4. Acid volatile sulfide (AVS) and simultaneously extracted metals
(SEM)........................................................................................................... 126
VI- DISCUSSION ............................................................................... 130
6.1. Geochemical Analysis ............................................................................. 130
6.1.1. Grain size distribution ..............................................................................130
6.1.2. Total carbonate ..........................................................................................131
6.1.3. Organic matter...........................................................................................132
6.2. Trace metal distribution ......................................................................... 133
6.2.1. Total metals ................................................................................................133
6.2.1.1. Assessment of sediment quality ......................................................................... 144
6.2.1.2. Estimation of pollution impact ........................................................................... 156
6.2.1.2.1. Enrichment factor .................................................................................................. 156
6.2.1.2.2. Geoaccumulation Index (Igeo) ................................................................................ 159
6.2.1.2.3. Contamination factor (Cf) and Degree of contamination (Cd) .............................. 162
6.2.1.2.4. Modified degree of contamination (mCd)............................................................. 162
6.2.1.2.5. Pollution Load Index ............................................................................................. 163
6.2.2. Leachable metals .......................................................................................166
6.2.3. Fractionation of trace metals ...................................................................176
6.2.3.1. Risk assessment code (RAC).............................................................................. 188
6.2.4. Toxicity assessment of trace metals in sediments based on AVS, SEM
models ...................................................................................................................190
6.3. Ecological risk assessment ...................................................................... 196
6.3.1. Problem formation ....................................................................................196
6.3.2. Analysis setup.............................................................................................197
6.3.2.1. Exposure characterization .................................................................................. 197
6.3.2.2. Effect characterization ........................................................................................ 197
6.3.3. Risk characterization ................................................................................200
6.3.4. Area of special concern based on the conducted SLERA ......................204
6.3.5. The potential Ecological Risk Index method ..........................................206
6.4. Multivariate Statistical analysis ............................................................ 209
6.4.1. Application of principal component analysis (PCA) ..............................209
6.4.2. Cluster analysis ..........................................................................................210
6.4.3. Correlation matrix.....................................................................................216
VII- English Summary ....................................................................... 225
References ........................................................................................... 235
LIST OF TABELS
Table
1.
2.
3.
4.
5.
6.
7.
8.
9.
10.
11.
12.
13.
14.
15.
16.
17.
Page
Description of the study area and sources of marine pollution
Concentrations of metals (µgg-1) obtained from the analysis of the
Standard Reference Materials (IAEA-405) and their recovery
Physical parameters of the surface seawater of the study area.
Calcium carbonate (CaCO3%), total organic matter (OM%),
Water Content (WC%), pH, grain size and type of sediments along
the Egyptian Mediterranean Coastal area.
Total trace metals concentration (µgg-1 dry weight) in surfacial
sediments along the Egyptian Mediterranean Coastal area.
Leachable trace metals concentration (µgg-1 dry weight) of
surfacial sediments along the Egyptian Mediterranean coastal
area.
Fractionation of Fe concentration (µgg-1) and its relative
percentage from the total in sediments of the Egyptian
Mediterranean coastal area.
Fractionation of Mn concentration (µgg-1) and its relative
percentage from the total in sediments of the Egyptian
Mediterranean coastal area.
Fractionation of Zn concentration (µgg-1) and its relative
percentage from the total in sediments of the Egyptian
Mediterranean coastal area.
Fractionation of Ni concentration (µgg-1) and its relative
percentage from the total in sediments of the Egyptian
Mediterranean coastal area.
Fractionation of Cr concentration (µgg-1) and its relative
percentage from the total in sediments of the Egyptian
Mediterranean coastal area.
Fractionation of Pb concentration (µgg-1) and its relative
percentage from the total in sediments of the Egyptian
Mediterranean coastal area.
Concentration of AVS and SEM (µmole/g, dry weight) in
surfacial sediments of the Egyptian Mediterranean coastal area
from El-Salloum to Rafah
Average concentrations (µgg-1) of total trace metals in sediments
along the Egyptian Mediterranean coastal area compared with
nocontaminated sediments
Concentrations of trace metals (µgg-1) determined in sediments of
the present study compared to other parts along the
Mediterranean coast.
Comparison of the surfacial sediments from the Egyptian
Mediterranean coast with Sediment quality guidelines
Estimated mean ERMq and PELq of surface sediments along the
Egyptian Mediterranean coastal area
77
80
87
89
96
103
112
113
114
115
116
117
127
142
143
148
154
Table
18.
19.
20.
21.
22.
23.
24.
25.
26.
27.
28.
29.
30.
31.
32.
33.
34.
35.
36.
37.
38.
39.
40.
Page
The metal threshold values of some different criteria used to
distinguish marine sediment quality (µgg1-)
Classification of sediments according to the background and
classification system from the Hong Kong environmental
Protection Department (EPD, 1992)
Enrichment factor (EF) in the surface sediments of the Egyptian
Mediterranean coastal area
Geoaccumulation Index Igeo
Igeo values of trace metals in the surface marine sediments along the
Egyptian Mediterranean coastal area
The classification and description of the modified degree of
contamination (mCd)
Modified degree of contamination (mCd) and contamination
factors (Cf) for the surfacial sediments along the Egyptian
Mediterranean coastal area
The relative percentage of leachable Metals from total metals in
the sediments of the Egyptian Mediterranean coastal area
Classification of Risk assessment code (RAC)
Risk assessment code (RAC) (%) and risk rank for trace metals in
the sediments of the Egyptian Mediterranean coastal area
The concentrations of AVS, SEM (µmole/g), and the ratio of
SEM and AVS in surface sediments of the Egyptian
Mediterranean coastal area.
The concentrations of AVS, SEM (µmole/g), and the SEM-AVS
in surface sediments of the Egyptian Mediterranean coastal area.
The concentrations of AVS, SEM (µmole/g), and the SEMAVS/fOC in surface sediments of the Egyptian Mediterranean
coastal area.
Table 31: Assessment and measurement endpoints for the SERA of
the Egyptian Mediterranean coast.
Summary of the effect concentration levels from different SQGs
applied in the SERA of the Egyptian Mediterranean coast
Summary of the information obtained from the risk
characterization process for pollutants in sediments of the
Egyptian Mediterranean Coast.
The calculated PEC HQ of the Egyptian Mediterranean coastal
sediments
Pollution index, integrated pollution index and pollution grade in
sediments along the Egyptian Mediterranean coastal area
Relation between RI and pollution levels
Evaluation on potential risk of trace metal pollution in sediments
along the Egyptian Mediterranean costal area
Factor loadings on elements in surfacial sediments samples along
the Egyptian Mediterranean coastal area (n=20)
Factor loadings on leachable metals in surfacial sediments samples
along the Egyptian Mediterranean coastal area (n=20)
Factor loadings of variable (5 metals) using PCA technique of the
data set in the Exchangeable fraction of the surfacial sediments
along the Egyptian Mediterranean coastal area.
155
155
158
159
160
163
164
172
188
189
194
195
195
197
201
205
206
207
208
208
212
212
213
Table
41.
42.
43.
44.
45.
46.
47.
48.
49.
50.
51.
52.
Page
Factor loadings of variable (5 metals) using PCA technique of the
data set in the Carbonate fraction of the surfacial sediments along
the Egyptian Mediterranean coastal area.
Factor loadings of variable (5 metals) using PCA technique of the
data set in the Oxides fraction of the surfacial sediments along the
Egyptian Mediterranean coastal area.
Factor loadings of variable (5 metals) using PCA technique of the
data set in the Organic fraction of the surfacial sediments along
the Egyptian Mediterranean coastal area.
Factor loadings of variable (5 metals) using PCA technique of the
data set in the Residual fraction of the surfacial sediments along
the Egyptian Mediterranean coastal area.
Correlation matrix between heavy metals, CaCO3 and OM in
sediments (n=20) along the Egyptian Mediterranean coast from
Salloum to Rafah
Correlation matrix between leachable trace metals , CaCO3 and OM in
sediments from Salloum to Rafah
Correlation coefficient matrix between chemical characteristics
and different species of Iron.
Correlation coefficient matrix between chemical characteristics
and different species of Manganese.
Correlation coefficient matrix between chemical characteristics
and different species of Zinc.
Correlation coefficient matrix between chemical characteristics
and different species of Nickel.
Correlation coefficient matrix between chemical characteristics
and different species of Chromium.
Correlation coefficient matrix between chemical characteristics
and different species of Lead.
-12-
213
213
214
214
217
222
223
223
223
224
224
224
LIST OF FIGURES
Figure
1.
2.
3.
4.
5.
6.
7.
8.
9.
10.
11.
12.
13.
14.
15.
16.
17.
18.
19.
Page
Figure (1): Map of Egyptian Mediterranean coast representing the
sampling sites of the study area
Flow chart of the sequential extration scheme used in this study
Distribution of CaCO3 % in surface sediments along the Egyptian
Mediterranean Coast.
Distribution of TOM% in surface sediments along the Egyptian
Mediterranean Coast.
Distribution of % Water Content in surface sediments along the
Egyptian Mediterranean Coast.
Distribution of Fe, Mn and Zn in surface sediments collected from
the Egyptian Mediterranean coastal area
Distribution of Cu, Ni, and Cr in surface sediments collected from
the Egyptian Mediterranean coastal area
Distribution of Co, Pb and Cd in surface sediments collected from
the Egyptian Mediterranean coastal area
Concentration of LFe, LMn and LZn in surfacial sediments along
the Egyptian Mediterranean coastal area.
Concentration of LCu, LNi, LCr in surfacial sediments along the
Egyptian Mediterranean coastal area.
Concentration of LCo, LPb, LCd in surfacial sediments along the
Egyptian Mediterranean coastal area.
Percent contribution of Fe, Mn and Zn fractions from the total
concentration in surfacial sediments of the Egyptian
Mediterranean coastal area
Percent contribution of Ni, Cr, and Pb fractions from the total
concentration in surfacial sediments of the Egyptian
Mediterranean coastal area
The relative percentage of Fe fractions from the total
concentration in the three regions, Egyptian Mediterranean
coastal area
The relative percentage of Mn fractions from the total
concentration in the three regions, Egyptian Mediterranean Sea
sediments
The relative percentage of Zn fractions from the total
concentration in the three regions, Egyptian Mediterranean Sea
sediments
The relative percentage of Ni fractions from the total
concentration in the three regions, Egyptian Mediterranean Sea
sediments
The relative percentage of Cr fractions from the total
concentration in the three regions, Egyptian Mediterranean Sea
sediments
The relative percentage of Pb Fractions from the total
concentration in the three regions, Egyptian Mediterranean Sea
sediments
-13-
76
82
90
90
90
97
98
99
104
105
106
118
119
120
121
122
123
124
125
Figure
20.
21.
22.
23.
24.
25.
26.
27.
28.
29.
30.
31.
32.
33.
34.
35.
36.
37.
38.
39.
40.
41.
Page
The relative percentage of simultaneously extracted metals in the
three regions, Egyptian Mediterranean Sea sediments
The average of SEM in the three regions, Egyptian
Mediterranean Sea sediments
The average of AVS in the three regions, Egyptian Mediterranean
Sea sediments
The average of total Fe, Mn and Cr levels in µgg-1 dry weight at
the studied regions
The average of total Ni, Zn and Pb levels in µgg-1 dry weight at the
studied regions
The average of total Cu, Co and Cd levels in µgg-1 dry weight at
the studied regions
Concentration of Fe, Mn and Cr, in sediments along the Egyptian
Mediterranean coastal area comparable to US SQG.
Concentration of Ni, Pb and Cu, in sediments along the Egyptian
Mediterranean coastal area comparable to US SQG.
Concentration of Cr, Ni and Zn, in sediments along the Egyptian
Mediterranean coastal area comparable to US NOAA's.
Concentration of Pb and Cu in sediments along the Egyptian
Mediterranean coastal area comparable to US NOAA's.
Estimated mean ERMq of surface sediments along the Egyptian
Mediterranean coast
Estimated mean PELq of surface sediments along the Egyptian
Mediterranean coast
Mean Enrichment factor for surface sediments along the Egyptian
Mediterranean coastal area
Igeo values distribution of trace metals in the surface marine
sediments along the Egyptian Mediterranean coastal area
Modified degree of contamination (mCd) for sediments along the
Egyptian Mediterranean coast
Pollution load index distribution at different stations along the
Egyptian Mediterranean coastal sediments.
Distribution of total trace metals concentrations (Fe, Mn and Pb)
and leachable fractions in sediments along the Egyptian
Mediterranean coast
Distribution of total trace metals concentrations (Zn, Ni and Cr)
and leachable fractions in sediments along the Egyptian
Mediterranean coast
Distribution of total trace metals concentrations (Co, Cu and Cd)
and leachable fractions in sediments along the Egyptian
Mediterranean coast
The average concentration of Fe speciation µgg-1 in surfacial
sediments from the Egyptain Mediterranean coast
The relative percentage of Fe species µgg-1 in surfacial sediments
from the Egyptain Mediterranean coast
The average concentration of Mn speciation µgg-1 in surfacial
sediments from the Egyptain Mediterranean coast
-14-
128
129
129
139
140
141
149
150
151
152
153
153
158
161
165
165
173
174
175
182
182
183
Figure
42.
43.
44.
45.
46.
47.
48.
49.
50.
51.
52.
53.
54.
55.
56.
57.
58.
59.
60.
61.
62.
Page
The relative percentage of Mn species µgg-1 in surfacial sediments
from the Egyptain Mediterranean coast
The average concentration of Zn speciation µgg-1 in surfacial
sediments from the Egyptian Mediterranean coast
The relative percentage of Zn species in surfacial sediments from
the Egyptian Mediterranean coast
The average concentration of Ni speciation µgg-1 in surfacial
sediments from the Egyptian Mediterranean coast
The relative percentage of Ni species in surfacial sediments from
the Egyptian Mediterranean coast
The average concentration of Cr speciation µgg-1 in surfacial
sediments from the Egyptian Mediterranean coast
The relative percentage of Cr species in surfacial sediments from
the Egyptian Mediterranean coast
The average concentration of Pb speciation µgg-1 in surfacial
sediments from the Egyptian Mediterranean coast
The relative percentage of Pb species in surfacial sediments from
the Egyptian Mediterranean coast
The relationship between OC and AVS
183
Conceptual site model for the SLERA in the Egyptian
Mediterranean Coast
SLERA approach applied in the Egyptian Mediterranean coast
TEC HQ and PEC HQ for trace metals in the Egyptian
Mediterranean coastal sediments
Cluster analysis of the selected trace metals in the surfacial
sediments along the Egyptian Mediterranean coastal area
Dendogram showing station groups formed by group averaging
cluster analysis of elemental concentrations
Relationship between Fe, Mn, Co, Cr and Ni
Relationship between Zn and Cu
Relationship between Zn and Pb
Relationship between Mn and Cr
Relationship between Co and Cr
Relationship between Ni and Cr, Mn and Co
199
-15-
184
184
185
185
186
186
187
187
192
202
203
215
215
218
219
219
219
219
220
Chapter I
Introduction
Marine sediments are not only the final depositories of pollutants, such as trace
metals, but have also to be considered as potential pollutant sources (Burges and Scott,
1992)(1). Through early diagenesis and benthic organism activity, but also because of natural
phenomena (mechanical re-suspension), trace metals are re-mobilized from sediments and are
re-distributed in the marine environment. The importance of sediments as a trace metals
source is more pronounced in the coastal zone. This is particularly relevant in a land-locked
marginal sea with a highly populated coastal zone, like the Mediterranean Sea. In the
Mediterranean Sea, anthropogenic trace elements are introduced into the marine environment
through atmospheric deposition, fluvial plumes and direct discharges from coastal urban and
industrial development. Settling particles scavenge the greater part of trace metals introduced
in the water column and transport them into the coastal as well as deep-sea sediments. After
deposition, sediments are not inert depositories of trace metals. On the contrary, through
benthic fluxes at the sediment/water interface, they may represent a very important trace
element secondary source to the marine Environment (Angelidis, 2005)(2).
Trace metals occur naturally as they are components of the lithosphere and are
released into the environment through volcanism and weathering of rocks (Fergusson,
1990)(3). Heavy metals and trace elements are by-products of many industrial processes,
contributing varying amounts of different metals and trace metals and as such are discharged
as waste into the marine environment (Robson and Neal, 1997)(4). They enter the marine
environment through atmospheric and land based effluent sources (Gonzalez- Macias et al.,
2006)(5).
The metals considered toxic and which are of concern have been restricted largely, but
not exclusively, to the ten which appear to be most poisonous to marine life. These include, in
order of decreasing toxicity: mercury, cadmium, silver, nickel, selenium, lead, copper,
chromium, arsenic and zinc (Davies, 1978)(6).
Sediments can be valuable indicators for monitoring contaminants in aquatic
environments (Atgin et al., 2000)(7). The sediments have been found polluted with various
types of hazardous and toxic substances, including trace metals which accumulate via
disposal of liquid effluents, terrestrial runoff, and leachates carrying chemicals originating
from numerous urban, industrial, and agricultural activities, as well as atmospheric deposition
(Mucha et al., 2003)(8).
The amount of heavy metals in sediments also affects their characteristics,
particularly, the type and quantities of organic matter, grain size, cation exchange capacity
and mineral constituents (Tam and Wong, 1995)(9). Most trace metals are bound in the finegrained fraction (<63μm) mostly because of its high surface area to grain size ratio and
humus substance content (Moore et al., 1989)(10) where they have a potentially greater
biological availability than those in the larger (2 mm–63μm) sediment fraction (Everaat and
Fischer, 1992)(11). Al-Abdali et al. (1996)(12) found that positive correlations exists between
increasing metal concentration and decreasing grain size, suggesting that adsorption onto mud
is the primary mechanism of metal accumulation in marine sediments.
-16-
Over the last decade, much concern has been focused on the measurement of total
concentration in sediment. However, it cannot provide sufficient and proper information on
mobility, bioavailability and toxicity of trace metals (Wei et al., 2007)(13).
Pollution by metals is a serious problem because of their toxicity and ability to
accumulate in biota with a negative impact on the environment and human health. It is now
widely accepted that the role of aquatic sediments as a sink or as a source of pollutants cannot
fully be assessed by measuring total metal concentrations. In addition, the determination of
total elements does not give an accurate estimate of the likely environmental impact. Instead,
it is desirable to have information on the potential availability of metals (whether toxic or
essential) to biota under various environmental conditions. Since the mobility of trace metals
and their bioavailability and related ecotoxicity to plants depend critically on the chemical
form in which a metal is present in the sediment, considerable interest exists in element
speciation (Davidson et al. 2004)(14).
Sequential extraction schemes provide information on evaluation of metal
bioavailability and identification of binding sites of metals for assessing metal accumulation,
pollution and transport mechanisms. The most widely used sequential extraction procedures
was proposed by Tessier et al. (1979)(15) and the Community Bureau of Reference (BCR)(Ure
et al., 1993; Rauret et al., 1999)(16, 17). To overcome the problem of re-adsorption and
redistribution among the trace elements during the extraction procedure, some modified BCR
schemes have also been proposed to eliminate this problem, which have been applied to soil,
sewage sludge and sediment (Rauret et al.,1999; Rauret, 1998; Quevauviller, 1998, Filgueiras
et al., 2002)(17-20). In these application metal partitioning in sediments was divided into acid
soluble (metal precipitated or co-precipitated with carbonate), reducible (hydrous oxides of
Mn and Fe), oxidizable (complexation with organic or bound to sulfide) and silicate residuals
(Davidson et al., 2004)(21). Sulfide, which was among the oxdiziable phase during sequential
extraction, was chosen for controlling factor on partitioning of toxic metals between aqueous
and solid phases (Di Toro et al., 1990 ; Ankley et al., 1996)(22, 23). However, sulfide was not
distinguished from organic in sequential extraction, which obscures the true speciation of
metals in reduced sediments (Wei et al., 2007)(13).
SEM (simultaneously extracted metals) /AVS (acid volatile sulfide) was endorsed as
the best technology for assessing the bioavailability of five metals (Pb, Zn, Cu, Cd, and Ni).
Different from BCR, it laid emphasis on toxicity and bioavailability of metals through the
comparison of sulfide and SEM. However, it has been recognized that this measurement was
applicable only to reduced sediments with measurable AVS and other tests of metal toxicity
were needed for oxidized sediments. Although SEM were mainly controlled by sulfide in
anoxic sediments, other minerals such as Fe-Mn oxide, carbonate and silicate also affect
speciation of the metals as well (O’ Day et al., 2000)(24).
The effectiveness of sequential extraction with AVS analysis was evaluated and it was
found that sequential extraction significantly underestimated the amount of Zn associated
with sulfide phases (Peltier et al., 2005)(25). Obviously, BCR or SEM/AVS measurement
alone cannot comprehensively assess the toxicity and bioavailability of trace metals in
sediments (Fang et al., 2005)(26). BCR results cannot specify the effect of sulfide on metal
toxicity and bioavailability although it provides enough information on metal speciation.
However SEM/AVS method neglects other metal speciations and only targets reduced
sediment. Comparison of sequential extraction and SEM/AVS make up for the drawbacks that
SEM/AVS only applied in anoxic sediments and overestimate metal bioavailability if
nonsulfide was also extracted other than sulfide phases. Therefore, it will contribute to
-17-
quantifying the possible chemical partitioning patterns in sediments, thus have a deep insight
in interpreting the speciation of trace metals (Wei et al., 2007)(13).
The Problem
During the last decades, there has been an extensive increase in the level of
urbanization and industrialization along the Egyptian coastal area. At the same time, little data
are available for the concentrations of trace metals along the Egyptian Mediterranean Coast.
Studies on the environmental state of the Egyptian Mediterranean Coast have concentrated in
the past mostly on simple chemical and bulk analysis of sediments. Most of the studies
investigated metals in the total form and no comprehensive study was performed to
investigate the most important metal bearing fractions in sediments and their mobility,
bioavailability and thereby their toxicity. In addition, no data are available for the AVS/SEM
as an important tool for evaluating metal toxicity in sediments.
Aim of the study
Industrial and human activities have increased dramatically in the last 20 years.
Consequently, updated information is needed as basis for decision makers to establish
effective management plans. Thus, the purpose of this study was set to:
1. Evaluate the levels and the spatial distribution of trace metals in the surfacial
sediments of the Egyptian Mediterranean Coast.
2. Investigate the possibility of assessing the sediment toxicity based on analysis of
simultaneously extracted metals (SEM) and acid volatile sulfides (AVS).
3. Evaluate metals mobility, bioavailability and toxicity by applying the sequential
extraction procedure on the sediment samples.
4. Comparison between the above mentioned techniques to determine the best available
methodology for assessing the sediment toxicity of the Egyptian Mediterranean Coast
and to try to develop site specific sediment quality guidelines.
5. Formulate a detailed level ecological risk assessment on sediments to evaluate the
possibility of occurrence of adverse ecological effects to benthic species based on the
site specific characteristics.
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Chapter II
Literature Review
2.1. Contaminants in the environment
Human activities have introduced synthetic substances into the environment, and have
increased the concentrations of many naturally occurring substances.
When the
concentrations of these substances rise above the background levels, they are regarded as
contaminates. The term anthropogenic is used to describe contaminants derived from man
activities. Contaminants are termed pollutants when concentrations are high enough to cause
toxic effects in exposed organisms. Organisms are not necessarily impacted negatively by the
presence of contaminants but as concentrations increase, the likelihood of organism
experiencing negative effects from exposure increase (Campbell et al., 2006)(27).
Growing social concern about environmental quality can be observed in recent years,
both on a global and local scale. Emission of harmful substances has negative effects on the
natural environment, human health and agricultural production efficiency. When the
consequences of environmental pollution become visible, it is often too late to prevent them
and chronic toxic effects, impossible to notice at the initial stage of the process, may manifest
after many years. Toxic chemical substances introduced into the environment may be
transported by the air, water and living organisms and may become a part of the natural
biogeochemical cycle and accumulate in the food chain (Adu, 2010)(28).
Due to constant technological progress the natural environment undergoes numerous
changes, deteriorating its quality, which often results in negative interactions between
particular ecosystem components (Duffus, 2002)(29). Various contaminants released into the
sea may significantly affect marine ecology, and in extreme cases, may lead to the destruction
of whole ecosystems. The contaminants of major concern are sewage, nutrients, metallic
compounds, substances disrupting endocrine functions, persistent organic pollutants and
petroleum hydrocarbons (GESAMP, 2001)(30). Many contaminants such as trace metals (also
referred as heavy metals), though occur at extremely low concentrations in seawater, are
accumulated by marine organisms and concentrations in their body tissue can be hundreds of
times greater than the levels in seawater (Riley and Chester, 1971)(31).
The study of metals in the marine environment has received considerable attention
because of their biological significance as well as the possibility of their transfer to man
through the food chains in quantities that can be harmful. The available information indicates
that though the open oceans suffer from some contamination and ecological damage, they are
still in a relatively healthy state as compared to some coastal areas. It is the waters nearest to
the shore, particularly in estuaries and in semi-enclosed seas and bays bordering highly
industrialized and urbanized zones that have suffered ecological degradation over the past few
decades (GESAMP, 2001)(30). Environmental pollution with toxic metals is becoming a
global phenomenon. As a result of the increasing concern with the potential effects of the
metallic contaminants on human health and the environment, the research on fundamental,
applied and health aspects of trace metals in the environment is increasing (Vernet, 1991)(32).
2.2. Trace metals
Metals are discovery of the old world. Humans began using metals in about 6000
B.C. Metals are everywhere, from air to water and to the deepest places in Earth. Metals are
useful but they can be hazardous and toxic (Dahrazma, 2005)(33). The term “heavy metal” is
-19-
used to describe metals having density greater than 5-6 g cm-³ (Sedgwick, 2005)(34). Heavy
metals are often referred to as trace metals, occurring naturally in low concentrations in
organisms, although the term trace metal might imply the presence of an essential requirement
by organisms for that particular metal (Furness and Rainbow, 2000)(35). Heavy metals can be
divided into two groups namely; i) metals essential to at least some organisms
(micronutrients) such as As, Cr, Cu, Ni, and Zn and ii) nonessential heavy metals with no
known biological function which include Hg, Pb and Cd (Furness and Rainbow, 2000)(35).
Even though micronutrients are essential to some forms of life, they become toxic at high
concentrations (Sedgwick, 2005)(34).
The contamination of sediments by trace metals is of particular importance because of
their widespread use and their non degradable nature (Calmano et al., 1988)(36). The trace
metals of greatest environmental concern are those that are both industrial and toxicological
importance, such as Cd, Cr, Cu, Ni, Pb, Sn and Zn (Salbu and Steinnes, 1995)(37). Chromium
(Cr VI) is classified as priority pollutant by the United States Environmental Protection
Agency (USEPA) with a carcinogenicity classification A (human carcinogen), while Cd and
Pb are classified in the same list with a carcinogenicity classification B (probable human
carcinogen) (USEPA, 1999)(38). Trace metals are ubiquitous in the environment, and can enter
aquatic systems through natural processes as well as anthropogenic loadings (Drever,
1997)(39). Metals participate in various biogeochemical mechanisms, have significant
mobility, can affect the ecosystems through bioaccumulation processes and are potentially
toxic for the environment and for human life (Manahan, 2000)(40).
Metals are natural constituents of rocks, soils, sediments, and water. However, over
the 200 years following the beginning of industrialization huge changes in the global budget
of critical chemicals at the earth's surface have occurred, challenging those regulatory systems
which took millions of years to evolve (Wood and Wang, 1983)(41).
Advances in information of the distributions and concentrations of trace metals in the
marine environment have occurred since the mid 1970s (Burton and Statham, 2000)(42). This
is mainly due to developments in procedures for contamination free sampling, the adoption of
clean methodologies for handling and analysis of samples, and increased application of
improved analytical methods such as inductively coupled plasma-mass spectrometry (ICPMS) (Plant et al., 2003)(43).
The trace metal content of sediments comes from natural sources (rock weathering,
soil erosion, dissolution of water-soluble salts) as well as anthropogenic sources such as
municipal wastewater treatment plants, manufacturing industries, and agricultural activities
(Guven and Akinci, 2008)(44). Trace metals, which include Cr, Mn, Fe, Co, Ni, Cu, Zn, As,
Pb, Cd, Hg, Ti and V enter the oceans as a result of natural processes and human activities via
rivers, land runoff, dumping, atmosphere and the seabed. Major natural sources of these
elements are rock weathering, degassing, emissions from terrestrial and submarine volcanoes
and dissolution from marine sediments. Hence, in areas characterized by metal-bearing
formations these metals occur at elevated concentrations in water and bottom sediments of the
marine zones receiving the runoff (Munsgaard and Parry, 2002)(45). The dominant inputs of
most trace metals to coastal areas are through rivers and land runoff but for a few elements
such as Hg, As, and Pb the atmospheric route is also important, particularly in the open ocean,
although even for these elements local discharges and rivers often dominate the coastal input.
The important anthropogenic sources of metals like Cr, Mn, Co, Ni, Cu, Zn, Pb, Cd, Hg and
As to coastal and inshore waters are industrial processing of ores and metals; ferrous and
nonferrous metal industries including metal plating; industries producing both organic and
inorganic chemicals; use of metal and metal components; leaching of metals from solid waste;
-20-
and offshore dumping of domestic sewage, sludge and industrial wastes (Owens et al.,
1997)(46).
Coastal environments continue to attract heavy investment in various economic
activities such as tourism, aquaculture, industrialization, agricultural, mining and
urbanization. Therefore coastal zone pollution problems are becoming increasingly prevalent,
with potentially adverse environmental and socio-economic effects. Consequently in global
context, anthropogenic fluxes of contaminants such as trace metals now exceed the natural
flux in many areas (Onyari et al., 2003)(47). Other potential sources include ports, harbors,
marinas and mooring sites, also subjected to trace metal inputs associated with recreational,
commercial, and occasionally, military, boating, and shipping activities (Denton et al.,
1997)(48).
Trace metals have certain unique properties and characteristics, distinct from organic
substances, which create difficulty in determining ecological risk. Because they are naturally
occurring and ubiquitous in the environment, metals can concentrate at a particular site (e.g.
related to a specific point source discharge) or on a regional scale (e.g. related to a particular
watershed) due to natural processes. Metals can occur in the environment in various chemical
forms. These different forms influence availability, hence extent of toxic effect, to organisms.
The bioavailability and persistence of metals are generally controlled by external
environmental conditions. These characteristics create challenges when assessing risk of
metals in ecosystem (Kabata-Pendia and Pendia, 2000)(49). In aquatic systems trace metals
play a dual role as toxicant and essential nutrient (Morel and Hering, 1993)(50). Metal risk
assessment in aquatic system is considerably more complicated than that associated with
organic chemicals. Unlike many organic chemicals, metals cannot be degraded biologically;
they are, in a sense, infinitely persistent. Using a more scientifically defensible approach, one
defines persistence in terms of bioavailable metal transport, fate, and bioavailability. These
include chemical speciation (including redox chemistry), adsorption/desorption, and
precipitation/dissolution (Adams et al., 2000)(51).
The distribution of metals within the aquatic environments is governed by complex
processes of material exchange affected by various anthropogenic activities or natural
processes including riverine or atmospheric inputs, coastal and seafloor erosion, biological
activities, water drainage, discharge or urban and industrial wastewaters (Ip Carman et al.,
2007)(52). The pollutants from industrial activity and urban sewage commonly reach their
final destination in the coastal waters through river mouths or by direct discharge into the sea.
Metals are important among these pollutants because many of them are highly toxic and that
they are not biodegradable in the environment (Morillo et al., 2002)(53).
2.2.1. Sources, uses and toxicity of trace metals
Metals have many sources from which they can flow into the water body, these sources are:
Natural sources: Metals are found throughout the earth, in rocks, soil and are introduced into
the water body through natural processes, weathering and erosion (Rashed, 2004)(54). Trace
metals occur naturally as they are components of the lithosphere and are released into the
environment through volcanism and weathering of rocks (Fergusson, 1990)(3).
Rivers: Trace metals can enter coastal and estuarine waters via riverine input, non point
source runoff from land and direct point source discharges. Most rivers make a major
contribution to trace metal loadings in estuarine and coastal waterways. A large quantity of
trace metals in river water originates from weathering of rocks and leaching of soils. For this
reason, the trace metal loading depends largely on the natural occurrence of metals and ore
-21-
bearing deposits in the catchment area. Where the river passes through industrial, agricultural
or urbanized areas, stormwater runoff contributes to the dissolved or suspended metal burden
within the river (Burton, 2005)(55).
Atmospheric pollution : Acid rains containing trace metals as well as suspended particulate
matter input to the water body will cause the pollution of water with metals (Rashed,
2004)(54).
Mining and processing ores: Digging a mine, removing ore from it, and extraction and
processing of the minerals may destroy habitats, farmland, and homes; produce soil erosion;
and pollute waterways via toxic drainage. Ore processing, smelting, and refining operations
can cause deposition of large quantities of trace metals, such as lead, zinc, copper, arsenic,
and silver into drainage basins or direct discharge into aquatic environments (Csuros et al.,
2002)(56).
Fossil fuel combustion: Fossil combustion is a major source of airborne metal contamination
of natural waters (Csuros et al., 2002)(56).
Agricultural sources: Agricultural discharge contains residual of pesticides and fertilizers
which contains metals (Rashed, 2004)(54). The metal content of agriculture runoff originate in
sediments and soils saturated by animal and plant residues, fertilizers, specific herbicides and
fungicides, and use of sewage and sludge as plant nutrients (Csuros et al., 2002)(56). Metal
containing pesticides (such as Zn salts, Cu and Pb arsenates, and organometallic compounds)
are used to control pests in horticultural and vegetable crops, and in turf cultivation. Major
nutrient fertilizers often contain appreciable quantities of trace metals as impurities derived
from the raw materials. Agricultural reuse of urban and farm wastes, such as pig and poultry
waste slurries and sewage sludge, can also cause appreciable enrichment of agricultural soil
with a range of metals. Under certain situations these metals may entre waterways via
leaching and surface runoff (Burton, 2005)(55).
Domestic wastewater effluents: Large amounts of heavy metals copper, lead, zinc, and
cadmium, can be found in metabolic waste products, corrosion of water pipes from the
domestic wastewater effluents while iron, manganese, chromium, nickel, cobalt, zinc, and
arsenic are often present in household products, such as detergents. Although wastewater
treatment can removes metals from the influent, more than 50% of metal content in the
influent still remain in the effluent. Moreover, the sludge resulting from wastewater treatment
is also one of the major artificial sources of cadmium, chromium, copper, iron, lead, and
mercury pollution (Csuros et al., 2002)(56).
Storm water runoff: Many activities such as city planning, traffic, road construction land
use, can contribute to the metal pollution in the receiving waters via storm water runoff from
the urbanized areas (Csuros et al., 2002)(56). In Heavily urbanized areas, stormwater runoff
(generally regarded as a diffuse source) can make a significant contribution for a wide range
of metals. Trace metals of greatest concern in urban stormwater runoff are Pb, Cu, Zn, Cd,
As and Be. The principal diffuse urban sources are emissions from motor vehicles and metal
release from weathering of surface material (Makepeace et al., 1995)(57).
Industrial effluents and emissions: Trace metals are employed during the manufacturing of
a wide variety of products. The most important activities are metal processing and finishing
industries, such as coating, smelting and refining of non-ferrous metals, and iron/steel
manufacturing. However, a number of other industries also use metals in their processes.
These include paint/ink production, leather tanning/finishing, timber preservation, battery
manufacturing and petroleum refining. Regulatory efforts to restrict the discharge of metalbearing effluents have lead to declines in industrial loadings of trace metals in natural
waterways. However, there are many examples of areas where sediments remain
contaminated as a result of past industrial discharges (Burton, 2005)(55). Trace metal
-22-
contributions from past mining activities have results in sever sediment contamination, with
levels of contamination similar to those found in the mine tailings themselves (Teasdale et al.,
2003)(58).
Coastal waste disposal: other forms of disposal of solid and liquid waste in or near the
marine environment contribute to the direct pollution of the sea in various ways, depending
on the type and amount of the material disposed off.
Maritime (offshore) sources: the pollutant discharged in greatest amounts is ballast oil from
tankers. In many cases, beaches suffer a variable amount of damage. Ships also dispose of
waste materials overbroad. In very general terms, the majority of the social economic
consequences of marine pollution can be expressed as immediate of long-term effects on
human health. In this context, the two main types of human exposure to pollutants in the
marine environment are through direct contact with polluted seawater and/or beach sand, and
consumption of contaminated seafood. In some cases where the seawater, after being treated
in desalinization plants, is used for drinking purposes, pollution may constitute a potential
health risk (WHO/UNEP, 1995)(59).
Summary of past and present uses of some of the examined trace metals
Metal
Cadmium
Chromium
Copper
Lead
Nickel
Zinc
Uses of Metals and Compounds
Electroplating (anticorrosion coatings); thermoplastic stabilizers, e.g.
in PVC; Ni-Cd batteries; alloys; solders; catalysts; engraving; semiconductors; TV tube phosphorus; pigments in paints and plastics;
glass ceramics; biocides (Denton et al., 1997) (48)
Metallurgy-ferrochromium alloys, refractory bricks; electroplating;
industrial dyes; ink; tanning; paint; wood preservative; glass making;
cement production (Denton et al., 1997) (48)
Electrical industry; alloys, e.g. brass; chemical catalyst; anti-fouling
paint; algaecide; wood preservative (Denton et al., 1997) (48)
Storage batteries; leaded gasoline; pigments; read lead paint;
ammunition; solder; cable covering; anti-fouling paint; glazing; PVC
stabilizers (Denton et al., 1997) (48)
Hydrogenation oil processes, rechargeable batteries (nickelcadmium), ambient air at very low levels as a result of releases from
oil and coal combustion, nickel metal refining, sewage sludge
incineration, manufacturing facilities (Dean et al., 1972)(60)
Cleaners, cosmetics, fillers, pesticides, inks, oils, medicines, paints,
polish, preservatives, water treatment, fertilizers, oil refining,
inorganic chemicals (Dean et al., 1972)(60).
Cadmium (Cd)
Cadmium is widely spread by human activity, volcanic activities and erosion. Large
quantities of Cd are mainly used in the production of nickel-cadmium batteries or in welding
(El Nemr, 2007a)(61). Cadmium is a heavy metal that is widely distributed in the environment.
Its concentration in the earth's crust is generally estimated to be 0.15 to 0.20µgg-1, whereas
that in uncontaminated soils is 0.40 µgg-1 (Fleischer et al.,1974)(62). The concentration in
normal animal foods seldom exceeds 0.5 µgg-1 (Doyle et al., 1974)(63). Industrially, Cd
is used extensively in the mining and electroplating industries and is found in fertilizers
and fungicides. All chemical forms are toxic (Fleischer et al., 1974)(62). Cadmium is
usually present as complex oxides, sulfides, and carbonates in zinc, lead and copper ores
(Finkelman, 2005)(64). Cadmium in several aspects is similar to zinc and therefore, it is always
associated with zinc in mineral deposits (Callender, 2003)(65). Cadmium is extremely toxic to
-23-
most plants and animal species particularly in the form of free cadmium ions (Denton et al.,
1997)(48). The major sources of cadmium include metallurgical industries, municipal effluents,
sewage sludge and mine wastes, fossil fuels and some phosphorus containing fertilizers
(Denton et al., 2001)(66).
Nonferrous metal mines represent a major source of cadmium to the aquatic
environment. Contamination can arise from mine drainage water, waste water from the
processing of ores, overflow from the tailings pond, and rainwater runoff from the general
mine area. The release of these effluents to local watercourses can lead to excessive
contamination downstream of the mining operation (Morley et al., 1997)(67).
In sediments, cadmium does not appear to be absorbed to colloidal material, but
organic matter, such as humic substances and organic debris, appear to be the main sorption
material for the metal. Cadmium levels tend to increase with decrease in size and increase in
density in terms of partition of sediment samples by size and density. The sorption of
cadmium to sediments, and to the clay content, increases with pH. The release of cadmium
from the sediment is influenced by a number of factors including acidity, redox conditions
and complexing agents in the water. Under alkaline conditions, cadmium is less mobile
(Fergusson, 1990)(3). Also under saline conditions, Cl¯ ion affects the mobility of cadmium
(Morrison et al., 1996)(68).
The average cadmium level of sea water has been given as about 0.1 µg L-1 or less
(Owen and Sandhu, 2000)(69). WHO (1995)(70) reported that current measurements of
dissolved cadmium in surface waters of the open oceans gave values of <5 ng L -1. The
vertical distribution of dissolved cadmium in ocean waters is characterized by a surface
depletion and deep water enrichment, which corresponds to the pattern of nutrient
concentrations in these areas (Shen et al., 1987)(71). This distribution is considered to result
from the absorption of cadmium by phytoplankton in surface waters and its transport to the
depths, incorporation to biological debris, and subsequent release. In contrast, cadmium is
enriched in the surface waters of areas of upwelling and this also leads to elevated levels in
plankton unconnected with human activity. Oceanic sediments underlying these areas of high
productivity can contain markedly elevated cadmium levels as a result of inputs associated
with biological debris (Simpson, 1981)(72).
Cadmium can potentially harm human health depending upon the form of cadmium
present, the amount taken in, and whether the cadmium is eaten or breathed. If the levels are
high enough, the cadmium in the kidney will cause kidney damage and also may cause bones
to become fragile and easily broken (Finkelman, 2005)(64). Moreover, when ingested by
humans, cadmium accumulates in the intestine, liver and kidney (WHO, 2004)(73). The
kidney cortex is regarded as the most sensitive organ. The health effects of chronic exposure
of Cd include proximal tubular disease and osteomalacia (Jambe and Nandini, 2009)(74). It
accumulates in the body, especially in the kidney and the liver, over many years because the
body has no homeostatic mechanism to keep Cd at a constant safe level, such as those that
function for zinc. Hence, Cd is a cumulative poison (Miller, 1971)(75). The major effects of
cadmium poisoning are experienced in the lungs, kidneys and bones. Acute effects of
inhalation are bronchitis and toxemia in the liver. Chronic inhalation of cadmium compounds
as fumes or dust produce pulmonary emphysema, where the small air sacs of the lungs
become distended and eventually destroyed reducing lung capacity (Ansari et al., 2004)(76).
Cadmium is a serious environmental contaminant that is also transported atmospherically. In
fish, it can cause anaemia and vertebral fractures, osmoregulatory problems, decreased
digestive efficiency, haematological and biochemical effects, erratic swimming, and mortality
(Burger, 2002)(77).
-24-
Chromium (Cr)
Chromium like zinc, is one of the most abundant heavy metals in the lithosphere with
an average concentration of about 69 and mercury level in carbonate sediments is reported to
be 0.03 µgg-1 (Callender, 2003)(65). Chromium is moderately toxic to aquatic organisms.
Major coastal marine contributors of chromium are dominated by input from rivers, urban
runoff, domestic and industrial wastewaters and sewage sludge (Denton et al., 1997)(48). Also
other major sources in the aquatic environment include the waste stream from electroplating
and metal finishing industry (Finkelman, 2005)(64).
Chromium is found chiefly in Chrome-Iron Ore (FeO-Cr2O3). Chromium is
considered non-essential for plants, but an essential element for animals. The average
abundance of Cr in the earth's crust is 122µgg-1; in soils Cr ranges from 11-22 µgg-1. It is
used in alloys, in electroplating and in pigments. Chromium and its salts are used in the
leather tanning industry, the manufacture of catalyst, pigments and paints, fungicides, the
ceramic and glass industry and in photography and for chrome alloy and chromium metal
production, chrome plating and corrosion control (Abbasi et al., 1998)(78). Hexavalent
compounds are carcinogenic by inhalation and are corrosive to tissues.
There is an
environmental cycle for chromium, from rocks and soils to water, biota, air, and back to the
soil. However, a substantial amount (estimated at 6.7 x 106 kg per year) is diverted from this
cycle by discharge into streams, and by run-off and dumping into the sea (Schulz and Zabel,
2000)(79). Cr has been considered a metal with low biogeochemical mobility which should
reduce its toxicity potential (Muniz et al., 2004).(80)
Chromium is carcinogenic to humans and long term exposure has been associated with
lung cancer in workers exposed to levels in air that in the order of 100 to 1,000 times higher
than usually found in the environment (Finkelman, 2005)(64). Trivalent chromium (Cr3+) may
be essential in human nutrition, but hexa-valent chromium (Cr+6) is highly toxic. Intake of
hexa-valent chromium can cause hemorrhaging in the liver, kidneys, and respiratory organs.
When people are exposed to hexa-valent chromium, dermatitis and ulceration and perforation
of the nasal septum have been developed. Also, gastric cancers, presumably from excessive
inhalation of dust containing chromium, have been reported (Fergusson, 1990)(3).
Copper (Cu)
The average abundance of Copper is 68 µgg-1; in soils it is between 9 and 33 µgg-1.
Copper is widely used in electrical wiring, roofing, various alloy, pigments, cooking utensils,
piping and in the chemical industry. Copper is present in munitions, alloys (brass, bronze)
and coatings. Copper compounds are used as or in fungicides algicides, instecticides, and
wood preservatives and in electroplating, azo dye manufacture, engraving, lithography,
petroleum refining and pyrotechnics. Copper compounds can be added to fertilizers and
animal feeds as a nutrient to support plant and animal growth. Copper compounds are also
used as food additives (Abbasi et al., 1998)(78). In addition, copper salts are used in water
supply systems to control biological growths in reservoirs and distribution pipes and it forms
a number of complexes in natural waters with inorganic and organic lignads (WHO, 2004)(73).
The aqueous species of copper include Cu 2+, Cu(OH)2 and CuHCO3+(Jumbe and Nandini,
2009)(74).
Inputs of copper into the natural waters come from various sources including mining,
smelting, domestic and industrial wastewaters, steam electrical production, incinerator
emissions, and the dumping of sewage sludge. Algaecides and antifouling paints are identified
as major contributors of copper to harbor areas whereas coastal waters are generally receiving
inputs from rivers and atmospheric sources (Denton et al., 1997)(48). Copper has a high
-25-
affinity for clay mineral fractions, especially those rich in coatings containing organic carbon
and manganese oxides (Callender, 2003)(65) and as a result, residues are often elevated in
sediments near localized sources of inputs (Denton et al., 1997)(48). Callender (2003)(65)
reported that the adsorption behavior of copper in natural systems is strongly dependent on
the type and concentration of inorganic and organic ligands.
Copper exists in construction materials, pesticides, fertilizers, sprays, and agricultural
and municipal wastes as well as industrial wastes. As these materials are exposed to the
environment the levels of copper increase. The average level of copper in plants is in the
range of 5 to 30 µgg-1 while in soil it varies between 2 to 100 µgg-1 (Cameron, 1992)(81).
Because Cu is an essential trace nutrient, most marine organisms have evolved
mechanisms to regulate concentrations of this metal in their tissues in the presence of variable
concentrations in the ambient water, sediment and food (Mohammed, 1991)(82). Although
essential for life due to its major role in enzyme functions, copper in large amounts is quite
toxic. For example, copper salts are used to kill bacteria, fungi, and algae, and paints
containing copper are used on ship hulls to prevent fouling by marine organisms. Acute
exposure overdose causes an immediate metallic taste, followed by epigastric burning,
nausea, vomiting, and diarrhea (Fergusson, 1990)(3). Regulation of copper in the human body
is not sufficient with very low amounts of dietary copper (0.38 mg/day) and it appears to be
delayed when copper intake is high (Turnlund, 1998)(83). Pro-oxidant activity of copper ions
can make it extremely toxic at elevated levels. Copper related inherited disorders such as
Menkes disease (due to copper deficiency) and Wilson disease (due to copper toxicosis)
demonstrate the necessity and potential toxicity of copper. There is a good chance of
involving copper induced oxidative damage in other pathogenic or neurodegenerative
conditions such as Alzheimer's disease and prion diseases (LIanos and Mercer, 2002)(84).
Copper is essential for good health. However, exposure to higher doses can be fatal. Long
term exposure to copper results in nose irritation, mouth, and eyes, and cause headaches,
dizziness, nausea, and diarrhea (Finkelman, 2005)(64). It is an essential element in the normal
metabolism of plants, animals and human life, however in high levels, causes many diseases
(Gratten et al., 2003)(85). In high levels, it accumulates in blood, liver, and kidneys and causes
anemia, renal and intestinal irritations, coma death, and Wilson's disease (Gratten et al.,
2003)(85).
Cobalt (Co)
Cobalt is a silvery grey solid at room temperature. It is the 33rd most abundant
element and has been found in a variety of media, including air, surface water, leachate from
hazardous waste sites, groundwater, soil, and sediment. Sources of exposure to cobalt and
inorganic cobalt compounds are both natural and anthropogenic (Barceloux, 1999)(86). Natural
sources include wind-blown dust, seawater spray, volcanoes, forest fires, and continental and
marine biogenic emissions. Anthropogenic sources include the burning of fossil fuels, sewage
sludge, phosphate fertilizers, mining and smelting of cobalt ores, processing of cobalt alloys,
and industries that use or process cobalt compounds. Cobalt released into the atmosphere is
deposited on soil, and cobalt released to water may sorb to particles and settle into sediment
or sorb directly to sediment (Erdogan, 2009)(87). The average cobalt abundance in earth's
crust is 29 µgg-1; in soils it is 1.0-14 µgg-1. Cobalt is widely used as alloy for various steel, in
electroplating, in fertilizers, porcelain and glass making. Cobalt is considered essential for
algae and bacteria but not essential to higher plants. In animals it is only requires in trace
basis (WHO, 2004)(73).
The distribution coefficient of cobalt varies due to pH, redox conditions, ionic
strength, and dissolved organic matter concentrations. Factors affecting the speciation and
-26-
fate of cobalt in water, sediments, and soil include organic ligands such as humic acids,
anions, pH, and redox potential. The soil mobility of cobalt is inversely related to the strength
of adsorption by soil constituents. Although plants may take up cobalt from the soil, the
translocation of cobalt from the roots to other parts of the plant is not significant. Mean cobalt
concentrations in seawater have been reported to be less than 1 µgL-1. Cobalt concentrations
in drinking water are generally <1–2 µgL-1. In rainwater, mean concentrations are 0.3–1.7
µgL-1. The earth’s crust contains an average cobalt concentration of 20–25 µgg-1. Near some
anthropogenic sources, the concentration of cobalt in soil may be several hundred milligrams
per kilogram. The major anthropogenic sources of environmental cobalt include mining and
processing of cobalt-bearing ores, the use of cobalt-containing sludge or phosphate fertilizers
on soil, the disposal of cobalt containing waste, and atmospheric deposition from activities
such as the burning of fossil fuels and smelting and refining of metals (Smith and Carson,
1981)(88).
Cobalt is beneficial for humans because it is a part of vitamin B12, which is essential
for human health. Cobalt is used to treat anemia with pregnant women, because it stimulate
the production of red blood cells. However, too high concentrations of cobalt may damage
human health. When we breath in too high concentrations of cobalt through air we
experience lung effects, such as asthma and pneumonia. This mainly occurs with people that
work with cobalt. Soils near mining and melting facilities may contain very high amounts of
cobalt, so that the uptake by humans through eating plants can cause health effects. Health
effects that are a result of uptake of high concentrations of cobalt are: vomiting and nausea,
vision problems, heart problem, and thyroid damage. Health effects may also be caused by
radiation of radioactive cobalt isotopes. This can cause sterility, hair loss, vomiting, bleeding,
diarrhea, coma and even death. This radiation is sometimes used with cancer patients to
destroy tumors. Cobalt dust may cause asthma like disease with symptoms ranging from
cough, shortness of breath and dyspnea to decreased pulmonary function, nodular fibrosis,
permanent disability and death. Exposure to cobalt may cause weight loss, dermatitis, and
respiratory hypersensitivity (Greenwood and Earnshaw, 1984)(89).
Lead (Pb)
Lead is one of the oldest known metals. Its emission into the atmosphere has much
increased during the last century. Its anthropogenic inputs greatly exceed those from natural
sources. The atmospheric fallout of lead constitutes its most important source in the aquatic
environment (Moore and Ramamoorthy, 1984)(90). It is present in all tissues of organisms
causing hazardous effects, as it is a cumulative poison (Saad, 2004)(91). In fish, lead can cause
deficits or decreases in survival, growth rates, development, and metabolism, in addition to
increased mucus formation (Burger et al., 2002)(77).
Lead is one of the most abundant toxic metals in earth's crust. It has been used since
prehistoric times and has become widely distributed and mobilized in the environment.
Exposure to lead in the environmental and occupational settings continues to be a serious
public health problem (WHO, 1995)(92). The average abundance in Earth's crust is 13µgg-1; in
natural soils background level ranges from 2.6-25 µgg-1; the common aqueous species are
hydroxides and carbonates of Pb2+. Lead in water comes from industrial, mines and smelter
discharges before being deposited in the sediment sinks. Lead is non essential for plants and
animals and is toxic by ingestion-being a cumulative poison. Lead is also used in the
production of lead acid batteries, solder, alloys cable sheathing, pigments, rust inhibitors,
ammunition, glazes and plastic stabilizers. Tetraethyl and tetramethyl lead are important
because of their extensive use as antiknock compounds in petrol (Sharma and Pervez,
2003)(93). Lead is reported to be in the 15-50 µgg-1 range for coastal and estuarine sediments
around the world with < 25 µgg-1 in clean coastal sediments (Denton et al., 1997)(48).
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The main sources of Pb in natural waters include manufacturing processes (especially
metals), atmospheric deposition (e.g. from pyrometallurgical nonferrous metal production; the
combustion of leaded fuels; the burning of wood and coal; and the incineration of municipal
refuse). Other sources include domestic wastewaters, sewage and sewage sludge (Denton et
al., 1997)(48). Metallic lead does occur in nature, but it is rare. Lead is usually found in ore
with zinc, silver and copper, and is extracted together with these metals. The main lead
mineral is galena (PbS), which contains 86.6% lead. Inorganic lead is moderately toxic to
aquatic flora and ranks behind mercury, cadmium, copper and zinc in the order of toxicity to
invertebrates. On the other hand, organolead compounds, particularly the alkyl-lead
compounds are considered toxic to any forms of life (Denton et al., 1997)(48).
Inorganic lead is moderately toxic to aquatic flora and ranks behind mercury,
cadmium, copper and zinc in the order of toxicity to invertebrates. Organolead compounds, on
the other hand, particularly the alkyl-lead compounds used as antiknock agents in gasoline,
are considered toxic to any forms of life (Denton et al., 1997)(48). Lead is a non essential
element, being a toxic metal that can affect humans when ingested or inhaled in high doses,
causing encephalopathy, colic, renal diseases, and anemia. In particular, children are
susceptible to lead toxicity, with numerous epidemiological studies reporting neurocongitive
functions to be inversely correlated with blood or tooth lead levels (USCDC, 1991)(94). Lead
enters the body through breathing or swallowing and the main target for lead toxicity is the
nervous system. Lead exposure may also cause weakness in fingers, wrists, or ankles and
miscarriage for pregnant women (Finkelman, 2005)(64). Lead is toxic to the nervous system of
human beings, especially children. It is readily absorbed from the intestinal tract and
deposited in the central nervous system. High lead levels in blood (more than 10µg/dl) may
contribute to learning disabilities, nervous system damage, and stunted growth (Fergusson,
1990)(3). The chronic effect of lead on man includes neurological disorders, especially in the
foetus and children. This can lead to behavioral changes and impaired performance in IQ
tests (Needleman, 1987)(95). Lead toxicity leads to anaemia both by impairment of haemobiosynthesis and acceleration of red blood cell destruction. Both are dose related. Lead also
depresses sperm count (Anglin-Brown et al., 1995)(96). In addition, Pb can also produce a
damaging effect on the kidney, liver, male and nervous system, blood vessels and other
tissues (Sharma and Pervez, 2003)(93).
Nickel (Ni)
Nickel is a ubiquitous trace metal and occurs in soil, water, air, and in the biosphere.
Levels in natural waters have been found to range from 2 to 10 µg L-1 (fresh water) and from
0.2 to 0.7 µg L-1 (marine). The prevalence ionic form is nickel (II) (WHO, 1991)(97). The
common aqueous species found in water is predominantly Ni 2+. It is suspected to be essential
trace elements for plants and animals. Nickel may be present in some ground waters as a
consequence of dissolution from nickel ore-bearing rocks. Nickel is used principally in its
metallic form combined with other metals and nonmetals as alloys. Nickel alloys are
characterized by their hardness, strength and resistance to corrosion and heat. Nickel is also
used mainly in the production of stainless steels, non-ferrrous alloys and super alloys. Other
uses of nickel-cadmium batteries, in coins, in welding products. It is estimated that 8% of
nickel is used for household appliance. Nickel is also incorporated in some food supplements,
which can contains several microorganisms of nickel per tablet (WHO, 2004)(73).
Nickel is moderately toxic to most species of aquatic plants, though it is one of the
least toxic inorganic agents to invertebrates and fish. The major source of discharge to natural
waters is municipal wastewater followed by smelting and the refining of nonferrous metals
(Denton et al., 2001)(66). Also mine drainage effluents are known to be major contributors due
to high concentrations of nickel found in the discharges (Finkelman, 2005)(64). Typically,
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nickel residues in sediments can be up to 100 µgg-1 or higher but may fall below 1 µgg-1 in
some clean coastal waters (Denton et al., 1997)(48) with the average concentration of nickel in
the lithosphere of 55 (Callender, 2003)(65).
Major uses of nickel include its metallurgical use as an alloy, plating and
electroplating, a major component of Ni-Cd batteries, and as a catalyst for hydrogenating
vegetable oils (Callender, 2003)(65). Most nickel is used for the production of stainless steel
and other nickel alloys with high corrosion and temperature resistance. Nickel alloys and
nickel plating are used in vehicles, processing, machinery, armaments, tools, electrical
equipment, household appliances, and coinage. Nickel compounds are also used as catalyst,
pigments, and in batteries. The primary sources of nickel emissions into the ambient air are
the combustion of coal and oil for heat or power generation, the incineration of waste and
sewage sludge, nickel mining and primary production, steel manufacture, electroplating, and
miscellaneous sources such as cement manufacturing. Nickel from various industrial
processes and other sources finally, reaches waste water. Residues from waste-water
treatment are disposed of by deep well injection, ocean dumping, land treatment, and
incineration (WHO, 1991)(97).
Entry into the aquatic environment is by removal from the atmosphere, by surface runoff, by discharge of industrial and municipal waste, and also following natural erosion of soils
and rocks. In rivers, nickel is mainly transported in the form of a precipitated coating on
particles and in association with organic mater (Schuz and Zabel, 2000)(79). In the bottom
sediments of estuaries in which anaerobic conditions often occur, sulfide tends to control the
mobility of nickel. However, under aerobic conditions, the solubility of nickel is mainly
controlled by either the co-precipitate NiFe2O4 or Ni(OH)2(s) (Callender, 2003)(65).
Although the role of nickel in the human body is not known, it is an essential nutrient
for other mammalian species and has been suggested to be essential for human nutrition as
well. Following data from animals, it is estimated by extrapolation that a 70kg person would
have a daily requirement of 50 µg per kg nickel. Since nickel is an element of the crust of the
earth, it exists in food, air, and water. Food is believed to be the major source of nickel
exposure. In the US, an average intake for adults of approximately 100 to 300 micrograms
per day (µg/d) is estimated. Nickel production, processing, and use, contact with everyday
items such as nickel-containing jewelry and stainless steel cooking and eating utensils, and
smoking tobacco are major sources of exposure of nickel to humans (USEPA, 2001)(98).
Nickel and its compounds have little toxicity. Nickel itch or contact dermatitis is the most
commonly seen reaction to nickel compounds especially in women due to use of nickel in
costume jewelry, especially earrings. Chronic exposure to nickel causes cancer in the
respiratory tract and the lungs (Fergusson, 1990)(3). Among Nickel's properties, its
mutagenic, immunotoxic, carcinogenic and teratogenic properties are the most important.
Due to efficient phagocytosis into the cells, solid nickel compounds are the most toxic, at
least in regard to carcinogenicity and mutagenicity. In addition, water soluble salts, like
NiCl2, are also mutagenic. Nickel salts have been found to be harmful for reproduction of
laboratory rodents. Embryonal and fetal mortality have been found to increase when females
have been exposed to nickel (Kakela et al., 1999; Snodgrass, 1980)(99, 100). Some of the most
serious health effects due to exposure to nickel include reduced lung function some nickel
compounds are reported to be carcinogenic to humans and metallic nickel may also be
carcinogenic (Finkelman, 2005)(64).
Zinc (Zn)
The average abundance of Zn in Earth's crust is 76µgg-1; in soils it is 25-68
µgg-1. Zinc is used in a number of alloys including brass and bronze, batteries, fungicides and
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pigments. Zinc is an essential growth element for plants and animals but at elevated levels it
is toxic to some species of aquatic life (Eaton and Franson, 2005)(101). In addition, Zn is
involved in a variety of enzyme systems which contribute to energy metabolism, transcription
and translation. Zinc is also potentially hazardous and excessive concentrations in soil lead to
phytotoxicity as it is a weed killer (Abbasi et al., 1998)(78). Zinc is used in galvanizing steel
and iron products. Zinc oxide, used in rubber as a white pigment, for example, is the most
widely used zinc compound. Peroral zince is occasionally used to treat zinc deficiency in
humans. Zinc carbamates are used as pesticides (Anglin-Brown, 1995)(96).
Zinc is one of the most ubiquitous and mobile of the trace metals and it transported in
natural waters in both dissolved forms and associated with suspended particles (Mance,
1987)(102). In river waters, zinc is predominantly present in the dissolved form. In estuaries,
where concentrations of suspended particles are greater, a greater proportion of the zinc is
adsorbed to suspended particles (CCREM, 1987)(103). In low salinity areas of estuaries, zinc
can be mobilized from particles by microbial degradation of organic matter and displacement
by calcium and magnesium. In the turbidity maximum, zinc associated with suspended
sediment will be deposited with flocculated particles where it can accumulate particularly in
anaerobic sediments (Schulz and Zabel, 2000)(79). In seawater, much of the zinc is found in
dissolved form as inorganic and organic complexes. However, zinc (and nickel) are less toxic
metals. Zinc is the most abundant trace metals in human body. It functions as a cofactor
where many enzymes depends upon it as well as body cells (Al-Ouran, 2005)(104).
Zinc is widely used in modern society, most commonly to coat or galvanise iron to
prevent corrosion. It is also mixed with other metals to form alloys such as brass. Particles
released from vehicle tyres and brake linings are a major source of zinc in the environment
(WHO, 2001)(105). Zinc is a very common environmental contaminant and it is commonly
found in association with lead and cadmium (Finkelman, 2005)(64). It is used in coating to
protect iron and steel, in alloys for die casting, in brass, in strips of dry batteries, in roofing
and in some print processes (Al-Ouran, 2005)(104). Although it is not regarded as particularly
toxic, it is sometimes released into the sea in substantial quantities (Denton et al., 2001)(66). It
may enter the aquatic environment through natural or anthropogenic sources, including
sewage and industrial discharge (Al-Ouran, 2005)(104). Major sources of Zinc to the aquatic
environment include the discharge of domestic wastewaters; coal-burning power plants;
manufacturing processes involving metals; and atmospheric fallout (Denton et al., 2001)(66).
Approximately one third of all atmospheric zinc emissions are from natural sources, the rest
come from nonferrous metals, burning of fossil fuels and municipal wastes, and from
fertilizer and cement production (Callender, 2003)(65).
Zinc is an essential micronutrient in all marine organisms, being a cofactor in nearly
300 enzymes. Therefore, marine animals are able to regulate tissue Zn at the concentrations
in sea water and sediment from normal ambient levels to incipient lethal levels (Franca et al.,
2005)(106). Several crustacean are able to regulate the uptake of zinc but, at higher
concentrations, this process appears to breakdown leading to an influx of zinc (Schulz and
Zabel, 2000)(79).
Zinc is the second most abundant trace element in the body (after iron), and the brain
has its share, with about 10 mg of zinc in a gram of wet tissue (Choi and Koh, 1998)(107). The
required daily intake for the human body ranges from 10 to 20 mg (Smith et al., 1976)(108).
Zinc is an essential structural element of many proteins. Neural activity releases zinc as a
signaling messenger at many central excitatory synapses. New evidence suggests that zinc
may also be a parameter in neuronal death associated with transit global ischemia, sustained
seizures, and perhaps other neurological disease states (Choi and Koh, 1998)(107). Excessive
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zinc intake may inhibit copper absorption and lead to copper deficiency. Acidic beverages
packaged in galvanized containers may produce toxic zinc concentration levels, causing
nausea, vomiting, stomach cramps, and diarrhea (Fergusson, 1990)(3). Taking excess zinc into
the body through food, water and dietary supplements can have adverse effects on health.
Ingesting high levels of zinc for several months may cause anemia, damage to pancreas, and
decrease levels of high-density lipoprotein (HDL) cholesterol (Finkelman, 2005)(64).
Zinc is an essential nutrient for the human body and has an importance for health
(Hotz et al., 2003)(109). Zinc acts as a catalytic or structural component in many enzymes that
are involved in energy metabolism and in transcription and translation of RNA (Moolenaar,
1998)(110). Zinc also has a prominent role in determining the outcome of pregnancies
and supporting neurobehavioral development (Hotz et al., 2003)(109). However, like other
metals, it can be toxic in high concentrations (ANZECC/ARMCANZ, 2000)(111). Although
uncommon, gastrointestinal distress and diarrhoea have been reported following ingestion of
beverages standing in galvanized cans or from use of galvanised utensils (WHO, 2001)(105).
Other symptoms of Zn toxicity are slow reflexes, shakes, paralyzation of extremities,
anaemia, metabolic disorder, terratogenic effects and increased mortality (Klaassen,
1996)(112). Taking excess zinc into the body through food, water and dietary supplements can
have adverse effects on health. Ingesting high levels of zinc for several months may cause
anemia, damage to pancreas, and decrease levels of high-density lipoprotein (HDL)
cholesterol (Finkelman, 2005)(64).
Manganese (Mn)
Mn in Earth's crust is 1600 µgg-1; in soils it is 61-1060 µgg-1. The common aqueous
species found in water is predominantly M2+ and M4+. Manganese is essential for plants and
animal. Manganese dioxide and other manganese compounds are used in products such as
batteries, glass and fireworks. Potassium permanganate is used as an antioxidant for cleaning,
bleaching and disinfection purposes. Manganese greensands are used in some locations for
potable water treatment. An organic manganese compound, Methylcyclopentadienyl
Manganese tricarbonyl (MMT), is used as an octane-enhancing agent in unleaded petrol.
Other manganese compounds are used in fertilizers, varnish and fungicides and as livestock
feeding supplements. Manganese can be adsorbed onto soil, the extent of adsorption
depending on the organic content and cation exchange capacity of the soil. It can
bioaccumulate in lower organisms (e.g. phytoplankton, algae, mollusks, and some fish) but
not in higher organisms; bio-magnification in food chains is not expected to be very
significant (WHO, 2004)(73). Manganese is the 11th element in terms of abundance in the earth
crust (Anschutz et al., 2005)(113). Oxidized forms of Mn are very reactive and have a strong
capacity for adsorption of trace metals and play and important role as both an electron donor
and acceptor in redox processes of aquatic environments (Chester et al., 1988)(114).
Manganese is a naturally occurring element in the environment. It comprises about
0.1% of the earth’s crust (NAS, 1973)(115), and found in rock, soil, water, and food. Thus, all
humans are exposed to manganese, and it is a normal component of the human body. Food is
usually the most important route of exposure for humans. Manganese is released to air mainly
as particulate matter, and the fate and transport of the particles depend on their size and
density and on wind speed and direction. Some manganese compounds are readily soluble in
water, so significant exposures can also occur by ingestion of contaminated drinking water.
Manganese in surface water can oxidize or adsorb to sediment particles and settle to the
bottom. Manganese in soil can migrate as particulate matter to air or water, or soluble
manganese compounds can be leached from the soil (Erdogan, 2009)(87).
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Manganese does not occur naturally as a base metal but is a component of more than
100 minerals, including various sulfides, oxides, carbonates, silicates, phosphates, and
borates. The most commonly occurring manganese-bearing minerals include pyrolusite
(manganese dioxide), rhodocrosite (manganese carbonate), and rhodonite (manganese
silicate) (NAS, 1973)(115). Natural levels of manganese in soil range from 40 to 900 µgg-1,
with an estimated mean of 330 µgg-1 (Rope et al., 1988)(116). Accumulation of manganese in
soil usually occurs in the subsoil and not on the soil surface (WHO, 1981)(117).
Manganese is an essential trace element for both animals and man necessary for the
formation of connective tissue and bone, and for growth, carbohydrate and lipid metabolism,
the embryonic development of the inner ear, and reproductive functions. Some specific
biochemical functions of manganese have been discovered such as the catalysing of the
glucosamine-serine linkages in the synthesis of the mucopolysaccharides of cartilage.
Terrestrial mammals however, may concentrate available manganese up to a factor of 10,
whereas fish and marine plants concentrate it by factors of 100 and 100 000, respectively
(Preston et al., 1972)(118). The accumulation of manganese in living organism has the potential
of reaching toxic levels. Symptoms of the Manganese toxicity in man include dullness, weak
muscles, headaches and insomnia. In chronic inhalation exposure to manganese, the main
organ systems affected are the lungs, nervous system, and reproductive system, although
effects on other organ systems have also been observed (Erdogan, 2009)(87).
Iron (Fe)
Iron is the most abundant metal, and is believed to be the tenth most abundant element
in the universe. Iron is a metal extracted from iron ore, and is hardly ever found in the free
(elemental) state. Iron is the most used of all the metals, comprising 95 percent of all the
metal tonnage produced worldwide. Its combination of low cost and high strength make it
indispensable, especially in applications like automobiles, the hulls of large ships, and
structural components for buildings. Steel is the best known alloy of iron (Greaney, 2005)(119).
Iron is the fourth most abundant element in the Earth's crust. While it is naturally released into
the environment from weathering, it may also be released into the aquatic environment
through human activities, such as burning of coke and coal, acid mine drainage, mineral
processing, sewage, iron related industries and the corrosion of iron and steel (CCREM,
1987)(103).
Iron is essential to all organisms, except for a few bacteria. It is mostly stably
incorporated in the inside of metalloproteins, because in exposed or in free form it causes
production of free radicals that are generally toxic to cells. Iron binds avidly to virtually all
biomolecules so it will adhere nonspecifically to cell membranes, nucleic acids, proteins etc.
Iron distribution is heavily regulated in mammals. The iron absorbed from the duodenum
binds to transferrin, and carried by blood it reaches different cells. It is strongly advised not
to let the chemical enter into the environment because it persists in the environment. Excess
iron in the body causes liver and kidney damage (haemochromatosis). Some iron compounds
are suspected carcinogens (Greaney, 2005)(119). Iron toxicity is due to its rapid absorption by
the body. Drinking water, iron pipes, cookware, and preparations are the main sources of iron
and its target organs are the liver, cardiovascular system, and kidneys (Almoray and Belhadj,
2007)(120).
Iron, one of the most abundant metals on Earth, is essential to most life forms and to
normal human physiology. Iron is an integral part of many proteins and enzymes that
maintain good health (Institute of Medicine, 2001)(121). In humans, iron is an essential
component of proteins involved in oxygen transport (Dallman, 1986)(122). It is also essential
for the regulation of cell growth and differentiation (Andrews, 1999)(123). A deficiency if iron
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limits oxygen delivery to cells, resulting in fatigue, poor work performance, and decreased
immunity (Institute of Medicine, 2001)(121). On the other hand, excess amount of iron in man
can result in toxicity and even death (Corbett, 1995)(124). There is considerable potential for
iron toxicity because very little iron is excreted from the body. Thus, iron can accumulate in
body tissues and organs when normal storage sites are full. For example, people with
hemachromatosis are at risk of developing iron toxicity because of their high iron stores.
Symptoms of Alzheimer's and Parkinson's disease may also be iron-related (Corbett,
1995)(124). High iron concentrations affect vital organs in humans including the liver,
cardiovascular system and kidneys (Alabaster and Lloyd, 1980)(125).
2.2.2. Pollution of the Aquatic Environment with Trace Metals
The aquatic environment with its water quality is considered the main factor
controlling the state of health and disease in both man and animal (Rashed, 2004)(54).
Nowadays, the increasing use of the waste chemical and agricultural drainage systems
represents the most dangerous form of chemical pollution particularly heavy metal pollution.
The most important heavy metals from the point of view of water pollution are Zinc (Zn),
Copper (Cu), Lead (Pb), Cadmium (Cd), Mercury (Hg), Nickel (Ni) and Chromium (Cr)
(Rashed, 2004)(54).
When trace metals enter the aquatic environment, the metal ions can react with
constituents of the water or settle to the bottom and react with the sediments. Trace metals
have a greater chance of remaining in solution when complexed to chelating ligands such
as specific anions whose concentrations are described by the pH of the surrounding
environment. Metals precipitate as oxides/hydroxides at different pH regions and the
amphoteric elements return to solution at higher pH. The hydroxide concentration (or pH) is
therefore of great importance for the mobility of metals. Other factors also affect the fate of
the metal ions like redox conditions and the presence of adsorbent sediments (Alloway and
Ayres, 1998)(126).
There are basically three reservoirs of metals in the aquatic environment: water,
sediment and biota. Metal levels in each of these three reservoirs are dominated by a complex
dynamic equilibrium governed by various physical, chemical and biological factors (Murray
and Murray, 1973)(127). Among these three reservoirs, sediment is the major repository for
metals, in some cases, holding over 99% of the total amount of metal present in the system
(Renfro, 1973)(128). The analysis of sediment is a useful method of studying aquatic pollution
with trace metals (Batley, 1989)(129). Contaminated sediments, in both freshwater and marine
systems, are a significant issue worldwide. Contaminants can persist for many years in
sediments, where they have the potential to adversely affect human health and the
environment. Some chemicals continue to be released to surface waters from industrial and
municipal sources and polluted run-off streams from urban agricultural areas and build up
harmful levels of contamination in sediments (Mackeviciene et al., 2002)(130). The
enrichment of metals in sediments is influenced by allocthonous influence which is made up
of natural and anthropogenic effects and autochthonous influences comprising of
precipitation, sorption, enrichment of organism and organometallic completion during
sedimentation as well as the post depositional effects of digeneses (Forstner and Witlmann,
1979)(131).
2.2.2.1. Trace metals in the water column
Sources of trace metals in the water column
Trace metals in the water column are contributed by natural and anthropogenic
sources. They enter the water body naturally from the atmosphere and via run-off and, in the
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case of larger water bodies such as lakes and oceans, from smaller streams and rivers. Some
of the anthropogenic sources of trace metals include industrial wastes and by-products
generated by mining and smelting, production and use of materials containing the heavy
metals, burning of fossil fuels, leaching of waste dumps, urban run-off, sewage effluent,
shipping, waste dumping and agriculture run-off.
Trace metal partitioning – dissolved and particulate phases
In the marine environment, trace metals are partitioned amongst dissolved phases,
suspended and bottom sediments and biota in the water column. According to Elder
(1988)(132) the fractionation of trace metals depends on many factors including organic matter
composition, pH, salinity and binding affinities of heavy metals. The dissolved fraction that
represents the principal source of bio-available metals is favored under conditions of low pH,
low particulate loads and high concentrations of dissolved organic matter. Low pH is
particularly important because:
(i) The solubility of metal hydroxides increases as pH decreases;
(ii) The adsorption capacity of solid surfaces decreases; and
(iii) H+ ions compete with metals for coordination sites on organic molecules.
More trace metals may also enter solution as water hardness increases since cations
(especially Ca2+ and Mg2+) also compete with metals for binding sites. However, increasing
salinity usually results in reduced dissolved metal concentrations due to clay-organic particles
forming flocs with a high settling velocity. High pH and Eh as well as elevated particulate
organic matter concentrations favor metal partitioning to bottom sediments, or to suspended
particulate phases if hydraulic energy is high enough (Elder, 1988)(132).
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Metal partitioning in water column among 3 major phases (inside triangle) and some
environmental conditions that favor each phase (outside triangle)(Elder, 1988)
Distribution and behavior of trace metals in the water column
A number of studies have been carried out in order to investigate the distribution and
fate of trace metals in the water column. Zeri and Voutsinou-Taliadouri (2003)(133) indicated
that the distribution and behavior of trace metals in the seawater column and sediments were
controlled by many complicated physiochemical processes such as hydrodynamic mixing,
adsorption onto both inorganic and organic phases, complexation, precipitation, biological
uptake and diffusion from bottom sediments. Leivuori (1998)(134) described in a study of
trace metals in water column, that 11 m above the bottom, 77% of suspended particulate
matter was originated from re-suspension of sediments. It is well known that decomposition
of sinking organic matter, oxygen depletion in surface sediments and oxic/anoxic conditions
play an important role in heavy metal burying in sediments and re-mobilisation processes
from sediments into the water column (Sullivan and Drever, 2001)(135). Sokolowski et al.
(2001)(136) showed that particulate organic matter, chlorophyll a and iron and manganese
oxyhydroxides govern the behavior of trace metals in the water column.
Suspended particulate matter in the water column is one of the main sources of trace
metals in the marine ecosystem, and plays an important role in the transport and storage of
potentially hazardous metals. The processes controlling trace metal concentrations of
suspended particulate matter are generally well known, but the relative importance of the
different processes is poorly quantified. In marine ecosystems, particulate organic and
inorganic toxic pollutants including trace metals enter the water column via atmospheric
input, river runoff, local point sources and bottom sediment re-suspension (Nguyen et al.,
-35-
2005)(137). Trace metals are mainly bound to fine-grained particles of mineral or organic
origin as well as to iron and manganese sulphides and oxyhydroxides. Adsorbed on
particulate matter, trace metals will ultimately reach the sediments where they become
permanently buried or released to the water column, either as re-suspended material or in a
dissolved form after geochemical transformation in the sediments (Leivuori et al., 2000)(138).
The interaction of dissolved trace metals with suspended particulate matter in seawater
has been suggested as the major controlling factor affecting the concentration and distribution
of trace metals throughout the water column (Sherrell and Boyle, 1992)(139). In general, it is
clear that vertical and horizontal variation in dissolved trace metal concentrations result from
particle formation, decomposition, and transport superimposed upon physical mixing and
advection processes. Due to the abundance of the suspended particulate matter and their large
available surface area, it has been suggested that they control exchange with dissolved metals.
Several studies have described the distribution of trace metals that are relatively abundant in
suspended particulate matter (Bishop and Fleisher, 1987)(140).
2.2.2.2. Trace metals in seawater
Metals occur naturally in waters due to geologic leaching and diffusion from minerals
(Aguilera and Amils, 2005)(141). However, anthropogenic contamination has resulted in
increased concentrations of metals, including Ag, Al, Cd, Cr, Cu, Fe, Hg, Mn, Ni, Pb, Sn, and
Zn, and metalloids, including As and Se, which will hereafter be referred to as metal
contaminants. Anthropogenic metal contaminants are released into all three compartments of
the environment (air, water, and soil/sediment), and may cause metal concentrations to rise
greatly above natural background levels. Sources of anthropogenic metal contamination in
waters include industrial effluent discharges, urban runoff (e.g. roads, building roofs), sewage
outfalls, leaching from municipal waste, leaching of ship antifouling coatings, mining
discharges, sediment dredging activities, atmospheric desorption of emission (e.g. industrial,
transport, mining, power station), and from the use of fertilizers and pesticides (Warnken et
al., 2004)(142). The concentrations of metals in aquatic systems may vary greatly between
locations depending on the proximity to natural and anthropogenic sources and may decrease
rapidly with time and distance from a source, due to dilution and dispersion (Aguilera and
Amils, 2005)(141). Spatial and temporal fluctuations in metal concentrations may also be
caused by changing environmental processes/parameters, such as water flow rate, pH,
salinity, turbidity, temperature, and organic matter (Baeyenes et al., 2005)(143).
Metal contaminations are distinctive as they are persistent in the environment and are
not removed by degradation. Instead they undergo conversions between different chemical
forms. These conversions may be mediated by both biotic (micro-and macro-organsims) and
abiotic (environmental) mechanisms (Hirose, 2006)(144). The distribution of metals between
different chemical forms or species is referred to as metal speciation (Apte et al., 2005)(145).
Chemical species may vary due to differences in isotopic composition, electronic or oxidation
state, and/or molecular structure (Smith et al., 2002)(146).
Metals can exist in seawater at least in 4 different forms namely in true solution, as
colloidal particles, adsorbed on other colloidal particles, and as a part of living organisms.
Their dispersion in seawater is affected by properties such as their volatility, tendency to form
insoluble complexes with inorganic and organic compounds, and their ability to forms either
soluble complexes with ions common in seawater or oxides or sulphides of low solubility
(Tessier et al., 1979)(15). Trace metals dissolved in seawater are present at very low
concentrations and contamination during sampling, storage and analysis often gives erroneous
results (Forstner, 1979a)(147).
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2.2.2.3. Trace metals in marine biota
Essential metals such as Cu, Zn, Fe, Co, have important biochemical functions in the
organism and form either an electron donor system or function as ligands in complex
enzymatic compounds (Fostner, 1979b)(148). In natural marine environment their enrichment
in organisms does not exceed the level to interfere with the enzyme system functioning
through their concentrations in organisms are generally higher than in seawater. However, if
the ambient water or food contains high concentration of these metals, the homeostatic
mechanism ceases to function and the essential trace metals act in an either actually or
chronically toxic manner (Nelson and Donkinn, 1985)(149).
Simkiss and Taylor (1989)(150) have discussed the pathways of metal accumulation by
aquatic organisms and suggested six possible types of uptake. The most common is by a
passive process of transfer from seawater down a concentration gradient into the tissues. In
some cases uptake may also occur through ion pumps because of energy dependency. For
many metals including Cd, Cu and Zn, the free metal ion is the most bioavailable form of the
element (Goldberg et al., 1976)(151). In case of metals such as Hg and As which can be
alkylated in marine environment, the biological uptake is likely to occur in the alkylated form
(Turner and Southworth, 1999)(152). Autotrophic organisms probably accumulate trace metals
directly from seawater. Hence, environmental parameters such as temperature, salinity, light
and pH; and chemical factors such as speciation and complexation of the metal are of
considerable significance. The presence of several potentially toxic trace metals can also
affect their preferential uptake (FAO, 1976)(153).
Variations in concentration of trace metals between species particularly in
heterotrophs can be due to trophic level relationships. Though magnification of trace metals
through the food web is important, there are several other causes affecting the transfer of trace
metals in marine organisms. There is also evidence of seasonal variations in the trace metal
content with species. The biological uptake, retention and translocation of trace metals in
marine biota are related to chemical changes in storage tissues with the organism (FAO,
1976)(153). Thus, methyl-Hg assimilated through water or from metabolic conversion of Hg,
accumulates in the muscle tissues of carnivore fish while the ionic Hg is preferentially
transferred in their liver or spleen. Unlike Hg, Cd does not appear to concentrate in fish flesh,
but accumulates in the gills, liver and gastro intestinal tract. Much of the As in marine
organisms is in the form of organic As compounds (Culler and Reimer, 1989)(154).
2.2.2.4. Trace metals in marine sediments
Marine sediments constitute part of the contaminants in aquatic environments. The
bottom sediment serves as a reservoir for trace metals, and therefore, deserves special
consideration in the planning and design of aquatic pollution research studies. Trace metals
such as cadmium, mercury, lead, copper, and zinc, are regarded as serious marine pollutants
because of their toxicity, tendency to be incorporated into food chains, and ability to remain
in an environment for a long time (Puyate et al., 2007)(155).
Sediments are known to act as the main sink for trace metals in coastal ecosystems
that are impacted by anthropogenic activities. The concentration of trace metals in sediments
can be influenced by variation in their texture, composition, reduction/oxidation reactions,
adsorption/desorption, and physical transport or sorting in addition to anthropogenic input
(Basaham and El-Sayed, 1997)(156). Potentially, toxic compounds, especially trace metals, are
adsorbed on mineral or organic particles either in their organic or inorganic forms (Forstner
and Wittman, 1983)(157). Studies on the distribution of trace metals in sediments and other
media are of great importance in the context of environmental pollution (Howari, 2005)(158).
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Metals in marine sediments can originate from several sources. In inshore and coastal
areas the major source of metals is lithogenous associated with weathering products from the
source rock. Hydrogenous formations that include precipitation products and metals adsorbed
on particles formed due to physico-chemical changes in water also contribute to the metal
burden in marine sediments (Riley and Chester, 1971)(31). Anthropogenically introduced
metals are ultimately transferred to the bed sediments (Marchand et al., 2006)(159). Hence the
environmental geochemistry requires knowledge of the naturally occurring metals in order to
assess man's environmental impact. Although sediment analyses do not represent the extent
of toxicity, they are useful to assess the burden of anthropogenic component over and above
the lithogenic background and also in some instance, trace the sources of pollution long after
input has taken place (Buccolieri et al., 2006)(160).
Sediments of rivers, lakes and estuaries in a large number of locations have been
contaminated by inorganic and organic materials. Among the inorganic materials, metals are
frequent and important contaminants in aquatic sediments. They are involved in a number of
reactions in the system including sorption and precipitation, and they are greatly influenced
by redox conditions in the sediments (Allen, 1995)(161). Heavy metals are transported as either
dissolved species in water or an integral part of suspended solids. They may be volatilized to
the atmosphere or stored in riverbed sediments. They can remain in solution or suspension
and precipitate on the bottom or can be taken up by organisms (Topcuoglu et al., 2002)(162).
The heavy metal content of sediments comes from natural sources (rock weathering, soil
erosion, dissolution of water-soluble salts) as well as anthropogenic sources such as municipal
wastewater-treatment plants, manufacturing industries, and agricultural activities (Güven and
Akıncı 2008)(44).
Sediments accumulate hazardous trace elements to levels many times higher than
water column concentrations, causing a serious problem due to their toxicity and their ability
to accumulate in biota (Morillo et al., 2007)(163). The concentrations of trace metals present in
bottom sediments typically exceed the concentrations in the overlying water by between three
and five orders or magnitude. In such situations, the contaminated sediments can represent a
significant, long term source of trace metals to the overlying water column and the aquatic
ecosystem. The sediment particles themselves can also represent a bioavailable source of
trace metals, thereby causing direct toxicity to benthic biota or providing a pathway for trace
metal entry into aquatic food chains (Burton, 2005)(55).
Sediments are heterogeneous assemblages of minerals and organic components and
therefore exhibit a wide degree of variability in terms of composition and physico-chemical
properties (Berner, 1981)(164). Trace metals may be distributed within a range of geochemical
solid phases, as well as being dissolved in the associated sediment pore waters (Hong et al.,
1995)(165). Consideration of the partitioning of trace metals between pore water and the
various solid phases in natural sediments provides a far more realistic and accurate
description of environmental risk compared to total metal concentrations (Berry et al.,
1996)(166). Trace metal distribution amongst the various sediment phases is controlled by a
range of sorption-desorption reactions, which include precipitation-dissolution of mineral
species and adsorption-desorption on mineral and organic surfaces (Tessier et al., 1994)(167).
When environmental conditions change (pH, sediment redox potential, organic matter, etc.);
some of the sediment bound metals can remobilize and be released back into the water, where
they can have adverse effects on living organisms (Song et al., 2009)(168). Metals are
associated with different phases (crystalline minerals, carbonates, hydrous metal oxides,
organic substances, etc.) that determine their behavior in the environment, mobility,
bioavailability, and toxicity (Usero et al., 1998)(169). The most crucial property of metal ions
is that they are bioavailable and not biodegradable in the environment and that their uptake by
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benthic organisms depends largely on their mobility, total concentration and chemical forms
(Morillo et al., 2007)(163). Thus, to assess the environmental impact of polluted sediments, the
measurement of total trace metals is not enough (Kersten and Forstner, 1987)(170).
2.2.3. The importance of studying sediments
Sediment is the loose sand, clay, silt and other soil particles that settle at the bottom of
body of water (Davies and Abowei, 2009)(171). The erosion of bedrock and soils leads to
accumulation of sediments of past or on-going natural and anthropogenic processes and
components. Data from sediments can provide information on the impact of distant human
activity on the wider ecosystem. The composition of sediment sequences provides the best
natural archives of recent environmental changes (Olubunmi, 2010)(172).
Sediment is a habitat and major nutrient source for aquatic organisms. Sediment
analysis is important in evaluating qualities of total ecosystem of a body of water in addition
to water sample analysis practiced for many years because it reflects the long term quality
situation independent of the current inputs (Adeyemo et al., 2008)(173) and it is the ultimate
sink of contaminants in the aquatic system (Davies and Abowei, 2009)(171). Monitoring of the
contamination of soil and sediment with trace metals is of interest due to their influence on
ground water and surface water and also on plants, animals and humans (Suciu et al.,
2008)(174). Accumulation of trace metals occur in upper sediment in aquatic environment by
biological and geochemical mechanisms and become toxic to sediment-dwelling organisms
and fish, resulting in death, reduced growth, or in impaired reproduction and lower species
diversity (Praveena et al., 2007)(175). Trace metals also occur naturally in rock forming
minerals and ore minerals; hence they can reach the environment from natural processes
(Akinmosin et al., 2009)(176). The occurrence of metals in aquatic ecosystems in excess of
natural background loads has become a problem of increasing concern. Trace metals in
environment may accumulate to toxic levels without visible signs. This may occur naturally
from normal geological phenomenon such as ore formation, weathering of rocks and leaching
or due to increased population, urbanization, industrial activities, agricultural practices,
exploration and exploitation of natural resources (Ajayi and Osibanjo, 1981)(177).
After the Industrial Revolution, point sources from mining, municipalities, industries
and non-point sources from both agriculture and urban storm water runoff have accumulated
in water resources, settling to the bottom areas of water bodies. Contaminated sediments pose
a risk to the environment in two basic mechanisms: (1) ecological risk to aquatic and
piscivorous animals and (2) toxic risks on terrestrial habitat when contaminated area is
dredged and placed on land (Khan et al., 2000)(178).
The study of marine sediments provides useful information in marine, environmental
and geochemical research about pollution of the marine environment (Calace et al., 2005)(179).
Urban developments and industrial activities contribute to the introduction of significant
amounts of contaminants (among them trace metals) into the marine environment and affect
directly the coastal systems where they are often deposited (Dassenakis et al., 2003)(180). The
growing apprehension about the potential effects of sediment toxicity poses to marine fauna
and flora and the risk posed to the environment by the contaminants accumulated by the
sediments have aroused an increase in research interests in marine sediments (Calace et al.,
2005)(179).
Study of marine sediments represents a useful tool for determining the actual state of
environmental pollution and for understanding the origin and mechanism of the phenomena.
Sediments could be regarded as a historical reflection to changes occurring in the overlying
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water system and can be used as good indicators for metals in any study area. Meanwhile,
sediments are major repository of trace metals in the coastal marine environment (Salomonos
and Forstner, 1984)(181) and sometimes become a major potential source of trace metals,
which can be eventually returned to the biosphere (Saad and Badr, 2003)(182).
Sediment is a matrix of materials, made up of detritus, inorganic and organic particles,
and is relatively heterogeneous in terms of its physical, chemical and biological
characteristics (Sarkar et al., 2004)(183) and consist of a wide range of particle sizes, including
gravels, sand, silt and clay (Fergusson, 1990)(3). Sediments are well known to act as a major
sink for many of the more persistent organic and inorganic chemicals introduced into the
aquatic environment by atmospheric deposition, erosion of the geological matrix, or from
anthropogenic sources (such as industrial effluents, mining wastes, etc.) (Calace et al.,
2005)(179). Sediments also act as potential sources of pollution for the surrounding water and
benthic flora and fauna by releasing sorbed contaminants back to the overlying water column
should remobilization occur through any disturbance (Adamo et al., 2005)(184). Contaminants
of major concern found in the marine coastal environment include trace metals, persistent
organic pollutants (POPs), oils (hydrocarbons), nutrients and radioactive substances (UNEP,
1999)(185).
Contamination of the environment, and in particular the marine environment by trace
metals has received considerable attention as evidenced by an increase of publications
appearing in the literature (Chandra, 2002)(186). Environmental contamination from trace
metals is of major concern because they show behavior similar to those of persistent toxic
chemicals. Metals are partitioned amongst soluble phases, suspended and bottom sediments
and biota in aquatic systems (Gangaiya, 1994)(187).
Sediments play a useful role in the assessment of trace metal contamination (Gangaiya
et al., 2001)(187). The partitioning behavior of trace metals is such that they tend to accumulate
in sediments to levels that are several orders of magnitude higher than in the surrounding
water (Denton et al., 1997)(48). Further, their deposition rates are generally related to their
rates of input in the surrounding water (Forstner, 1990)(188). The analysis of trace metals in the
sediments permits detection of contaminants that may be either absent or in low
concentrations in the water column, and their distribution in coastal sediments provides a
record of the spatial and temporal history of pollution in a particular area or ecosystem
(Binning and Baird, 2001)(189). Therefore, the chemical analysis of sediments, is very
important from the environmental pollution point of view because sediment concentrates
metals from aquatic systems, and represents an appropriate medium in monitoring of
environmental pollution (Sarkar et al., 2004)(183).
Sediments are essential and integral parts of water systems. They provide the
substrate for organisms and through interaction with the overlying waters play an essential
role in the aquatic ecosystem (Burden et al., 2002)(190). They are increasingly recognized as
both a carrier and a possible source of contaminants in aquatic systems, and these materials
may also affect groundwater quality and agricultural products when disposed on land.
Contaminants are not necessarily fixed permanently by the sediment, but may be recycled via
biological and chemical agents both within the sedimentary compartment and the water
column. Bioaccumulation and food chain transfer may be strongly affected by sedimentassociated proportions of pollutants. Benthic organisms, in particular, have direct contact with
sediment, and the contaminant level in the sediment may have greater impact on their survival
than do aqueous concentrations (Malins et al., 1984)(191).
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They are also an important repository for metal pollutants that enter the sea. Sediments
provide habitats for many aquatic organisms and also a major repository for many of the more
persistent chemicals that are introduced into surface waters. In the aquatic environment, most
anthropogenic chemicals and waste materials including toxic organic and inorganic chemicals
eventually accumulate in sediments. In aquatic environments, many trace metals are
transported predominantly in association with particulate matter, and consequently, a high
concentration of trace metals is often detected in sediments in many industrialized harbors
and coastal regions around the world (Wang et al., 2007)(192).
Contaminated sediments are crucial indicators of pollution in aquatic environments
and can be defined as soils, sand, organic matter, or minerals accumulated at the bottom of a
water body (USEPA, 1998)(193). Contaminants contained in sediments can be released to
overlying waters and sediments can be important sources of contaminants in waters (Güven
and Akinci, 2008)(44). Many of the sediments in seas, rivers, lakes, and oceans have been
contaminated by pollutants. These pollutants are directly discharged by industrial plants and
municipal sewage treatment plants, others come from polluted runoff in urban and
agricultural areas, and some are the result of historical contamination (Begum, et al.,
2009)(194). Contaminated sediments can threaten creatures in the marine environment. Some
kinds of toxic sediments kill benthic organisms, reducing the food available to larger animals
such as fish. Some contaminants in the sediment are taken up by benthic organisms in a
process called bioaccumulation. When larger animals feed on these contaminated organisms,
the toxins are taken into their bodies, moving up the food chain in increasing concentrations
in a process known as bio-magnification. As a result, fish and shellfish, waterfowl, and
freshwater and marine mammals may accumulate hazardous concentrations of toxic
chemicals (Begum et al., 2009)(194).
Contaminated sediments do not always remain at the bottom of a water body.
Anything that stirs up the water, such as dredging, can re-suspend sediments. Re-suspension
may mean that all of the animals in the water, and not just the bottom-dwelling organisms,
will be directly exposed to toxic contaminants. Different aquatic organisms often respond to
external contamination in different ways, where the quantity and form of the element in water,
sediment, or food will determine the degree of Accumulation (Begum et al., 2008)(195). Many
dangerous chemical elements, if released into the environment, accumulate in the soil and
sediments of water bodies. Under certain conditions, chemical elements accumulated in the
silt and bottom sediments of water bodies can migrate back into the water. Silt can become a
secondary source of trace metal pollution (Begum et al., 2009)(194).
2.2.4 Effects of Trace Metals Contamination in Sediments
2.2.4.1. Metal polluted Sediments
Trace metals are preferentially transferred from the dissolved to the particulate phase
(Gangaiya et al., 2001)(196) and this results in the elevation of metal concentrations in estuaries
and marine sediments. Therefore, concentrations often exceed those in overlying water by
several orders of magnitude (Langston, 2000)(197). Because sediments can accumulate trace
metals, concentrations can be high and ultimately become potentially toxic (Williamson et al.,
2003)(198). Exposure and uptake of even a small fraction of sediment-bound metal by
organisms could have significant toxicological significance, in particular where conditions
favor bioavailability. In addition, increased metal concentrations in pore water may contribute
significantly to sediment toxicity (Langston, 2000)(197).
Evidence of fatal effects of metal-polluted sediments can be determined by the
absence of sensitive species or by the development of resistance mechanisms and adaptation
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in tolerant forms such like efficient excretory features in organisms (Langston, 2000)(197).
Binning and Baird (2001)(189) reported that the density and diversity of nematode communities
in the Swaetkops River estuary, South Africa, are influenced by the degree of trace metal
contamination in the sediments. Many of the metals have no known biological function in the
marine environment, but can act synergistically with other chemical species to increase
toxicity (Binning and Baird, 2001)(189). The potential impacts of accumulating levels of trace
metals can be estimated by comparing the concentrations of contaminants of interest present
in sediments with sediment quality guidelines (SQGs) (Williamson et al., 2003)(198).
The Sediment Quality Guidelines (SQGs) have been derived from large databases,
which correlate the incidence of adverse biological effects with the concentrations of
contaminants in sediments (Long et al., 1995)(199) and are used in the evaluation of sediment
contamination and potential ecotoxicological effects (Casado-Martinez et al., 2005)(200). One
of the most widely used SQGs for estuarine and marine environments, which will be applied
in this study as well, is a set of effects-range guidelines developed by the National Oceanic
and Atmospheric Administration (NOAA) (McCready et al., 2006)(201) in the U.S. (Long et
al., 1995)(199). There are two levels of risks considered under this SQG – effects low range
(ERL) and effects range-moderate (ERM) (Guerra-Garcia and Garcia-Gomez, 2005)(202).
Concentrations below the ERL value are rarely associated with biological effects while
concentrations above the ERL but below the ERM indicate a possible range in which effects
would occur occasionally. Concentrations above ERM are associated with frequent adverse
ecological effects on the benthic communities (Guerra-Garcia and Garcia-Gomez, 2005)(202).
2.2.4.2. Ecology
Given that metal concentrations in sediment are normally higher than those in the
overlying water, macrophytes, at the primary production level, rooted in these enriched
sediments tend to have higher concentrations than the sediment (Mance, 1987)(102). This is
due to the uptake of metals not just only from the sediment via the roots but also from the
surrounding water (Redfern, 2006)(203).
The effect of trace metal contaminants in the sediment on benthic organisms can be
either acute or chronic (cumulative) (Binning and Baird, 2001)(189). No matter whether metals
are essential or not, all trace metals form an important group of enzyme inhibitors when
natural concentrations are exceeded. Therefore, organisms living in or adjacent to metal
contaminated sediments may suffer toxic effects that can be fatal in highly contaminated
situations to relative abnormal metabolic adjustments at the sub-lethal level (Denton et al.,
2001)(66). In addition, metal enrichment in estuaries and coastal environments is a major
concern as heavy metals have the ability to bio-accumulate in the tissues of various biota, and
can ultimately affect the distribution and density of benthic organisms, as well as the
composition and diversity of infaunal communities (Binning and Baird, 2001)(189).
A wide range of criteria to assess the impact of metals on marine organisms have been
developed during the last few years and consequently, effects are now recognized at a much
lower levels than in earlier LC50 studies (Langston, 2000)(197). Growth, reproduction, and
recruitment are usually the processes most susceptible to metal stress (Redfern, 2006)(203).
The toxicity of a trace element to an organism depends on the metal chemical species,
its concentration and the organism being affected (Sedgwick, 2005)(34). As for the organism,
toxic effects occur when excretory, metabolic, storage, and detoxicification mechanisms are
no longer have the capacity to match uptake rates. This capacity may vary between phyla,
species, populations, even individuals and can also depend on the stage in the life history of
the organism (Langston, 2000)(197).
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2.2.4.3. Human health
The consumption of marine food is the principal path to human exposure to trace
metals. Effects on humans can be observed after either a one-off exposure to a large nonlethal dose (acute) or after repeated exposure to lower doses (chronic) (Sedgwick, 2005)(34).
Trace metal contamination has become a subject of public interest because humans
have been harmed by metal contamination (Chandra, 2002)(186). One of the well-documented
incidents of metal contamination that attracted greater public interest was an industrial
mercury contamination on Minimata Bay in Japan caused by severe health effects in
consumers of fish and shellfish from the area and led to the identification of methyl mercury
as an extremely toxic compound (Chandra, 2002)(186).
Methyl mercury can bio-accumulate in the food chain and its accumulation in some
fish represents a risk to those people who consume a large proportion of fish in their diet
(Williamson et al., 2003)(198).
2.2.5. Factors affecting trace metal distribution, concentration and
fate in sediments
Enrichment of trace metals from anthropogenic sources in the estuarine and marine
environments has become a serious human health and environmental concern. Once trace
metals are deposited in sediments, they undergo a series of physical, biological and chemical
processes. Trace metals occur in a number of different forms, mainly in the dissolved and in
the solid state (adsorbed onto surfaces of clays, element oxides, organic material, coprecipitated with sediment phases and incorporated into organic matter). The highest
proportion is usually in the solid phase. How an element is bound in sediment determines the
biologically active fraction and its fate and cycling (Erdogan, 2009)(87).
Understanding the processes affecting trace metals concentration and their fate in
sediments is important in gauging appropriate designs for sampling and monitoring programs
and planning for appropriate remediation options (Williamson and Wilcock, 1994)(204).
It has been validated that each environmental factor presents unique influence on
metals distribution in sediment. The influences of some factors, such as pH, (Oxidation
reduction potential) ORP and (Organic matter) OM are more crucial than the others. Only a
slight variation of them, the distributions of metals would be producing some significant
variations. Correspondingly, some other factors (e.g. salinity, temperature) can only alter
metals distribution to a less extent (Peng et al., 2009)(205). The relatively important factors are
introduced in the following:
2.2.5.1. Trace metals input into sediment
Increased inputs of metals in such forms available for association with sediments
ultimately result in increases in metal concentrations in sediments (Luoma, 2000)(206). Metal
enrichment in the sediments is usually located close to past and present sources of pollution
(Williamson and Wilcock, 1994)(204) and in particular the sheltered environments where
discharge is restricted (Williamson et al., 1992)(207). This is supported by numerous studies
which have demonstrated that metal inputs from human activities are reflected by metal
concentrations in sediments (Luoma, 2000)(206).
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Simply, concentrations drop with distance away from the source of input as
determined by the physical processes that dilute and redistribute the particles with which
metals associate such like the dilution with less contaminated sediments (Luoma, 2000)(206).
However, the rate of decline in concentration with distance from sources of pollution is
extremely variable (Williamson and Wilcock, 1994)(204).
2.2.5.2. The influence of grain size
Metal concentrations in marine and estuarine sediments are determined not only by
metal inputs but also by sediment characteristics. Grain size distribution has been recognized
to affect trace metal concentrations in estuarine and marine sediments (Luoma, 2000)(206).
Generally, trace metals are mainly concentrated in the silt/clay sediment fraction, consisting
of particles with a grain size <0.063 mm (Krumgalz et al., 1992)(208). The enrichment of the
silt/clay fraction by anthropogenic trace metals is due to the large specific area of this fraction
and to the strong adsorptive properties of clay minerals (Krumgalz et al., 1992)(208).
Fine grained sediment of this particular fraction has a high absorption potential
because the sediment has a large surface area to ratio, and contain large amounts of interstitial
water (Loomb, 2001)(209). The feature of clay that is of significance regarding the trace metals
is their ability to absorb metal ions by their outer sheath of hydroxyl groups (Fergusson,
1990)(3). The surface of clays may be also negatively charged which is crucial in providing
potential absorption sites for metal ions (Williamson et al., 1992)(207). Clay minerals also have
a higher surface area to volume ratio and can absorb material into their lattice framework
(Loomb, 2001)(209).
Luoma (2000)(206), however, found that the relationship of metal concentration and
grain size distribution is sometimes complicated. For example, silt/clay sizes particles in
estuarine and marine sediments are usually aggregates and particle aggregates can complicate
the relationship of metal concentration and surface area. Another potential factor that can
affect this relationship is the physical and chemical processes that may affect the degree of
accumulation of metal-reactive sediment components on the sediment surface (e.g.,
movement of water across a redox boundary may facilitate accumulation of iron and
manganese oxides) (Luoma, 2000)(206).
The accumulation of trace metals in sediment is controlled by the granular
compositions of sediment. Grain size affects the surface area, settling velocity and deposition
rate of suspended solids in the water column, as well as the degree of chemical partitioning
onto the sediment. Metals are generally found to be associated largely with the fine grain
fraction of sediment that has been traditionally used to study the pollution of trace metals in
sediments. However, trace metals sometimes accumulate in the coarse grain fraction of
sediment (Qian et al., 1996)(210). Fine-grained bottom sediments tend to accumulate
contaminants due to their sorptive nature, and thus act as an important reservoir by reducing
the toxicity potential to aquatic organisms (Fan et al., 2002)(211). Changes in sediment
chemistry, due to seabed disturbance, can result in contaminant remobilization. Subsequently,
exposure to a different chemical environment could result in desorption and transformation of
contaminants into more bioavailable or toxic chemical forms (Sturm et al., 2002)(212).
Sediments are able to store and accumulate metals to some extent, due to their
adsorption capacity. Sediments exist as complex and dynamic mixtures of mineral particles,
particular organic material (detritus) and microbes. The flocculated chemicals and
contaminants in sediments may be associated with any of these sediment fractions. The grain
size of sediments affects the accumulation of metals through both their chemical and physical
properties. The fine-grained sediments tend to show higher concentration of metals, due to
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their higher surface area. Fine sediments are also often associated with the anoxic nature,
higher concentration of sulfide, as well as elevated metals and organic contaminants
(Tabata et al., 2009)(213).
2.2.5.3. Sediment composition
One of the most important sediment characteristics is the concentration of sediment
components. The concentrations of the important metal-reactive components of sediments (Fe
and Mn hydrous oxides, sulphides and organic materials) can vary among estuarine and
marine environments (Luoma, 2000)(206) and are found to increase as sediment textures
become finer (Williamson and Wilcock, 1994)(204). The sediment components (particulate
matter) contain many phases that strongly adsorb metals such as amorphous iron oxides,
manganese oxides, and polar organic matter (humic acid, carbohydrates, proteinous material)
(Williamson et al., 2003)(198).
The nature of Fe oxides, organic materials, or Mn oxides will determine site densities
for sediment components; and the relative concentrations of Fe, Mn and organic materials will
determine aggregate differences in binding site density or binding intensity (Luoma,
2000)(206). Amorphous iron hydrous oxide (FeOOH) has a large surface area per unit weight.
It is abundant and it may be an important medium for trace metal adsorption. Concentrations
do not change greatly with depth or redox conditions, though are strongly dependent on
sediment texture (Williamson and Wilcock, 1994)(204). Amorphous manganese oxide
(MnOOH) concentration is lower than iron oxides, though its rapid cycling between reduced
and oxidized forms, and between sediments and water mean that it is important in trace metal
mobilization from sediments (Williamson and Wilcock, 1994)(204). Generally organic matter
content of sediments increase as the sediment texture becomes finer (Denton et al., 2001)(66).
The presence of organic matter can potentially increase metal concentrations in sediment by
adsorption of metals from surrounding environment onto organic material (Loomb, 2001)(209).
Also dead organisms in sediments may carry the trace metals with them, either taken in by the
organism while alive or sorbed on to the animal before or after death (Fergusson, 1990)(3) and
this contribute directly to the metal levels in the sediments. Organic compounds containing
metal ions may also be sorbed onto Fe-Mn oxides. Organic materials can affect metal species
solubilization by complexing the metal ions, but they can also take metal ions out from the
solution and contribute to the sediments. Decomposition of organic material produces organic
ligands that may extract metals from the sediments which can effectively mobilize metals by
increasing their concentration in the water (Taulis, 2005)(214). Interestingly, the IrvingWilliams series of increasing stability of metal complexes, is the same order as increasing
mobilization (Fergusson, 1990)(3); Mn < Fe < Co < Ni < Cu > Zn. Therefore, the subsequent
fate of metal-reactive components of sediment (Fe and Mn hydrous oxides, and organic
materials) will determine the fate of the associated heavy metals (Williamson et al., 2003)(198).
2.2.5.4. The influence of pH values
The pH is a key parameter controlling trace metal transfer behavior in sediment.
Normally, with pH decreasing in sediment, the competition between H and the dissolved
metals for ligands (e.g. OH-,CO32-,SO42-,Cl-,S2- and phosphates) becomes more and more
significant. It subsequently decreases the adsorption abilities and bioavailability of the metals,
and then increases the mobility of trace metal. Sometimes, only with a few lower of pH units,
the fixation percentage of trace metals on sediment particles may range from almost a 100%
to none (Gundersen and Steinnes, 2003)(215). In sediment, due to the OM degradation and the
acid volatile sulfide oxidation, the pH of sediment usually decreases from the neutral in the
initial to acid, sometimes even decreasing to pH 1.2, which often results in some metals being
released into water again even under stable water conditions (Uta and Jens, 2006)(216). In
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sediment, there exists a limit pH controlling heavy metal mobility. And the trace metal would
be released only as they reaching such pH value. For different metals, this limit pH is
different correspondingly, which can be found as the following:
The limit pH values of different metals in sediment
Metal species
Zn
Cd
Ni
As
Cu
Pb
Al
Fe
pH limit
6.0-6.5
6.0
5.0-6.0
5.5-6.0
4.5
4.0
2.5
2.5
Therefore, under similar pH value, the potential mobility of heavy metals in sediment
is different significantly. For example, when pH was controlled at 4.0, the potential mobility
of metal decrease as follows: Zn>Cd>Ni>As>Cu>Pb (Valerie et al., 2004)(217).
2.2.5.5. Influence of organic matter
Organic compound in sediment, frequently existing in considerable amounts in
particle form, play an important role in trace metal transformation. For example, in the
sediment of some rivers or lakes, the trace metal bound to organic matter (OM) generally
takes up the largest fraction. Additionally, in sediment, the solubility of organic matters
usually directly determines the mobility of trace metals. Normally, the complexation of metal
ions with insoluble organic compounds can strongly lower their mobility, whereas the
formation of soluble metal complexes with dissolved organic compounds would enhance their
mobility (Sekaly et al., 1999)(218).
In natural rivers or lakes, OM is mainly composed of humic and fulvic substances.
The complexation reaction between trace metals and organic complexants is usually
recognized as the most important reaction pathway, due to this reaction determining, to a
large extent, the speciation and bioavailability of metal, and then influencing the mobility of
trace metal in natural water environment. However, in severely polluted river, due to the
complexity of organic matter, the reaction types between organic complexes and metals are
difficult to predict. In most conditions, precipitation, coprecipitation or flocculation usually
plays the most important role in heavy metal fixation (Peng et al., 2009)(205).
Total organic carbon added to the sediments, primarily through the decomposition of
plant and animal matters, can directly binds and/or absorbs trace metals and also
contains trace metals accumulated by plants that have been exposed to contaminated sediment
during their life time. Nonetheless, high percentages of organic matter and /or small grains in
sediments, reduce the metals availability, and undoubtedly play a major role in controlling
metals' concentration in the pore water (Tabata et al., 2009)(213).
2.2.5.6. The influence of Oxidation reduction potential
It is generally accepted that sediment oxidation reduction potential (ORP) is also a
most important factors controlling heavy metal mobility (Salomons and Stigliani, 1995)(219).
In anaerobic sediment, Acid volatile sulfide (AVS), a key component controlling the activities
of some divalent cationic metals, usually present naturally (Di-Toro et al., 1990)(22). Initially,
the majority of AVS contained in the anaerobic sediment is bound to iron as solid iron
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monosulfide (FeS), crystalline mackinawite (FeS), pyrrhotite (FeS), greigite (Fe3S4),or exits
As free sulfides. However, if divalent metals, such as cadmium, copper, chromium, lead or
zinc are present, the iron in iron-sulfide are displaced and one of these heavy metals rapidly
bind to AVS with stronger affinity (Hansen et al., 1996)(220). Finally, in those sediment
contaminated, the metal bound with sulfide usually takes up a rather high proportion (Peng et
al., 2009)(205).
When ORP in sediment increases, the oxidization rate of metal sulfides and the
degradation rate of organic compounds will increase correspondingly. Both can accelerate the
release of the adsorbed/complexing heavy metal (Calmano et al., 1993)(221). The reaction can
be expressed as: MS2 + (15/4)O2 +(7/2)H2O = M(OH)3 +2SO42 - +4H+. The release of H+ ions
into porewater would decrease the pH of sediment and then cause a secondary release of
heavy metals. Part of this released material will be re-adsorbed, especially onto the more
labile binding fractions. For example, with ORP in sediment increase, the Cd bound to
organic sulfide, a stable metal forms, would decrease from 65% to 30% and form a more
mobile form (Tneofanis et al., 2001)(222). Accordingly, with the annual variations of ORP in
sediment, the trace metal presents seasonal release and fixation. These phenomena are very
significant in some seasonally flooding rivers. For example, in Mulde reservoir,
approximately 18t of Zn can be released from sediment into water only due to the sediment
being disturbed and oxidized in flooding, which led to a significant increase of Zn
concentration in water (Tneofanis et al., 2001)(222). Therefore, in the dredging process of river
or lake, for decreasing the release of metal from sediment, oxidation of sediment should be
avoided (Peng et al., 2009)(205).
Biological oxidation of metabolisable forms of organic carbon has been recognized as
the most important factor in early diagenesis (Williamson and Wilcock, 1994)(204). In the
surface sediments, organic matter is decomposed by organisms in the presence of oxygen. In
finer sediments such as clay, with the exception of coarse sediments and other high-energy
areas where there is rapid advection of oxygenated water, this decomposition uses up oxygen
in the sediment quickly. The decomposition occurs faster than the rate of diffusion of oxygen
into the sediments and, as a result, most sediments are anaerobic just below the surface
(Williamson et al., 2003)(198). The decomposition of organic matter proceeds under anoxic
conditions using alternative electron acceptors to oxygen, such as nitrate, manganese and iron
oxides, and sulphate to oxidize (metabolise) organic carbon (Williamson et al., 2003)(198).
This oxidation, together with the resulting anoxic conditions, produces large changes to the
form of iron, manganese and sulphur, which are important in ‘binding’ trace metals in
sediment and releasing them to the overlaying water (Williamson and Wilcock, 1994)(204).
For instance, when sulphate is used, sulphide is produced and heavy metals react with
the sulphides and produce soluble precipitates (Williamson et al., 2003)(198). Sulphide also
reacts with Fe to form black ferrous sulphide and pyrites (FeS2) to give the subsurface
sediments their characteristic black and grey color. Mercury is different. While it undergoes
similar reactions to other metals, microbiota changes a proportion to organo-mercury
(methylmercury) a toxic compound that has the ability to bio-accumulate in the food chain
(Williamson et al., 2003)(198).
In sediments with high inputs of organic matter, reactions are concentrated close to the
sediment-water interface. In addition, iron, manganese and sulphur in many oxidized and
reduced forms can react with each other abiotically. All these reactions contribute to the
complex sediment chemistry (Williamson and Wilcock, 1994)(204).
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2.2.5.7. Adsorption / desorption reactions
The release of metals from oxic sediments can happen as the salinities increase or
reduction/oxidation condition changes due to the high ionic strength of seawater and potential
competition for binding sites with Ca and Mg (Luoma, 2000)(206). The flux of oxygen from
the water to the sediments controlled the reaction at the interface; hence the depletion of
oxygen consequently limits the release of the metals into the water. The extent of losing the
metals from oxic sediments can be reduced by readsorption to precipitating Fe and Mn oxides
(Luoma, 2000)(206). Metal removal from estuarine sediments can occur due to desorption of
metals from particulates to solution when freshwater and seawater mix and this has been
demonstrated in the laboratory experiments. However, readsorption also takes place at the
freshwater/seawater interface in estuaries due to the formation of new particles, increased
turbidity, increased pH and the occurrence of complex recirculation patterns (Luoma,
2000)(206).
The above processes are not only fueled by the changing redox environment and pH
but also by the metabolism of carbon organic matter, changing environmental conditions
brought about by bioturbation (muds and sands) or waves (sand) and irrigating the sediment
(Williamson and Wilcock, 1994)(204). The following diagram illustrates how the pollutants
move within the hydrosphere.
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Movement of pollutants in the hydrosphere (Fergussion, 1990)(3)
2.2.5.8. Physical transport
Sediment type and dynamics are known to affect contaminant concentration and fate
(Williamson and Wilcock, 1994)(204). In coastal environments with low hydrodynamic energy,
fine particulates tend to be trapped, while in areas where hydrodynamic energy is high, fine
particulates are ‘moved on’ (Williamson et al., 2003)(198). There are 3 main zones that can be
distinguished based on three types of processes (Williamson et al., 2003)(198).
i) Areas of accumulation (deposition). This is where fine materials are continuously being
deposited. Wave and current energies are very low in this area;
ii) Areas of transportation. This is where fine materials are deposited discontinuously, i.e.
periods of accumulation are interrupted by periods of periods of remobilization (generally of
short duration and associated with storms); and
iii) Areas of erosion. This is where there is little deposition of fine materials. Wave and/or
current energies are high in this area.
Bare rock, gravels and sand characterize areas of erosion while sediments in areas of
transport are diverse; ranging from muds to sands. The deposits within areas of accumulation
are typically muddy, with a high water and organic matter content (Williamson and Wilcock,
1994)(204).
For most metals, the intermediate fate is the deposition area where the finest sediments
generally accumulate (Williamson et al., 2003)(198). These areas tend to be sheltered estuaries
and embayments, and deeper water offshore. The ultimate fate of trace metals is burial given
that metals do not break down. When buried, trace metals become immobilized as insoluble
sulphide precipitates due to the decomposition of organic material (dead animals, plant
material) (Williamson et al., 2003)(198).
2.2.5.9. Bioturbation Movement of pollutants in the hydrosphere
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Numerous studies have suggested that bioturbation strongly affects the physical,
chemical and biological properties of sediments (Williamson and Wilcock, 1994)(204).
Biological mixing is more important in muddy sediments than sandy sediments, where strong
physical processes occur. There are two groups of animals based on the magnitude of
sediment disturbance; i) large individual predators, such as rays and crabs, may shift large
amount of sediment when searching for food, and ii) smaller animals such as worms and
gastropods. The latter burrow and ingest sediment, shift only small amounts of sediments, but
their high population densities make them important in sediment turnover (Williamson and
Wilcock, 1994)(204).
Bioturbation has a major impact on the chemistry of muddy sediments. Oxygen
penetration in unbioturbated sediments is highly restricted, and the redox boundary occurs a
few millimeter below the surface (Williamson and Wilcock, 1994)(204). Marine organisms,
such as polychaetes and crabs, create extensive ‘honeycombs’ in mud flats up to 60 cm in
depths allowing oxygen to be transported deep within the sediments via burrows (Williamson
and Wilcock, 1994)(204). These ‘honeycombs’, in turn, provide low tide irrigation of burrow
water, which maintains oxic conditions when tidal flats are submerged (thus facilitating the
exchange of solutes between sediments and water) (Redfern, 2006)(203).
Burrowing organisms also can alter the levels and the speciation of trace metals in
sediments by affecting the sediment stratigraphy to become blurred, particle size altered, pore
spaces and pH changed (Fergusson, 1990)(3). Consequently, metal concentrations do not
appear to change rapidly with depth (Williamson and Wilcock, 1994)(204). The organisms may
also bio-accumulate trace metals and remove them from the sediment profile (Fergusson,
1990)(3) hence decreasing the metal concentrations in the sediments (Redfern, 2006)(203).
2.2.5.10. The influence of other species
Except pH, OM and ORP, some other factors, such as temperature, salinity, metal
species and retention time, can also affect the distributions of trace metal in sediment. For
example, due to the differences of cation exchange capacity among different metal, their
mobility capacity varies correspondingly, and usually ranges as follows:
Cs>Zn>Cd>Fe>Ag>Co>Mn (Emile and Jonathan, 1998)(223); with temperature increasing, the
adsorption content of trace metal on sediment often decreases gradually; with salinity in pore
water increasing, the total adsorption content of heavy metal would decrease ascribed to the
competition among heavy metal and some other cations (Garnier et al., 2006)(224).
Additionally, long time kinetic adsorption–desorption experiments also show that the metal
freshly associated with particles presents less stable and higher potential bioavailability than
those associated for a long time (Peng et al., 2009)(205).
2.2.6. Analysis of trace metals in sediments
Metals in sediments are generally considered to be present in the following forms:
water soluble, exchangeable, carbonate bound, bound to hydrous metal oxides (as Fe, Mn),
bound to organic matter and sulfides, and finally in the crystal lattice of the different minerals
composing marine sediments. Many terms have been used to describe metals bound to
sediments. Metals that are loosely bound to sediments and can be released back to the water
column are known as labile metals. Another name used to describe this fraction is non-detrital
metals as they are bound to non-detrital particles (Loring et al., 1985)(225). One more term
used to describe this portion is non-residual metals. The other portion that is found inside the
lattice of sedimentary minerals is known as residual metals, detrital or non-labile metals, as
they can’t move back to the overlying water column (Khairy, 2008)(226).
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Concentrations of total metals in sediments can be obtained by the digestion of the
sediment sample in a mixture of acids to release all metals bound to the different fractions of
the sediments. But the determination of total metals will not be useful in environmental
pollution assessment studies in a given area as it implies that all forms of a given metal have
an equal impact on the environment. Such assumption is clearly untenable (Tessier et al.,
1979)(15). Also, the use of the total concentration of a trace metal in sediment as a measure of
its toxicity and its ability to bioaccumulate is problematic because different sediments exhibit
different degrees of bioavailability for the same total quality of metal. These differences have
been reconciled by relating organism toxic response (mortality) to the contaminant
concentration in the sediment interstitial water (Di-Toro et al., 1990)(22). AVS is one of the
major chemical components that control the activities and availability of metals in the
interstitial waters of sediments (Prica et al., 2008)(227).
A large reservoir of sulfide exists as iron sulfide in anoxic sediment. Sulfide reacts
with several divalent transition metal cations (cadmium, copper, mercury, nickel, lead and
zinc) and predominantly monovalent silver to form highly insoluble compounds that are not
bioavailable. It follows in theory, and with verification (Di-Toro et al., 1990)(22), that divalent
transition metals do not begin to cause toxicity in anoxic sediment until the reservoir of
sulfide is used up (i.e. the molar concentration of metals exceeds the molar concentration of
sulfide), typically at relatively high dry weight metal concentrations. This observation has led
to a laboratory measurement technique for calculating the difference between simultaneously
extracted metal (SEM) concentration and acid volatile sulfide (AVS) concentration from field
samples to determine potential toxicity (Ankley et al., 1991)(228). The presence of AVS in
sediments provides a mechanism for partitioning metals from the pore water to solid phase,
reducing the potential solubility and bioavailability (Ankely et al., 1996)(23).
Methods for measuring metals in each fraction of the sediment are known as
sequential extraction techniques. The results obtained by the sequential extraction technique
can provide a wider and deeper insight into the metal pollution in the environment. However,
the analytical procedure of the sequential extraction techniques is tedious and time
consuming. It is not an effective method, which can be easily applied in environmental
investigations (Chen et al., 1989)(229).
Numerous extraction schemes for soils and sediments have been discussed in literature
(Tessier et al. 1979; Sposito et al. 1982; Wette et al. 1983; Clevenger, 1990; Howard and
Vandenbrink, 1999)(15, 230-233). The procedure of Tessier et al. (1979)(15) is one of the
thoroughly researched and widely used procedures to evaluate the possible chemical
associations of metals in sediments and soils. Another two important schemes of sequential
extraction are those of Kersten and Forstner, (1986)(234) and the Bureau Communautaire de
Reference (BCR) sequential extraction method (Ure et al. 1993)(16).
The sequential extraction procedure is also often used to assess the possible chemical
partitioning of trace metals, particularly on the mobility and bioavailability of metals in
sediments (Li et al., 2001)(235). A comparison of sequential extraction procedure and SEM
results may provide new insights for interpreting the SEM/AVS measurement (Fang et al.,
2005)(26).
2.2.6.1. Acid Volatile Sulphide (AVS)
Trace metals in the overlying water can be adsorbed by sediments through the main
metal binding phases of total organic carbon (TOC), and acid volatile sulphide (AVS) (Song
and Muller, 1999)(236). The organic matter accumulated in sediments is decomposed to
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sulfide in reduced environment. Therefore, the presence of sulfide may be an indication of
high TOC concentration (Griethuysen et al., 2002)(237). AVS is a reactive pool of solid-phase
sulfide that is available to bind metals and render that portion unavailable and nontoxic to
biota (Simpson et al., 2000)(238). In anoxic sediments, the availability of divalent metals to
organisms living nearby oxidative surface sediments has been related to AVS (Hansen et al.,
1998)(239). The metals that associated with AVS may be released via oxidation within the
sediments (Daskalakis and O'Connor, 1989)(240). The oxidative release of trace metals during
the mixing or re-suspension of anoxic sediments in well oxygenated overlaying waters is an
obvious concern (Yu et al., 2001)(241). According to the geochemistry theory, divalent metals
first bind to AVS among those competitive metal-binding phases since the Ksp values of
metal sulfides are very low. The sulfides of most trace metals are highly insoluble, and are
considered to be less mobile and less bio-available to benthic organisms (Argeses et al.,
1997)(242). It has been verified that divalent transition metals don not begin to cause toxicity
in sediment until the reservoir of sulfide is used up (i.e. until the molar concentration of
metals exceeds the molar concentration of sulfide), typically at relatively high dry weight
metal concentrations (Di Toro et al., 1990)(22). It has been pointed out by many authors that,
metals which are easily extracted in dilute acid form sediments, are said to be bio-available
(Forstner, 1989)(243).
AVS is operationally defined as the amount of sulfides released during cold acid
extraction, while SEM is defined as the amount of metals (normally including Cu, Pb, Zn, Cd,
and Ni) liberated during the extraction of AVS (Di Toro et al., 1990)(22). Acid volatile
sulphide (AVS) has been used to predict the toxicity in sediments of divalent metals,
including copper (Cu), cadmium (Cd), nickel (Ni), lead (Pb), and zinc (Zn) (Di-Toro et al.,
1992)(244). The rational is that the AVS present in sediment reacts with the simultaneously
extracted metal-SEM, the reactive metal fraction that is measured in the cold acid extract used
to measure AVS. This reaction forms an insoluble metal sulphide that is relatively nonavailable for uptake by benthic organisms. The amount of AVS present in sediments will
therefore serve a critical role in setting the limits of metal bioavailability and toxicity in
sediments (McGrath et al., 2002)(245).
The use of different relationships between AVS and SEM to establish mechanical
models such as the ratio of SEM and AVS (SEM/AVS), the difference between SEM and
AVS (SEM-AVS) or the organic carbon normalized difference between SEM and AVS
(SEM-AVS)/ foc to assess metal toxicity has been widely applied (Di Toro et al., 1990;
Burton et al., 2005;Yin et al., 2008)(22, 246, 247). SEM/AVS approach is one of the current
methods to take bioavailability of trace metals in sediments into account (e.g., Di Toro et al.,
1992)(244), which assumes precipitation of trace metals (measured as SEM) with sulfides
(measured as AVS) in reduced (anoxic) environments. If SEM exceeds AVS, trace metals
will potentially be present in the pore water, depending on the presence of other sorption sites
in the sediment. In contrast, if SEM is smaller than AVS, virtually all metals will be
precipitated. This concept has been tested in laboratory and field settings as an evaluation tool
of toxicity risks. SEM-AVS has been found to give good predictions about pore water
concentrations and absence of toxicity in test organisms (e.g., Ankley et al., 1991; Di Toro et
al., 1992)(228, 244).
2.2.6.2. Sequential extraction procedure
Trace metals are widely dispersed in the aquatic environment, and ultimately
deposited in the sediment, which is therefore of particular interest concerning its metal
content (Gomes et al., 2010)(248). Sediments have a high storage capacity for pollutants; in no
part of the hydrological system more than 1% of these substances are actually dissolved in the
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water, while more than 99% are stored in the sediments (Salomons and Stigliani, 1995)(219).
Metals can exist in different forms in the environment, where they may be transformed from
one forms into another or exists in different forms simultaneously (Martinez-Sanchez et al.,
2008)(249). Aquatic sediments consist of several different geochemical phases that act as
reservoirs of trace metals. These phases include carbonates, sulfides, organic matter, iron and
manganese oxides, and clays, all of which may occur in a variety of structural forms (Gao et
al., 2010) (250).
The form or "chemical speciation" of metals varies widely depending on
environmental conditions. Such differences in chemical speciation affect the environmental
fate, bioavailability and environmental risk of the metals. Knowledge of metal's speciation
can help to assess the metal retention in soil or sediments and how easily it may be released
into the soil solution (Martinez-Sanchez et al., 2008)(249).
The terms “species” and “speciation”, in a chemical sense, have become widely used
in the literature, and it is now well established that the occurrence of an element in different
compounds and forms is often crucial to understanding the environmental and occupational
toxicity of that element (Cornelis et al., 2005)(251). A number of definitions of speciation can
be found in the literature. In the past, the term “speciation” has been used to refer to “reaction
specificity” (rarely); in geochemistry and environmental chemistry, to changes taking place
during natural cycles of an element (species transformation); to the analytical activity of
measuring the distribution of an element among species in a sample (speciation analysis); and
to the distribution itself of an element among different species in a sample (species
distribution) (WHO, 2006)(252).
Chemical extraction is employed to assess operationally defined metal fractions,
which can be related to chemical species, as well as to potentially mobile, bioavailable, or
ecotoxic phases of a sample. According to Verloo et al. (1980) (253), the mobile fraction is
defined as the sum of the amount dissolved in the liquid phase and an amount, which can be
transferred into the liquid phase. It has been generally accepted that the ecological effects of
metals (e.g., their bioavailability, ecotoxicology, and risk of groundwater contamination) are
related to such mobile fractions rather than the total concentration (Adriano, 1986)(254). Shortterm effects have been related to metal concentrations, frequently referred to as the intensity
factor (Lindsay, 1979)(255), while medium- to long-term effects are mainly governed by the
kinetics of desorption and dissolution of metals from solid-phase species, representing a
capacity factor of metal solubility. The use of selective extraction methods to distinguish
analytes, which are immobilized in different phases of soils and sediments, is also of
particular interest in exploration geochemistry for location of deeply buried mineral deposits.
Fractionation is usually performed by a sequence of selective chemical extraction techniques,
including the successive removal, or dissolution of these phases and their associated metals
(Hlavay et al., 2004)(256).
The concept of chemical leaching is based on the idea that a particular chemical
solvent is either selective for a particular phase or selective in its action. Although a
differentiated analysis is advantageous over investigations of bulk chemistry of soil and
sediments, verification studies indicate that there are many problems associated with
operational fractionation by partial dissolution techniques. Selectivity for a specific phase or
binding form cannot be expected for most of these procedures. There is no general agreement
on the solutions preferred for the various components in sediment or soils to be extracted, due
mostly to the matrix effect involved in the heterogeneous chemical processes (Martin et al.,
1987)(257). All factors have to be critically considered when an extractant for a specific
investigation is chosen. Important factors are the aim of the study, the type of solid materials,
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and the elements of interest. Partial dissolution techniques should include reagents that are
sensitive to only one of the various components significant in trace metal binding. Whatever
extraction procedure is selected, the validity of selective extraction results primarily depends
on the sample collection and preservation prior to analysis (Hlavay et al., 2004)(256).
In sediments, metals can be present in a number of chemical forms, and generally
exhibit different physical and chemical behavior in terms of chemical interaction, mobility,
biological availability and potential toxicity (Wang et al., 2010)(258).
Therefore, it is
necessary to identify and quantify the forms in which a metal is present in sediment to gain a
more precise understanding of the potential and actual impacts of elevated levels of metals in
sediments and to evaluate processes of downstream transport, deposition and release under
changing environmental conditions (Horowitz, 1991)(259). Determination of the total
concentration of a metal in sediment is not a particularly useful indicator of sediment toxicity,
as it does not distinguish between the natural and anthropogenic components of the metal.
Quantifying the available or labile metals in the sediment provides a better indicator of
sediment toxicity (Larner et al., 2008)(260). Leaching techniques are widely used for the
assessment of heavy metal mobilization (Gleyzes et al., 2002)(261). Extraction procedures are
undertaken to evaluate the metal availability and bioavailability (Kersten and Forstner,
1995)(262). In sequential extraction several selective reagents are used consequently to extract
"operationally defined phases" from the sediment in a set sequence (Passos et al., 2011)(263).
Methods for the determination of different forms of metals in sediments include sequential
extraction, whereby a series of reagents is used to extract operationally defined discrete
phases from sediments in an outlined sequence. The overall behavior of metals in an aquatic
environment is strongly influenced by the associations of metals with various geochemical
phases in sediments (Horowitz, 1991)(259).
Sequential extraction techniques have been applied to study the geochemical
partitioning of metals in contaminated soils and sediments (Zakir and Shikazono, 2011)(264).
Geochemical distribution results have also been used as an aid in predicting potential
contaminant mobility and bioavailability (Zakir et al., 2008)(265). This could help us to
understand the geochemical processes governing metal mobilization and potential risks
induced (Zakir and Shikazono, 2011)(264). There are five major mechanisms of accumulation
of metals in sediments, namely exchangeable, bound to carbonate, bound to reducible phases
(amorphous Fe oxyhroxide and crystalline Fe-oxide), bound to organic matter (oxidizable
fraction) (Singh et al., 2005)(266). These categories have different behaviors with respect to
remobilization under changing environmental conditions. The fractions introduced by
anthropogenic activity include the adsorptive, exchangeable and bound to carbonate fractions.
These are considered to be weakly bounded metals, which may equilibrate with the aqueous
phase and thus become more rapidly bioavailable (Pardo et al., 1990)(267).
Exchangeable fraction
The fraction of exchangeable metals includes the portion, which is held by the
electrostatic adsorption as well as those specially adsorbed. The amount of metals in this
phase indicated the environmental conditions of the overlying water bodies. Metals in this
fraction are the most mobile and readily available for biological uptake in the environment
(Zakir et al., 2008)(265). Exchangeable fraction, usually extracted with magnesium chloride
solution or sodium acetate solution (1M) at pH 8.2 for 1h, refers to the metals directly
adsorbing on sediments (Krishna et al., 2001)(268). Through some typical sorption–adsorptions
processes, these metals can be exchangeable and are inequilibrium with the ionic content in
water. Generally, this fraction is usually used to represent the environmentally available
components (Peng et al., 2009)(205).
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This fraction comprises metals adsorbed on the surface of the sediment. It is the most
accessible, and metal migration from the solid phase to water can occur, e.g. after a change of
ionic composition of water or as a result of a shift of adsorption equilibrium in the system
(Zerbe et al., 1999)(269). It has been widely accepted that the exchangeable metal in sediments
is labile, highly toxic and the most bioavailable fraction (Wang et al., 2010)(258).
Carbonates Fraction
Metals associated with this fraction are not strongly bound to the sediment solids and
can be released to the sediment pore water in acidic conditions (pH< 5). Among the chemical
reagents, acetic acid best represented the fraction that was likely to be bioavailable to
sediment ingesting and benthic organisms. Finally association of metals with this fraction is
probably the best example of human-induced influence in the sediment of the study area
(Zakir and Shikazono, 2011)(264).
Continuously extracted with NaOAc or HOAc solution (1M) at pH 5.0 for 5h, mainly
refers to the metals that are precipitated or co-precipitated with carbonate. This fraction is
sensitive to pH variations (Peng et al., 2009)(205). The adsorption of trace metals is related to
changes in water ionic composition which probably affect sorption-desorption processes. It is
known that the carbonates of sediment contain significant trace metal concentrations and the
concentrations are sensitive to changes of pH (Tokalioglu et al., 2003)(270). They can be
released as a result of pH decrease (Zerbe et al., 1999)(269).
Reducible or bound to Fe and Mn oxides fraction
Fe–Mn oxides fraction includes the soluble metal oxides/hydroxides under slightly
acidic pH as well as the metal associated with reducible amorphous Fe–Mn oxyhydroxides,
which was extracted with 25% (v/v) acetic acid containing some NH2OH·HCl at 960C for 6h.
This fraction can be dissolved with oxidation–reduction potential (ORP) varying (Peng et al.,
2009)(205). This fraction is sensitive to the redox potential changes and in anaerobic
conditions it is thermodynamically unstable (Zerbe et al., 1999)(269). It is well known that Feand Mn-oxides are present as cement, concentrations or nodules between particles or only as
coating on particles. These oxides bind the trace metals and have strong scavenging
efficiency for trace metals, but they are thermodynamically unstable under the anoxic
circumstances (Tokalioglu et al., 2000)(270). Due to the large surface area, amorphous hydrous
Fe-Mn oxides are one of the most important geochemical phases impacting the mobility and
behavior of trace meals (Turner, 2000)(271). Large numbers of studies have been conducted to
remove organic or inorganic pollutants by utilizing the storng adsorption capacity of hydrous
Fe or Mn oxides (Root et al., 2007)(272).
Oxidizable or bound to organic fraction
Organic fraction may be associated with various forms of organic material such as
living organisms, detritus or coatings on mineral particles through complexation or
bioaccumulation process. It is extracted mainly with 0.02 M nitric acid and 30% hydrogen
peroxide at pH 2.0 and 850C. This kind of metals can exist in sediment for longer periods, and
can also be released with OM decomposition (Peng et al., 2009)(205). Organic matter and
sulfides are important factors controlling the mobility and bioavailability of heavy metals
(Wang et al., 2010)(258). The trace metals may be associated with various forms of organic
material such as living organisms, detritus, or coatings on mineral particles. The
complexation and peptization characteristics of the natural organic substances are well
known. The organic substances may be broken up freeing the soluble trace metals in the
natural waters under oxidizing conditions (Tokalioglu et al., 2000)(270).
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Degradation of organic matter under oxidizing conditions can lead to the release of
soluble metals bound to this fraction (Purushothaman and Chakrapani, 2007)(273). The affinity
of trace metals for organic substances and their decomposition products are of great
importance for the release of the metals into water. Trace metal bound to this fraction is
assumed to reflect strong association with sediment organic material (Sharmin et al.,
2010)(274).
Residual fraction
Residual fraction, namely the metals still remained in sediment after the above
extraction procedures, usually presents as consolidated oxides, co-precipitates, and strongly
held complexes, which keeps relatively stable and does not show significant transformation in
various conditions (Peng et al., 2009)(205). This residual solid contains mainly primary and
secondary minerals which include the trace metals within their crystal structures (Tokalioglu
et al., 2000)(270). In natural conditions they are practically inaccessible for living organisms
and can be treated as permanently immobile (Zerbe et al., 1999)(269). The residual or
lithogenic fraction is a major carrier of transition metals in most aquatic system. The
concentration of trace metals in the crystalline fraction is largely controlled by the mineralogy
and extent of weathering trace metals in the form which are not soluble under experimental
conditions and hence may be considered to be held within the mineral matrix (Sharmin et al.,
2010)(274). Residual phases of metal are generally much less toxic for organisms in aquatic
environment (Wang et al., 2010)(258).
Fractionation is not only very useful for determining the degree of association of the
metals in the sediments, and to what extent they may be remobilized into the environment
(Jain, 2004)(275), but also for distinguishing those metals with a lithogenic origin from those
with an anthropogenic origin (Passos et al., 2011)(263). According to Rubio et al. (1991)(276)
metals with an anthropogenic origin are mainly present in the first three extraction fractions
(soluble in acid, associated with Fe and Mn oxides and associated with organic matter and
sulfides), while in the last stage of the process the residual fraction is obtained corresponding
to metals with lithogenic origin. Normally, the summation of the mobile and the
exchangeable fractions can be used to assess the environmentally available components. The
fractions bound to Mn oxides and organic materials are supposed to represent the potentially
mobile component under changing conditions, which are reviewed as the most important
components in sediments for metals binding. While the residual fraction represents the more
stable metal forms associated with anthropogenic or geogenic components, the influence of
which to ecological system is much less than the others in major conditions (Peng et al.,
2009)(205).
2.3. Environmental problems and previous studies on the coastal
area along the Egyptian Mediterranean Sea
Coastal areas provide important benefits to humans in terms of food resources and
ecosystem services. At the same time, human activities here may have significant negative
impacts on the health of ecosystems and the vitality of resources. Therefore, coastal and
marine pollution control is required to predict and monitor the consequences of human
activities on marine and estuarine ecosystems (Erdogan, 2009)(87). From an environmental
point of view, the coastal zone can be considered as the geographic space of interaction
between terrestrial and marine ecosystems that is of great importance for the survival of a
large variety of plants, animals and marine species (Castro et al., 1999)(277).
Most of the coastal areas of the world have been reported to be damaged from
pollution, significantly affecting commercial coastal and marine fisheries. therefore, control of
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aquatic pollution has been identified as an immediate need for sustained management and
conservation of the existing fisheries and aquatic resources (González-Macias et al., 2006)(5).
Coastal regions are some of the most sensitive environments and yet they are subject
to growing human pressures (David, 2003)(278). Since the industrial revolution, tremendous
amounts of toxic pollutants have been discarded into coastal environment and the sediments
of bays and estuaries have represented huge sinks of trace metals (Fan et al., 2002)(279).
Urban and industrial activities contribute to the input of significant amounts of pollutants
(among them trace metals) into the marine environment and directly affect the coastal systems
in which they are often deposited (Dassenakis et al., 2003)(180). Increasing residential
populations and tourism threaten coastal landscapes, ecosystems, resources and the rich
biodiversity inhabiting these environments (UNEP, 1996)(280). Therefore, pollution levels are
often elevated in the coast because of nearby land based pollution sources (Wang et al.,
2007)(192).
Mediterranean Sea appears to suffer from high anthropogenic pressure due to inputs
from; industrial, sewage effluents, storm water drains, shipping activities, spillage, rivers,
atmospheric fallout, coastal activities and natural oil seeps (El Nemr et al., 2007b)(281).
Mediterranean coastal water environments are impacted more and more by
anthropogenic activities resulting in pollution of marine sediments by persistent pollutants
especially the trace metals (MESAEP, 2003)(282).
Intense human development in the Mediterranean region may have negative
consequences on the marine environment. Inland activities such as industrial production,
agriculture and tourism, combined with population growth and heavy urbanization, originate
increasing pressures onto the coastal areas (EEA, 1999)(283), involving the degradation of
coastal habitats and a heightened risk for inhabitating organisms and human population.
Moreover, the particular oceanographic characteristics of the Mediterranean basin (relatively
shallow, semi-enclosed and with limited natural water exchanges) favour the accumulation
rather than the dispersion of contaminant inputs, rendering it a particularly vulnerable and
potentially threatened ecosystem (Fowler, 1986)(284).
The Mediterranean coast of Egypt of the Pleistocene age, extends about 900 km
inlength from Sallum to Rafah and is cut by the two branches of the Nile river (El-Wakeel et
al., 2006)(285). It hosts a number of activities beside a long stretch of undeveloped areas. The
Mediterranean coast of Egypt with different inland and offshore activities, the estuaries of the
Nile, at Rosetta and Damietta, and the coastal lakes Maryut, Idku, Burullus, Manzalah and
Bardawil pour ca. 8,000 x 106 m3/y of water some of which carries heavy load of pollutants.
Major cities on the coast used to get ride of their waste through sea outfalls like Alexandria,
and into lakes that pour its water into the sea like Port Said. The lakes of Maryut and
Manzalah are heavily polluted, and parts of them are actually dead.
The maritime transport in the eastern Mediterranean, including oil tankers,
commercial ships and passenger ships, affect the coast to a large extent. The entire beaches
are permanently polluted by oil lumps, litter and plastic debris even in the very far remote
areas of the coast where there are no known activities there.
The marine environment (as a part the coastal zone) is of great economic and
environmental significance. This zone in Egypt is currently under sever and ever increasing
pressure. A number of factors contribute to this situation: a) rapid urbanization of the coast;
b) pollution from residential commercial and industrial activities, c) tourism development, d)
resource users; e) continuous development in hazards prone areas. Pollution originated from
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the discharge of huge amounts of wastewater (sewage, industrial and agricultural discharge)
into the coastal water of the Mediterranean Sea of Egypt (UNEP, 2002)(286).
Degradation of water quality due to land based pollution is a serious problem in the
Mediterranean coastal areas. The countries of the Mediterranean Sea basin face a variety of
shared environmental problems that are trans-boundary in nature (EEAA, 2009a)(287). Coastal
and marine resources contribution to sustainable development is often overlooked.
Population growth, agricultural and urban development is taking place at rapid rates at many
coastal areas resulting in excessive pollution loads that are allowed to enter coastal waters
(Nessim et al., 2010)(288).
Measurement of trace metal concentrations in sediments is the first step in evaluating
their potential health or ecological hazard. Several studies were performed on the total
concentration of trace metals in sediments in different areas along the Egyptian Mediterranean
Coast to evaluate their ecological risk.
Concentrations of eight heavy metals (Cd, Cr, Cu, Fe, Mn, Ni, Pb, and Zn) in surface
sediment from El-Sallum to Sidi-Kreer along western part of Egyptian Mediterranean Coast
were determined to evaluate their levels and spatial distribution by Ahdy and Khaled
(2009)(289). The ranges of the measured concentrations in the sediments are as follows: 0.5240.924 μgg-1 for Cd, 16.248-34.164 μgg-1 for Cr, 26.529-33.332 μgg-1 for Cu, 846.4261433.933 μgg-1 for Fe, 32.371-108.915 μgg-1 for Mn, 31.703-43.592 μgg-1 for Ni, 20.67235.624 μgg-1 for Pb, and 26.267-112.73 μgg-1 dry weights for Zn. There were no significant
correlations among most of these metals, indicating they have different anthropogenic and
natural sources. To assess metal concentrations in sediment, Numerical Sediment Quality
Guidelines (SQGs) were applied. The concentrations of Cd, Cr, Pb, and Zn in all sediment
samples are lower than the proposed TECs indicated that there are no harmful effects from
these metals. On the other hand, concentration of Ni exceeded TEC in all samples while Cu
was exceeded the TEC at El-Sallum and Sidi-Barrani indicated that these stations were in
potential risk. The metal contamination in the sediments was evaluated by applying Index of
geo-accumulation and contamination factor.
El Nemr et al. (2007a)(61) assessed the marine pollution due to metals for surfacial
sediments sampled from 20 sites along Mediterranean coast of Egypt. The samples were
analyzed for leachable and total heavy metal concentrations (Cd, Co, Cr, Cu, Fe, Mn, Ni, Pb,
and Zn). Evaluation of the heavy metals pollution status was carried out using enrichment
factors, the effect range-low (ERL) and the effect range median (ERM). The study showed
high concentrations of Cd, Co, Pb, Ni and moderate concentrations of Cr, Cu and Mn were
contaminated in the sediments of studied sites. They found that the main source of
contamination is the offshore oil field and industrial wastes, which arise due to the ineffective
and inefficient operation equipments, illegal discharge and lack of supervision and
prosecution of offenders.
The coastal strip along Alexandria city from Agami in the west to Maamoura in the
east was studied by El-Sammak and Aboul-Kassim (1999)(290). They used the pollution load
index to find out the mutual pollution effect at the different stations by the different metals
(Co, Zn, Ba, Mn, Sr, Ni, and Fe). They found that there was an increase of PLI toward the
western side. They concluded that these two zones was considered to be the most polluted
areas at Alexandria beach.
El-Mex bay:
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El-Mex bay, lies in the western part of Alexandria coast between longitude 29 45 and
54 E and latitude 31 07 and 31 15 N, and extends for about 7 km between El-Agami
(west) and the Alexandria commercial Western Harbor (east) with maximum depth 20 m.
The bay is one of Alexandria coasts but it differs in having several industrial plants that
directly discharge their effluents into it. These are petrochemical, cement, chemicals,
tanneries, industrial dyes, ink, petroleum refining, meat processing, and fish production beside
the Alexandria iron and steel factory (Rifaat, 1982)(291). In addition, the bay is an estuarine
zone of huge agricultural drain called Omoum Drain its discharge rate 2,547.7 x 106 m3/year.
The downstream part of the last drain receives considerable amounts of Alexandria domestic
and industrial wastes prior discharging to the bay at a site lies just west of the Western Harbor
(Abdallah, 2007a)(292). Under prevailing hydrographic conditions and a sheltered geographical
position, El-Mex Bay estuary is two layered estuary, the fresh water from the drain spreads in
the estuary after mixing with seawater as surface brackish layer that overlies the other layer of
almost proper Mediterranean seawater (El-Gindy et al., 1986)(293). El-Mex Bay is considered
one of the most important hot spot areas in Alexandria. It is subjected to large quantities of
untreated industrial and domestic sewages (Masoud et al., 2007)(294).
29
The distribution of several heavy metals (Cr, Cu, Cd, Pb, Zn, and Hg) was
investigated by Massoud et al. (2007)(294) in muscle, gill, and liver in two different fish
species seasonally collected in El-Mex Bay (autumn 2004–summer 2005). In order to
evaluate the pollution status of the Bay, the concentrations of the selected metals in the labile
and total fractions were analyzed in sediment samples collected from eight sites in El-Mex
Bay during autumn 2004. Also, the Index of Geoaccumulation (Igeo) for the sediment was
estimated. The total and labile fractions of the selected metals in sediment samples were 15.2
and 62.8 μgg-1 dw for Cu, 1.8 and 5.0 μgg-1 dw for Cd, 79.1 and 130.3 μgg-1 dw for Zn, 0.2
and 1.2 μgg-1 dw for Hg, 35.8 and 93.0 μgg-1 dw for Pb, and 13.9 and 31.0 μgg-1 dw for Cr.
Abdallah (2008)(295) studied the distribution of heavy metals in El-Mex bay located in
Northern Egypt. Eight metals (Cd, Cr, Co, Cu, Zn, Pb, Mn and Fe) were determined in water,
suspended particulate matter and sediments. The bay presents higher metal concentrations in
aqueous Cd, Co and Pb are above the chronic freshwater quality criteria for aquatic life. Most
of the dissolved trace metals displayed a negative association with salinity, indicating Omoum
Drain as a source of inputs for them. In the bay water, Fe, Mn and Zn are the most abundant
elements in suspended particulate matter whereas Co the less abundant. El-Mex bay
exhibited relatively lower concentrations of all the examined metals in sediments, comparing
with other regions in the world.
The western harbour
The western harbor of Alexandria is situated between longitude 29 50 and 29 53 E
and latitude 31 9 and 31 12 N. Its length is 7 km, maximum width is 2 km and water depth
varies from 5.5-14 m (Saad et al., 2004)(91). It is the largest and oldest harbour on the
Egyptian Mediterranean coast, serving about three quarters of Egypt’s international trade. The
harbour receives imported and exported materials such as coal, manufactured iron, cement,
fertilizers, grains, food, textiles, chemicals, timber, as well as crude and refined oil. There is
also an old dry dock and workshop for ship building and repairs. Located on the southern and
eastern sides of the harbour, the quays for the various maritime activities divide these sides
into several small semi-enclosed or open basins. Wastewaters of varying quantity and quality
are discharged into the harbour, mainly through the Umoum Drain (76.4 m3 s-1 ) and the
Noubaria Canal (1 m3 s-1 ) (Dorgham et al., 2004)(296).
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The western harbour of Alexandria, situated to the west of the city of Alexandria
along the Mediterranean coast of Egypt, is a complex harbor with an area of about 31 km2.
The harbor is connected to the open sea by a narrow strait, which protects the bay from the
prevailing NW winds and renders the harbor a safe anchorage. The water body inside the
harbor has an average depth of 7m. The harbor was naturally formed during the PreHolocene subsidence of the coast and the subsequent transgression of the sea (Butzer,
1960)(297). Calcareous marine deposits cover the continental shelf surrounding the harbor
(Mostafa and Barakat, 2004)(298).
The western harbour is connected to the sea by El-Boughaz. It is protected by two
water breaks and is divided into two main basins; the inner and outer ports having areas of
200 and 600 acres, respectively. The Mahmoudia canal used to discharge into the inner port
has recently ceased to flow in El-Noubaria Navigation canal, crossing the heavily polluted
lake Mariut, opens into the outer port. The western harbour is under stress from various
pollutants from different external and internal sources. The external pollution results from
domestic, industrial and agricultural wastes. Thousands of cubic meters of contaminated
waters are discharged daily into the western harbour via El-Nobaria Canal and from El-Mex
Pumping Station to the west of the western harbour and are partially introduced into it by the
prevailing wind. Besides, considerable amounts of untreated sewage and industrial wastes are
dumped directly into the western harbour from several outfalls. The internal pollution
originates from different shipping wastes beside the discharges resulting from loading and
unloading processes of the imported and exported industrial raw materials (Saad et al.,
2004)(91).
The western Harbour of Alexandria is the major trade port of the Northern Territory of
Egypt. Industrial activities have increased dramatically in this area over the past 20 years.
The harbor handles approximately 75% of all ship-borne cargo of the country. As a result
aggressive urbanization and industrialization of Alexandria region, the coastal waters in
general and the Western Harbour in particular have received considerable amounts of treated
and untreated industrial, agricultural and domestic wastes (Salem and Sharkawi, 1981)(299).
These wastes are derived principally from effluent discharge from the El-Mex pumping
station to the west of the harbor. The effluent consists of overflow from lake Maryut (a
coastal lagoon heavily polluted mostly by domestic and industrial wastes), and the drainage
water from the El-Umum drain and El-Noubariya Canal. Industries that benefit from their
quay-side location include a chlor-alkali plant (Misr Chemical Industries). Portland cement
factories discharge unknown quantity of wastes into the harbor. Minor amounts of industrial
wastes are directly discharged into the harbor through tanneries (Mostafa and Barakat,
2004)(298).
According to Saad and Hemeda (1992)(300), the discharged pollutants cause hazardous
effects on the compartments of the western harbour. The harbor’s sediments are mostly
covered with sludge containing high levels of organic matter, nutrients and heavy metals
(Riffat, 1982)(291). The problem of pollution in the western harbour increases with progress of
time, due to the successive increase in population and in the number of visiting ships (Saad
and Hemeda, 1992)(300).
Because the harbor is a semi-closed basin with restricted water circulation, it may
serve as an entrapment of the wastes introduced from land-based sources as well as from the
harbor itself due to shipping activities. Future development and continued operation of
Alexandria Harbour is of great economic importance to this region, but these activities may
also impact its ecological functioning (Mostafa and Barakat, 2004)(298).
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Mostafa and Barakat (2004)(298) determined the concentrations of heavy metals in
surfacial bottom sediments of the Western Harbor to assess their potential biological effects
and to identify their possible sources. Sediment samples from 21 stations throughout the
harbor were analyzed for grain size, total organic carbon content, and metals (Al, As, Ba, Be,
Cd, Cr, Cu, Fe, Mg, Mn, Ni, Pb, Se, Sn, V, and Zn) to assess the extent of contamination in
the area. The results indicated that concentrations of metals in the sediments varied widely
depending on the location. High levels of metals were observed in the Arsenal Basin and the
outfall area of El Mahmoudiya Canal in the inner harbor.
El-Mamoney and Mohamed (2006)(301) collected some short cores from the Western
Harbor, Eastern Harbor and Abu Qir bay to investigate the penetration of heavy metals into
the bottom sediments. Every core interval was analyzed for granulometric characteristics,
water content, organic carbon, carbonate, Cd, Co, Cu, Cr, Ni, Zn, Mn and Fe concentrations.
They found that the upper most layers of the cores show the highest heavy metal
accumulations with the maximum levels in the Western Harbor.
Eastern Harbour
The Eastern Harbour of Alexandria is located in the southern Mediterranean Sea
between latitudes 31 12 and 31 13 N and longitudes 29 53 and 29 54 E ; it is a semienclosed basin (mean depth 6m), and has been strongly affected by urbanization (Abdallah
and Abdallah, 2008)(302). The harbour has a surface area of about 2.53 km2; accordingly its
volume is about 15.2 x 106 m3. It is sheltered from the sea by artificial break water leaving
two openings (left and right). The left opening is the main navigation entrance to the harbour
(300 m width). Through these openings the exchange of water between the harbour and
Mediterranean water takes place (Abdallah, 2007b)(303).
Sources of pollution include untreated domestic discharges through several submerged
minor sewage outfalls distributed along the harbour coast. There are various small industries
such as; photo development shops, car cleaning and repairing, some foodstuff plants, gas
stations, small dairy plants and some foundries. Also the harbour water receives additional
waste effluents from fishing ships and the ship yard situated on its western side. Other
wastewater discharges occur through two marine outfalls lying at the outer sides of the harbor
the Qait Bey and Silsila marine outfalls (El-Rayis et al., 2003)(304). The importance of this
harbour relates to the discovery of ancient buildings on the peninsula of Timonium and
Antirhodos Island (located in the eastern part of the harbour) and artificial dykes constructed
by the Ptolemies and by Marcus Antonius. In addition, about 1000 different artifacts
including columns, basins, sphinx statues, and parts of obelisks with hieroglyphs and
ceramics were discovered during the excavations. This basin is also considered as one of the
main fishing grounds for the city of Alexandria and its adjacent areas (Abdallah, 2007b)(303).
The Eastern harbour has been subjected to high levels of pollution due to municipal
waste disposal in addition to the waste dumping from ships and shipping activities. Similarly,
El-Mex bay having several industrial plants situated close to the coast and directly discharges
its effluents into it. In addition this bay is an estuarine zone of huge agricultural drain
(Omoum drain), its discharge rate about 2,547x 106 m3/year (El-Rayis and Abdallah,
2006)(305).
The distribution, enrichment and accumulation of heavy metals in the surfacial
sediments of the Alexandria City Eastern Harbour (Mediterranean coast of Egypt) were
investigated by Abdallah (2007b)(303). Surface sediments (in the <63µm fraction) were
collected from 12 sites representing the entire area of the harbour, were analyzed by AAS for
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Cd, Cu, Zn, Cr, Pb and Al. Metal levels were compared with literature data to assess the
pollution status of sediments. Enrichment factors (EFs) and the geoaccumulation Index (Igeo)
were calculated as a criterion of possible contamination. It was found that Zn represents the
highest median concentration, followed by Pb. According to calculations of enrichment
factors, Pb and Cd are of moderately severe enrichment, while Zn has the highest level of
enrichment. Regarding other elements, Cr is at background levels and Cu is of minor
enrichment.
Abdallah and Abdallah (2008)(302) determined the concentrations of Cd, Cu, Co, Zn,
Mn and Fe in biota and sediment samples collected form the Eastern Harbour and El-Mex
Bay in the Mediterranean Sea, Egypt. Results showed that the two species of bivalves, Donax
trunculus and Paphia textile had different amounts of metals in their tissue. The abundance
of heavy metals concentrations in the mussel samples was found in the order Fe> Zn >Mn >
Cu> Co> Cd> and Fe> Zn> Mn > Cu> Cd> Co, respectively for the two species. The metals
concentrations were generally higher compared with the previous studies in the same area.
The levels of metals accumulated in the investigated fish samples, were higher than those of
Marmara Sea (Turkey), for Co and Cd and lower for Cu, Zn, Mn and Fe. El-Mex bay having
the highest metals concentration in sediments as their order of abundance were Fe> Zn> Mn>
Cu> Cd> Co.
The concentrations of heavy metals were determined in sediments of the Eastern
Harbour by Nasr et al. (2011)(306). The range and average concentrations (μgg-1) were
2026.79- 2799.11 (2457.59) for Fe; 95.18-285.53 (192.69) for Mn; 55- 164 (111.85) for Zn;
33.33- 180.95 (87.08) for Pb; 24.22-173.70 (67.53) for Cu; 36.47- 127.66 (65.43) for Cr and
122.36- 168.78 (141.56) for Ni. The study concluded that the Eastern Harbor of Alexandria is
subjected to serious contamination sources of heavy metals, and reflect the anthropogenic
input as a source of heavy metals and normalize them with grain size, total organic matter and
total carbonate content of the sediment of the Eastern Harbor.
Rossetta and Damietta
The Rosetta estuary receives annually controlled outflow of the Nile at certain times
through Edfina barrage. The most northern part of this estuary runs through a narrow sandy
cape previously formed by the previously gradual sedimentation of mud and silt transported
by the Nile flood. This cape was subjected to gradual erosion by seawater currents after
construction of the Aswan High Dam, as most of the suspended matter transported by the Nile
flood is deposited in the High Dam Lake. The highest discharge from Edfina Barrage occurs
in January and the lowest during the period from April to November (Saad and Hassan,
2002)(307).
The new Damietta Harbor was constructed in 1982, and is located about 9.7 km west
of the Damietta Nile Branch. The harbor basin was constructed inland and its entrance was
protected by two breakwaters. The western breakwater extends about 1500 m parallel to the
navigation channel, attaining the 7m depth contour. The eastern one is about 500m long,
perpendicular to the shoreline and tends to about 3m water depth contour. The navigation
channel extends offshore to the middle shelf or about 15m. Since January 1984, the channel
of the harbor has experienced sedimentation and subsequently threatening the navigation
activities (Shereet, 2009)(308).
New Damietta Harbour is considered as semi closed water body influenced by
loading/unloading activities, municipal and agricultural waters resulting from Damietta
Governorate (Shereet, 2009)(308). Mostly Damietta is an industrial center known for its
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furniture, leathers, textile and sweets industries in addition to dairy products and rice mills
and for its agricultural heritage. It is also a fishing industry town, with one of the largest fleets
on the Mediterranean which accounts for fully half of the fishing boats of Egypt. Finally, it is
well known for the port. Damietta Harbor is considered as semi closed water body affected
mainly from loading/unloading operations, municipal and agricultural wastes resulting from
Damietta governorate. It is generally agreed that, the pollution in near-shore waters of the
Mediterranean sea has reached a critical level. This is chiefly due to the total absence of
control on toxic components. River runoff (Damietta branch) has the direct effect of reducing
the salinity of the surface layer of the Mediterranean sea, while tidal currents have a
considerable influence on the vertical mixing of shallow water near the coast (Said and
Hamed, 2001)(309). Untreated domestic wastewater with agricultural and industrial wastes is
still released through a number of drainages and outfalls along the coastal area of study (Said
and Hamed, 2006)(310).
Faragallah et al. (2009)(311) investigate the hydrographic characteristics, nutrient salts
and some heavy metals (Fe, Cu, Zn, Pb and Ni) as well as chemical oxygen demand (COD),
biological oxygen demand (BOD) and chlorophyll-a (Chl-a) in six vertical profiles in the
open Mediterranean sea water far about 60 Km from Damietta harbor (Egypt). Water samples
were collected during April 2007. The results revealed that most of nutrients and heavy
metals are concentrated in the surface layer and decreased with increasing the depth. N/P ratio
and abundance of the N-ions revealed that the area is mostly N-limiting. Enrichment factor
(EF) of the metals gives low values, less than 1 indicating to the enrichment and advection of
heavy metals counted each other. The results indicated that the impact of anthropogenic
inputs was limited in the distribution of nutrient and heavy metals, the values of metals were
similar or lower than that reported for water quality criteria except that the Zn level was
slightly higher. The relationships between the different heavy metal concentrations and
the other parameters (salinity, chlorophyll-a and suspended particulate matter) were
discussed. The values of BOD/ COD ratio indicated that the water of the study area has a
biodegradable nature. Relatively high levels of Chl-a concentrations was observed in the
surface layer during the period of study and negative correlation was found between Chl-a
and both NO3-N, PO4-P and SiO4 (r= -0.58, -0.38 and -0.58, respectively).
El-Dekheila Harbour
The newly constructed El-Dekheila harbor lies to the west of Alexandria city, between
El-Mex and El-Agami at Latitude 29 47 and Longitude 31 10, with surface area of about
12.5 km2 and a depth ranging from 6 to 19 m with an average of 12.4 m. El-Dekheila harbor
was constructed to serve basically Alexandria Iron and Steel factory. The harbor as a part of
El-Mex bay is subjected to several sources of wastewater (Shata and Deghedy, 2004)(312).
This is a commercial harbour. El-Dekheila Harbour is a semi-enclosed basin constructed
recently, after 1986, on the western side of El-Mex bay (Fahmy et al., 1997)(313). According to
Abdalla et al. (1995)(314) and Fahmy et al. (1997)(313), the harbour’s water is subjected to
several sources of wastewater. A huge volume of brackish water is discharged into El-Mex
bay through the El-Umoum drain. On its western side, near the harbour, this bay also receives
industrial wastes from several sources. The degree of water contamination of the harbour
water from the above mentioned sources depends on water circulation in the bay. El-Dekheila
Harbour, like other harbours, is affected by shipping activities. Its water depth ranges between
6 and 19 m with an average of 12.4 m. The exposed panels were dangled at the container pier.
Abu Qir Bay
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Abu-Qir bay is one of the Mediterranean coastal bays. It is a shallow (mean depth
12m) semi-circular bay, lies 36 km east of Alexandria city, the area of the bay is about 360
km2 with shore line of about 50 km long. This bay receives a continuous run off (totally of
about 5 million m3/d) mainly from three land-based sources, namely from west to east Tabia
pumping station (mainly industrial and agricultural waste waters), Maadia outlet (mainly
agricultural drainage water) and Rosetta mouth (mainly river water). Many reefs, shoals and
patches of foul ground could be found in the south west part of Abu Qir bay. These ridges
cause a rather limited water exchange with the open sea (Faragallah, 2004)(315). It is
simultaneously used as a marine habitat, source of food, site for recreation and disposal of
industrial, agricultural and domestic wastes (Aboul Dahab, 1992)(316).
Abu Qir Bay is boarded from the eastern side by the Rosetta estuary of the Nile and
from the western side by Abu Qir headland. The area of the bay is 360 Km2, with an average
depth of 12 m and a volume of 4.32 Km3. The seaward limit of the bay is considered as the
line adjoining the Rosetta mouth and Abu Qir headland. Abu Qir bay receives fresh Nile
water from Rosetta estuary at its eastern side, brackish water from Lake Edku through Maadia
Channel at its southern side and highly polluted waters via Tabia Pumping Station at its
south-western side and the rain water was insignificant. After damming of the Nile in 1964,
the quantity of the discharged Nile waters into the sea through Rosetta mouth was much
reduced to 3 x 109 m3 annually (Irrigation Department). The discharged Nile water did not
show an effect on the western region of Abu-Qir Bay, but decreased the surface salinity of its
eastern part to about 37% during winter (El Deeb, 1977)(317). About 1400 million m3 of
brackish lake water (less than 2%) were discharged from Lake Edku to Abu Qir Bay and its
flushing rate was 2.5 months (Abdel-Moati, 1991)(318). The industrial wastes of several
factories discharged into Abu Qir Drain result from industries including fertilizer, textile,
paper, food processing and canning. Abu Qir drain flows through cultivated area till Tabia
pumbing station. This heavily polluted water, which spreads out into the bay through Tabia
pumping station, amounted to 2.0 x 106 m3/d and contained mainly organic and inorganic,
pollutants (Saad and Badr, 2005)(319).
In an attempt to evaluate the environmental quality of Abu Qir Bay, an important
highly productive area in Alexandria, Egypt, an environmental risk assessment was performed
by Khairy (2008)(226); including a screening level ecological risk assessment (SLERA) and a
human health risk assessment (HHRA). To fulfill the goals, 30 surfacial sediment samples
were collected from different locations covering the region from the harbour area till the
Maadeya Outlet. Concentrations of total Ni, Co, Cr, Pb and V showed close pattern of spatial
distribution in the bay sediments. Cu and Zn concentrations were much higher in front of the
Abu Qir Drain and the Fertilizer Company. Hg and Sn total concentrations were higher in
front of the Maadeya Outlet and in the offshore direction too reflecting the influence of the
discharged agricultural wastes and shipping activities respectively. Concentrations of all
metals were lower in the western part of the bay. Most of the investigated metals (except Se
and Sn) showed significant positive correlation with clay content and the Fe %, which may
indicate that metals could have entered the bay either adsorbed on Fe oxyhydroxides and/or
scavenged by the Fe oxyhydroxides when they entered the bay waters.
Burullus lagoon
Burullus lagoon is one of four shallow brackish water coastal lagoons (namely:
Manzala, Burullus, Edku and Maryout) of Nile Delta. It is the second largest one which has
an area of about 410 km2. It lies between Longitudes 30 30 and 31 10 E and Latitudes 31
21 and 31 35 N. The lagoon width varies between 4 and 16 km, and its mean water depth
varies from 60 to 150 cm. It is separated from the Mediterranean Sea by sand bars and sand
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dunes of different widths and heights and is connected to the sea with Boughaz El-Brullus.
Burullus lagoon is affected mainly by agricultural drainage water mixed with different types
of wastes from fish farms (Terra drain, drain 7, drain 8 and drain 11), wastewaters effluents
(Terra drain, drain 7 and El-Gharbia drain) as well as domestic drainage water discharged
mainly from El-Gharbia drain and drain 11 (Khalil et al., 2007)(320).
Lake Burullus has an area of 370 km2. It occupies a central position along the
Mediterranean coast of Egypt. It is connected to the Mediterranean through El-Boughaz ElBurg. Water depth ranges from 50 to 160cm. The lake receives 3.6 x 106 m3 of water
annually, almost from domestic and agricultural sewage (El-Sammak and El-Sabrouti,
1995)(321).
El-Sammak and El-Sabrouti (1995)(321) determine the heavy metals contents in the
sediments of Lake Burullus. The contamination factors were found to fall in the following
sequence: Cd>Pb>Cu>Zn.
Fathi and Abdelzaher (2005)(322) studied some characteristics of Lake Burullus. Data
from sediment core analysis showed that, the maximum concentration of Zinc (6.2 µgg-1) and
Cu (1.0 µgg-1) in sediment were observed at 6 cm and 26.5 cm sediment depth, respectively.
The highest concentration Pb (7.5) µgg-1 was found at 0.5 cm depth.
Twelve samples from surface sediments of Brullus lagoon and adjacent Mediterranean
Sea were collected by Khalil et al. (2007)(320). They found that the variations of the measured
metals (Fe, Mn, Zn, Cu, Ni, Co, Cd, and Pb) in sediments are varied depending on the
locations, whereas the high levels were observed in the western area of the lagoon. They
were positively correlated with each other and organic carbon suggested that the distributions
of these metals are associated with the organic matter accumulation.
Saeed and Shaker (2008)(323) determined concentrations of heavy metals including
Iron, Zinc, Copper, Manganese, Cadmium and Lead (Fe, Zn, Cu, Mn, Cd and Pb) in water
and sediments in northern Delta Lakes (Edku, Borollus and Manzala) and their accumulation
in Nile tilapia (Oreochromis niloticus) organs (muscle, gills and liver) were investigated.
Water, sediments and fish organs from Lake Manzala showed greater concentrations of most
of the studied metals than those from Lake Edku and Lake Borollus. Fe, Mn, Cd and Pb (in
Lake Manzala) and Mn and Pb in Lake Borollus showed levels above the international
permissible limits in water. In sediment samples Mn (in Lake Edku) and Cd (in Lake
Manzala) recorded higher values than the sediment quality guidelines. Gills and Liver of O.
niloticus contained the highest concentration of most the detected heavy metals, while
muscles appeared to be the last preferred site for the bioaccumulation of metals. The edible
part of O. niloticus showed higher levels of Cd (in Lake Edku and Manzala) and Pb (in Lake
Manzala). Nile tilapia caught from these two Lakes may pose health hazards for consumers.
Port Said
El Tokhi et al. (2008)(324) determined the concentration of Cu, Zn, Ni, Pb, Cd and V in
the bottom sediment of the east Port Said area. They found that Cu, Zn, Ni, Pb, Cd and V
concentration above the accepted level. The contamination levels were found due to
anthropogenic origin. These include intensive maritime activities, disposal of waters from
anchorage vessels that are waiting the crossing of the Suez Canal, lubricating oil, waters from
the ships yard at the entrance of the Suez Canal and wastes from the urban centers (Port Said
and Port Fouad cities)
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There are very limited studies on the geochemical forms of trace metals in the
sediments along the Mediterranean coastal area. In this section we will represent some of this
studies that performed sequential extraction procedure to identify the potential impact of trace
metals along the Egyptian Mediterranean Sea sediments.
Nasr et al. (1990)(325) applied a sequential chemical extraction technique for the
partitioning of particulate heavy metals (iron, manganese, zinc, copper and lead) into five
fractions: exchangeable, bound to carbonate, bound to Fe-Mn oxides or easily reducible,
bound to organic matter including sulphides and residual. This study reveals high levels of
zinc, copper and lead pollution in the coastal water of Alexandria.
Saad and Badr (2003)(182) used a five step sequential extraction technique to describe
the chemical association of copper with major sedimentary phases; exchangeable (F1),
carbonates (F2), Fe/Mn oxides (F3), organic/sulfides (F4), and residual associations (F5).
They found that the Cu fractions in the Eastern harbour’s sediments decreased in the
following order: F5>F4>F3>F2>F1. The distribution pattern of these fractions is mostly
controlled by the mineral composition of sediments and the anthropogenic wastes introduced
to the study area. The low association of copper with F1 confirms that metal was low in the
soluble form. The higher value of F2 at the center of the Eastern harbour coincided with
abundance of the biogenic clastics as major source of carbonates in the sediments of this area.
The high concentrations of F3 in the suspended matter and sediments at the boatyard area and
the Eastern harbour’s opening indicate the interaction between both phases. The high level of
F4 is not surprising, due to continuous accumulation of organic matter from large amounts of
domestic sewage discharging.
Saad et al. (2004)(91) studied the local distribution of total lead and its species in the
western harbor sediments. A five step sequential extraction scheme was applied to illustrate
the contribution shared by each individual fraction in the total lead concentration in the
western harbor sediments. The exchangeable fraction (F1) was very low. The bound to
carbonate fraction (F2), the bound to iron-manganese oxide fraction (F3) and the bound to
organic matter- sulfide fraction (F4) ranked fourth, second and third in abundance.
The sediments of El-Mex Bay estuary on the southern Mediterranean Sea have been
analyzed by Abdallah (2007)(292) for trace metals after sediment fractionation by sequential
leaching. A sequential extraction procedure was applied to identify forms of Mn, Cu, Cd, Cr,
Zn and Fe. The five steps of the sequential extraction procedure partitioned metals into:
extractable (F1); carbonate extractable (F2), reducible extractable (F3); organic extractable
(F4) and acid soluble residue (F5). Extracted concentrations of trace metals analyzed after all
five steps, were found to be (µg/g) for Mn: 1930, Cu: 165.3, Cd: 60.9, Cr: 386.3, Zn: 2351.3
and Fe: 10895. Most of elements were found in reducible fraction except Fe found in acid
soluble residue, characterizing stable compounds in sediments.
Khairy (2008)(226) applied a sequential extraction procedure to surface sediments from
Abu Qir Bay. Results obtained revealed that the majority of Ni and Co were incorporated in
the residual fraction of the sediments, thus chemically inactive. Although the residual fraction
of Fe, Cr, Mn and Zn showed the highest percentage compared to the other 4 fractions, the
sum of the first four non-residual fractions were higher than the residual fraction of each
element, especially the reducible fraction, which showed a significant portion of the total
metal concentration in Fe and Cr. At the same time, the reducible fraction showed the highest
concentrations of Pb and V. The oxidizable fraction of Cu was found to be the dominant
fraction which may indicate the association of Cu with organic matter in the bay sediments.
High significant positive correlation was observed between concentrations of the reducible Fe
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and concentrations of the labile fractions of metals (F1 + F2 + F3), which confirm the strong
association between metals and the reactive Fe in the bay sediments. The anthropogenic
influence on the investigated trace metals concentrations was identified in most of the
investigated sediment samples, especially for Se, Cd, Hg and Pb. These elements were highly
enriched in relation to Al, which was used as the normalizer.
The distribution of total and different chemical forms of copper and manganese has
been studied by Faragallah and Khalil (2009)(326) using sequential extraction techniques.
Chemical analysis of the sediment show that, in Abu-Qir bay, cupper is bonded more with the
residual and organic fractions and manganese with residual and carbonate fractions. In ElMex bay, cupper is enriched in the labile fractions (exchangeable, carbonate and oxides) and
manganese with carbonate fraction. The mobility and bioavailability of cupper and
manganese in El-Mex bay is higher than that in Abu-Qir bay.
El-Anany (2008)(327) identified the geochemical forms of aluminium (Al) in core
sediments from Lake Burullus. The distribution mapping of aluminium level of the surface
and subsurface sediments of different fractions were also studied. The results of the vertical
distribution of Al-exchangeable fraction revealed irregular variations through the four cores.
Generally, the residual fraction is the first highest percentage reached to more than 90% of the
total aluminium and followed by the organic fraction which has Al level more than 2% for
most intervals of the four cores. The relative percentage of Al bound to the five fractions
from the total concentrations can be arranged as follows: carbonate (0.01-0.05%) <
Exchangeable (0.01-0.41%) < Fe-Mn oxides (0.03-0.58%) < organic (0.41-3.12%) < Residual
(more than 90%).
There are very limited studies on the relationships of trace metals and AVS in the
sediments along the Mediterranean coastal area. Studies on the environmental state of the
Egyptian Mediterranean Coast have concentrated in the past mostly on simple chemical and
bulk analysis of sediments. Okbah et al. (2011)(328) studied the relationship between Zn, Cu,
Ni, Pb, Cd, and amorphous iron sulphides to predict the bioavailability and toxicity of metals
in Damietta Port and offshore sediments. Sulphides were measured as acid volatile sulphides.
The average values of SEM in Damietta Port sediments were 11.21±1.95, 5.16±1.18,
0.23±0.08, 1.30±0.58, 0.71±0.17, 0.11±0.04, 1.23±0.71 and 0.18±0.09 µmoleg-1, dry weight
respectively for the offshore samples. Generally, the concentrations of trace metals in both
areas were in the order of Zn >Pb>Cu>Cd>Ni. The difference value between AVS and
SEM is an important tool of trace metals bioavailability and ecological risk. The range and
average values of SEM concentration ranged from 1.96 to 29.94 (14.70 ± 8.16) µmoleg-1,
the results revealed significantly between the two regions of the study area. The levels of
AVS concentrations showed spatial variations, the range and average value of AVS were
between 7.71 and 40.57 (24.12±9.53) µmoleg-1. The ratio of SEM/AVS was lower than 1 at
all the offshore sediments except at two stations as well as at five stations located inside the
port (SEM/AVS>1), which suggests that the metals have toxicity potential found in these
sediments.
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2.4. Ecological risk assessment
Ecological risk assessment is a broad framework for estimating the likelihood and
severity of potential impacts to the structure, functioning, components or processess of
ecological systems (ecosystems); such assessments are used routinely by many countries.
Typically, these assessments follow the general ecotoxicology paradigram, identifying and
assessing the exposure and effects of human and other populations to such stressors as
contaminants or pathogens or habitat change. The approach has been formalized by the
United States "Environmental Protection Agency" among other bodies. Significantly, the
approach applies equally well to analysis of the potential ecological effects of invasive
species and biological control agents (Andresen and Draggan, 2007)(329). Ecological risk
assessments are developed within a risk management context to evaluate human-induced
changes that are considered undesirable. As a result, these guidelines focus on stressors and
adverse effects generated or influenced by anthropogenic activity. Defining adversity is
important because a stressor may cause adverse effects on one ecosystem component but be
neutral or even beneficial to other components. Changes often considered undesirable are
those that alter important structural or functional characteristics or components of ecosystems.
An evaluation of adversity may include a consideration of the type, intensity, and scale of the
effect as well as the potential for recovery. The acceptability of adverse effects is determined
by risk managers. Although intended to evaluate adverse effects, the ecological risk
assessment process can be adapted to predict beneficial changes or risk from natural events
(USEPA, 1998)(330).
2.4.1. Ecological Risk Terms
Some important terms in ecological risk assessment:
-
Hazard: A situation that could potentially lead to ecological harm.
-
Stressor: The component or components of a system that may lead to a hazard or
undesirable outcome.
-
Receptor: The component or components of a system that may be negatively
influenced by a stressor.
Pollutant risk assessments are always directed at such particular assessment endpoints
as response of growth or fecundity to toxicity. These endpoints are focused typically on
effects of chemical contaminants on individuals. However, since the scale of impacts of
interest are usually population level effects (i.e. changes in abundance), ecological models are
often used to perform the necessary extrapolation from the measurement endpoint (e.g. a
toxic, teratogenic or developmental effect) to the assessment endpoint (the ecological
characteristics or feature that is to be protected from risk, such as the abundance of some
sensitive species) (Andresen and Draggan, 2007)(329).
2.4.2. Phases of Ecological Risk Assessment
The ecological risk assessment process is based on two major elements:
characterization of effects and characterization of exposure. These provide the focus for
conducting the three phases of risk assessment: problem formulation, analysis, and risk
characterization. The three phases of risk assessment are enclosed by a drak solid line. Boxes
outside this line identify critical activities that influence why and how a risk assessment is
conducted and how it will be used.
1- Problem formulation: the first phase is shown at the top. In problem formulation,
the purpose for the assessment is articulated, the problem is defined, and a plan for
analyzing and characterizing risk is determined. Initial work in problem formulation
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includes the integration of available information on sources, stressor, effects, and
ecosystem and receptor characteristics. From this information two products are
generated: assessment endpoints and conceptual models. Either product may be
generated first (the order depends on the type of risk assessment), but both are needed
to complete an analysis plan, the product of problem formulation.
2- Analysis: shown in the middle box, is directed by the products of problem
formulation. During the analysis phase, data are evaluated to determine how exposure
to stressors is likely to occur (characterization of exposure) and, given this exposure,
the potential and type of ecological effects that can be expected (characterization of
ecological effects). The first step in analysis is to determine the strenghts and
limitations of data on exposure, effects, and ecosystem and receptor characterization.
Data are then analyzed to characterize the nature of potential or actual exposure and
the ecological responses under the circumistances defined in the conceptual model(s).
The products from these analyses are two profiles, one for exposure and one for
stressor response. These products provide the basis for risk characterization.
3- Risk characterization: during risk characterization, as shown in the third box, the
exposure and stressor-response profile are intergrated through the risk estimation
process. Risk characterization includes a summary of assumptions, scientific
uncertainities, and strenghths and limitations of the analyses. The final product is a
risk description in which the results of the integration are presented, including an
interpretation of ecological adversity and descriptions of uncertainity and lines of
evidence (USEPA, 1998)(330).
The framework for ecological risk assessment proposed by the (USEPA 1998)(330).
2.4.3. Ecological risk assessment in a management context
Ecological risk assessments are designed and conducted to provide information to risk
managers about the potential adverse effects of different management decisions. Attempts to
eliminate risks associated with human activities in the face of uncertainties and potentially
high costs present a challenge to risk managers. Although many considerations and sources
of information are used by managers in the decision process, ecological risk assessments are
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unique in providing a scientific evaluation of ecological risk that explicitly addresses
uncertainty (USEPA, 1998)(330).
Detailed steps of the ERA as proposed by (USEPA 1998)(330).
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2.4.4. Contributions of Ecological Risk Assessment to Environmental Decision
Making
At EPA, ecological risk assessments are used to support many types of management
actions, including the regulation of hazardous waste sites, industrial chemicals, and pesticides,
or the management of watersheds or other ecosystems affected by multiple nonchemical and
chemical stressors. The ecological risk assessment process has several features that
contribute to effective environmental decision making:
-
Through an iterative process, new information can be incorporated into risk
assessments, which can be used to improve environmental decision making. This
feature is consistent with adaptive management principles used in managing natural
resources.
-
Risk assessments can be used to express changes in ecological effects as a function of
changes in exposure to stressors. This capability may be particularly useful to the
decision maker who must evaluate tradeoffs, examine different alternatives, or
determine the extent to which stressors must be reduced to achieve a given outcome.
-
Risk assessments explicitly evaluate uncertainty. Uncertainty analysis describes the
degree of confidence in the assessment and can help the risk manager focus research
on those areas that will lead to the greatest reductions in uncertainty.
-
Risk assessments provide a basis for comparing, ranking, and prioritizing risks. The
results can also be used in cost-benefit and cost-effectiveness analyses that offer
additional interpretation of the effects of alternative management options.
-
Risk assessments consider management goals and objectives as well as scientific
issues in developing assessment endpoints and conceptual models during problem
formulation. Such initial planning activities help ensure that results will be useful to
risk managers (USEPA, 1998)(330).
2.4.5. Factors Affecting the Value of Ecological Risk Assessment for
Environmental Decision Making
The wide use and important advantages of ecological risk assessments do not mean
they are the sole determinants of management decisions; risk managers consider many
factors. Legal mandates and political, social, and economic considerations may lead risk
managers to make decisions that are more or less protective. Reducing risk to the lowest level
may be too expensive or not technically feasible. Thus, although ecological risk assessments
provide critical information to risk managers, they are only part of the environmental
decision-making process (USEPA, 1998)(330).
In some cases, it may be desirable to broaden the scope of a risk assessment during the
planning phase. A risk assessment that is too narrowly focused on one type of stressor in a
system (e.g., chemicals) could fail to consider more important stressors (e.g., habitat
alteration). However, options for modifying the scope of a risk assessment may be limited
when the scope is defined by statute (USEPA, 1998)(330).
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Chapter III
Socioeconomic and Environmental Aspects
3.1. Description of the Egyptian Mediterranean coastal
environment
3.1.1. History and culture
Some of the most ancient civilizations flourished around the Mediterranean. It was
opened as a highway for commerce by merchants trading from Phoenicia. Carthage, Greece,
Sicily and Rome were rivals for dominance of its shores and trade; under the Roman Empire
it became virtually a Roman lake and was called Mare Nostrum (our sea). The city of
Alexandria is located immediately west of the Nile delta and was founded by Alexander the
Great 332-331 B.C. Alexandria is considered today one of the oldest, continuously existing
cities of the world. It is the second largest city in Egypt and one of the major ports in the
Mediterranean Sea, besides being an important commercial, business, industrial and cultural
center. Furthermore, a number of submerged archaeological sites have recently been
discovered off the Alexandria coastline, which is also the principle seaside resort of Egypt,
attracting about two million visitors every summer. The city hosts four harbours: El-Diekhila
Harbour, the Western and Eastern Harbour and Abu Qir Harbour (Frihy et al., 2004)(331).
Alexandria has a deep rooted culture and heritage. The eastern harbor is separated
from the western harbor, which is the main Egyptian commercial seaport; by a causeway
leading to the Pharos Island that currently hold Kayet-Bay Fort, which is built on the base of
the devastated old Alexandria Lighthouse (Morcos, 2003)(332) . With the opening of the Suez
Canal in1869, the Mediterranean resumed its importance as a link with the East. The
development of the northern regions of Africa and the oil fields in the Middle East has
increased trade in the Mediterranean. Its importance as a trade link and as a route for attacks
on Europe resulted in European rivalry for control of its coasts and islands and led to
campaigns in the region during both world wars. Since World War II the Mediterranean
region has been of strategic importance to the United States and, until its dissolution, the
Soviet Union.
3.1.2. Climate
The Mediterranean climate is subjected to both subtropical and mid-latitude weather
systems, and climate is also partly influenced by the northern mountain ranges. The region is
characterized by a windy, mild, wet winter and a relatively calm, hot and dry summer.
Generally air temperature differences between winter and summer are limited to about 15,
although local meteorological and geographic factors can result in extreme conditions, such as
on the coasts of Libya and Egypt where the air temperature can reach 50 0C (EEAA,
1999)(283). Egypt has an area of 1.109 million km2. The climate is warm and dry. The rain is
limited to the northern coastal area, while upper Egypt and the Red Sea Sinai mountainous
area are flooded from time to time (UNEP, 2002)(286). Precipitation along the Mediterranean
coast varies between 130 and 170 millimeters yearly and decreases gradually to the south.
The tidal range is about 30 to 40 centimeters (EEAA, 2001)(333). The Mediterranean coast of
Egypt experienced a successive increase in the amount of annual rainfall during the last three
decades. The mean trend over the area is 0.76 millimeters per year. Rainfall has increased
over the western coast of Egypt by up to 3 millimeters per year. The changes in the general
circulation of the atmosphere and effects of some incidents, such as the North Atlantic
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Oscillaiton phenomenon, seem responsible for this change. Increased intensity and frequency
of marine storms will necessarily increase the risks of transportation accidents and health
risks in the coastal zone (Susmita et al., 2009)(334). The amount of cloud cover has declined
over most parts of Egypt except over western part of the North Coast. This decline has
reached 0.02 Oct/ year. Relative humidity has increased in all parts of Egypt. The upward
trend has reached 0.35% / year especially over southern part of Egypt. Mean-sea- level
pressure has increased in recent years over most parts of Egypt. The positive trend has
culminated to +0.05 hPa / year over southwest part of Upper Egypt (EEAA, 1999)(335).
The Mediterranean coastal zone of Egypt is situated at the northern margin of the large
hot and dry subtropical desert area of the Sahara and the Arabian Deserts. Winds from the
Mediterranean are relatively cool and bring some moisture during the winter of the northern
hemisphere. The combination of high temperature and low humidity leads to a high
evaporation (especially during the summer) which far exceeds the spare precipitation. This
clearly illustrates the importance of irrigation water from the Nile for agriculture in the delta
region. Air pressure and sea water temperature correlate well with the seasonal variation of
the mean sea level over a range of 0.17-0.19 m (Dewidar, 1997)(336).
Depending on the atmospheric circulation in the Eastern Mediterranean including the
Egyptian coast, three seasons have been distinguished: a-Winter season occurs from
November to March, b - spring season extends from April to May and c-Summer season
which covers the period from June to October (Hamed, 1983)(337).
3.1.3. Topography
The Mediterranean zone is characterized by a wide flat coastal area along the sea and
existence of Maryout plateau in the northwest area (Shata, 2000)(338). The shoreline between
Salloum and Abu-Qir Bay is more or less straight with slight undulations forming small
embayment. The depths of the continental shelf edge generally increase with increasing shelf
width. The widest continental shelf in the southern Mediterranean is found in front of the
Nile Delta, where a shelf more than 70 km wide has been built up by the sediments of the
River Nile (Kamel, 2010)(339).
The beach and coastal landscape west of Alexandria are one of the most attractive
recreational sites in Egypt. The coastline is characterized by wide beaches consisting of white
oolitic carbonate sand. The shore is generally linear with few protective configurations.
However, the beach is not suitable for swimming because of the steep beach slope ranging 1:3
to 1:5. The beach is also characterized by hazardous rip currents (Nafaa and Frihy, 1993)(340).
The shoreline of the Nile delta, from Abu Quir to Port Said, is typically of an arcuate
coast. The delta beach and its contiguous coastal flat are backed by coastal dunes or by large
lagoons. The three main Nile promontories at Rosetta, Burullus, and Damietta interrupt the
sandy shoreline of the delta. The near-shore area is a hydrologically active zone characterized
by a gentle slope varying from 1:50 to 1:100, and a dissipative wide beach (Nafaa and Frihy,
1993)(340). Meanwhile, the coastal zone of the Nile delta is presently undergoing extensive
changes caused by both natural and anthropogenic influences (Stanley and Warne, 1993)(341).
The Nile Delta shore consists of sandy arcuate beaches, approximately 240 km in total length.
The beaches are backed by coastal flats followed by coastal dunes and three lagoons (from
east to west, Manzala, Burullus, and Idko). With the exception of the sand dunes, the coastal
plain is characterized by a low-relief surface <3.0 m above mean sea level. It has been
designated as a vulnerable zone to the impacts of climate change and the expected rise in sealevel (El-Raey et al., 1999)(342). Assessment presumed that by the year 2100, the coastal belt
with elevation below 1 m above sea level will be submerged (Stanley, 1988)(343).
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The coastline of the Sinai Peninsula is sandy, smooth and interrupted by two
protruding headlands at Port Said and the Bardawil-bulge coast. These headlands are
separated by two large embayments, backed by sand dunes and lowlands made of saltpans
(Sabkha). The bardawil lagoon is separated from the sea by a long, curved, narrow sand
barrier. Three inlets connect it to the Mediterranean Sea. The whole coast is essentially
unprotected except for the jetties built to stabilize the lagoon inlets, the fishing port at El
Arish and the breakwaters of El Arish power plant (Frihy, 2001)(344).
3.1.4. Subsidence and sea level rise
Local subsidence along the delta coast indicates that the entire northern Nile Delta
coastal plain is subsiding with preferential northeastward tilting. A general trend of natural
subsidence occurs along the northern Egyptian coast ranging from 0.4 to 5 mm/yr. This has
been indicated from archeological evidence and from date core sections. On the basis of tide
gauge records, recently reported a rate of subsidence of 2 mm/yr at Alexandria and 2.4 mm/yr
at Port Said (Dewidar, 1997)(336).
3.1.5. Current
The general pattern of circulation in the offshore area of the southern Mediterranean is
from west to east, while in the nearshore area there are some deviations from this rule and
some gyres and convergence zone exist (UNEP, 2002)(286). The surface circulation is
dominated by an anticyclonic circulation off Salum Bay in winter, spring and summer. In
nearshore areas, the current flows eastwards at the shallower levels and it flows westwards at
the deeper levels. Off the Nile Delta, the current is almost eastward with a higher velocity in
summer and autumn, while in spring it is very weak. Off the area between Port Said and
Rafah, there is a clear cyclonic circulation appearing in all seasons except winter. At 50 and
75 m depth, the velocity of the circulation is weak. At 100 m depth, the circulation that
appeared between Matruh and Alamen in summer decreases in area and magnitude at the
former depths. At 200 and 300 m in winter, the current velocity is quite low. In spring the
current flows southwards off the area between Rafah and Port Said. In summer, the current
off the area between Port Said and Rafah is quite strong and flows to the south. The situation
in autumn is quite similar to that in summer, except in the eastern area, where the current is a
westward one. Investigation of the circulation pattern in the Egyptian Mediterranean water is
obviously of great importance for discovering the distribution of coastal pollution and the
possibility of intermediate water formation (Kamel, 2010)(339).
3.1.6. Wind movement
The normal wind regime along the Mediterranean coast of Egypt is controlled by
various atmospheric conditions that occur on a seasonal basis (Abdallah, 2006)(345). Along the
Egyptian Mediterranean coast, winds blow predominantly (60-65% of time) from the
directions from West to North. Speeds average 3-5 m/s in summer and spring; 5-7 m/s in
winter with the higher velocities from North-Westerly directions (Dewidar, 1997)(336).
According to Shata and Hanitsch (2006)(346) most stations have an annual mean wind
speed more than 3.1 m/s except Damietta and El-Arish and the main wind direction over the
Mediterranean Sea is north and northwest. The Mediterranean zone is windy. The wind
speed has a maximum value of 6.3 m/s at Mersa Matruh in March, and a minimum value of
2.0 m/s at El Arish in October. This zone characterized by sea land winds. High wind speed
occur in the winter and spring seasons. This may be due to the Mediterranean Sea secondary
depressions (Salem, 1995)(347).
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Mean monthly and annual wind speed (m/s) and its direction (at a height 10 m) (Shata and Hanitsch, 2006)(346)
Month
Annual mean
Station
Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec
Salloum
5.4
4.8
5.2
4.3
3.6
4.1
4.7
4.2
3.4
3.6
4.2 5.2
4.4
Sidi
5.6
5.7
5.9
5.7
4.7
4.6
5.2
4.5
4.0
4.2
4.6 5.6
5.0
Barrani
Mersa
6.1
6.1
6.3
5.7
4.9
5.2
5.2
4.7
4.4
4.3
4.8 5.9
5.3
Matruh
Eldabaa
5.5
5.8
6.1
5.8
5.1
6.0
6.0
5.6
5.0
4.2
4.5 5.5
5.4
Dekheila
4.5
4.6
4.9
4.8
4.3
4.5
4.6
4.5
4.2
3.7
3.8 4.1
4.4
Alexandria 4.4
4.4
4.6
4.3
4.0
4.0
4.4
4.0
3.7
3.1
3.4 4.1
4.0
Balteam
3.6
3.5
4.2
4.0
3.5
3.7
3.9
3.6
2.8
2.3
2.6 3.3
3.4
Damietta
2.8
3.1
3.7
3.6
3.1
3.1
2.8
2.5
2.3
2.5
2.5 2.9
2.9
Port Said
4.8
5.2
5.8
5.4
4.8
4.6
4.3
3.8
3.8
4.1
4.3 4.3
>4.6
El-Arish
2.5
2.9
3.0
2.5
2.4
2.4
2.3
2.1
2.2
2.0
2.1 2.4
2.4
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Wind direction
330 NW
330 NW
330 NW
330 NW
360 N
330 NW
330 NW
330 NW
360 N
330 NW
3.1.7. Waves
Information on Mediterranean waves is available from observations on ships plying
the Eastern Mediterranean (Hogben et al., 1986)(348) and from observations on a platform in
almost 20 m of water depth in Abu Qir bay (Nafaa and Frihy, 1993)(340). Dominant wave
periods are 5 to 8 seconds; extremes reach 10 seconds and more. The height and direction of
the wave are affected by refraction, dissipation of energy and breaking during their
propagation into shallow water near the coast. The effect depends highly on the local
bathymetry which may allow only a limited window of directions to reach a certain location.
These processes explain why about 70% of the waves arrive from between WNW and NNW
at the location in Abu Qir Bay. The directional distribution of the wave energy along the
Mediterranean Sea generally leads to a clockwise rotation of the local dominant wave
direction (and to a decrease of the wave energy) as the perpendicular to the local part of the
coast turns from NW to NE. The wave direction, however, always trails behind the
perpendicular. This cause a resultant eastward longshore direction of the wave energy along
the coasts, except in the eastern part of Abu Qir Bay, in the "shadow" of the Rosetta
promontory (Wobber, 1967)(349).
3.1.8. Tides and storm surges
The tidal variations of the Mediterranean Sea level are small with a range of about
0.15 m during neap tides and 0.3 m during spring. Extreme astronomical deviations from
mean sea level are 0.3 to 0.4 m. The tide is slightly higher in Alexandria than Port Said.
Winds and other meteorological phenomena cause deviations from the astronomical tides,
leading to the occurrence of higher levels. (Delft hydraulics, 1992)(350). The tides along the
Nile Delta coast are semi-diurnal in nature with two high and two low water levels in a tide
day (El-Fishawi and Fanos, 1989)(351). Storm surges during winter season and swell action
during summer season (Manohar, 1976)(352) may cause the increase of the water levels
considerably and hence accelerate erosion along the coasts (El Fishawi and Khafagy,
1991)(353).
3.1.9. Hydrology
Different water masses are found off the Egyptian coast: a surface water mass of high
salinity, a subsurface water mass of minimum salinity and maximum oxygen content of
Atlantic origin that extends below 50-150 m, an intermediate water mass of maximum salinity
extending below 150 m to about 300-400 m depth, and deep waters of Eastern Mediterranean
origin (Said and Karam, 1990)(354).
Generally, the eastern part of the Egyptian Mediterranean has higher surface
temperature vary between 18.12 0C in spring and 24.3 0C in summer, decreasing with depth.
The surface salinity varies between 39.05 psu in summer and 38.61 psu in winter. The
highest values of surface salinity observed in the offshore water. The maximum average of
vertical salinity observed in summer (39.13 psu) with the minimum value (38.63 psu) found
in winter. The vertical salinity values increased to 50-75 m depth, below which the salinity
was almost constant (Lascaratos et al., 1999)(355).
Annual evaporation exceeds rainfall and river runoff over most of the Mediterranean
Sea, so that in general the Sea is characterized as having very high salinity (around 38 psu).
This highly saline water flows along the bottom of the Strait of Gibraltar and out into the
Atlantic Ocean, where it can be traced throughout the central Atlantic. On the other hand,
relatively less dense Atlantic water flows into the Mediterranean Sea in a surface layer. This
incoming Atlantic water is eventually turned into dense Mediterranean waters through
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evaporation. The water sinks into the deep Mediterranean Sea, due to winter cooling, and
exits the Mediterranean through the bottom of the Strait of Gibraltar. This cycle takes
between 100 and 300 years. Tidal amplitudes in the Mediterranean Sea are small and the
narrow continental shelves prevent tidal amplification along the coast (Tomczak, 1994)(356).
3.1.10. Biological diversity
The Mediterranean Sea can be characterized as having low biomass per-unit volume,
due to low nutrient levels, but high diversity, with over 10,000 marine species observed and
with a large proportion (28%) of them endemic. Although no species disappearance has been
reported in the Mediterranean Sea, changes in species composition and richness have been
determined for some areas, including the introduction of exotic species (e.g. the massive
introduction of tropical species from the Red Sea after the opening of the Suez Canal).
Currently, there are several endangered species reported in the Mediterranean Sea, including
the Monk seals, red coral, sea turtles and colonial water birds. One of the most important
habitats in the Mediterranean Sea is the large seagrass meadows (Posidonia oceanica) that
occur at depths down to 40m in the western and eastern basins (EEA, 1999)(283).
3.1.11. Ecosystem features
The main features that characterize the ecosystem of the Egyptian Mediterranean
coastal areas are:
1. Seasonal variations in water salinity ahead of Nile Delta River, during the autumn and
winter seasons. The water-salinity values are mostly lower (average surface-water salinity =
38.4 psu) than for the surrounding offshore area (average surface water salinity = 39.1psu).
2. Estimated N:P ratios have higher values during winter (N:P range = 45–60) compared with
other seasons (N:P range = 5–15), indicating phosphorus as a limiting factoring in winter. In
contrast, the western side of the Nile Delta River is nitrogen limited during the other seasons.
3. Both algae biomass (chlorophyll-a), and zooplankton standing crop values have higher
concentrations (averages=4.5mg gm-3, average=3×103 Ind. m-3, respectively) during the
winter season compared with the other seasons (range = 0.5–1.5 mg m-3, range = 1–2 × 10-3
Ind. m-3, respectively).
4. A marked algae bloom in winter is mainly localized around the Nile Delta River and
extends eastward. It is also important to mention that, seasonal primary productivity
distribution in this area have high values during the autumn and winter seasons (average = 6.0
and 3.25 mg C m-3, respectively) compared with the other two seasons (range = 0.5–1.5 mg C
m-3) (Dowidar, 1988)(357).
3.1.12. Water quality
Egyptian Coastal Water Environmental Monitoring Program is one of the sustainable
programs which the MSEA and EEAA are adopting to monitor marine and coastal
environment area status. Within this program marine samples from selected stations along the
Egyptian coasts are collected and analyzed four times a year in March, May, July and
September so as to represent physical, chemical and biological conditions in the four seasons.
The program, implemented in collaboration with specialized stakeholders in Egypt, focuses
on the monitoring of bacterial counts and nutrient salt concentrations (ammonia, nitrates,
nitrites, phosphorus and chlorophyll). In addition, other hydro-geographic measurements are
also made they might help in explaining phenomena affecting marine environment conditions.
According to the results of the 2010 (the study year) monitoring program there was an
increase in the concentration of dissolved oxygen in sea water observed (higher than 4 mg/l)
from El-Salloum to Rafah. The results were higher than the results obtained during the cruise
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of 2009. The Salinity varied between 28.15 psu to 39.28 psu. Regarding water transparency,
the highest transparency level was found in the western zone extended from El-Salloum to El
Bitash (west of Alexandria) and the eastern zone extended from El Arish to Rafah, whereas,
the lowest level was found in front of Nile Delta as well as drainage and lakes exists. The
hydrogen ion concentrations in sea water, samples results ranged from (6.3 to 7.2) which was
within the normal level of sea water. Temperature ranges from 19.8 to 31 C0 which was
within the normal level of sea water.
The results of nutrient salts and chlorophyll-a concentration in the Mediterranean Sea
were as follows: There was a relatively decrease in the concentration of total nitrogen during
the four crises. The highest value (0.29 mg/l) was observed at Port Said station during May
2010 and the lowest value (0.054 mg/l) was observed at Baghoush Station during September
2010. Comparing the results of total nitrogen during (2007-2010) there is a relatively
decrease in total nitrogen concentration. The concentration of ammonia was relatively low at
all stations during the year. The highest value was observed at Port Said Station (0.44 mg/l)
and the lowest value was (0.003 mg/l) at Baghoush station. There was a decrease in the
concentration of ammonia during 2010 comparing with the results of ammonia obtained
during the period from 2008-2010. The lowest value (0.23 µg/l) of chlorophyll was obtained
at the reference station in Baghosh and the highest values (15.05 and 13.03 µg/l) were
obtained at Port Said and El-Gamil west stations, respectively. There was a decrease in the
concentration of chlorophyll a at the year 2010 comparing with the concentrations obtained
during the period from 2008-2010. The concentration of nitrite was very low at all stations.
The highest value (0.03 mg/l) was found at the Eastern Harbour and the lowest value (0.001
mg/l) was found at El-Salloum station. The concentration of nitrates ranged from (0.007 to
0.115 mg/l). The total phosphorus was very low and ranged from (0.015 to 0.078 mg/l). The
highest value was observed at El-Mex station and the lowest value was observed at the
reference station at Baghoush.
Different patterns of fecal contamination ranging between very clean stations such as
Salloum, Baghoush and Rafah and other clean stations where the number of bacteria did not
exceed the permissible limits. El-Dikhaila, El-Mex, Western Harbour, Eastern Abu Qir,
Rashid, El-Bourg showed high level of Faecal contamination caused by the discharge of
agricultural run off or sewage discharge or both in these areas (EEAA/EIMP, 2011)(358).
3.1.13. Sediment characteristics
The Egyptian Mediterranean coast contains a wide variety of sediments. Coastal
sediments are mainly composed of two principal types: carbonate sands and quartz-dominant
sands (Hilmy, 1951)(359). The western region from Mersa Matrouh to Alexandria is mainly
composed of pure oolitic carbonate (Anwar et al., 1984)(360). The oolitic grains constitute an
average of 78% and 89% of the bottom and beach sediments, respectively. Further westward,
sand at Salloum showed an average oolite content in the nearshore of 58% (Anwar et al.,
1981; Nasr et al., 1978)(361, 362).
Beach sand from Alexandria to Rashid primarily consists of quartz grains with
common shell fragments (Hilmy, 1951)(359). The Nile River has been identified as the major
source of quartz-rich sediments and heavy minerals on the Nile Delta beaches (Frihy,
1994)(363). As a result of the current erosion affecting beaches along the Nile delta, the eroded
shorelines are characterized by high density minerals, whereas the accreted shoreline is
characterized by low density minerals (Frihy et al., 1997)(364). Black sand deposits are
extensive along the beaches of Rashid on the Egyptian Mediterranean coast (Saleh et al.,
2004)(365). These black sands contain some important minerals such as zircon and monazite
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(El-khatib and Abou El-khier, 1988)(366), both of which contain uranium and thorium in their
chemical structures. High background radioactivity values at Rashid have been detected as a
result of these black sands (El-Gamal et al., 2004)(367). Grain size analysis of El-Salloum
yellow sand documents discrete size groups extending from gravel to fine sand. This sand is
characterized as bimodal sediments predominantly in the fine sand size class. Alexandria,
Matrouh, and Rashid sands are all characterized as unimodal sediments in the medium sand
class size (El-Gamal et al., 2010)(368).
3.2. Economic Activities of the Egyptian Mediterranean coastal
area
The coastal area of Egypt on the Mediterranean Sea extends for about 1,200 km. It
hosts a number of important residential and economic centers, like the cities of Alexandria,
Port Said, Damietta, Rosseta, Matruh, and Al-Arish. The coastal strip between Alexandria
and Matruh hosts tens of tourist villages, which are usually crowded by visitors during
summer. Many activities are known in the coastal area, including fishing, industrial, tourism,
trading and agricultural, oil and gas production, and transportation (Dorgham, 2011)(369).
Highest population centers (25 million, i.e. 38% of the total population) are found at
the coastal zone of Alexandria and Nile delta. The Nile river delta contributes 30–40% of
agricultural production and 60% of fish catch (marine and lagoonal). Most of the fertile soils
of the Nile delta are used for agriculture, irrigated by the water supplied from the Nile system.
Lands near the lagoon margins, especially at locations with low elevation, have been
reclaimed for the purpose of agriculture and fish farms exist. Half of Egypt’s industrial
production comes from the delta, mainly from Alexandria. Main commercial ports are located
at El Diekhila, Alexandria, Abu Quir, Idku, Damietta, Port Said and east of Port Said. The
beaches of the Nile delta are little used whereas Alexandria coast is totally used as public
resort summer beach. Main delta resort beaches are located at Gamasa, Baltim, Ras El Bar, El
Gamil and Port Said. A coastal road was recently completed to connect the delta region with
west of Alexandria and Sinai. In the Nile delta few people live below1 m contour, whereas
populations are usually concentrated on ground more than 2 m above sea level, especially on
the raised present or past alluvial channels (Sestini, 1990)(370). Commercial and industrial
centers are situated above the 3 m-contour line. Economic activities in the coastal zone
include agriculture, industry, fishes/aquaculture and recreation beaches are denoted in the
following map:
Map of the lower Nile delta showing (A) main anthropogenic features below and
above mean sea level up to 4 m contour and (B) distribution of main
socioeconomic elements across lower Nile delta coastal plain.
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The coastal area may be roughly divided into three units, in accordance to the presence
of the hot spots, sensitive areas and main geomorphologic characterization. The main unit is
the central unit between Alexandria (west) and Port Said (east), including the Nile Delta,
coastal lakes (lagoons) and wetlands, and stretched along the boundaries of six governorates.
The second unit, to east from the central unit, extends from Port Said to Rafah (north Sinai
Governorate), while the third unit, extends from Alexandria westwards to Salloum (Mersa
Matruh Governorate). Associated with the Nile Delta are the coastal lakes and wetlands,
which represent another important feature of the Mediterranean coast of Egypt. These lakes
and wetlands cover a total area of at least 280,000 hectares, representing more than 25% of all
natural and semi-natural wetlands of the entire Mediterranean region. Many areas in the
Mediterranean coast are used for recreation by the public. The Mediterranean Coastline and
coastal lakes are also major sources of fishery resources in Egypt (Barakat et al., 2011)(371).
3.2.1. Population
Egypt is located in the north-eastern corner of the African continent with an area about
million square kilometres. It is considered a developing country burdened by the scarcity of
natural resources associated with extreme population growth (over 70 million people in total).
The Egyptian terrain consists of a desert plateau interrupted by the Nile valley and
delta. It is located in an arid - to semi-arid zone. The coastal zone of Egypt extends for more
than 3,500 km and 40% of the population live there. Most of these people live in and
around a number of major industrial and commercial cities: Alexandria, Port Said,
Damietta, Rosetta, and Suez (El-Raey, 1999)(372). The coastal zone is home to several highly
populated economic centers, such as the cities of Alexandria, Damietta, Hurgada, Port Said,
Suez, and Sharm El-Sheik (EEAA, 2006)(373).
Given the country's geography and the concentration of population and economic
activities in the Nile Valley and the Nile Delta (almost 99%), Egypt faces environmental
pollution problems, affecting the health of the population, and the poorest sections are
exposed the most. Inadequate sewage disposal, uncontrolled industrial effluents, extremely
uneven population distribution, and shortage of arable land, have created major water
pollution problems (DG Environment European Commission, 2006)(374).
The pressure of the human activity on the coastal resources of the Mediterranean and
Red Seas are very intense. The Egyptian coasts receive the impact of the major part of the
country's population. The enormous urban population and adjacent already formed
agricultural lands, all, contribute to the pollution load reaching coastal waters, weather direct
like the Alexandria region or via coastal lagoons (such as lake Manzala which receives the
major part of the Cairo mixed waste water). Egypt industry also contributes to the pollution
load in the coastal waters (UNEP, 2002)(286).
Alexandria Governorate extends for about 75 km along the Mediterranean coast of
Egypt. It is the main harbor of Egypt were 60% of the country's exports and imports going
through it (Beltagy, 2009)(375). Alexandria covers an area of about 2758.4 km2 and the widest
part of the city ranging from 4 to 5 km length (Abdrabo et al., 2002)(376). The population of
Alexandria, according to the 1996 Census, reached a total of 3.33 millions (CAMPS,
1996)(377), compared to 2.917 millions in 1986 (CAMPS, 1986)(378). This means that a
population increase rate of about 14.1% over a decade. The population of Alexandria,
according to the 2004 Census, reached a total of 4.30 millions (CAMPS, 2004)(379), compared
to 3.33 millions (CAMPS, 1996)(377). This means that a population increase at a rate of about
30% through 8 years. Being the first summer destination in the country, if receives an extra
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of 1, 500, 5000 visitors in this season. Administratively, the Governorate is divided into six
districts (Beltagy, 2009)(375).
Port Said Governorate lies at the northern end of Suez Canal with a total area of 1,
344, 96 km2 and a populated area of 1,320.68 km2. It's bounded by North Sinai Governorate
to the east; Damietta Governorate to the west; Lake Manzalah, Dakahlia, Sharkia and Ismailia
Governorates to the south. The total population was 529,684 inhabitants in 2004.
Administratively, the Governorate includes 7 districts. Activities in the Governorate are
diversified as they range from agriculture, industry, to maritime services and shipyards.
(Beltagy, 2009)(375).
3.2.2. Main activities sectors
Agriculture
More than 95% of Egypt's population and all of its agriculture are concentrated in <
5% of Egypt's land, along the banks of the Nile and throughout the 25,000 km2 Nile delta.
For more than 5,000 years, Egyptians depended on the annual fall flooding on the Nile, which
irrigated and fertilized the floodplain and eventually discharged to the Mediterranean Sea
(Oczkowski et al., 2009)(380).
Agriculture contributes 17% of the country's GDP and is the largest source of
employment, constituting 30% of the labor force. About three fifths of the country's
agricultural production is the low lying delta in close proximity to the Idku, Burullus, and
Manzalla lagoons (Agrawala et al., 2004)(381). In the northeastern part of Nile delta the most
important crops by cultivated area: clover, maize, cotton, wheat, rice, sunflower, sugar beets
and soya beans, are the predominant cops of the lower coastal zone (< 2 m) (Sestini,
1992)(382).
Tourism
Tourism currently represents 11.3% of Egypt's GDP, 40 percent of total
noncommodity exports, and 19.3% of foreign currency revenues (EEAA, 1999)(335). The
Northwestern Coast of Egypt extends for about 500 km along the Mediterranean Sea from
Alexandria to El-Salloum, near the Libyan border. This coast is well known in Egypt as a
splendid summer resort, including numerous new tourist resorts. On the level of government
policy, tourism is placed very high in development goals where the tourism sector represents
about 25 % of economy sectors in Egypt. In opening up the Northwestern Coast, Egypt would
enter the highly competitive beach-oriented Mediterranean tourist market, and an important
part of population could be absorbed in this zone (Rabenau., 1994)(383).
The northwestern coast of Egypt is distinct by white oolitic carbonate sand, clear blue
water, mild weather and sun prevailing most of the year. All these privileges made this coast
an attractive site to be developed rapidly for tourist and entertainment activities and for
recreational uses. El-Allamien Marina large scale recreation summer tourist center is one of
these rapidly developing areas on this coast at kilo 94 on Alexandria Matrouh coastal road
(Iskander et al., 2008)(384). In addition to international tourism, the coastal cities of
Alexandria, Bulteem, Gamasa, Port Said, and Ras El-Bar are popular destinations for local
tourists, and many middle and low income Egyptian spends their summers in these towns (ElRaey, 2010)(385). The main resort at the northern part of the Nile Delta is Ras El Bar resort
(Dewidar, 1997)(336).
Port Said Governorate considered free trade zone and play a good role in import and
export business. Also, there are some touristic sites such as, Delecips monument, Army
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Museum, resort beaches and Suez Canal (Dewidar, 1997)(336). The coastal zone of Port Said
area is socioeconomically important to most of the population in this area. Tourism is
primarily oriented toward swimming and sunbathing. Therefore, the coast, its slope, and the
quality of beach and sea are of prime importance to this industry. Most tourist facilities such
as hotels and youth camps are located within 200 to 300 m of the coast. There are also
important archaeological sites along the northern part of the Suez Canal (El-Raey et al.,
1999)(342).
Industrial and commercial activity
Half of Egypt’s industrial production comes from the delta, mainly from Alexandria.
Main commercial ports are located at El Diekhila, Alexandria, Abu Quir, Idku, Damietta, Port
Said and east of Port Said (Sestini, 1990)(370). Many industrial activities, including petroleum
and chemical production, and important tourism centers are located along the coasts. Trading
and transportation networks and a large number of harbors are also found on the coasts
(EEAA, 2006)(373). The Mediterranean coastal shoreline of Egypt includes five large lakes,
which constitute about 25 % of the total wetlands in the Mediterranean region. This coastal
zone has a large number or economic and industrial centers, as well as important beaches and
tourist resorts (EEAA, 2001)(333). The Mediterranean coastal zone is the site of Egypt's second
largest city, Alexandria, which is the country's main harbor. Located on the western side of
the Nile Delta, the city sits partly on low elevation land. Alexandria is home to about 40
percent of the country's industrial capacity, in addition to being a prominent summer resort.
The western harbour is heavily polluted receiving several extended and internal pollutants
from various sources. The major plants and other sources discharging their wastes into the
western harbor are Misr chemical Industry, Portland cement factory, discharges from
Noubaria canal, effluents from Mex pumping station, cement packing unit discharged wastes
from vessels during unloading and loading of the exported and imported industrial raw
materials and domestic wastes discharged from the city of Alexandria (Saad et al., 2003)(386).
Abu Qir drain recieves and discharges industrial wastewatres from 22 different factories of
food processing and canning, paper industry, fertililzer industry and textile manufacturing
(Abdeloneim and Shata, 1993)(387). El-Mex bight recieves the industrial wastes from the
nearby chemical industries, chlor-alkali plant and Alexandria petroleum company (Ezzat et
al., 1985)(388).
Other large cities in the northern low-lying Delta region include Rosetta and
Damietta. Port Said is an important regional trading center on the Suez Canal to the eastern
side of the delta (El-Raey, 2010)(385). The main income of this governorate depends on
revenue from Suez Canal, tourism, free trade zones and industrial activities. The industrial
activities of Port Said include food canning, cloth making, carpet weaving, and the leather
industry (IDSC, 1995)(389). The Nile Delta region in the Mediterranean coastal zone represents
the major industrial, agricultural, and economic resource of the country. It is home to over 50
% of Egypt's population of 80 million and to about 70 % of the nation's industrial and
commercial activities (EEAA, 2009)(390).
Damietta is an industrial center known for its furniture, leathers, textile and sweets
industries in addition to dairy products and rice mills and for its agricultural heritage. It is also
a fishing industry town, with one of the largest fleets on the Mediterranean which accounts for
fully half of the fishing boats of Egypt (Faragallah et al., 2009)(311).
Fishing activity
The coastal zone is an important source for fisheries, providing income and food
security, the annual production of fish is about 875,990 tons, of which 116.560 tons (13.3 %
-82-
of the overall production) are from the coastal waters (EEAA, 2006)(373). Approximately 60%
of Egypt's annual fish catch are from three main Delta lagoons, namely Idku, Burullus and
Manzalla. Fishing is mainly done by trammel net and various primitive methods (e.g.
catching by hand and collecting fishes under vegetation using a cone shaped net).
Aquaculture is the largest single source of fish supply in Egypt accounting for almost 51% of
the total fish production of the country with over 98% produced from privately owned farms;
most aquaculture activities are located in the Nile Delta Region with fish farms usually found
clustered in the areas surrounding the four Delta lagoons. Fish hatcheries are also generally
located in the vicinity of the fish farms. Pollution, reclamation, fragmentation, over fishing
and illegal harvesting of fish fry are the major environmental issues threatening the fragile
ecosystem of the northern lagoons (UNDP, 2009)(391). The Egyptian Mediterranean coast is
about 1100 km, extending from El-Sallum in the West to El-Arish in the East. The mean
annual fish production from this area does not exceed 45,000 tons (1984-2005) General
Authority for Fish Resources Development). There are 3 fishing methods conducted in the
Egyptian Mediterranean: trawling, purse-seining, and long-lining by hand. The fishing
grounds along the Egyptian Mediterranean coast are divided into 4 regions: Western region
(Alexandria and El-Mex, Abu-Qir, Rasheed, El- Maadiya, and Mersa Matruh), Eastern
region (Port Said and El-Arish), Demietta region, and Nile Delta region. The Port Said
region is one of the most productive fishing grounds, constituting 24% of the total fish
production in the Egyptian Mediterranean (Mehanna, 2007)(392).
The main fishing ground in the Egyptian Mediterranean water is the continental shelf
in front of the Nile Delta. Inshore fisheries widely exist with artisanal fishermen along the
coast. Fishing fleet is consisted from 3129 boats from them 1095 boat are trawlers varies in
length from 15 to 28m and in engine power from 50 to 850hp. Italian type bottom trawl net is
common with cod end mesh size about 2cm, they targeting shrimp, cuttlefish, red mullet and
different bream species. The continental shelf is generally fairly heavily exploited, trawl catch
target small size fishes that requested by the market. Discards and by-catch species are the
majority of the catch. Managements measurements are necessary include increasing mesh size
and modify of net design for conservation of the Egyptian fish stocks (Gaber and El-Haweet,
2011)(393). Fish farming is growing in importance and is sponsored both by local government
and Ministry of Agriculture. In Manzala lagoon there is also a considerable number of private
frame, consisting of seasonally opened and closed basins (hoshas). Aquaculture farm is
concentrated around the southern margin of Idku and Manzala lagoons and Maryut valley.
Aquaculture production accounted for about 15% (62,000 t) of total fish production in 1995.
Tilapia represents about 38% (129,000 t) of total fish production in 1994 (CAPMS, 1994)(394).
The Egyptian Mediterranean fishing fleet
Official statistics for the year 2004 (source: The Egyptian Fisheries Statistics 2004,
The General Authority for Fish Resources Development, The Egyptian Ministry of
Agriculture) reported 3027 fishing vessels with license registered in ports of the
Mediterranean coast. However, no data were available about the western area. So, the real
number of vessels is higher. Vessels are typically small, with exception of trawlers. During
the interviews with artisanal fishermen in the western and eastern regions, they stated that
they set the nets perpendicularly to the beach. The length of the net might range 80 to 400
meters. The net is left in the water from 8 hours to 2-3 days.
-83-
Number of vessels by fishing gear in 2004.
Region
Western
Alexandria
Central
Eastern
(Sinai)
Fishing
community
El salloum
Marsa Matruh
Sidi Barani
El Hamam
El-Mex
Anfoushi
Abou Keer
Total
El Maadiya
Edko
Rousata
Borg El Borolos
Azzbat El Borg
Domiat
Total
Port Said
Lake Bardawil
Areash
Total
Total
Trawlers
Longliners
Purse
seiners
Set
net
Others
total
No licenses were given n the year 2004
42
192
8
48
1
291
48
90
115
107
16
614
852
208
208
1150
209
401
87
69
190
207
553
155
11
166
1120
15
23
36
28
22
16
102
39
56
95
220
140
188
101
75
1
9
186
0
347
1
5
5
157
157
163
412
703
339
0
279
229
851
0
1698
402
0
224
626
3027
3.2.3. Land use
The construction of the International Coastal Highway which is located parallel to the
Mediterranean shoreline, is expected to cause changes in land use /land cover. A remarkable
set of wetlands is located near the coast, which, include wet Sabkhas, dry Sabkhas, and Lake
Maryiout splinters. Land use in the hinterland has different trends. The local people
(Bedouin) living in this region utilize natural vegetation for pasturing within some scattered
cultivation in the terrestrial inland (Mohammed et al., 2000)(395). Some urban growth takes
place around Lake Maryiout shorelines, mostly at the expense of the wetlands. A trend to
future expansion and land use changes due to tourist sites and summer resorts along the coast
is also obvious (Kaffas et al., 2000)(396).
Urban and industrial expansion in Alexandria is physically restricted, by the sea on
one side, and by Lake Maryut and agricultural land on the other side. This narrow strip of
land is now a continuous sequence of high rise buildings. Port Said has a great potential as a
commercial and industrial center at the north entrance of the Suez canal, but expansion has
been limited by lack of land, being squeezed between the sea and two lagoons (Dewidar,
1997)(336). The coastal zone west of Alexandria is characterized by intensive development of
new settlements of touristic villages. These recreational projects create potential problems
resulting from increasing human pressure and man made interventions. These recreational
projects were built without environmental impact assessment. Unexpected impact have
changed the shoreline stability and expected to alter the water quality of sea (Dewidar,
1997)(336).
The Port Said Governorate has a unique geographical location. It is located in a prime
position for national and regional development. In addition, it is considered an extension of
the Gulf of Suez and is close to the development in Sinai. Port Said has valuable resources
suitable for tourism development such as the Mediterranean Sea and the Read Sea beaches,
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lakes, protected areas and historical and archaeological sites. The economic policy and
potential use of the land accelerates the growth of the governorate. Therefore, Port Said has
quickly become the third largest urban governorate in Egypt with respect to population size.
By the beginning of the 1980s, the East Port Said harbour was constructed to help
maintain fuelling stations for ships passing through the Suez Canal waterway as well as a
summer resort for Egyptians. El-Malha Lake represents a valuable natural resource for
tourism development which will require infrastructure and productive wetlands. However,
land use conflicts already constrain tourism development of the El-Tina plain. These conflicts
prevent establishment of any significant comprehensive tourism activities along coastline and
adjacent areas. These conflicts may also discourage investors from supporting development
plans (El-Bastawisy et al., 2006)(397).
3.2.4. The economic importance of the coastal lakes
Five water bodies lie adjacent to the Mediterranean Sea on the fringe of the Nile
Delta; Lake Mariut, Nozha Hydrodrome, Lake Edku, Lake Brollus and Lake Manzalah. The
current configuration of these lakes is changing rapidly, due to the natural processes and
man's activities (Saad, 1990)(398). Contrary to the first two lakes, the latter three ones are
connected to the Mediterranean Sea. The main water supply to the Delta lakes comes from
the contaminated agricultural drains. However, the Hydrodrome is fed with the Nile water
(Saad, 1974)(399). Lake Mariut, located SW of Alexandria, is the heavily polluted lake in
Egypt and its water is pumped to the sea. It is now divided artificially into four basins and the
area of the lake proper reaches 27.3 km2 and its depth ranges from 90-150 cm. The Nozha
Hydrodrome was isolated from Lake Mariut. It has an area of 5 km2 and average water depth
of 3 m. Lake Edku, situated at 30 km NE of Alexandria, has an area of 126 km2 and water
depth ranging from 50-150 cm. Lake Brollus, situated between the two Nile branches, has an
area of 420 km2 and water depth varying from 70-240 cm. Lake Manzalah, the largest Delta
lake, is lying in the eastern region of the Nile Delta. Its area had been reduced to 1200 km2
by 1980 (Meininger and Mullie, 1981)(400).
The coastal Delta lakes are used for fishery, water supply for agriculture and industry.
The fish population in these lakes is comprised from fresh and marine water fish species.
Many of the latter migrate for feeding in these lakes through their connections with the sea
(Alsyes and Soliman, 1993)(401).
These lakes are situated on the Mediterranean Coast of the Delta and cover about 6%
of the non-desert surface area of Egypt. The lakes are an important natural resource for fish
production in Egypt, until 1991, these lakes have always contributed more than 40% of the
country's total fish production, but at present this has decreased to less than 12.22%
(GAFRED, 2006)(402). Tilapia species including Oreochromis niloticus, Oreochromis aureus,
Sarotherodon galilaeus and Tilapia zillii ranked first followed by Clarias gariepineus in the
fish production of the lakes (Saeed and Shaker, 2008)(323).
In the meantime, the lakes were subjected to a gradual shrinkage during the past few
decades due to land reclamation and transformation of significant parts of the lakes to fish
farms, particularly along the southern regions. In addition, large parts of the lakes are
overgrown with aquatic vegetation which reduces the open water to nearly half of its total
area, speeding up the process of land transformation (Saeed and Shaker, 2008)(323).
These lagoons are considered as reservoir for agricultural, industrial, and municipal
wastes, which are discharged from surrounding cities and cultivated lands. The Egyptian
-85-
Mediterranean coast receives huge volumes of waste waters every year through coastal
lagoons and from other land based effluents. These wastes are loaded by variable amounts
and types of pollutants, in addition to great amount of nitrogenous and phosphorus
compounds, which in turn cause high level of eutrophication along a significant part of the
Mediterranean coast, particularly of both the Nile Delta region and Alexandria coast
(Dorgham, 2011)(369).
3.2.5. The economic importance of the Egyptian Mediterranean ports
There are three main Egyptian ports on the Mediterranean; Alexandria ports, Port Said
and Damietta port, beside a number of terminals for oil transportation. Alexandria hosts three
harbors: namely the Western Harbor, which is the main harbor of the country, El'Dekhiela
Harbor just west of the Western Harbor, and the Eastern Harbor (fishing and yachting). It
also hosts about 40% of the country's industries (Beltagy, 2009)(375). Alexandria port
overlooks the Mediterranean in the north east of Africa and is considered as the main port of
the Arab Republic of Egypt. It is 4.8 km long and is 2 km wide at its widest point. It covers
8.5 km2 of water surface area (EMDB, 2003)(403).
Alexandria Port is situated on the north coast of Egypt, west of the Nile Delta,
approximately 200 km from Cairo. It is the main port in Egypt (Abdrabo et al., 2002)(376).
Alexandria port covers capacity of production 23.7 million ton / year, ship size 15000 gross
weight and 14 berths for petroleum, oil, fertilizers, cement, coal, and molase. The tankers of
oil terminal specified for the supply of the ships operating the port is considered as one of the
possible main sources of oil pollution of the port (APA, 2005)(404).
The Western Harbor receives the vast majority of Egyptian external trade. It is
considered as one of the most important and biggest harbors in the Mediterranean Sea from
the economical point of view (Saad et al., 2003)(386). The western Harbor of Alexandria is the
major trade port of the Northern Territory of Egypt. Industrial activities have increased
dramatically in this area over the past 20 years. The harbor handles approximately 75% of all
ship-borne cargo of the country. As a result of aggressive urbanization and industrialization
of the Alexandria region, the coastal waters in general and the Western Harbor in particular
have received considerable amounts of treated and untreated industrial, agricultural and
domestic wastes (Salem and Sharkawi, 1981)(299).
The Eastern Harbor is the most important embayment on the Mediterranean coast for
Egypt as well as for the world heritage, since it contains thousands of the ancient Egyptian
artefacts which encourage the idea of establishing a natural museum of underwater
archeology in the harbour. It is a relatively small semi-closed basin on the central part of
Alexandria coast, occupying an area of 2.53 km2, with an average depth of 6m. The harbour is
isolated from the open sea by a high thick breakwater, but connecting to the sea through two
narrow openings, Qayet Bey (Boughaz) and Silsila. The harbour is exposed to several kinds
of human activities, including fishing, yachts port, anchoring offishing ducks and ships, landbased effluents, boat building workshops, and recreation (El Rayis and Hinckley, 1999)(405).
El-Dekheila Port is the second largest port in Alexandria city. The port of El-Dekheila
is situated at 5 km west of Alexandria harbour. Dekheila port is a natural extension to
Alexandria port; constructions started in 1980 and the port became operational in 1986 with
the operation of the minerals berth. Dekheila port takes the place of the former airport
westward Alexandria Port, 7 km distanced from sea and 10 k m from land. El-Dekheila is an
extension to the port of Alexandria, it is newly constructed to accommodate and operate
bigger draft and size ships. It started to operate in 1986 and still under completion (EMDB,
-86-
2003)(403). El-Dekheila port was planned to serve Alexandria Iron and Steel factory. It lies to
the east of El-Agami head and to the west of El-Mex Bay (Abd-Alla et al., 1995)(406). The
area of the port is situated on the north coast of Egypt, west of the Nile Delta, approx. 210 km
from Cairo. El-Dekheila and Alexandria Port are the main port in Egypt. El-Dekheila is
located at the north east of the Mediterranean and is considered as the main port of
Alexandria. It covers total area of 6 km2, 2.74 km2 of water and land area of 3.2 km2. The
capacity of productivity is 15 million ton/year, ship size is 17000 GRT and 15 berths for
minerals, oil, grain, containers, multi-purpose (EMDB, 2005)(403).
Port Said is situated at the end of the Suez Canal on the Mediterranean Sea and is
around 4.8 km long and approximately 0.8 km wide with a water surface of about 2.8 km2.
Suez port is situated at the end of the Suez Bay with a water surface of about 2.85 km 2
including Adabia port. The Suez Canal is connected to the water surface in the south and the
ship's waiting area for those crossing the canal in the north. Port Said Port is situated on the
Northern entrance of the Gulf of Suez. It is considered one of the main Egyptian ports due to
its distinguished location at the crossroad of the most important world sea trade route between
East and Europe via Suez Canal, and the most extensive transshipment port in the world
(EMDB, 2003)(403).
Damietta Port is situated 8.4 km west of the Nile estuary of Damietta branch westward
Raas El-Bar, and it is 70 km distance from Port Said with a water surface of about 3.2 km2
(EMDB, 2003)(403).
The economic contribution of the ports and economic impact value of trade passing
through Egyptian ports consists of four components:
1- foreign trade (imported and exported cargo)
2- ship calling and passengers
3- containers handling
4- customers duties
Egypt's major ports handled 95.4% of Egyptian's waterborne trade in 2004 and 91.8%
of all Egypt container trade. Between 1994 and 2004 the containers volume at Egypt major
ports increased by 435% while total Egyptian ports volume increased by 473.8% (EMDB,
2005)(407). Foreign trade (imports) exceeded 46.1 million tons at 2004 while the total foreign
trade was 82.2 million tons. In 2003 the total foreign trade exceeded 73.7 million ton while
imports were 42.5 million ton. The volume of foreign trade was $ 24.1 billion (free on board)
imports and $ 14.33 billion (free on board) exports through the ports of Alexandria, ElDekheila, Damietta, Port Said and Suez Port (EMDB, 2005)(407).
The volume of maritime foreign trade was 90% of Egypt's total import/export cargo.
Alexandria and El-Dekheila ports handled 35.4 billion tons, with a trade value of 18.7 billion
dollars. These figures rank the port at the top position among Egypt's five major ports
(Alexandria and El-Dekhaila, Port Said, Damietta, Suez and Adabia and El-Sokhna) (EMDB,
2005)(407). The second component, were ship calling and passengers through the major ports
(14.2%) and (1%) respectively. Services provided to vessels as they call at the ports is a
major part of the impact of the ports on the local area economy. These services include
navigation services; the loading and unloading of the cargo itself; local arrangements,
accounting, purchase of supplies; the supplies themselves; fuel that is sold to the vessels,
cargo packing services, and customer services. The number (32776) of ships passes through
Egyptian ports including Suez Canal during 2004, while in 2003 that reached (29194).
-87-
However, the total number of ships in Egyptian ports have been increased with a percentage
of 12.3% where as the total number of passengers was (2,810,734) (EMDB, 2003)(403).
The third component was the growth of foreign container trade (27.3%). Containers
trade includes exports (1,422,008) and imports (1,477,702). They are illustrating the strong
competitiveness and the magnitude of the consumer population. Domestic container trade
increased with a percentage of 27.5% in relation to foreign trade. In 2004, containers activity
was 569867 while import one was 557528 (EMDB, 2005)(407). The forth components is the
vital yearly flow which is around $4440 million into the regional economy from port
operations, in addition to $1100 million from traffic in Suez Canal and 12100 million from
trade related. The port's activities would significantly impact Egyptian's economic growth
and add considerable costs to import and help in creating job opportunities (EMDB,
2005)(407).
3.3. Environmental Aspects
The Egyptian coastal zone bordering the Mediterranean Sea has high economic,
ecological, aesthetic, and recreational importance. Developments in this zone may affect the
terrestrial, coastal, and marine ecosystems. The vast majority of Egypt's coastal areas are
under increasing stress from both natural (erosion, dune quarrying, and subsidence and rising
sea level) and anthropogenic influences caused by population growth and increasing
development (Stanley and Warne, 1993)(341). The whole of the Egyptian Mediterranean coast
is polluted with Non-biodegradable Ocean borne debris such as plastic, rubber and nylon,
most beaches are also polluted to a greater or lesser degree with oil and tar the most polluted
areas were in the western region, the majority of beaches being classified as "moderately"
"heavily" or "very heavily" polluted (Clarke et al., 2000)(408).
The pressure on the environment of the coastal zone of Egypt is very intense, as it
combines intensive socio-economic activities and urbanized areas. Along the Mediterranean
coast of Egypt, there are eight coastal governorates. These are from west to east Matruh,
Alexandria, Behaira, Kafr El-Sheikh, Damietta, Daqahliya, Port Said, and North Sinai. The
enormous urban population and adjacent agricultural areas, all contribute to the pollution load
reaching coastal waters. These derived either directly from coastal cities discharge points; the
Rosetta branch of the River Nile, the Mahmudiya and Nubariya irrigation canals, drainage
canals discharged directly to the sea, such as "El-Tabia and El -Ummum", or from coastal
lagoons "lakes" Maryut, Idku, Burullus and Manzala. These sources discharge about 8 billion
m/y into the Mediterranean. This includes heavy loads of pollutants from various sources. The
maritime transport in the eastern Mediterranean, including oil tankers, commercial ships and
passenger ships, affect the coast to a large extent. The entire beaches are frequently polluted
by oil lumps, litter and plastic debris; even in the very far remote areas of the coast where
there are no related activities. Large parts of the Nile Delta suffer from severe coastal erosion,
although adequate protection and mitigation measures have been considered. Most of the
coastal lagoons "lakes" are however in crisis, suffering from the excessive discharge of
industrial ,agricultural and domestic sewage flow. Large parts of the Nile Delta suffer from
severe coastal erosion, although adequate protection and mitigation measures have been
considered. (UNEP, 1999((185).
Environmental issues of the Nile delta coast have become more prominent recently
due to increasing population and intensifying industry (El-Rayis, 2005)(409). Four large
coastal lagoons are being ecologically degraded owing to the discharge of untreated
wastewaters into the lakes. For example, sewerage output has extended to Mariut lagoon
from Alexandria on the northwestern Nile coast and that from Port Said, Demietta and
-88-
Matariya to Manzala lagoon on the northeastern coast, where heavy metals are significantly
enriched in the water and sediments. Cu and Cd in Manzala Lake have gone up almost 60%
and Zn has increased almost twofold. Nowadays, Mariut and Manzala lagoons are the most
polluted and Idku and Burullus follow behind. Presently these two lakes are facing more
environmental pressures from local industries and urbanization with the increasing pollutants
being expelled towards the lake coast (Kamal and Magdy, 2005)(410). Heavy metals and
related environmental conservation of the Nile coast calls for more attention. Many projects
implemented have targeted heavy metal distribution and transportation in relation to
aquacultural health and societal response. Abdel-Moati and El-Sammak (1997)(411) found that
Cd and Pb expelling into the lagoons has increased by 8-70 times during the past 25 years.
Pancreatic risk in the Manzala region seems to be closely associated with cadmium
concentration (Soliman et al., 2006)(412). Siegel et al. (1994)(413) stated that Pb, Zn, Hg and Cu
enriched in Manzala lagoon were primarily due to cheaper power generators after High
Aswan dam emplacement in 1964. This has considerably degraded aquacultural products
both in quantity and quality in the Manzala region, where it has long been the most important
aquacultural base, by providing more than 50% of aquacultural products for Egyptian.
Alexandria governorate coastal zone receives a large amount of metal pollution form
the principle industries of this region include fertilizers, agrochemicals, pulp, paper, power
plant, food processing, detergents, fibers, dysestuffs, textile, and building materials where, the
daily average industrial discharge amounts to 30,000 and 128-261,000 m3 per day domestic
sewage and 1-2 million cubic meters per day of agricultural wastes (El-Nemr et al., 2007a)(61).
-89-
Chapter IV
Material & Methods
4.1. Study area
The Egyptian Mediterranean coast extends between longitude 250 30/ E and 340 15/ E
and extends northward to latitude 330 N, Figure 1. It has a surface area of about 154840 km2
and its water volume is 224801 km3 (Said and Rajkovic, 1996)(414). The coastal zone of Egypt
on the Mediterranean extends from Rafah in the eastern region to El-Salloum to in the
western region for over 1200 km. It hosts five large lakes; namely Bardawil, Manzala,
Burullus, Edku and Maryut which represent about 25% in area of the total wetland of the
Mediterranean. It also hosts a number of important residential and economic centers of the
country including the cities of Alexandria, Matruh, Damietta, Rosetta, Port Said, and Al
Arish. Activities on the coastal zone including fishing, industrial activities, tourism, trading
and agricultural activities in the delta region (SMART, 2005)(415).
4.2. Sample collection, preservation and storage
Twenty surfacial sediment samples were collected from different selected stations
from El-Salloum at the west to Rafah at the east along the northern coast of Egypt
(Southeastern Mediterranean Sea) during July 2010. Egypt's Mediterranean coast is naturally
divided into three regions, the Eastern (Sinai) region from Rafah to Port Said, the Central
(Delta) region from Port Said to Alexandria, and the Western region from Alexandria to ElSalloum (Clarke et al., 2000)(408). The study area was divided into three regions as following:
region I located in the western area of the coast, represented by 11 stations from El-Salloum
to El-Maadia. Region II located in the Delta area, represented by 7 stations from Rashid west
to El-Gamil east and Region III located in the eastern region of the coast, represented by 2
stations from Port Said to Rafah (Figure 1 and Table 1).
The samples from the top (2-3) cm layer of sediments were collected using Peterson
grab sampler. The samples was placed into sealed polyethylene bags, carried to the laboratory
in an ice box and stored at -200C in the dark until analysis.
Separate samples were collected at the same time for the measurement of SEM/AVS.
Samples were collected in such a manner that exposure to atmospheric O2 is minimized, or
sulfides will be oxidized. Samples were collected in wide mouth jars with a minimum of
head space and stored at -20 0C. Under these conditions, no significant loss of AVS is
observed during storage for up to 2 weeks. Sample holding time must not exceed 14 days.
4.3. Sample pre-treatment
The samples, which were previously deep frozen at polyethylene bags at
approximately -20C0 in order to limit biological and chemical activities, were removed from
the freezer, thawed and placed in glass pettri-dishes then dried in an oven at 40 C0 until total
dryness. Subsamples of the dried samples were sieved through a clean sieve <2mm mesh size
to remove coarse materials such as pebbles, coarse organic pebbles and shells. Samples then
were homogenized with a pestle and mortar in order to normalize for variations in grain size
distribution. The dried homogenized sediment samples were packed in a clean well Stoppard
plastic containers and stored in a cool dry place until analysis. To remove any contamination,
all glassware and plastic vials used were initially cleaned with soap, rinsed thoroughly with
tap water and distilled water, and then soaked in 14% HNO3 (v/v) for 24 hr, cleaned
thoroughly with deionized water and dried. All solutions were prepared using Milli Q water.
-90-
1

El-Saloum
2

13

8 9  11
6  
3 4 5   10 12
1
 7
1 1 1 11
111111
14
111
16
 
15 1718
19
1
1 11 1
11 11
Figure (1): Satellite Map of Egyptian Mediterranean coast representing the sampling sites of the study area
-91-
20 
Rafah
Table (1): Description of the study area and sources of marine pollution
Region Station
Latitude
Longitude
Location
Western
Region
1
2
3
4
5
6
7
8
9
10
Middle
Region
11
12
13
14
15
16
17
18
Eastern
Region
19
31 33 30 N
31 10 40 N
31 05 15 N
31 08 10 N
31 09 21 N
31 09 18 N
31( 12( 52((
N
31( 12( 20((
N
31( 18( 29((
N
31( 15( 53((
N
31 16 40 N
31 26 56 N
31 27 46 N
31 27 46 N
31 3010 N
31( 30( 20((
N
31( 17( 10((
N
31( 18( 16((
N
31( 16( 44((
N
25 04 42 E
27 40 07 E
29 42 51 E
29 49 17 E
29 50 42 E
29 49 17(( E
29( 52( 59(( E
El-Salloum
Baghoush
El-Nobarreya
El-Dikhaila
El-Mex
Western Harbour
NIOF
29( 52( 59(( E
Eastern Harbour
30( 04( 11(( E
Abu Qir
30( 08( 55(( E
Power Station
30 09 12 E
30 21 33 E
30 21 53 E
30( 58( 28(( E
31( 47( 22(( E
31( 48( 36(( E
Maadia
Rashid West
Rashid East
Burullus
New Damietta
Ras El-Barr
31( 45( 58(( E
El-Gamil West
32( 10( 28(( E
El-Gamil East
32( 15( 29(( E
Port Said
Sources of pollution
Main sites of pollution
Industrial waste water and domestic sewage
from residential areas and tourist resort
areas.
Much of the wastewater is discharged into
the coastal lakes which are connected to the
sea.
Shipping activities
El-Nobarreya drain
El-Mex Out fall
Abu Qir Drain (El-Amiaa)
Lake Edku outlet
Western harbour
El-Dikhaila harbour
Eastern Harbour
Nile water and agricultural drains
contaminated with hazardous industrial
waste, domestic sewage, organic matter,
fertilizers and pesticides.
The Lake Manzalah receive the sewage
from Cairo via El-Baqar canal
Shipping activities
Resetta branch of the Nile
Damietta branch of the Nile
The outlet from Lake ElManzala (El-Gamil outfall)
Burullus outlet
New Damietta Port
Oil production and oil terminals
Shipping activities
Port said
-92-
20
31( 17( 07((
N
34 09 47 E
Rafah
-93-
4.4. Physicochemical characterization of sediment
4.4.1. Determination of water content (%WC)
Water or moisture content is the ratio of the weight of water to the weight of sediment
in a given volume of sediment, expressed as a percentage. Water content is useful when
contaminant concentration is to be reported on a dry weight basis, even if determined in wet
samples. Water content was determined according to Loring and Rantala, (1992)(416). For
water content determination, a known weight of each sample was placed in pre-cleaned and
weighed Petri dish. Petri dishes were then placed in an oven previously heated to 105 оC and
samples were left for 24 hours. Petri dishes were then removed from the oven and placed in a
desiccator until reaching room temperature. Finally, Petri dishes were weighed and these steps
are repeated until constant weight is reached. The percentage (%) of water content was
calculated from the following equation:
%WC 
WB  WA
100
WB
Where,
% WC= percent of water content
WB= weight of sample before drying
WA = weight of sample after drying
4.4.2. Determination of pH
Approximately 10g of dried sediment sample were suspended in 50 ml distilled water
and agitated for 10 min. The suspension was allowed to rest for about 1 hour with occasional
shaking until the pH was measured. A combined glass electrode connected to a pH meter
(Schott-Gerate GmbH) was used for pH measurements (Yan et al., 2010)(417).
4.4.3. Determination of grain size analysis
Grain size determination was made on the dried samples by the conventional sieving
method. About 50 gm split of each of the quartered sample was placed into the top most
sieve and the entire column was shaken on a mechanical shaker for about 15 min. The sieve
meshes give the class intervals 1000, 500, 250, 125, 63, 37 µm. These correspond to the phi
classes of 0, 1, 2, 3, 4, 5, respectively (Folk and Ward, 1957)(418).
4.4.4. Determination of total carbonate
Total carbonate content of the sediment samples was determined by titration technique
(Black, 1965)(419).
Procedure
1- Weigh 1 g of sediment and transfer into a 250 ml Erlenmeyer flak.
2- Add 30 ml of standardized 0.5 N HCl
3- The released amount of CO2 was determined by back titration with previously
standardized 0.25 N NaOH solution using phenolphthaline indicator.
4- The CaCO3 percentage was determined by the following equation:
Calculation
% CaCO3 = 100 x 0.05 x [NHCl x VHCl] – [NNaOH x VNaOH]
-94-
4.4.5. Determination of total organic carbon (TOC)
Total organic carbon (TOC %) was determined according to Walkely-Blak’s wet
oxidation method (Baruah and Bathakur, 1997)(420). 0.5-1 g of oven dried and fine grained
sediment; 10 ml of 1 N K2Cr2O7 and 20 ml of concentrated sulfuric acid (H2SO4) were added.
The mixture was gently shaken and allowed to stand for 30 min until completion of the
oxidation reaction. The mixture was diluted with 200 ml of distilled water, then 10 ml of
85% phosphoric acid (H3PO4) and 1 ml of diphenylamine indicator were added. Finally, the
mixture was back titrated against 0.5 N ferrous sulfate (Fe (NH4)2(SO4)2.6H2O), until violet
color was observed. A blank titration was carried out in the similar manner, then TOC% was
calculated according to the following formula:
TOC%= (B-S) x 0.5 x 0.003x (100/W) x 1.3 x 1.724
Where,
B: volume in ml of ferrous sulfate needed for blank titration,
S: volume in ml of ferrous sulfate needed for sample titration,
W: weight in grams of sample,
0.5 is normality of potassium dichromate, 0.003 is grams of organic carbon, 1.3 is Walkely
correction factor and 1.724 is von Bemmlen factor.
4.5. Determination of total metals
4.5.1. Sample preparation
Sediment samples were oven dried at 40 0C for almost a week. After drying, sediments
were ground with an agate grinder and stored in a plastic vial until analysis. Total elemental
analyses of sediments were determined according to Oregioni and Aston (1984)(421).
1- Weigh out accurately between 0.25-0.5 g of dried sediment into a clean Teflon beaker
cleaned with concentrated nitric acid and deionized double distilled water and dried at 105 0C.
2- Add 2 ml of concentrated HNO3 to the Teflon beaker evaporate to dryness at 80 0C. Add 5
ml of HNO3/HClO4/HF (3:2:1) mixture and evaporate to dryness, gradually increasing the
temperature to 120 0C to remove the HClO4 residue.
3- Cool to room temperature and use 5 ml of 0.1M HCl to rinse the beaker and transfer to a 25
ml measuring flask.
4- Add another 5 ml of 0.1 M HCl to the beaker and transfer to the flask. Make up the
volume with 0.1 M HCl.
5- Check that the entire residue has dissolved and, if necessary, repeat the digestion
procedure.
6- Total heavy metals (Fe, Mn, Zn, Cu, Pb, Co, Cr, Cd and Ni), were determined using Flame
Atomic Absorption Spectrophotometer (AAS). The concentrations of each trace metals was
calculated from the following equation:
Metal concentration (µgg-1) = (25/W) x µg ml-1
Where,
25 = final volume of each digested sample
W = dry weight of digested sample
µg ml-1 = concentration of each metal in the resulting solution measured by AAS
4.5.2. Measuring system
The determination of the metals in the sediment samples were performed with a
SHIMADZU AA6650 atomic absorption spectrophotometer.
An atomizer with an
air/acetylene burner was used for determining all the investigated elements. All instrumental
settings were those recommended in the manufacture's manual book. The wavelengths (nm)
-95-
used for the determination of the analytes were as follows: Cr (357.9), Cd (228.8), Co (240.7),
Ni (232.0), Cu (324.8), Zn (213.9), Pb (283.3), Mn (279.5), Fe (248.3)
4.5.3. Analysis of certified reference materials
In order to check for the quality of the method applied for the analysis of trace metals,
certified reference materials (IAEA-405: Estuarine sediment, International Atomic Energy
Agency, Vienna, Austria) were analyzed with the developed method. Replicate analysis of
this reference material showed good accuracy, with recovery rates for metals between 91.11%
and 104.99 % (Table 2). To determine the precision of the analytical processes, three samples
were analyzed in triplicate. The average values of the variation coefficients obtained (in
general, < 10%) can be considered satisfactory for environmental analysis.
Table 2: Concentrations of metals (µgg-1) obtained from the analysis of the Standard
Reference Materials (IAEA-405) and their recovery
Element
Cd
Co
Cr
Cu
Fe
Mn
Ni
Pb
Zn
Certified value
(gg-1)
0.73
13.7
84
47.7
37400
495
32.5
74.8
279
Determined
value (gg-1)
0.70
14.38
78.02
50.01
38334.35
460.08
31.45
77.44
254.2
Recovery %
95.89
104.99
92.88
104.84
102.50
92.95
96.76
104.99
91.11
4.6. Determination of leachable metals
Acid leachable metals were determined according to the method described by Tanner
and Leong (1999)(422). About 1 g of each dry sediment sample was extracted at room
temperature with 25 mL of 1 N HCl with continuous agitation on a shaker (Mistral MultiMixer) for 16 hours. Samples were centrifuged at 2000 rpm for 15 minutes and the
supernatant was separated by means of a pipette and kept in polypropylene bottles for
analysis. The resulting solutions were analyzed by Atomic Absorption Spectrophotometer
(Shimadzu AA 6650 Spectrophotometer). The concentration of metals (μgg-1) in each sample
was calculated using the following equation:
25 x µg mL-1
Metal concentration (µgg-1) =
W
Where,
25 = volume of the acid added
W = dry weight of the extracted sediment sample
μgmL-1 = Concentration of each metal in the resulting solution measured by AAS
-96-
4.7. Sequential extraction procedure for trace metals
Use of total metal concentrations as a criterion to assess the potential effects of
sediments contamination, implies that all chemical forms of a given metal have equal impacts
on the environment, which is highly improbable (Tessier et al., 1979)(15). Sequential
extraction procedures, as opposed to bulk sample analysis, allow for the differentiation of the
relative binding strengths of the trace metals to the various solid phases by successive
leaching with chemicals of increasing strength and selectivity (Kersten and Forstner,
1985)(423). This procedure provides much more data on the origin of pollutants, their reaction
pathways, biological and physico-chemical availability and possible remobilization, than does
bulk sample analysis (Abdallah, 2007a)(292).
In order to define the fractions of the trace metals (exchangeable, carbonate, Mn and
Fe-oxides, organic matter and residual fraction) and explore the metals bioavailability, the
sequential extraction technique represented in Figure 2 was performed according to the
procedure reported by Tessier et al. (1979) (15).
(i)
Exchangeable: The sediment was extracted at room temperature for 1 h with 8 ml
of sodium acetate solution (1 M NaOAc, pH 8.2) with continuous agitation.
(ii)
Bound to carbonate: The residue from (i) was leached at room temperature with
8 ml of 1 NaOAc adjusted to pH 5 with acetic acid (HOAc). Continuous agitation
was maintained for 5 hrs.
(iii)
Bound to Fe-Mn Oxides: The residue from (ii) was extracted with 20 ml 0.04 M
NH2OH.HCl in 25% (v/v) HOAc. This was performed at 96 ± 3 0C with
occasional agitation for 6 hrs.
(iv)
Bound to Organic Matter: To the residue from (iii) were added 3 ml of 0.02 M
HNO3 and 5 ml of 30% H2O2 adjusted to pH 2 with HNO3, and the mixture was
heated to 85 ± 2 0C for 2 h occasional agitation. After cooling, 5 ml of 3.2 M
NH4OAc in 20% (v/v) HNO3 was added for 30 min.
(v)
Residual: The residue from (iv) was transferred to Teflon beaker and digested
with HNO3-HClO4- HF mixture according to Oregioni and Aston (1984)(421).
The selective extractions were conducted in centrifuge tubes (polypropylene, 50 ml) to
minimize losses of solid material. Between each successive extraction, separation was
effected by centrifuging at 10 000 rpm for 30 min. The supernatant was removed with a
pipette and analyzed for trace metals, whereas the residue was washed with 8 ml of deionized
water; after centrifugation for 30 min, this second supernatant was discarded. The volume of
rinse water used was kept to a minimum to avoid excessive solubilization of solid material,
particularly organic matter.
The metal concentrations in each fraction were measured using Atomic Absorption
Spectrophotometer (Shimadzu AA 6650 Spectrophotometer)
Deionized water used in preparing stock solutions and in each step of the leaching
procedure was obtained from a Millipore Milli-Q3RO / Milli-Q2 system. All glassware used
for the experiments was previously soaked in 14% HNO3 (v/v) and rinsed with deionized
water. All reagents used in this study were of analytical grade.
-97-
Dry sediment (1 g)
8 ml of 1 M (1 M NaOAc, pH=8.2) continuous agitation. 1 h
Centrifugation
Liquid extract
Exchangeable phase
Residue
8 ml of 1 M (1 M NaOAc, pH=5 with HOAc) continuous agitation. 5 h
Centrifugation
Liquid extract
Carbonate phase
Residue
20 ml 0.04 M NH2OH.HCl in 25% (v/v) HOAc, at 96 ± 3 0C, occasional
agitation for 6 hrs.
Centrifugation
Liquid extract
Fe-Mn Oxide phase
Residue
3 ml of 0.02 M HNO3 and 5 ml of 30% H2O2 adjusted to pH 2 with HNO3,
heated to 85 ± 2 0C for 2 h occasional agitation. Cooling, 5 ml of 3.2 M
NH4OAc in 20% (v/v) HNO3 was added for 30 min.
Centrifugation
Liquid extract
Sulphides & Organic phase
Residue
Digestion with digested with HNO3-HClO4- HF mixture
Silicates & residual phase
Figure 2: Flow chart of the sequential extration scheme used in this study
-98-
4.8. SEM/AVS analysis procedure
The procedures for AVS; acid volatile sulphide and SEM; simultaneously extracted
metals analysis were stemmed from the method described by Allen et al. 1993(424) with some
modifications in the flow rate of nitrogen.
4.8.1. Reagents
Hydrochloric acid
Sodium hydroxide
Sodium sulfide standard (Na2S. 9H2O)
Mixed diamine reagent: 660 ml H2SO4, 340 ml reagent water, 2.25 g N-N-dimethyl-pphenylenediamine oxalate mixed with 5.4 g FeCl3.6H2O, 100 ml HCl, 100 ml reagent water.
4.8.2. Generation of H2S
Sample reaction train
-
The apparatus for the colorimetric method utilized a series of four flasks connected to
one another with Tygon tubing. The first was gas washing bottle, a 500 ml round
bottom flask containing reagent water immediately after the N2 cylinder.
-
The second flask was the reaction vessel, a 250 ml round bottom flask with a side arm
and septum.
-
Following were two 125 ml flat bottom boiling flasks containing the trapping
reagents.
4.8.3. Procedure
-
To release AVS from the sediment sample, 100 ml of reagent water, minus the volume
of water expected to be in the wet sediment sample, is added to the round bottom
flask.
-
Traps are filled with 80 ml of 0.5 M NaOH and the system is purged for 2 minutes at
280 ml/min with nitrogen gas. Make sure that the flow rate is not so high that the 0.5
M NaOH trapping solution is displaced from the tubes.
-
Transfer about 10 g of wet sediment to a strip of parafilm on a tared top pan balance.
Record the wet weight of the sample.
-
Transfer the weighed sample with the minimum of delay to the reaction flask.
-
Purge for 2 minutes at 280 ml/min.
-
Turn off the nitrogen flow (at the flowmeter), and using a clean plastic disposable
syringe add 20 ml of 6 N hydrochloric acid through the sidearm.
-
Start the magnetic stirrer and adjust the speed to enable moderate stirring of the
sample.
-
Bubble nitrogen through the sample for 30 minutes at 280 ml/min.
-99-
4.8.4. Analysis of sulfide
Colorimetric method
Trapped sulfide is measured by adding 10 ml of mixed diamine reagent to the contents
of the first NaOH trap and diluting to 100 ml with reagent water. Absorbance is measured at
670 nm after color develops for 30 minutes. Samples are compared with calibration standards
made up in 0.5 N NaOH (low range = 0-80 µg S2-, high range = 0-640 µg S2-). If sample
absorbance exceeds 0.6, the sample should be diluted with 1 M NaOH.
4.8.5. Calculation of AVS
-
The sediment dry weight/wet ratio (R) must be determined separately; acid volatile
sulfide can be oxidized or altered to non volatile forms during drying.
-
Transfer an aliquot of the sediment to a 100 ml evaporating dish. Weigh the dish plus
the wet sediment. Calculate the wet weight of the sample. Dry the sediment at 103105 0C and weigh. Calculate the dry weight of sediment.
-
Determine the ratio of dry weight to wet weight for the sediment sample:
R = Wd/Ww
Where R = ratio of dry weight to wet weight,
Wd = dry weight of sediment sample (g), and
Ww= wet weight of sediment sample (g)
-
Compute the sulfide concentration per gram dry weight of sediment
S
AVS (µmole/g) =
R x Ww
Where S = the amount of AVS in sediment (µmoles)
4.8.6. Determination of simultaneously extracted metals (SEM)
-
After AVS generation has been completed the contents of the sample flask is filtered
through an acid resistant membrane filter and diluted to 120 ml with milli Q water.
-
Metals (Cd, Cu, Pb, Ni, and Zn) are determined by atomic absorption
spectrophotometer model: AA6650 Shimadzu, comparing samples with calibration
standards.
-
Individual metals are reported as µmol/g dry sediment, calculated as
conc
µmol
= conc sol'n x 120
g dry weight R x W x AWSEM
where,
conc sol’n = the concentration of the element of interest in solution, in units of µg/ml,
R = the ratio of dry to wet weight;
-100-
W = the weight of wet sediment added to the sample flask, and
AWSEM = the atomic weight of the element of interest.
-
The toxicity of metals to organisms is estimated from the ratio of SEM: AVS, where
both are in units of µmol/g dry sediment. SEM is the sum of all metals analyzed (e.g.
C, Cu, Pb, Ni, and Zn). Previous work has indicated that ratios greater than unity
were associated with sediment toxicity.
-101-
Chapter V
Results
5.1. Physico-chemical Parameters
The results of physical-chemical parameters of various sites along the Egyptian
Mediterranean coastal area were tabulated in Table 3. The measurements were carried out
parallel to the sediment sampling of the present study.
5.1.1. Seawater Temperature
Temperature is affected by several factors including weather, removal of shading
stream bank vegetation and storm water (Faragallah et al., 2009)(311). The Egyptian
Mediterranean seawater temperature during the sampling time were in the range from 28.24 to
31.00 C (average: 29.34 C). The highest value was found at electrical power station
(Station 10), while the lowest value found at staion 1 in El-Salloum and station 4 in ElDikhaila (Table 3).
5.1.2. pH value
The pH values in surface seawater of the Egyptian Mediterranean coastal area from
El-Salloum in the West to Rafah in the East are listed in Table 3. It ranged between 7.21 and
8.30. The highest value was reported at station 1 (El-Salloum) and the lowest value was
reported at station 20 (Rafah). The wide variety of aquatic animals prefers of pH ranged
between 6.50 and 8.00. Outside this range the diversity may be reduced because it stresses
the physiological systems of most organisms and can reduce reproduction. Low pH can also
allow toxic elements to become more available for uptake by aquatic plants and animals. This
can produce conditions that are toxic to aquatic life (Faragallah et al., 2009)(311).
5.1.3. Salinity
Salinity, as temperature, is an important factor which affects the marine environment.
The results of surface salinity of the seawater along the Egyptian Mediterranean coast are
shown in Table 3. The results showed that salinity of seawater during the sampling time
ranged between 28.15-38.3 PSU with an average: 36.32 PSU. The highest value was found at
Rafah and Baghoush and the lowest value was found at El-Gamil West. A relatively decrease
of salinity was observed in the Mediterranean seawater in front of drainages and the Nile
delta. For example, it was 35.64 PSU at Rashid west (station 12) and 35.31 PSU at Rashid
east (station 13), 28.15 PSU at El-Gamil west (station 17) and 30.95 PSU at El-Gamil east
(station 18) in front of Manzala lake.
5.1.4. Dissolved oxygen
Dissolved oxygen (DO) is considered as one of the most important and useful
parameters in identification of different water masses and in assessing the degree of
pollution especially with organic pollutants which affects fish and other marine life through
oxygen reduction or depletion (Nessem et al., 2005)(425). The distribution pattern of DO along
the Egyptian Mediterranean coast is presented in Table 3. Dissolved oxygen in seawater
fluctuated between a minimum value of 4.99 mg/l at station 7 (NIOF) and a maximum value
of 7.56 mg/l at station 17 (El-Gamil west).
-102-
Table 3: Physical parameters of the surface seawater of the study area.
Region
Western Region
Middle region
Eastern region
Min.
Max.
Average
Station
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
Temperature
(oC)
28.24
28.25
29.67
28.24
28.47
28.44
29.45
28.98
29.29
31.00
29.66
29.16
29.07
29.40
30.30
30.33
29.24
29.17
29.97
30.46
28.24
31
29.34
-103-
DO
(mg/l)
6.64
5.57
6.76
5.02
5.03
5.86
4.99
5.73
5.89
6.13
5.74
6.26
6.06
6.80
7.19
7.14
7.56
7.13
5.97
7.19
4.99
7.56
6.23
Salinity
(%)
38.18
38.30
36.49
33.83
35.86
36.07
38.06
37.83
37.63
38.09
37.87
35.64
35.31
36.30
38.02
38.12
28.15
30.95
37.37
38.30
28.15
38.30
36.32
pH
8.30
8.24
8.05
8.04
8.14
8.09
8.10
8.16
8.19
8.22
8.05
8.14
8.22
8.09
8.18
8.11
8.07
8.16
7.95
7.21
7.21
8.30
5.2. Geochemical Analysis
The Mediterranean sea is an area where sediments have different geochemical
composition (Abdallah, 2007b)(303). The results of sediment characteristics including
sediment texture, organic carbon, and carbonate of the study area are given in Table 4.
5.2.1. pH
The pH values in surface sediments of the Egyptian Mediterranean coastal area from
El-Salloum in the West to Rafah in the East are listed in Table 4. It ranged between 7.20 and
8.38. The highest value was reported at station 8 in the Eastern Harbour (western region) and
the lowest value was reported at station 14 in Burullus (Middle region). The pH values
indicated that the surface sediments are slightly alkaline. It is worth to mention that, the
variation of the pH value influences the release or adsorption of each metal into the sediment
fraction (Smith, 1994)(426).
5.2.2. Calcium Carbonate (CaCO3)
Carbonate is often an important component of marine sediments and has been found to
be an important indicator of provenance and dispersal of terrigenous material (Loring and
Rantala, 1992)(416). The co-precipitation with carbonate minerals is of importance for a
number of metals, such as cadmium and zinc (Forstner and Wittman, 1983)(427). The results
of the analysis of carbonate content of the sediments are listed in Table 4 and Figure 3. The
range and the average of total carbonate percentage in the surface sediments of the Egyptian
Mediterranean coast were found to vary from 2.85 to 95.57 (37.78%). The highest value of
CaCO3 was observed at station 6 in the Western harbour (western region) and the lowest
value was observed at station 13 in Rashid East (Middle region).
5.2.3. Total Organic Matter (TOM%)
Total organic matter (TOM) is one of the most important collectors of pollutants in the
marine sediments. Organic matter tends to form strong organometallic complexes with
metals, rendering them immobile. An increase in TOM content may results in an increase in
the levels of metals in marine sediment (Massoud et al., 2007)(294). Table 4 and Figure 4
represented the distribution of organic matter content in the study area. The range and the
average of TOM in the surface sediments of the study area were found to vary from 0.082 to
1.72 % (0.56%). The highest value of TOM was observed at station 7 in NIOF (western
region) and the lowest value was observed at station 12 in Rashid West (Middle region).
5.2.4. Water Content (%WC)
The results of water content of all samples from different locations are presented in
Table 4 and Figure 5. In general, surface sediments of the Mediterranean Coast of Egypt
were characterized by moderate water content (average: 20.43%). Figure 5 illustrates the
same trend except NIOF and Western harbour. The maximum value of water content
(35.45%) was reported at station 7 in NIOF (western region) and the minimum value of water
content (7.49%) was reported at station 6 in the Western harbour (Western region).
5.2.5. Grain Size
The grain size of the sediment is a specific parameter, this is one of the major or
controlling factors for the distribution of trace metals in coastal areas (Ip et al., 2007)(428).
The percent of sand, silt and clay together with the texture of all samples are shown in Table
4. Textural characteristics of the sediments are showed according to the classification of
-104-
Shepard (Shepard and Moore, 1954)(429). Result showed that fine sand (0.250-0.125 mm) and
very fine sand (0.125-0.062 mm) were the dominated fractions of all sediment samples. The
range and average of sand were 86.85 to 100% (99.15%). Course sand (1.0-0.5 mm) was
observed only at station 6 in the Western harbour (Western region). The highest value of
sand content (100%) was observed at stations 3, 5, 6 at Nobareya, El-Mex and Western
harbour, respectively in the Western region and station 12 in Rashid west (Middle region)
and station 20 at Rafah (Eastern region) and the lowest value (86.85%) was observed at
station 1 in El-Salloum (Western region).
Table 4: Calcium carbonate (CaCO3%), total organic matter (OM%), Water Content
(WC%), pH, grain size and type of sediments along the Egyptian Mediterranean Coastal
area.
Region
Western
region
Middle
region
Station
CaCO3 %
OM%
W.C.%
pH
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
34.79
94.45
95.49
85.42
67.24
95.57
91.64
93.56
22.11
14.69
15.64
3.83
2.85
5.09
6.08
4.87
4.85
5.64
7.32
6.33
95.57
2.85
37.87
0.46
0.55
1.23
0.59
0.40
0.54
1.72
0.97
0.38
1.36
0.48
0.08
0.18
0.31
0.51
0.24
0.32
0.35
0.24
0.35
1.72
0.08
0.56
24.60
19.92
18.03
24.83
17.81
7.49
35.45
19.003
18.92
17.72
19.57
19.28
19.64
19.97
22.23
24.78
21.5
23.77
19.28
15.39
35.45
7.49
20.43
8.26
8.21
8.26
8.31
7.98
7.82
7.84
8.38
8.19
7.80
8.02
7.93
7.41
7.20
7.97
7.63
7.56
7.77
7.48
7.79
8.38
7.20
7.89
Eastern
region
Max.
Min.
Average
-105-
Sand
86.85
99.99
100
99.96
100
100
98.26
99.99
99.98
99.89
99.96
100
99.93
99.93
99.67
99.57
99.72
99.76
99.60
100
100
86.85
99.15
Grain size analysis
Mud
Texture
(Silt & clay)
13.15
Very fine sand
0.01
Fine sand
0.0
Fine sand
0.04
Fine sand
0.0
Coarse sand
0.0
Very corse sand
1.74
Fine sand
0.01
Fine sand
0.02
Fine sand
0.11
Fine sand
0.04
Fine sand
0.0
Fine sand
0.07
Fine sand
0.07
Fine sand
0.33
Very Fine sand
0.43
Very Fine sand
0.28
Very Fine sand
0.24
Very Fine sand
0.40
Very Fine sand
0.0
Fine sand
13.15
0
0.85
-
% CaCO3
100
90
80
70
60
50
40
30
20
10
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
Figure 3: Distribution of CaCO3 % in surface sediments along the Egyptian
Mediterranean Coast.
% TOM
2
1.8
1.6
1.4
1.2
1
0.8
0.6
0.4
0.2
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
Figure 4: Distribution of TOM% in surface sediments along the Egyptian
Mediterranean Coast.
% Water Content
40
35
30
25
20
15
10
5
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
Figure 5: Distribution of % Water Content in surface sediments along the Egyptian
Mediterranean Coast.
-106-
5.3. Trace Metals Distribution
The study of marine sediments represents a useful tool for determining the actual state
of environmental pollution and for understanding the origin and mechanism of the
phenomena. Sediments could be regarded as a historical reflection to changes occurring in
the overlying water system and can be used as good indicators for metals in any study area
(Salomons and Forstner, 1984)(181).
5.3.1. Total Metals
The total concentrations of metals (Fe, Mn, Zn, Cu, Ni, Cr, Co, Pb and Cd) in surface
sediments of the Egyptian Mediterranean Coast are given in Table 5. The range and average
± SD concentrations (µgg-1) were 243.48–38045.05 (13255.69± 12911.36) for Fe, 17.251085.72 (380.71 ±305.38) for Mn, 2.05-62.21 (22.19 ±15.84) for Zn, 0.46-26.26 (8.46 ±6.22)
for Cu, 1.65-60.25 (25.93 ±20.96) for Ni, 4.08-297.95 (82.74 ±90.18) for Cr, 0.43-26.39 (8.24
±8.40) for Co, 3.34-53.67 (13.17 ±11.90) for Pb, 0.04-0.47 (0.22 ±0.15) for Cd.
5.3.1.1. Total Iron (TFe)
Most ferrous compounds in aquatic environments resulting in iron precipitate in
alkaline and oxidizing conditions (Abdulla et al., 1973)(430). Iron behaves similarly to
manganese in seawater, as both elements undergo oxidation. In Oxygenated seawater, Fe+2
oxidized to Fe+3, which is strongly hydrolyzed in the marine environment, forming various
(oxy) hydroxide phases of very low solubility (Millero, 1996)(431).
The results of the total iron concentration in the Egyptian Mediterranean coastal
sediments are listed in Table 5 and illustrated at Figure 6. The study area was divided into
three regions, which revealed differences in the distribution of total Fe concentration. In the
Western region, the lowest level of Fe (243.48 µgg-1) was found at station 2 in Baghoush,
while the highest value (11478.38 µgg-1) was found at station 9 in Abu Qir Bay. The
distribution of total Fe concentration in the Middle region revealed low level at station 12 in
Rashid west (9653.67 µgg-1) and high concentration (38045.05 µgg-1) at station 15 in New
Damietta. Finally, the lowest level (1422.39 µgg-1) of Fe in the Eastern region was found at
station 20 in Rafah, while the highest concentration of Fe (33971.5 µgg-1) was found at station
19 in Port Said. In general, the highest level of total iron in the surfacial sediments along the
Egyptian Mediterranean Coast was observed at station 15 in New Danietta (the Middle
region) and the lowest level was observed at station 2 in Baghoush (the Western region). The
average concentration of iron in the three regions along the Egyptian Mediterranean Coast can
be arranged as follows:
Middle region (25516.22±10090.14µgg-1) > Eastern region (17696.95±23015.7µgg-1) >
Western region (4646.03±3459.61µgg-1)
5.3.1.2. Total Manganese (TMn)
Manganese is found in minerals mostly as carbonates, oxides, silicates and sulphides.
It may occur as a particulate manganese or precipitating coating on mineral species, organic
matter, and iron manganese hydroxides (Masoud et al., 2005)(432).
The results of the analysis of total manganese concentrations in surface sediments of
Egyptian Mediterranean Coastal area, during the study area are given in Table 5 and
illustrated at Figure 6. The manganese concentration in sediment of the Western region was
fluctuated between a minimum of (17.25 µgg-1) at station 3 in El-Nobarreya and a maximum
of (526.53 µgg-1) at station 9 in Abu Qir Bay (average: 208.60±163.08 µgg-1). The lowest
concentration of Mn in the Middle region was found at station 12 in Rashid west
(257.30µgg-1) while the highest concentration (822.60µgg-1) was found at station 15 in New
-107-
Damietta. The average of Mn in the Middle region was (597.02±203.84µgg-1). The
distribution of total Mn concentration in the Eastern region of the Egyptian Mediterranean
coast revealed low value (54.76 µgg-1) at station 20 in Rafah and high level (1085.72 µgg-1) at
station 19 in Port Said. The average value of the Mn in the Eastern region was
(570.24±729.00µgg-1). The average concentration of manganese in the different regions along
the Egyptian Mediterranean Sea coast can be arranged as follows:
Middle region (597.02±203.84µgg-1) > Eastern region (570.24±729.00µgg-1) > Western
region (208.60±163.08 µgg-1).
5.3.1.3. Total Zinc (TZn)
Zinc is essential for growth of marine organisms and its concentration affected by
plankton communities (Hayward, 1969)(433). The kinetics of removal of zinc in aquatic
systems is strongly affected by the rates of biological processes. It is also a common pollutant
found in several industrial effluents such as those produced by textile, basic chemicals,
electroplating, motor and other industries (Masoud et al., 2005)(432).
The concentrations of total Zinc in the surface sediments of the Egyptian
Mediterranean coastal area are listed in Table 5 and illustrated at Figure 6. The results of Zn
concentration in the Western region ranged from 2.05 to 62.21 with an average
23.43±20.07µgg-1. The highest concentration (62.21µgg-1) was found at station 7 in NIOF
and the lowest value (2.05µgg-1) was found at station 2 in Baghoush. In the Middle region,
the minimum concentration of Zn was 10.49 µgg-1 at station 12 in Rashid west, while the
maximum concentration (29.82 µgg-1) was found at station 16 in Ras El-Barr. The average
value of total Zn in this region was 21.90±6.81µgg-1. Zinc level in the sediment of the
Eastern region was fluctuated between a minimum of 2.61 µgg-1 at station 20 in Rafah and a
maximum of 30.05 µgg-1 at station 19 in Port Said (16.33±19.40 µgg-1). The average
concentration of zinc in the different regions along the Egyptian Mediterranean coast can be
arranged as follows:
Western region (23.43±20.07µgg-1)> Middle region (21.90±6.81µgg-1) > Eastern region
(16.33±19.40 µgg-1).
5.3.1.4. Total Copper (TCu)
Copper is an essential element in the normal metabolism of plants, animals and human
life, however in high levels, cause many disease (Graten et al., 2003)(434). It is uses in
electrical equipments, alloys, antifouling paint for ships's shull, as an algaecide and wood
preservative (WHO, 1998)(435).
The concentrations of total Copper in surface sediments of the Egyptian
Mediterranean Coastal area are given in Table 5 and illustrated at Figure 7. In the Western
region, the range and average concentration of Cu was between 0.46 and 26.26
(7.93±7.36µgg-1). The highest concentration (26.26 µgg-1) was found at station 7 (NIOF) and
the lowest concentration (0.46 µgg-1) was found at station 2 in Baghoush. The minimum
concentration of Cu in the sediments of the Middle region was (3.44µgg-1) at station 12 in
Rashid West, while the maximum concentration (17.27µgg-1) was found at station 16 in Ras
El-Barr. The average value of this region was (9.86±4.47µgg-1). In the Eastern region, the
lowest level of Cu (1.57 µgg-1) was observed at station 20 in Rafah, while the highest level of
Cu (11.39µgg-1) was found at station 19 in Port Said. The average value of this region was
(6.48±6.94µgg-1). The average concentration of copper in the different regions along the
Egyptian Mediterranean Sea coast can be arranged as follows:
Middle region (9.86±4.47µgg-1)> Western region (7.93±7.36µgg-1)> Eastern region
(6.48±6.94µgg-1)
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5.3.1.5. Total Nickel (TNi)
Nickel (Ni) is one of the major elements on the earth, constituting about 2% by weight
(Nriagu,1980)(436). However, nickel is a common pollutant resulting from various industrial
activities like mining and refining, electroplating, production of Ni-Cd batteries, waste
incineration etc., domestic wastewater, and to a lesser from natural weathering (El-Nemr et
al., 2006)(437). The results of total Nickel concentration in surface sediments of the Egyptian
Mediterranean coast during this study are listed in Table 5 and illustrated at Figure 7. The
range and average concentrations of Ni was 1.65-60.25 (25.93 µgg-1) respectively. The
lowest concentration of Ni (1.65 µgg-1) was found at station 2 in Baghoush in the Western
region and the highest concentration (60.25 µgg-1) was found at station 17 in El-Gamil West
in the Eastern region.
In the Western region, the lowest value of Ni (1.65 µgg-1) was found at station 2 in
Baghoush, while the highest level (19.36 µgg-1) was observed at station 9 in Abu Qir Bay.
The average value in this region was 8.81±5.72µgg-1. The nickel level in sediment of the
Middle region was fluctuated between a minimum of (36.38 µgg-1) at station 12 in Rashid
west and a maximum of (60.25 µgg-1) at station 17 in El-Gamil West (average: 49.48±8.64
µgg-1). The lowest concentration of Ni (27.79 µgg-1) in the Eastern region was found at
station 20 in Rafah, while the highest concentration of Ni in this region (47.42µgg-1) was
observed at station 19 in Port Said; (average: 37.61±13.88 µgg-1). The average concentration
of nickel in the different regions along the Egyptian Mediterranean coast can be arranged as
follows:
Middle region (49.48±8.64 µgg-1) > Eastern region (37.61±13.88 µgg-1)> Western region
(8.81±5.72µgg-1)
5.3.1.6. Total Chromium (TCr)
Chromium is one of the biochemically active transition metals in aquatic environment
(Masoud et al., 2005)(432). The primary sources of chromium includes domestic wastewater,
manufacturing processes, involving metals and the dumping of sewage sludge (Beltagy,
1973)(438). The results of total Chromium (Cr) concentrations in surface sediments of the
Egyptian Mediterranean Coastal area are listed in Table 5 and illustrated at Figure 7. Its
concentration gave minimum and maximum values fluctuated between 4.08 µgg-1 and 297.95
µgg-1 at stations 2 (Baghoush) and 13 (Rashid East), respectively.
The lowest concentrations of Cr (4.08µgg-1) in the sediment of the Western region was
found at station 2 in Baghoush, while the highest levels of Cr (85.48 µgg-1) was observed at
station 9 in Abu Qir Bay. The average level in this region was 24.76±22.97µgg-1. In the
Middle region, the minimum concentration of Cr (79.04µgg-1) was found at station 15 in New
Damietta, while the highest level (297.95µgg-1) was found at station 13 in Rashid East. The
average of Cr in this region was 146.53±84.69µgg-1. Finally, the lowest level (72.43 µgg-1) of
Cr in the Eastern region was found at station 20 in Rafah, while the highest concentration of
Cr (284.30 µgg-1) was found at station 19 in Port Said (average: 178.37±149.81µgg-1). The
average concentration of chromium in the different regions along the Egyptian Mediterranean
coast can be arranged as follows:
Eastern region (178.37±149.81µgg-1)> Middle region (146.53±84.69µgg-1)> Western region
(24.76±22.97µgg-1).
5.3.1.7. Total Cobalt (TCo)
Cobalt is a very important element in aquatic environment. It is moderately toxic to
most aquatic species more than Cr3+ and Cr6+ but much less toxic than Cd2+ and Cu2+. Higher
levels of cobalt have been found in the open ocean, but these are probably due to
-109-
contamination with phytoplankton and other organisms that metabolize vitamin B12 (Masoud
et al., 2005)(432).
The results of total Cobalt in the surface sediments of the Mediterranean Coast of
Egypt are listed in Table 5 and illustrated in Figure 8. The range and average concentrations
of Co were 0.43-26.39 µgg-1 (8.24 µgg-1) respectively. The highest concentration of Co 26.39
µgg-1 was found at station 16 in Ras El-Burr and the lowest concentration (0.43 µgg-1) was
observed at station 2 in Baghoush.
In the Western region, the lowest conent of Co (0.43 µgg-1) was found at station 2 in
Baghoush, while the highest level (6.63 µgg-1) was observed at station 11 in Maadia. The
average value of total Co in this region was 2.46±1.89 µgg-1. The minimum concentration of
Co in the Middle region (7.94µgg-1) was observed at station 12 in Rashid west and the
maximum level of Co (26.39µgg-1) was observed at station 16 in Ras El-Barr with an average
of 16.96±6.91µgg-1. The concentration of cobalt in the sediments of the Eastern region was
fluctuated between a minimum of 1.93 µgg-1 at station 20 in Rafah and a maximum of 17.07
µgg-1 at station 19 in Port Said (average: 9.50±10.71 µgg-1). The average concentration of
cobalt in the different regions along the Egyptian Mediterranean coast can be arranged as
follows:
Middle region (16.96±6.91µgg-1)> Eastern region (9.50±10.71 µgg-1)> Western region
(2.46±1.89 µgg-1).
5.3.1.8. Total Lead (TPb)
Lead in sediments is chemically precipitated from the surface water solution, and the
remainder has been transported in detrital particles. However, lead occurring in sediments
has two distinct mineral associations, one with the clay minerals and the other with authigenic
minerals and/or biogenous debris (Masoud et al., 2005)(432).
The concentrations of total Lead in surface sediments of the Egyptian Mediterranean
Coastal area are listed in Table 5 and illustrated at Figure 8. The concentration of Pb ranged
from 3.34 to 53.67 with and average value of 13.17±11.90µgg-1. The highest concentration of
Pb (53.67µgg-1) was observed at station 5 in El-Mex Bay, while the lowest level of Pb (3.34
µgg-1) was observed at station 20 in Rafah.
The lowest concentration of total Pb (4.41 µgg-1) in the Western region was found at
station 2 in Baghoush, while the highest level (53.67µgg-1) was observed at station 5 in ElMex Bay. The average value of total Pb in this region was 18.93±13.73 µgg-1. In the Middle
region, the lowest level of total Pb (4.61µgg-1) was observed at station 12 in Rashid West,
while the highest level (8.83µgg-1) was observed at station 15 in New Damietta. The average
of total Pb in this region was 6.70±1.37 µgg-1. The distribution of total Pb concentration in
the Eastern region revealed low value (3.34µgg-1) at station 20 in Rafah and high value
(6.67µgg-1) at station 19 in Port Said. The average value of Pb in this region was
5.01±2.35µgg-1. The average concentration of Pb in the different regions can be arranged as
follows:
Western region (18.93±13.73)> Middle region (6.70±1.37µgg-1)> Eastern region
(5.01±2.35µgg-1)
5.3.1.9. Total Cadmium (TCd)
The major specific sources on the worldwide basis are atmospheric deposition,
smelting and refining of nonferrous metals, manufacturing processes related to chemicals and
-110-
metals, and domestic waste water (Huzinger, 1980)(439). The concentrations of total cadmium
(Cd) in surface sediments of the Egyptian Mediterranean Coastal area are listed in Table 5 and
illustrated in Figure 8. The range and average concentrations of Cd in the sediments of the
Egyptian Mediterranean coast were 0.04-0.47 (0.22µgg-1). The data revealed that all stations
in the study area showed relatively low concentrations of cadmium.
The distribution of total cadmium levels in the western region revealed low
concentration (0.06µgg-1) at station 3 in El-Nobarreya and high concentration (0.47µgg-1) at
station 6 in the Western Harbour. The average value of Cd in this region was 0.29±0.15µgg-1.
In the Middle region, the lowest concentration of Cd (0.04µgg-1) was found at station 15 in
New Damietta, while the highest level (0.23µgg-1) was found at station 14 in Burullus. The
average value of Cd in the Middle region was 0.12±0.07µgg-1. The lowest concentration of
Cd (0.04µgg-1) in the Eastern region was found at station 19 in Port Said, while the highest
concentration (0.18µgg-1) was observed at station 20 in Rafah. The average value of Cd in
this region was 0.11±0.10µgg-1. The average concentration of Cd in the three regions can be
arranged as follows:
Western region (0.29±0.15µgg-1) > Middle region (0.12±0.07µgg-1) > Eastern region
(0.11±0.10µgg-1)
-111-
Table 5: Total trace metals concentration (µgg-1 dry weight) in surfacial sediments along
the Egyptian Mediterranean Coastal area.
Region
Western
Region
Middle
Region
Eastern
Region
All
Regions
Station
1
2
3
4
5
6
7
8
9
10
11
Min.
Max.
Average
SD
12
13
14
15
16
17
18
Min.
Max.
Average
SD
19
20
Min.
Max.
Average
SD
Min.
Max.
Average
SD
Fe
Mn
7166.6
116.31
243.48
26.13
277.34
17.25
3946.02 175.32
3565.22 104.33
1527.85 230.61
4813.8
178.81
3440.82
136.3
11478.38 526.53
6620.29 413.32
8026.49 369.73
243.48
17.25
11478.38 526.53
4646.03 208.60
3459.61 163.08
9653.67
257.3
27044.33 818.74
25164.02 538.97
38045.05 822.6
37398.1 721.08
18271.33 508.01
23037.04 512.44
9653.67 257.30
38045.05 822.60
25516.22 597.02
10090.14 203.84
33971.5 1085.72
1422.39
54.76
1422.39
54.76
33971.50 1085.72
17696.95 570.24
23015.70 729.00
243.48
17.25
38045.05 1085.72
13255.69 380.71
12911.36 305.38
Zn
38.66
2.05
9.49
22.93
55.43
6.28
62.21
20.02
15.99
10.05
14.64
2.05
62.21
23.43
20.07
10.49
24.78
17.00
26.33
29.82
18.4
26.51
10.49
29.82
21.90
6.81
30.05
2.61
2.61
30.05
16.33
19.40
2.05
62.21
22.19
15.84
Cu
13.87
0.46
0.86
9.69
7.34
4.13
26.26
10.94
5.7
2.99
4.95
0.46
26.26
7.93
7.36
3.44
7.06
9.92
12
17.27
7.3
12.04
3.44
17.27
9.86
4.47
11.39
1.57
1.57
11.39
6.48
6.94
0.46
26.26
8.46
6.22
-112-
Ni
7.58
1.65
9.5
8.74
5.71
3.68
8.71
2.71
19.36
17.89
11.43
1.65
19.36
8.81
5.72
36.38
56.54
43.55
44.31
56.41
60.25
48.93
36.38
60.25
49.48
8.64
47.42
27.79
27.79
47.42
37.61
13.88
1.65
60.25
25.93
20.96
Cr
31.56
4.08
5.26
16.37
23.19
7.08
25.41
9.45
85.48
33.85
30.59
4.08
85.48
24.76
22.97
84.01
297.95
116.93
79.04
102.18
233.7
111.93
79.04
297.95
146.53
84.69
284.3
72.43
72.43
284.30
178.37
149.81
4.08
297.95
82.74
90.18
Co
2.03
0.43
1.35
1.73
1.34
1.38
1.87
1.35
4.65
4.3
6.63
0.43
6.63
2.46
1.89
7.94
14.86
12.32
24.91
26.39
12.53
19.75
7.94
26.39
16.96
6.91
17.07
1.93
1.93
17.07
9.50
10.71
0.43
26.39
8.24
8.40
Pb
16.36
4.41
14.68
15.94
53.67
20.32
29.91
24.46
9.46
12.06
6.92
4.41
53.67
18.93
13.73
4.61
6.41
5.8
8.83
6.96
6.45
7.86
4.61
8.83
6.70
1.37
6.67
3.34
3.34
6.67
5.01
2.35
3.34
53.67
13.17
11.90
Cd
0.29
0.07
0.06
0.34
0.16
0.47
0.44
0.38
0.19
0.39
0.45
0.06
0.47
0.29
0.15
0.14
0.09
0.23
0.04
0.05
0.15
0.13
0.04
0.23
0.12
0.07
0.04
0.18
0.04
0.18
0.11
0.10
0.04
0.47
0.22
0.15
40000
35000
Fe (µgg-1)
30000
25000
20000
15000
10000
5000
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
1200
1000
Mn (µgg-1)
800
600
400
200
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
80
70
Zn (µgg-1)
60
50
40
30
20
10
0
1 2
3 4 5
6 7 8
9 10 11 12 13 14 15 16 17 18 19 20
Stations
Figure 6: Distribution of Fe, Mn and Zn in surface sediments collected from the
Egyptian Mediterranean coastal area
-113-
40
35
Cu (µgg-1)
30
25
20
15
10
5
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
70
Ni (µgg-1)
60
50
40
30
20
10
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
300
Cr (µgg-1)
250
200
150
100
50
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
Stations
Figure 7: Distribution of Cu, Ni, and Cr in surface sediments collected from the
Egyptian Mediterranean coastal area
-114-
30
Co (µgg-1)
25
20
15
10
5
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
60
Pb (µgg-1)
50
40
30
20
10
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
1
Cd (µgg-1)
0.8
0.6
0.4
0.2
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
Stations
Figure 8: Distribution of Co, Pb and Cd in surface sediments collected from the
Egyptian Mediterranean coastal area
-115-
5.3.2. Leachable Metals
The concentrations of leachable metals (Fe, Mn, Zn, Cu, Ni, Cr, Co, Pb and Cd) in
surface sediments of the Egyptian Mediterranean Coastal area during July 2010 are given in
Table 6 and Figure 9, 10 and 11. The range and average concentrations (µgg-1) were: 94.986831.61 (1930.77±1777.05) for Fe, 10.83-307.27 (126.74±81.82) for Mn, 0.11-49.48
(7.40±10.94) for Zn, 0.23-8.20 (2.19±2.29) for Cu, 0.15-11.16 (3.04±3.23) for Ni, 0.56-8.22
(2.61±2.01) for Cr, 0.10-8.77 (2.35±2.82) for Co, 0.54-51.35 (8.14±11.97) for Pb, 0.02-0.23
(0.06±0.05) for Cd.
5.3.2.1. Leachable Iron (LFe)
Table 6 and Figure 9 show the results of leachable Iron (LFe) in the surface sediments
of the coastal area of the Egyptian Mediterranean Sea. The highest concentration of LFe
(6831.61 µgg-1) was found at station 15 in New Damietta and the lowest concentration of LFe
(94.98 µgg-1) was found at station 20 in Rafah.
The study area was divided into three regions, which revealed differences in the
distribution of leachable iron. The lowest level of LFe (106.63µgg-1) in the sediments of the
Western region was observed at station 3 in Nobarreya, while the highest value
(2100.42µgg-1) was observed at station 11 in Maadia. The average LFe concentration was
1023.40±614.80µgg-1. In the Middle region, the lowest concentration of LFe (1494.21µgg-1)
was observed at station 12 in Rashid west, while the highest value (6831.61µgg-1) was
observed at station 15 in New Damietta.
The average LFe level was
-1
-1
3370.81µgg ±2092.12µgg . The distribution of LFe concentration in the Eastern region
revealed low level (94.98µgg-1) at station 20 in Rafah and the high concentration was found at
station 19 in Port Said (3667.31µgg-1). The average LFe level in this region was
1881.14±2526.02µgg-1.
5.3.2.2. Leachable Manganese (LMn)
The results of the analysis of leachabel Manganese (LMn) in surface sediments at all
stations are represented in Table 6 and Figure 9. Data showed that the highest concentration
of LMn (307.27µgg-1) in the surfacial sediments of the Egyptian Mediterranean coast was
found at station 15 in New Damietta in the Middle region and the lower concentration of LMn
(10.83µgg-1) was found at station 3 in Nobarreya in the Western region.
In the Western region, the lowest concentration of LMn (10.83µgg-1) was observed at
station 3 (Nobarreya), while the highest concentration (261.85µgg-1) was observed at station 9
(Abu Qir Bay) with an average (113.86±76.59µgg-1). The lowest level of LMn (29.99µgg-1)
in the Middle region was observed at station 13 in Rashid east, while the highest level
(307.27µgg-1) was observed at station 15 in New Damietta. The average of LMn in the
Middle region was 154.39±91.30µgg-1. The distribution of LMn concentration in the Eastern
region revealed low value at Rafah (Station 20; 30.20µgg-1), and the high value at Port Said
(Station 19; 171.53µgg-1). The average measured concentration of LMn in this region was
100.87±99.94µgg-1.
5.3.2.3. Leachable Zinc (LZn)
The concentrations of leachable Zinc (LZn) in surface sediments at all stations of the
Egyptian Mediterranean coastal area are presented in Table 6 and Figure 9. Data showed that
the highest concentration of LZn in the western region (49.48 µgg-1) was observed at station 5
in El-Mex Bay and the lowest concentration of LZn (0.11 µgg-1) was observed at station 3 in
Nobarreya (average: 10.89±13.92µgg-1). In the Middle region, the minimum concentration of
LZn was (1.49 µgg-1) at station 17 in El-Gamil west and the maximum concentration of LZn
-116-
was (6.54µgg-1) at station 15 in New Damietta, with an average value of (3.51±2.15µgg-1).
The distribution of LZn concentration in the eastern region revealed that the lowest
concentration was observed at Rafah (Station 20; 0.15µgg-1) and the highest concentration
was observed at Port Said (Station 19; 3.63µgg-1). The average of LZn was 1.89±2.46µgg-1.
5.3.2.4. Leachable Cupper (LCu)
The results of the analysis of leachable Cupper (LCu) in surface sediments of the
Egyptian Mediterranean coastal area are shown in Table 6 and Figure 10. In the Western
region, the highest concentration of LCu (8.20 µgg-1) was found at station 7 in NIOF and the
lowest concentration of LCu (0.23 µgg-1) was found at station 2 in Baghoush; the average
value of LCu in this region was (2.98±2.85µgg-1). The minimum concentration of LCu in the
sediments of the Middle region of the Egyptian Mediterranean coast (0.56µgg-1) was found at
station 12 in Rashid west, while the maximum concentration (2.08µgg-1) was found at station
15 in New Damietta. The average value of this region was (1.29±0.57µgg-1). In the Eastern
region, the lowest value of LCu (0.27µgg-1) was observed at station 20 in Rafah and the
highest value (1.79µgg-1) was observed at station 19 in Port Said. The average value of LCu
in the Middle region was 1.03±1.07µgg-1.
5.3.2.5. Leachable Nickel (LNi)
Table 6 and Figure 10 showed the results of Leachable Nickel (LNi) in the surfacial
sediments of the Egyptian Mediterranean coast. Stations 15 and 16 in New Damietta and Ras
El-Barr in the Middle region considered having the highest concentrations of LNi (11.16 and
10.00 µgg-1), respectively. The leachable nickel concentration in sediment of the Western
region was fluctuated between a minimum of 0.27 µgg-1 at station 3 in El-Nobarreya and a
maximum of 3.29 µgg-1 at station 11 in Maadia (average: 1.17±0.85 µgg-1). In the Middle
region, the lowest value of LNi (2.70µgg-1) was observed at station 12 in Rashid west, while
the highest level (11.16µgg-1) was observed at station 15 in New Damietta, the average value
was 5.88±3.53µgg-1. The distribution of LNi concentration in the Eastern region revealed low
concentration (0.15µgg-1) at station 20 in Rafah and high concentration (6.72µgg-1) at station
19 in Port Said, the average value of LNi was 3.44±4.64µgg-1.
5.3.2.6. Leachable Chromium (LCr)
The concentrations of leachable Chromium (LCr) in surface sediments at all stations
are shown in Table 6 and Figure 10. The highest concentration of LCr (8.22 µgg-1) was
observed at station 5 in El-Mex Bay in the Western region and the lowest concentration of
LCr (0.56µgg-1) was found at station 12 in Rashid west at the Middle region of the study area.
The obtained results showed relatively low concentrations of LCr.
The lowest concentrations of LCr (1.26µgg-1) in the sediment of the western region
was observed at station 3 in Nobarreya, while the highest levels of LCr (8.22µgg-1) was
observed at station 5 in El-Mex Bay, the average level was 3.30±2.38µgg-1. In the midle
region, the minimum concentration of LCr (0.56µgg-1) was found at station 12 in Rashid west,
while the maximum level (2.97µgg-1) was found at station 16 in Ras El-Barr, the average of
LCr was 1.77±1.05µgg-1. The distribution of LCr concentration in the Eastern region
revealed low concentration (0.75µgg-1) at station 20 in Rafah and high concentration
(2.73µgg-1) at station 19 in Port Said, the average value of LCr was 1.74±1.40µgg-1.
5.3.2.7. Leachable Cobalt (LCo)
The concentrations of leachable Cobalt (LCo) in surface sediments are shown in Table
6 and Figure 11. The highest concentration of LCo (8.77µgg-1) was found at station 15 in
-117-
New Damietta at the Middle region, while the lowest concentration of LCo (0.10µgg-1) was
found at the western region at station 2 in Baghoush.
In the western region, the lowest level of LCo (0.10µgg-1) was found at station 2 in
Baghoush, while the highest level (2.85µgg-1) was observed at station 11 in Maadia, the
average value of LCo was 0.67±0.81µgg-1. The minimum concentration of LCo in the middle
region was (1.88µgg-1) at station 12 in Rashid west, while the maximum concentration was
(8.77µgg-1) in station 15 at New Damietta, the average level of LCo was 4.49±2.81µgg-1. In
the eastern region, the lowest concentration of LCo (0.34µgg-1) was found at station 20 in
Rafah and the highest concentration of LCo (7.84µgg-1) was observed at station 19 in Port
Said, the average value of LCo was (4.09±5.30µgg-1).
5.3.2.8. Leachable Lead (LPb)
The concentrations of leachable Lead (LPb) at all stations of the present study are
shown in Table 6 and Figure 11. The highest concentration of LPb (51.35µgg-1) was found at
the western region (station 5) in El-Mex Bay, while the lowest concentration of LPb (0.54µgg1
) was found at the Eastern region (station 20) in Rafah.
The lowest concentration of LPb (1.78µgg-1) in the Western region was observed at
station 11 in Maadia, while the highest level (51.35µgg-1) was observed at station 5 in El-Mex
Bay, the average of LPb was 13.36±14.30µgg-1. In the Middle region, the lowest
concentration of LPb was (0.77µgg-1) at station 17 in El-Gamil west, while the highest level
of LPb was (3.98µgg-1) at station 15 in New Damietta. The average level of LPb was
1.96±1.34µgg-1. The distribution of LPb concentration in the Eastern region revealed low
value of (0.54µgg-1) at station 20 in Rafah and high value of (1.56µgg-1) at station 20 in Port
Said, the average value of LPb was 1.05±0.72µgg-1.
5.3.2.9. Leachable Cadmium (LCd)
The concentrations of leachabel Cadmium (LCd) in surface sediments at all stations
are shown in Table 6 and Figure 11. The highest concentration of LCd (0.23 µgg-1) was
found at station 1 in El-Salloum in the Western region, while the lower concentration of LCd
(0.02µgg-1) was found at station 15 in New Damietta at the Middle region.
The LCd level in sediment of the Western region was fluctuated between a minimum
of 0.03µgg-1 at station 2 and 3 (Bahgoush and Nobarreya), respectively and a maximum of
(0.23µgg-1) at station 1 in El-Salloum (average: 0.08±0.05µgg-1). In the Middle region, the
minimum concentration of LCd (0.02 µgg-1) found at station 15 in New Damietta and the
maximum level of LPb (0.07µgg-1) was observed at station 17 in El-Gamil west. The average
concentration of LPb in this region (Middle) was 0.04±0.02µgg-1. The distribution of LPb
level in the Eastern region revealed low value (0.02µgg-1) at station 20 in Rafah and high
concentration (0.03µgg-1) at station 19 in Port Said.
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Table 6: Leachable trace metals concentration (µgg-1 dry weight) of surfacial sediments
along the Egyptian Mediterranean coastal area.
Region
Western
Region
Middle
Region
Eastern
Region
All
Regions
Station
1
2
3
4
5
6
7
8
9
10
11
Min.
Max.
Average
SD
12
13
14
15
16
17
18
Min.
Max.
Average
SD
19
20
Min.
Max.
Average
SD
Min.
Max.
Average
SD
Fe
1059.41
169.11
106.63
1565.96
946.81
626.00
1568.98
1417.76
1042.36
653.88
2100.42
106.63
2100.42
1023.40
614.80
1494.21
2026.41
2225.87
6831.61
5481.41
1658.24
3877.94
1494.21
6831.61
3370.81
2092.12
3667.31
94.98
94.98
3667.31
1881.14
2526.02
94.98
6831.61
1930.77
1777.05
Mn
55.11
19.59
10.83
149.06
95.63
162.54
94.43
114.94
261.85
82.41
206.08
10.83
261.85
113.86
76.59
118.89
29.99
110.42
307.27
239.82
130.70
143.61
29.99
307.27
154.39
91.30
171.53
30.20
30.20
171.53
100.87
99.94
10.83
307.27
126.75
81.82
Zn
5.23
0.52
0.11
11.44
49.48
4.27
17.42
14.72
4.01
5.32
7.31
0.11
49.48
10.89
13.92
1.66
2.17
2.61
6.54
6.40
1.49
3.69
1.49
6.54
3.51
2.15
3.63
0.15
0.15
3.63
1.89
2.46
0.11
49.48
7.41
10.94
Cu
1.96
0.23
0.32
4.53
6.34
1.28
8.20
6.21
0.51
2.29
0.86
0.23
8.20
2.98
2.85
0.56
1.01
1.05
2.08
1.93
0.88
1.51
0.56
2.08
1.29
0.57
1.79
0.27
0.27
1.79
1.03
1.07
0.23
8.20
2.19
2.29
-119-
Ni
0.74
0.34
0.27
1.33
1.23
0.79
1.74
0.48
1.54
1.09
3.29
0.27
3.29
1.17
0.85
2.70
3.89
3.77
11.16
10.00
2.70
6.90
2.70
11.16
5.88
3.53
6.72
0.15
0.15
6.72
3.44
4.64
0.15
11.16
3.04
3.23
Cr
2.67
2.51
1.26
3.17
8.22
1.79
7.40
4.04
1.35
1.61
2.30
1.26
8.22
3.30
2.38
0.56
2.73
0.78
2.81
2.97
0.98
1.58
0.56
2.97
1.77
1.05
2.73
0.75
0.75
2.73
1.74
1.40
0.56
8.22
2.61
2.01
Co
0.30
0.10
0.29
0.57
0.42
0.18
0.61
0.36
1.42
0.23
2.85
0.10
2.85
0.67
0.81
1.88
2.52
3.06
8.77
7.89
2.38
4.88
1.88
8.77
4.49
2.81
7.84
0.34
0.34
7.84
4.09
5.30
0.10
8.77
2.35
2.82
Pb
7.74
3.15
5.40
10.39
51.35
12.50
19.96
23.10
4.28
7.27
1.78
1.78
51.35
13.36
14.30
1.33
0.94
1.28
3.98
1.66
0.77
3.77
0.77
3.98
1.96
1.34
1.56
0.54
0.54
1.56
1.05
0.72
0.54
51.35
8.14
11.97
Cd
0.23
0.03
0.03
0.07
0.06
0.10
0.09
0.10
0.06
0.07
0.05
0.03
0.23
0.08
0.05
0.06
0.03
0.03
0.02
0.04
0.07
0.04
0.02
0.07
0.04
0.02
0.03
0.02
0.02
0.03
0.03
0.00
0.02
0.23
0.06
0.05
7500
6500
LFe (µgg-1)
5500
4500
3500
2500
1500
500
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
LMn (µgg-1)
250
200
150
100
50
0
1
2
3
4
5
6
7
8
9 10 11 12 13 14 15 16 17 18 19 20
5 6
7
8 9 10 11 12 13 14 15 16 17 18 19 20
50
LZn (µgg-1)
40
30
20
10
0
1
2
3 4
Stations
Figure 9: Concentration of LFe, LMn and LZn in surfacial sediments along the
Egyptian Mediterranean coastal area.
-120-
10
LCu (µgg-1)
8
6
4
2
0
LNi (µgg-1)
1
2
3
4
5
6
7
8
9 10 11 12 13 14 15 16 17 18 19 20
14
12
10
8
6
4
2
0
1
2 3 4
5 6
7 8 9 10 11 12 13 14 15 16 17 18 19 20
10
LCr (µgg-1)
8
6
4
2
0
1
2
3
4
5 6
7
8
9 10 11 12 13 14 15 16 17 18 19 20
Stations
Figure 10: Concentration of LCu, LNi, LCr in surfacial sediments along the Egyptian
Mediterranean coastal area.
-121-
10
LCo (µgg-1)
8
6
4
2
0
1
2
3
4
5 6
7
8
9 10 11 12 13 14 15 16 17 18 19 20
1
2
3
4
5
7
8
9 10 11 12 13 14 15 16 17 18 19 20
60
50
LPb (µgg-1)
40
30
20
10
0
6
0.3
0.25
LCd (µgg-1)
0.2
0.15
0.1
0.05
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
Stations
Figure 11: Concentration of LCo, LPb, LCd in surfacial sediments along the Egyptian
Mediterranean coastal area.
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5.3.3. Partitioning of Metals
It is now widely accepted that the role of aquatic sediments as a sink or as a source of
pollutants cannot fully be assessed by measuring total metal concentrations. In addition,
determination of total elements does not give an accurate estimate of the likely environmental
impact. Instead, it is desirable to have information on the potential availability of metals
(whether toxic or essential) to biota under various enviornmental conditions (Davidson et al.,
1994)(440). The total content of a given metal is not informative as it does not describe the
amount of the metal available for living organisms and thus may overestimate the real threat.
The most important from the ecological point of view is to know which part of the total
concentration of a given metal is available to living organisms as it determines the real toxic
effect of this metal (Bernhard and Neff, 2001)(441). This need demanded development of new
analytical methods of bottom sediment study based on fractionation (Glosinska et al.,
2005)(442). Speciation is very useful, not only for determining the degree of association of the
metals to the sediments but also for differentiating metals with a lithogenic origin from those
with an anthropogenic origin (Izquierdo et al., 1997)(443). In the present study, it has been
investigated the partitioning of six metals (Fe, Mn, Zn, Ni, Cr, Pb) among different fractions
in the sediments of the Egyptian Mediterranean Sea.
5.3.3.1. Fractionation of Iron (Fe)
The results of iron concentrations of the different chemical phases: exchangeable (F1),
bound to carbonate (F2), bound to Fe/Mn oxides (F3), bound to organic matter (F4) and
residual fraction (F5) in the Egyptian Mediterranean coast are listed in Table 7 and the
relative percent contribution from the total concentration are shown in Figure 12. The iron
concentrations associated with the exchangeable fraction ranged between (0.11µgg-1) at
stations 3, 15 in Nobarreya and New Damietta, respectively and (1.02µgg-1) at station 13 in
Rashid east with an average (0.29±0.22µgg-1). The regional distribution of Fe concentration
in the three zones ranged from a minimum (0.20 µgg-1) at the western region to a maximum
(0.43 µgg-1) at the middle region (Table 7 and Figure 14).
The amount of iron associated with the carbonate fraction varied from 0.54µgg-1 at
station 7 in NIOF and 5.30 µgg-1 in station 10 at electrical power station with an average
(2.194±1.34µgg-1) (Table 7). The regional distribution of iron level in the three regions along
the coastal area ranged between a minimum (2.00µgg-1) at the western region to a maximum
(3.39µgg-1) at the eastern region (Table 7 and Figure 14).
The amount of Fe associated with the Fe-Mn oxides fraction varies from 64.59
-1
µgg at station 20 (Rafah) in the eastern region to 1523.75µgg-1 at station 15 (New Damietta)
in the middle region with an average (625.60±430.11µgg-1) (Table 7). The regional
distribution of Fe level in the three regions ranged from 475.78µgg-1 at the eastern region to
825.74µgg-1 at the middle region (Table 7 and Figure 14).
The amount of Fe associated with the organic matter fraction varied from 15.82µgg-1
at station 2 (Baghoush) in the western region and 2655.46µgg-1 at station 16 in Ras El-Barr in
the middle region with an average (712.73±664.20 µgg-1) (Table 7). The regional distribution
of Fe level in the three regions ranged from 357.10µgg-1 at the western region and
1294.69µgg-1 at the middle region (Table 7 and Figure 14).
The amount of Fe associated with the residual fraction varied from 156.96 µgg-1 at
station 2 (Baghoush) at the western region and 31423.59 µgg-1 in station 16 (Ras El-Barr) in
the middle region with an average (10388.84±9815.05 µgg-1) (Table 7). The regional
-123-
distribution of Fe concentration in the three regions ranged from 3892.37µgg-1 at the western
region and 19967.46µgg-1 at the middle region (Table 7 and Figure 14).
5.3.3.2. Fractionation of Manganese (Mn)
The results of Mn concentration of the different chemical phases: exchangeable (F1),
bound to carbonate (F2), bound to Fe/Mn oxides (F3), bound to organic matter (F4) and
residual fraction (F5) in the surface sediments of the study area from El-Salloum to Rafah are
listed in Table 8 and the relative percent contribution from the total concentration are shown
in Figure 12. Data listed in Table 8 showed that the amount of Mn associated with the
exchangeable fraction varied from ND at station 3 in El-Nobarreya (Western region) to 1.16
µgg-1 at station 18 in El-Gamil east (Eastern region) with an average (0.39 ±0.40 µgg-1). The
regional distribution of Mn level in the three regions ranged from 0.15µgg-1 at the Western
region to 0.73µgg-1 at the Middle region (Table 8 and Figure 15).
Table 8 showed that wide range in the distribution of Mn associated with the
carbonate fraction, varied from 0.71 µgg-1 at station 3 in El-Nobarreya (Western region) to
101.38 µgg-1 at station 19 in Port Said (Eastern region) with an average (30.28±32.17 µgg-1).
The regional distribution of Mn concentrations in the three regions ranged from 17.46µgg-1 at
the Western region to 54.24µgg-1 at the Eastern region (Table 8 and Figure 15).
From the obtained data at Table 8, it is observed that the amount of Mn associated
with the Fe-Mn fraction varied from 11.43 µgg-1 at station 20 (Rafah; Eastern region) to
186.20 µgg-1 at station 11 (Maadia; Western region) with an average (82.43 ±57.81µgg-1).
The regional distribution of Mn concentrations in the three regions ranged from 30.96µgg-1 at
the Eastern region to 95.04µgg-1 at the Western region (Table 8 and Figure 15).
Table 8 represented that the amount of Mn associated with organic fraction varied
from 0.49µgg-1 at station 20 in Rafah (Eastern region) to 32.10 µgg-1at station 15 (New
Danietta; middle region) with an average (10.45 ± 9.35µgg-1). The regional distribution of
Mn concentrations in the three zones ranged from 6.46µgg-1 at the Western region to
16.00µgg-1 at the Middle region (Table 8 and Figure 15).
Data listed in Table 8 showed that the amount of Mn associated with the residual
fraction varied from 0.46µgg-1 at station 3 in El-Nobarreya to 599.23µgg-1 at station 19 in Port
Said (Eastern region) with an average (178.50±193.00µgg-1). The regional distribution of Mn
concentrations in the three regions ranged from 53.53µgg-1 at the Western region to
337.32µgg-1 at the Middle region (Table 8 and Figure 15).
5.3.3.3. Fractionation of Zinc (Zn)
The results of Zinc concentrations of the different chemical phases: exchangeable
(F1), bound to carbonate (F2), bound to Fe/Mn oxides (F3), bound to organic matter (F4) and
residual fraction (F5) in surface sediments of the study area from El-Salloum to Rafah are
listed in Table 9 and the relative percent contribution from the total concentration are shown
in Figure 12. Table 9 represented that the Zn concentrations associated with the exchangeable
fraction ranged from ND at station 16 in Ras El-Barr (Middle region) to 0.25µgg-1 at station 1
in El-Salloum (Western region) with an average 0.04±0.05µgg-1. The regional distribution of
Zn levels in the three regions ranged from 0.01µgg-1 at the Middle region to 0.06µgg-1 at the
Western region (Table 9 and Figure 16).
The amount of Zn associated with the carbonate fraction varied from 0.04µgg-1 at
station 18 in El-Gamil East (Middle region) to 2.49µgg-1 at station 5 in El-Mex Bay (Western
region) with an average 0.59±0.76µgg-1 (Table 9). The regional distribution of Zn levels in
-124-
the three zones ranged from 0.06µgg-1 at the Middle region to 1.02µgg-1 at the Western region
(Table 9 and Figure 16).
From the obtained data at Table 9, it is observed that the amount of Zn associated with
the Fe-Mn fraction varied from 0.08 µgg-1 at station 3 in El-Nobarreya (Western region) to
11.01 µgg-1 at station 7 in NIOF (Western region) with an average 3.14±3.48µgg-1. The
regional distribution of Zn levels in the three regions ranged from 0.62µgg-1 at the Eastern
region to 4.93µgg-1 at the Western region (Table 9 and Figure 16).
Table 9 represented that the amount of Zn associated with organic matter fraction
varied from ND at station 2, 3, 20 in Baghoush, El-Nobarreya and Rafah, respectively to
4.94µgg-1 at station 7 in NIOF (Western region) with an average 1.21±1.37µgg-1. The
regional distribution of Zn concentrations in the three regions ranged from 0.51µgg-1 at the
Eastern region to 1.42µgg-1 at the Western region (Table 9 and Figure 16).
Data listed in Table 9 showed that the amount of Zn associated with the residual
fraction varied from 0.70µgg-1 at station 6 in Western harbour to 35.77µgg-1 at station 7 in
NIOF (Western region) with an average 12.49±9.25µgg-1. The regional distribution of Zn
concentrations in the three regions ranged from 10.62µgg-1 at the Eastern region to 14.68µgg-1
at the Middle region (Table 9 and Figure 16).
5.3.3.4 Fractionation of Nickel (Ni)
The results of application of sequential extraction of Ni in five different chemical
phases: exchangeable (F1), bound to carbonate (F2), bound to Fe/Mn oxides (F3), bound to
organic matter (F4) and residual fraction (F5) in surface sediments of the study area from ElSalloum to Rafah are presented in Table 10 and the relative percent contribution from the
total concentration are shown in Figure 13. The levels of Ni concentrations associated with
the exchangeable fration of surface sediments of the Egyptian Mediterranean sea are
fluctuated between a minimum of 0.02µgg-1 at station 20 in Rafah (Western region) and a
maximum of 0.26µgg-1 at station 16 in Ras El-Barr (Middle region) with an average
0.13±0.06µgg-1 (Table 10). Table 10 and Figure 17 showed the nickel concentration in the
three zones, which was 0.11µgg-1 at the eastern region to 0.14µgg-1 at the Middle region.
The amount of nickel associated with the carbonate fraction varied from 0.011µgg-1 at
station 11 (Maadia; Western region) and station 15 and 17 (New Damietta and El-Gamil west
respectively; Middle region) to 0.12 µgg-1 at station 4 and 5 in Dekhaila and El-Mex Bay,
respectively with an average 0.06±0.04µgg-1 (Table 10). Table 10 and Figure 17 showed the
average of Ni concentration from the total in the three regions, it ranged between 0.04µgg-1 at
the Eastern region to 0.07µgg-1 at the western region.
Nickel associated with Fe-Mn oxides fraction varied from 0.20µgg-1 at station 2 in
Baghoush (Western region) to 6.05µgg-1 at station 15 in New Damietta (Middle region) with
an average (1.68±1.76µgg-1) (Table 10). Table 10 and Figure 17 showed the average of Ni
concentration from the total in the three regions from the Egyptian Mediterranean coast, it
ranged from 0.59µgg-1 (Western region) to 3.45µgg-1 (Middle region).
Data listed in Table 10 showed that the amount of Ni associated with the organic
matter fraction varied from 0.05µgg-1 at station 20 in Rafah (Eastern region) to 5.70µgg-1 at
station 16 in Ras El-Barr (Middle region) with an average 1.43±1.74µgg-1. Table 10 and
Figure 17 reperesented Ni concentration associated with the organic matter for the three
-125-
regions. These amounts represented portions varying from 0.45µgg-1 (Western region) to
3.00µgg-1 (Middle region) from the total.
The amount of Ni associated with the residual fraction varied from 0.88 µgg-1 at
station 2 in Baghoush (Western region) to 42.67 µgg-1 at station 17 in El-Gamil west (Middle
region) with an average (16.51±13.62µgg-1). The regional distribution of Ni levels in the
three regions ranged from 5.64µgg-1 at the Western region to 30.85µgg-1 at the Middle region
(Table 10 and Figure 17).
5.3.3.5. Fractionation of Chromium (Cr)
The results of application of sequential extraction of Cr in five different chemical
phases: exchangeable (F1), bound to carbonate (F2), bound to Fe/Mn oxides (F3), bound to
organic matter (F4) and residual fraction (F5) in surface sediments of the Egyptian
Mediterranean coastal area from El-Salloum to Rafah are listed in Table 11 and the relative
percent contribution from the total concentration are shown in Figure 13.
Table 11 represented that the chromium concentrations associated with the
exchangeable fraction ranged between 0.02µgg-1 at station 18 in El-Gamil east (Middle
region) and station 19 in Port Said (Eastern region) to 0.12µgg-1 at station 1 in Salloum
(Western region) with an average 0.04±0.03µgg-1. The average concentration of Cr
associated with the exchangeable phase in the three regions (Table 11 and Figure 18) ranged
from 0.04µgg-1 (Middle region) to 0.05µgg-1 (the Eastern and Western region) from the total.
From the obtained data at Table 11, it is observed that the amount of chromium
associated with the carbonate fraction varied from 0.03µgg-1 at station 14 in Burullus (Middle
region) to 0.89µgg-1 at station 8 in Eastern harbour (Western region) with an average
0.19±0.18µgg-1. Table 11 and Figure 18 showed the average concentration of Cr associated
with the carbonate fraction for the three regions, it fluctuated between 0.07µgg-1 (Eastern
region) to 0.28µgg-1 (Western region) from the total.
Table 11 represented that the amount of Cr associated with Fe-Mn oxides fraction
fluctuated between a minium of 0.19 µgg-1 at station 10 in Power station (Middle region) to
4.30 µgg-1 at station 7 in NIOF (Middle region) with an average 1.06±1.02µgg-1. The
average concentration of Cr from the total (Table 11 and Figure 18) in the three regions
ranged between 0.52µgg-1 (Middle region) to 1.48µgg-1 (Western region).
Data listed in Table 11 showed that the amount of Cr associated with the organic
fraction varied from 0.15µgg-1 at station 3 in El-Nobarreya (Western region) to 2.39µgg-1 at
station 16 in Ras El-Barr (Middle region) with an average 1.05±0.66µgg-1. The average
concentration of Cr from the total (Table 11 and Figure 18) in the three regions fluctuated
between 0.74µgg-1 (Western region) and 1.54µgg-1 (Middle region).
The amount of Cr associated with the residual fraction varies from 1.19µgg-1 at station
2 in Bagous (Western region) to 230.69 µgg-1 at station 13 in Rashid east (Middle region)
with an average 63.34±68.62µgg-1. The average concentration of Cr from the total was
represented in Table 11 and Figure 18, it ranged from 17.63µgg-1 (Western region) to 125.73
µgg-1 (Eastern region).
5.3.3.6. Fractionation of Lead (Pb)
The results of Lead concentrations of the different chemical phases exchangeable (F1),
bound to carbonate (F2), bound to Fe/Mn oxides (F3), bound to organic matter (F4) and
-126-
residual fraction (F5) in surface sediments of the Egyptian Mediterranean coastal area from
El-Salloum to Rafah are listed in Table 12 and the relative percent contribution from the total
concentration are shown in Figure 13.
Data listed in Table 12 showed that the amount of Pb associated with the
exchangeable fraction fluctuated from a minimum of 0.05µgg-1 at station 9 in Abu Qir Bay to
0.50µgg-1 at station 13 in Rashid East with an average 0.25±0.14µgg-1. The average
concentration of Pb from the total was represented in Table 12 and Figure 19, it ranged from
0.22µgg-1 (Western region) to 0.30µgg-1 (Middle region).
The amount of Pb associated with the carbonate fraction fluctuated between a
minimum of 0.04µgg-1 at station 12 in Rashid west (Middle region) to 25.88 µgg-1 at station 5
in El-Mex Bay (Western region) with an average 2.74±5.95µgg-1 (Table 12). The average
concentration of Pb from the total was represented in Table 12 and Figure 19, it ranged from
0.29µgg-1 (Middle region) to 4.72µgg-1 (Western region).
From the obtained data at Table 12, it is observed that the amount of Pb associated
with the Fe-Mn fraction varied from 0.37µgg-1 at station 17 in El-Gamil west (Middle region)
to 15.71 µgg-1 at station 5 in El-Mex Bay (Western region) with an average 3.70±4.12µgg-1.
The average concentration of Pb from the total was represented in Table 12 and Figure 19, it
ranged from 1.07µgg-1 (Middle region) to 5.84µgg-1 (Western region).
Table 12 represented that the amount of Pb associated with organic matter fraction
varied from 0.25µgg-1 at station 12 (Rashid west) to 3.41µgg-1 at station 8 in Eastern harbour
(Middle region) with an average 0.95±0.78µgg-1. The average concentration of Pb from the
total was represented in Table 12 and Figure 19 it fluctuated between 0.38µgg-1 (Eastern
region) to 1.18µgg-1 (Western region).
Data listed in Table 12 showed that the amount of Pb associated with the residual
fraction varied from 0.66µgg-1 at station 6 in Western harbour (Western region) to 7.49µgg-1
at station 10 in power station (Middle region) with an average 4.59±2.08µgg-1. The average
concentration of Pb from the total was represented in Table 12 and Figure 19 it fluctuated
between 3.42µgg-1 (Eastern region) to 5.16µgg-1 (Middle region).
-127-
Table 7: Fractionation of Fe concentration
coastal area.
Region
Station
F1
0.14
1
0.12
2
0.11
3
0.16
4
0.22
5
0.25
6
Western Region
0.16
7
0.37
8
0.21
9
0.26
10
0.24
11
0.20
Average
0.29
12
1.02
13
0.49
14
0.11
15
Middle region
0.27
16
0.69
17
0.12
18
0.43
Average
0.27
19
0.27
Eastern region
20
0.27
Average
0.29
Average
0.22
SD
All region
0.11
Minimum
1.02
Maximum
(µgg-1) and its relative percentage from the total in sediments of the Egyptian Mediterranean
%
0.002
0.041
0.038
0.003
0.007
0.019
0.004
0.011
0.002
0.003
0.003
0.012
0.003
0.005
0.003
0.000
0.001
0.004
0.000
0.002
0.001
0.022
0.012
0.009
0.012
0.000
0.041
F2
2.47
1.52
0.89
1.17
0.91
1.36
0.54
1.66
3.02
5.30
3.18
2.00
2.08
2.19
0.90
1.37
2.51
3.55
2.50
2.16
5.28
1.50
3.39
2.19
1.34
0.54
5.30
%
0.03
0.52
0.30
0.03
0.03
0.10
0.01
0.05
0.03
0.07
0.03
0.11
0.02
0.01
0.00
0.00
0.01
0.02
0.01
0.01
0.02
0.13
0.07
0.07
0.13
0.00
0.52
-128-
F3
441.43
118.49
70.67
723.10
485.16
482.64
1371.76
856.24
369.17
287.79
573.75
525.47
464.47
397.84
526.10
1523.75
1478.51
522.64
866.88
825.74
886.98
64.59
475.78
625.60
430.11
64.59
1523.75
%
5.18
40.45
24.25
15.47
16.28
35.61
32.65
26.48
4.02
3.58
5.88
19.08
5.23
1.82
2.73
5.51
4.16
3.09
3.55
3.73
3.38
5.40
4.39
12.24
12.58
1.82
40.45
F4
199.80
15.82
18.51
405.54
299.13
174.99
530.78
321.86
656.57
240.01
1065.11
357.10
700.67
578.10
1115.20
1878.56
2655.46
913.98
1220.87
1294.69
1098.11
165.58
631.85
712.73
664.20
15.82
2655.46
%
2.34
5.40
6.35
8.67
10.04
12.91
12.63
9.95
7.14
2.98
10.92
8.12
7.89
2.64
5.78
6.79
7.47
5.40
4.99
5.85
4.19
13.85
9.02
7.42
3.43
2.34
13.85
F5
7876.49
156.96
201.18
3545.21
2195.12
695.97
2298.38
2053.70
8164.81
7515.20
8113.02
3892.37
7708.55
20890.14
17662.76
24243.63
31423.59
15482.45
22361.07
19967.46
24224.97
963.68
12594.33
10388.84
9815.05
156.96
31423.59
%
92.44
53.59
69.05
75.83
73.65
51.35
54.70
63.51
88.81
93.37
83.17
72.68
86.85
95.52
91.49
87.69
88.37
91.49
91.45
90.41
92.41
80.60
86.50
80.27
14.50
51.35
95.52
Table 8: Fractionation of Mn concentration (µgg-1) and its relative percentage from the total in sediments of the Egyptian Mediterranean
coastal area.
Region
Station
F1
%
F2
%
F3
%
F4
%
F5
%
0.249
0.22
4.69
4.19
41.166
36.78
4.073
3.64
61.736
55.16
1
0.02
0.08
0.87
3.50
21.01
84.73
1.30
5.24
1.60
6.46
2
ND
0.01
0.71
4.44
13.99
87.31
0.86
5.36
0.46
2.87
3
0.04
0.02
14.55
8.68
114.68
68.41
8.55
5.10
29.81
17.78
4
0.07
0.06
7.05
6.08
100.40
86.52
5.08
4.38
3.44
2.97
5
0.17
0.10
10.69
6.52
133.29
81.27
7.17
4.37
12.68
7.73
6
Western region
0.08
0.05
9.53
6.29
102.11
67.35
10.18
6.72
29.71
19.60
7
0.01
0.01
3.55
2.78
105.43
82.57
5.84
4.57
12.85
10.07
8
0.36
0.09
65.61
16.83
173.52
44.52
12.75
3.27
137.54
35.29
9
0.17
0.05
14.79
4.76
53.60
17.26
3.85
1.24
238.04
76.68
10
0.45
0.14
60.04
18.82
186.20
58.36
11.39
3.57
60.97
19.11
11
0.15
0.08
17.46
7.54
95.04
65.01
6.46
4.31
53.53
23.06
Average
0.64
0.32
11.70
5.91
33.31
16.82
6.77
3.42
145.68
73.54
12
0.04
0.01
5.35
0.87
14.09
2.29
5.57
0.91
590.23
95.93
13
0.55
0.14
16.74
4.25
36.17
9.19
9.69
2.46
330.40
83.95
14
0.80
0.13
75.64
11.91
167.17
26.33
32.10
5.06
359.14
56.57
15
1.14
0.20
65.76
11.35
157.30
27.15
32.06
5.53
323.07
55.77
16
Middle region
0.78
0.18
57.35
13.53
63.81
15.06
9.87
2.33
291.94
68.89
17
1.16
0.24
72.51
15.11
69.42
14.47
15.97
3.33
320.77
66.85
18
0.73
0.17
43.58
8.99
77.32
15.90
16.00
3.29
337.32
71.64
Average
1.05
0.14
101.38
13.04
50.49
6.49
25.54
3.28
599.23
77.05
19
0.07
0.18
7.09
17.81
11.43
28.72
0.49
1.24
20.72
52.06
20
Eastern region
0.56
0.16
54.24
15.43
30.96
17.60
13.02
2.26
309.97
64.55
Average
0.39
0.12
30.28
8.83
82.43
43.08
10.45
3.75
178.50
44.22
Average
0.40
0.09
32.17
5.50
57.81
30.74
9.35
1.57
193.00
30.94
SD
All region
ND
0.01
0.71
0.87
11.43
2.29
0.49
0.91
0.46
2.87
Minimum
1.16
0.32
101.38
18.82
186.20
87.31
32.10
6.72
599.23
95.93
Maximum
ND= Not detected
-129-
Table 9: Fractionation of Zn concentration (µgg-1) and its relative percentage from the total in sediments of the Egyptian Mediterranean
coastal area.
Region
Station
F1
%
F2
%
F3
%
F4
%
F5
%
0.25
0.85
0.56
1.87
3.11
10.47
0.06
0.18
25.75
86.62
1
0.07
3.84
0.12
7.26
0.15
8.62
ND
0.12
1.36
80.17
2
0.03
0.50
0.08
1.11
0.08
1.22
ND
0.03
6.55
97.14
3
0.04
0.23
0.91
5.31
7.35
42.83
1.70
9.89
7.17
41.74
4
0.06
0.16
2.49
6.34
10.36
26.38
4.44
11.31
21.92
55.81
5
0.03
0.37
1.13
14.89
5.08
67.26
0.62
8.16
0.70
9.32
6
Western region
0.08
0.15
2.37
4.37
11.01
20.33
4.94
9.12
35.77
66.03
7
0.02
0.11
1.32
8.90
8.53
57.53
1.47
9.92
3.49
23.54
8
0.03
0.21
0.56
4.56
2.04
16.64
0.47
3.83
9.16
74.75
9
0.05
0.38
1.10
9.20
2.52
21.09
0.30
2.50
7.99
66.83
10
0.02
0.15
0.56
4.62
3.98
32.91
1.57
13.01
5.96
49.31
11
0.06
0.63
1.02
6.22
4.93
27.75
1.42
6.19
11.44
59.20
Average
0.02
0.26
0.05
0.67
0.56
6.99
0.50
6.28
6.81
85.80
12
0.02
0.11
0.07
0.42
0.84
4.82
0.33
1.86
16.24
92.78
13
0.01
0.10
0.05
0.36
0.90
6.96
0.75
5.82
11.16
86.76
14
0.01
0.03
0.06
0.29
1.56
7.31
2.00
9.35
17.72
83.03
15
Middle region
ND
0.00
0.08
0.34
2.03
8.28
2.28
9.29
20.13
82.09
16
0.01
0.10
0.07
0.50
0.55
4.01
0.59
4.36
12.39
91.03
17
0.01
0.06
0.04
0.20
0.92
4.47
1.24
6.02
18.32
89.25
18
0.01
0.10
0.06
0.40
1.05
6.12
1.10
6.14
14.68
87.25
Average
0.04
0.16
0.07
0.32
1.01
4.62
1.03
4.72
19.62
90.18
19
0.03
1.61
0.09
4.34
0.24
12.06
ND
0.10
1.62
81.89
Eastern region
20
0.03
0.89
0.08
2.33
0.62
8.34
0.51
2.41
10.62
86.03
Average
0.04
0.47
0.59
3.79
3.14
18.24
1.21
5.79
12.49
71.70
Average
0.05
0.87
0.76
4.01
3.48
18.58
1.37
4.11
9.25
24.17
SD
All region
ND
0.00
0.04
0.20
0.08
1.22
ND
0.03
0.70
9.32
Minimum
0.25
3.84
2.49
14.89
11.01
67.26
4.94
13.01
35.77
97.14
Maximum
ND= Not detected
-130-
Table 10: Fractionation of Ni concentration (µgg-1) and its relative percentage from the total in sediments of the Egyptian Mediterranean
coastal area.
Region
Station
F1
%
F2
%
F3
%
F4
%
F5
%
0.12
2.17
0.06
1.11
0.56
10.11
0.48
8.70
4.30
77.92
1
0.09
6.97
0.03
2.43
0.20
15.74
0.08
6.11
0.88
68.75
2
0.17
2.23
0.10
1.25
0.41
5.36
0.16
2.09
6.79
89.08
3
0.05
0.79
0.12
1.74
0.72
10.80
0.15
2.19
5.64
84.48
4
Western region
0.17
3.78
0.12
2.64
0.35
7.92
0.12
2.73
3.65
82.94
5
0.15
5.24
0.05
1.94
0.59
21.13
0.07
2.37
1.93
69.32
6
0.15
2.10
0.08
1.20
0.73
10.36
0.24
3.38
5.82
82.97
7
0.18
9.35
0.03
1.53
0.50
25.38
0.25
12.82
1.00
50.92
8
0.11
0.66
0.07
0.41
0.50
2.96
1.05
6.29
15.03
89.68
9
0.04
0.33
0.06
0.47
0.45
3.42
0.41
3.08
12.25
92.70
10
0.11
1.32
0.01
0.13
1.48
17.80
1.91
22.97
4.81
57.78
11
0.12
3.18
0.07
1.35
0.59
11.91
0.45
6.61
5.64
76.96
Average
0.05
0.19
0.09
0.32
1.49
5.29
1.12
4.00
25.32
90.20
12
0.18
0.41
0.02
0.05
1.86
4.30
0.94
2.17
40.21
93.07
13
0.13
0.37
0.06
0.17
2.54
7.23
1.93
5.49
30.42
86.74
14
Middle region
0.16
0.52
0.01
0.04
6.05
19.46
5.36
17.25
19.49
62.74
15
0.26
0.62
0.08
0.19
5.81
13.81
5.70
13.56
30.19
71.82
16
0.08
0.16
0.01
0.02
2.60
5.48
2.06
4.34
42.67
90.00
17
0.09
0.25
0.11
0.31
3.81
10.72
3.92
11.02
27.63
77.69
18
0.14
0.36
0.06
0.16
3.45
9.47
3.00
8.26
30.85
81.75
Average
0.20
0.58
0.06
0.18
2.49
7.37
2.54
7.52
28.45
84.34
19
Eastern region
0.02
0.09
0.01
0.05
0.42
1.75
0.05
0.22
23.65
97.90
20
0.11
0.34
0.04
0.12
1.46
4.56
1.30
3.87
26.05
91.12
Average
0.13
1.91
0.06
0.81
1.68
10.32
1.43
6.91
16.51
80.05
Average
0.06
2.55
0.04
0.85
1.76
6.63
1.74
5.90
13.62
12.75
SD
All region
0.02
0.09
0.01
0.02
0.20
1.75
0.05
0.22
0.88
50.92
Minimum
0.26
9.35
0.12
2.64
6.05
25.38
5.70
22.97
42.67
97.90
Maximum
-131-
Table 11: Fractionation of Cr concentration (µgg-1) and its relative percentage from the total in sediments of the Egyptian Mediterranean
coastal area.
F4
Region
Station
F1
%
F2
%
F3
%
%
F5
%
1
2
3
4
5
Western region
6
7
8
9
10
11
Average
12
13
14
Middle region
15
16
17
18
Average
19
Eastern region
20
Average
Average
All Region
SD
Minimum
Maximum
0.12
0.09
0.04
0.03
0.04
0.04
0.03
0.03
0.05
0.03
0.02
0.05
0.06
0.03
0.04
0.04
0.03
0.04
0.02
0.04
0.02
0.08
0.05
0.04
0.03
0.02
0.12
0.48
2.79
1.04
0.25
0.16
0.57
0.15
0.41
0.07
0.10
0.09
0.56
0.08
0.01
0.04
0.07
0.03
0.02
0.02
0.04
0.01
0.16
0.08
0.33
0.63
0.01
2.79
0.24
0.29
0.09
0.25
0.17
0.19
0.30
0.89
0.20
0.22
0.20
0.28
0.13
0.19
0.03
0.11
0.08
0.06
0.08
0.10
0.07
0.08
0.07
0.19
0.18
0.03
0.89
0.97
9.65
2.26
2.03
0.64
2.93
1.65
12.17
0.32
0.69
0.84
3.10
0.19
0.08
0.03
0.17
0.08
0.03
0.08
0.09
0.03
0.14
0.09
1.75
3.27
0.03
12.17
-132-
1.01
0.93
0.55
1.79
2.40
1.33
4.30
2.74
0.74
0.19
0.34
1.48
0.48
0.52
0.40
0.57
0.40
0.76
0.55
0.52
0.65
0.61
0.63
1.06
1.02
0.19
4.30
4.05
30.45
14.23
14.64
9.17
20.92
23.79
37.34
1.16
0.59
1.45
14.34
0.70
0.22
0.47
0.89
0.38
0.42
0.54
0.52
0.32
1.16
0.74
8.14
11.52
0.22
37.34
1.24
0.55
0.15
0.26
0.60
0.19
1.88
0.34
0.61
0.84
1.44
0.74
1.03
1.13
1.33
2.11
2.39
1.05
1.74
1.54
1.59
0.54
1.07
1.05
0.66
0.15
2.39
4.99
17.98
3.98
2.10
2.31
2.95
10.38
4.61
0.95
2.57
6.20
5.37
1.51
0.48
1.57
3.30
2.27
0.58
1.72
1.63
0.79
1.02
0.90
3.61
4.12
0.48
17.98
22.27
1.19
3.02
9.88
22.94
4.60
11.58
3.33
62.52
31.33
21.25
17.63
66.82
230.69
82.77
61.21
102.07
178.97
98.85
117.34
199.72
51.75
125.73
63.34
68.62
1.19
230.69
89.51
39.14
78.49
80.98
87.72
72.63
64.03
45.46
97.50
96.06
91.41
76.63
97.53
99.20
97.89
95.58
97.24
98.94
97.63
97.72
98.85
97.52
98.19
86.17
17.96
39.14
99.20
Table 12: Fractionation of Pb concentration (µgg-1) and its relative percentage from the total in sediments of the Egyptian Mediterranean
coastal area.
Region
Station
F1
%
F2
%
F3
%
F4
%
F5
%
0.09
0.64
1.97
14.30
3.72
26.96
0.64
4.64
7.38
53.46
1
0.23
4.91
0.69
14.93
1.66
36.05
0.39
8.52
1.64
35.59
2
0.23
2.31
4.55
45.63
3.23
32.34
0.53
5.34
1.44
14.38
3
0.21
1.54
1.14
8.33
4.51
32.84
0.74
5.35
7.14
51.94
4
0.38
0.80
25.88
54.40
15.71
33.03
2.32
4.87
3.28
6.90
5
0.17
0.92
9.98
54.51
7.00
38.28
0.49
2.68
0.66
3.61
6
Western region
0.32
1.45
5.01
22.60
10.45
47.12
1.73
7.80
4.66
21.03
7
0.05
0.24
1.36
6.41
10.54
49.83
3.41
16.13
5.79
27.38
8
0.05
0.54
0.46
5.42
1.68
19.75
1.47
17.25
4.85
57.04
9
0.45
3.53
0.58
4.59
3.51
27.68
0.65
5.11
7.49
59.09
10
0.27
3.31
0.31
3.87
2.24
27.89
0.64
8.00
4.57
56.93
11
0.22
4.72
4.72
21.36
5.84
33.80
1.18
7.79
4.45
35.21
Average
0.31
5.67
0.04
0.82
1.84
34.01
0.25
4.69
2.96
54.82
12
0.50
6.64
0.14
1.82
0.76
10.13
0.53
6.98
5.61
74.44
13
0.35
5.56
0.53
8.36
1.11
17.56
0.51
8.06
3.80
60.46
14
Middle region
0.28
3.04
0.46
5.08
0.65
7.09
1.07
11.73
6.67
73.07
15
0.40
4.89
0.27
3.27
1.49
18.26
1.28
15.68
4.72
57.90
16
0.17
2.27
0.26
3.43
0.37
4.92
1.01
13.23
5.79
76.15
17
0.07
0.74
0.31
3.58
1.25
14.29
0.52
5.90
6.60
75.49
18
0.30
4.12
0.29
3.76
1.07
15.18
0.74
9.47
5.16
67.47
Average
0.45
5.81
0.45
5.81
1.50
19.51
0.26
3.35
5.02
65.53
19
Eastern region
0.13
3.60
0.40
10.76
0.87
23.42
0.50
13.42
1.82
48.81
20
0.29
4.70
0.42
8.28
1.18
21.47
0.38
8.38
3.42
57.17
Average
0.25
2.92
2.74
13.90
3.70
26.05
0.95
8.44
4.59
48.70
Average
All region
0.14
2.07
5.96
17.09
4.12
12.32
0.78
4.52
2.08
22.83
SD
0.05
0.24
0.04
0.82
0.37
4.92
0.25
2.68
0.66
3.61
Minimum
0.50
6.64
25.88
54.51
15.71
49.83
3.41
17.25
7.49
76.15
Maximum
-133-
Fe
100
80
60
40
20
0
1
2
F1
3
4
5
6
F2
7
8
9 10 11 12 13 14 15 16 17 18 19 20
F3
F4
F5
Mn
100
80
60
40
20
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
F1
F2
F3
F4
F5
Zn
100
80
60
40
20
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
F1
F2
F3
F4
F5
Figure 12: Percent contribution of Fe, Mn and Zn fractions from the total concentration
in surfacial sediments of the Egyptian Mediterranean coastal area
-134-
Ni
100
80
60
40
20
0
1
2
3
F1
4
5
6
7
8
9 10 11 12 13 14 15 16 17 18 19 20
F2
F3
F4
F5
Cr
100
80
60
40
20
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
F1
F2
F3
F4
F5
Pb
100
80
60
40
20
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
F1
F2
F3
F4
F5
Figure 13: Percent contribution of Ni, Cr, and Pb fractions from the total concentration
in surfacial sediments of the Egyptian Mediterranean coastal area
-135-
Western region
0%
0%
19%
8%
73%
Middle region
0% 4%
6%
90%
Eastern region
0% 4%
9%
87%
F1
F2
F3
F4
F5
Figure 14: The relative percentage of Fe fractions from the total concentration in
the three regions, Egyptian Mediterranean coastal area
-136-
Western region
0%
23%
8%
4%
65%
Middle region
0%
9%
16%
3%
72%
Eastern region
0%
15%
18%
65%
2%
F1
F2
F3
F4
F5
Figure 15: The relative percentage of Mn fractions from the total concentration in
the three regions, Egyptian Mediterranean Sea sediments
-137-
Western region
1%
6%
28%
59%
6%
Middle region
0%
0%
6%
6%
88%
Eastern region
1% 2%
8%
2%
87%
F1
F2
F3
F4
F5
Figure 16: The relative percentage of Zn fractions from the total concentration in
the three regions, Egyptian Mediterranean Sea sediments
-138-
Western region
3% 1%
12%
7%
77%
Middle region
0%
0%
9%
8%
83%
Eastern region
0%
0%
5%
4%
91%
F1
F2
F3
F4
F5
Figure 17: The relative percentage of Ni fractions from the total concentration in
the three regions, Egyptian Mediterranean Sea sediments
-139-
Western region
1%3%
14%
5%
77%
Middle region
0%
1% 2%
0%
97%
Eastern region
0%
0%
1%
1%
98%
F1
F2
F3
F4
F5
Figure 18: The relative percentage of Cr fractions from the total concentration in
the three regions, Egyptian Mediterranean Sea sediments
-140-
Western region
2%
21%
35%
8%
34%
Middle region
4%
4%
15%
9%
68%
Eastern region
5%
8%
21%
58%
8%
F1
F2
F3
F4
F5
Figure 19: The relative percentage of Pb Fractions from the total concentration in
the three regions, Egyptian Mediterranean Sea sediments
-141-
5.3.4. Acid volatile sulfide (AVS) and simultaneously extracted metals
(SEM)
Acid volatile sulfide (AVS) is one of the major chemical components that control the
activities and availability of metals in the interstitial waters of sediments. Sulfide reacts with
several divalent transition metal cations (cadmium, copper, nickel, lead, and zinc) to form
highly insoluble compounds that are not bioavailable (USEPA, 2004)(444). The concentrations
of reactive sulfide in anoxic sediments play a major role in the bioavailability and toxicity of
specific cationic metals of environmental concern (e.g. Cu, Cd, Pb, Ni, and Zn). Because
these metals are capable of forming relatively insoluble sulfide minerals, the reaction between
soluble metal species with sulfide serves to partition metals from porewater into the solid
phase (Edenborn, 2005)(445). The concentration of AVS and SEM (µmoleg-1, dry weight) in
the sediments of the Egyptian Mediterranean coast from El-Salloum to Rafah are shown in
Table 13.
From the obtained data in Table 13, it is observed that the range and average
concentrations of SEM were 0.0001-0.0011 µmole/g (Average: 0.0002 µmole/g) for Cd,
0.002-0.017 µmoleg-1 (Average: 0.0095 µmoleg-1) for Cu, 0.003-0.029 µmole/g (Average:
0.016µmoleg-1) for Ni, 0.004-0.008 (Average: 0.006 µmoleg-1) for Pb and 0.003-0.025
(Average: 0.014 µmoleg-1) for Zn.
The distribution of SEM in the western region revealed low concentration
(0.016µmoleg-1) at station 2 and 3 in Baghoush and El-Nobarreya, respectively and high
concentration (0.241µmoleg-1) at station 5 in El-Mex Bay. The average value of SEM in
this region was (0.128±0.075µmoleg-1). In the Middle region, the lowest concentration of
SEM (0.045µmoleg-1) was found at station 17 in El-Gamil west, while the highest
concntration (0.133µmoleg-1) was found at station 15 in New Damietta. The average value of
SEM in the Middle region was (0.079±0.036µmoleg-1). The lowest concentration of SEM
(0.012 µmoleg-1) in the Eastern region was observed at station 20 in Rafah, while the highest
concentration (0.079µmoleg-1) was observed at station 19 in Port Said. The average value of
SEM in this region was (0.046±0.047µmoleg-1). Figure 21 showed that the average
concentration of SEM in the three regions can be arranged as follows:
Western region (0.128±0.075µmoleg-1) > Middle region(0.079±0.036µmoleg-1) > Eastern
region(0.046±0.047µmoleg-1).
The lowest concentration of AVS (0.015µmoleg-1) in the western region was observed
at station 2 in Baghoush, while the highest concentration (31.326µmoleg-1) was observed at
station 7 in NIOF. The average value of AVS in this region was (3.307±9.348 µmoleg-1). In
the Middle region, the lowest concentration of AVS (0.038 µmoleg-1) was found at station 18
in El-Gamil East, while the higest concentration (0.11µmoleg-1) was observed at station 15 in
New Damietta. The average of AVS in this region was (0.058±0.025µmoleg-1). The
distribution of AVS level in the Eastern region revealed low value (0.029µmole g-1) at station
19 in Port Said and high value (0.119 µmole g-1) at station 20 in Rafah. The average value of
AVS in this region was (0.074±0.064 µmoleg-1). Figure 22 showed that the average
concentration of AVS in the different regions can be arranged as follows:
Western region (3.307±9.348 µmoleg-1) > Eastern region (0.074±0.064 µmoleg-1) > Middle
region (0.058±0.025µmoleg-1)
-142-
Table 13: Concentration of AVS and SEM (µmoleg-1, dry weight) in surfacial
sediments of the Egyptian Mediterranean coastal area from El-Salloum to Rafah
Region
Station
1
2
3
4
5
6
7
Western
8
region
9
10
11
Average
SD
Min
Max
12
13
14
15
16
Middle
17
region
18
Average
SD
Min
Max
19
20
Eastern Average
region
SD
Min
Max
AVS
3.505
0.015
0.036
0.26
0.335
0.032
31.326
0.058
0.576
0.123
0.116
3.307
9.348
0.015
31.326
0.049
0.052
0.044
0.11
0.043
0.071
0.038
0.058
0.025
0.038
0.11
0.029
0.119
0.074
0.064
0.029
0.119
Pb
0.026
0.009
0.008
0.018
0.033
0.011
0.008
0.014
0.016
0.023
0.011
0.016
0.008
0.008
0.033
0.006
0.009
0.004
0.009
0.008
0.005
0.007
0.007
0.002
0.004
0.009
0.008
0.004
0.006
0.003
0.004
0.008
Cd
0.0011
0.0001
0.0001
0.0005
0.0006
0.0001
0.0003
0.0004
0.0003
0.0002
0.0002
0.0004
0.0003
0.0001
0.0011
0.0003
0.0001
0.0002
0.0001
0.0001
0.0003
0.0001
0.0002
0.0001
0.0001
0.0003
0.0001
0.0003
0.0002
0.0001
0.0001
0.0011
-143-
Zn
0.071
0.004
0.002
0.091
0.174
0.038
0.141
0.139
0.049
0.07
0.083
0.078
0.056
0.002
0.174
0.017
0.019
0.022
0.047
0.056
0.021
0.03
0.030
0.015
0.017
0.056
0.025
0.003
0.014
0.016
0.003
0.025
Cu
0.031
0.002
0.005
0.021
0.023
0.008
0.029
0.046
0.011
0.039
0.019
0.021
0.014
0.002
0.046
0.011
0.009
0.012
0.017
0.017
0.009
0.012
0.012
0.003
0.009
0.017
0.017
0.002
0.0095
0.011
0.002
0.017
Ni
0.022
0.002
0.002
0.016
0.011
0.009
0.024
0.007
0.014
0.008
0.021
0.012
0.008
0.002
0.024
0.016
0.028
0.011
0.059
0.041
0.01
0.038
0.029
0.018
0.01
0.059
0.029
0.003
0.016
0.018
0.003
0.029
SEM
0.151
0.016
0.016
0.147
0.241
0.066
0.201
0.206
0.089
0.14
0.134
0.128
0.075
0.016
0.241
0.05
0.064
0.049
0.133
0.122
0.045
0.087
0.079
0.036
0.045
0.133
0.079
0.012
0.046
0.047
0.012
0.079
Western region
10%
13% 0%
17%
60%
Middle region
9% 0%
37%
38%
16%
Eastern region
13%
35%
0%
31%
21%
SEMPb
SEMCd
SEMZn
SEMCu
SEMNi
Figure 20: The relative percentage of simultaneously extracted metals in the three
regions, Egyptian Mediterranean Sea sediments
-144-
0.15
0.1
0.05
0
Western region
M iddle region
Eastern region
Figure 21: The average of SEM in the three regions, Egyptian Mediterranean
Sea sediments
4
3
2
1
0
Western region
M iddle region
Eastern region
Figure 22: The average of AVS in the three regions, Egyptian Mediterranean Sea
sediments
-145-
Chapter VI
Discussion
The study of marine sediments represents a useful tool for determining the actual state
of environmental pollution and for understanding the origin and mechanism of the
phenomena. Sediments could be regarded as a historical reflection to changes occurring in
the overlying water system and can be used as good indicators for metals in any study area
(Salomons and Forstner, 1984)(181).
Marine sediment contamination by persistent anthropogenic chemicals poses one of
the worst problems to coastal and estuarine ecosystems around the world. The accumulation
of urban, industrial and agricultural contaminants in marine sediments enters the sea through
river mouths or by direct discharge (Usero et al., 2008)(446). Due to its ecological importance,
the evaluation of marine sediment quality constitutes an important area of research (Silva et
al., 2004)(447).
Sediments may accumulate various kinds of hazardous and toxic substances, including
trace metals levels many times higher than water column concentrations, causing a serious
problem due to their toxicity and their ability to accumulate in the biota (Usero et al.,
2008)(446). The most crucial property of heavy metals is that they are bioavailable and not
biodegradable in the environment. In aquatic ecosystem, sediments are the main sink for
trace metals (Ridgway and Shimmield, 2002)(448). Metal fractionations in sediments are
strongly dependent on substrata geochemical conditions, such as pH and redox potential.
When environmental conditions change, some of the sediment-bound contaminants can
remobilize and be released back into the water acting as a source of pollution that can have
adverse effects on living organisms (Li et al., 2001)(235).
The toxicity and fate of the metal in sediment depend on its chemical form, and
therefore, quantification of the different forms of a metal is more meaningful than the
estimation of its total concentration (Usero et al., 2008)(446). In this study, fractionation of
metals in sediment has been investigated to determine its speciation and ecotoxic potential, as
well as evaluation of metal potential toxicity based on the simultaneously extracted metals
(SEM) and acid volatile sulfides (AVS) analysis.
6.1. Geochemical Analysis
6.1.1. Grain size distribution
Trace metals introduced into aquatic environment can have a detrimental effect on the
organisms living in that environment, as well as, on food chain, sediments are able to store
and accumulate metals to some extent, due to their adsorption capacity (Stigliani, 1991)(449).
Sediment exists as complex and dynamic mixtures of mineral particles, particular organic
material (detritus) and microbes. The flocculated chemicals and contaminants in sediments
may be associated with any of these sediment fractions (Tabata et al., 2009)(213). Grain size
plays a significant role in determining elemental concentrations in sediments (Szefer et al.,
1996)(450). Studies on the grain size effect are important to both environmental studies and
environmental management (Zhang et al., 2002)(451). Trace metals may occur in both fine and
sand fractions of sediments. However, most natural and anthropogenic substances (metals
and organic contaminants) show a much higher affinity to fine particulate matter than the
coarse fraction. In marine sediments, anthropogenic trace metals are usually preferentially
associated with finer grains (Ewais et al., 2000)(452). Sediment grain size is therefore
-146-
important in determining how sediments are transported and deposited in the marine
environment, and in the distribution of both natural and anthropogenically introduced
concentrations of metals in the marine environment (Stevenson, 2001)(453). The accumulation
of trace metal in fine grained estuarine sediments may cause a significant long term
contamination problem, as reworking of contaminated sediments and possible early
diagenetic release of heavy metals may cause continued input into the environment even after
cessation or reduction in industrial discharges (Rosales Hoz et al., 2003)(454).
In the present study Particle size analysis was carried out by dry sieving through
screens; particles <63µm were not differentiated into their clay and silt components. The
percent of sand, silt and clay together with the texture of all samples are shown in Table 4.
The obtained results showed that the particle size distributions for sediments samples from
Rafah to El-Salloum were dominated by sand particles (>63µm) with high percentages 86.85100% (average: 99.15%). In general, the particle size distribution was specific to the
respective sediments. Fine sand (0.250-0.125 mm) and very fine sand (0.125-0.062 mm)
were the dominated fractions of all sediment samples. Course sand (1.0-0.5 mm) was
observed at station 5 in El-Mex Bay and station 6 in the Western harbour (Western region).
The highest value of sand content (100%) was observed at stations 3, 5, 6 at Nobareya, ElMex Bay, Western harbour, respectively in the Western region and station 12 in Rashid west
(the Middle region) and station 20 at Rafah (the Eastern region) and the lowest value
(86.85%) was observed at station 1 in El-Salloum (the Western region). These results are in
agreement with those obtained by El-Nemr et al. (2007a)(61).
The Middle region was dominated by fine and very fine sand. This is in agreement
with El-Fishawi et al. (1976)(455) who stated that the beach sands of the Nile delta are mostly
fine grained sand (0.13-0.25 mm) with some local medium sand (0.25-0.5 mm) and composed
of quartz and feldspares with smaller amounts of heavy minerals and shell fragments, being 13%.
6.1.2. Total carbonate
Carbonate is often an important component of marine sediments and has been found to
be an important indicator of provenance and dispersal of terreginous material (Loring and
Rantala, 1992)(416). The interaction between trace metals and carbonate minerals is of an
environmetnal interest since they influence the distribution of ions in the aquatic environment
and the association with carbonate minerals can provide important pathways for scavenging
toxic metals like Zn, Pb, and Cd (Korfali and Davies, 2004)(456).
The Egyptian Mediterranean coast contains a wide variety of sediments. Coastal
sediments are mainly composed of two principal types: carbonate sands and quartz dominant
sands (Hilmy, 1951)(359). The results of carbonate as % CaCO3 in the Egyptian Mediterranean
coastal sediments are listed in Table 4 and Figure 3. The study area was divided into three
regions, which revealed differences in the distribution of CaCO3. In the Western region, the
results showed that the sediments at station 1 (El-Salloum) are characterized by relatively
medium CaCO3 percentage (34.79%). Sediments from station 2 at Baghoush to station 8 at
the Eastern harbour (west of Alexandria) are characterized by high percentage of CaCO3
ranged from 67.24 at station 5 (El-Mex) to a maximum percentage (the western harbour)
(95.57%). The major sources of carbonate in the western harbor's sediments are the detrital
fragmental molluscal shells and skeletons in addition to the tubiform skeletons of Hydroids
norvegicus (Saad et al., 2004)(91). These results are in agreement with the results of Anwar et
al. (1984)(360) who stated that the western region from Mersa Matrouh to Alexandria is mainly
composed of pure oolitic carbonate. The non-arrival of riverine muds or terrestrial discharge
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also plays a leading role in keeping the sediments of the western part of the Egyptian shelf as
highly carbonate content (El-Wakeel et al., 2006)(285). The oolitic grains constitute an average
of 78% and 89% of the bottom and beach sediments, respectively. Further westward, sand at
El-Salloum showed an average oolite content in the nearshore of 58% (Anwar et al.,
1981)(361). Also, Nasr (1978)(362) found that the carbonate content of beach sand west of
Alexandria varies between 44 and 99%
The relative percentage of CaCO3 at eastern region of Alexandria was characterized
by relatively low % (14.69, 15.64, 22.11%) at Power station, Maadia and Abu Qir Bay,
respectively. The lowest value of carbonate were found in the marine sediments where quartz
sand was predominates. It is thus evident that the shells of organisms are the principle
contributors of carbonate in sediments (Khalil et al., 2007)(320). This result is in agreement
with those of Selim (1974)(457) who found that the sediment along the coast west of
Alexandria consists mainly of calcareous material, which creates from limestone rocks and
ridges. These sediments differ from those found east of Alexandria, which creates from the
Nile Delta. The seaward calcareous ridges consist of Pleistocene carbonates with a marine
nature.
In the middle region, the relative percentage of CaCO3 was low ranged from 2.85% at
station 13 (Rashid east) to 6.08% at station 15 (New Damietta). This is in agreement with
Hilmy (1951)(359) who mentioned that beach sand from Alexandria to Rashid primarily
consists of quartz grains with common shell fragments. The Nile River has been identified as
the major source of quartz-rich sediments and heavy minerals on the Nile Delta beaches
(Frihy, 1994)(363).
In the eastern region, the relative percentages of CaCO3 are low, 6.33% at station 20
(Rafah) and 7.32% at station 19 (Port Said).
6.1.3. Organic matter
Organic carbon is determined to assess the role played by the organic fraction of
sediments in the transport, deposition and retention of trace metals (Loring and Rantala,
1992)(416). Organic matter in the sediment played an important role in the adsorption of trace
metals. Further, it was suggested that the organic matter content in general could be used as a
simple pollution index of the sediment (Ottosen and Villumsen, 2006)(458).
The composition and structure of the organic matter in the sediment are varying due to
its origin and geological history in the marine and aquatic environment. Phytoplankton and
zooplankton are the most abundant sources of the organic material in the sediments
(Grathwohl, 1990)(459). The organic matter content of the sediment is a result of the
contribution of terrigenous materials and the decomposition of plants and animals by the
action of bacteria (Draz, 1983)(460).
Total organic matter (TOM) is one of the most important collectors of pollutants in the
marine sediments. Organic matter tends to form strong organo-metallic complexes with
metals, rendering them immobile. An increase in TOM content may result in an increase in
levels of metals in the marine sediment (Massoud et al., 2007)(294).
Total organic carbon added to the sediments, primarily through the decomposition of
plant and animal matter, can directly binds and/or adsorb heavy metals and also contains
heavy metals accumulated by plants that have been exposed to contaminated sediment during
their life time (Peltier et al., 2003)(461). Nonetheless, high percentage of organic matter
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and/or small grains in sediments, reduces the metals availability, and undoubtedly plays a
major role in controlling metal's concentration in the pore water (Ankley et al., 1996)(23).
Organic carbon has a significant role in geochemical cycles of major and trace metals that
accumulate in sediments; it may be used as an index of depositional environment and
sedimentary processes (Seralathan et al., 1993)(462).
The results of organic matter contents in the Egyptian Mediterranean coastal
sediments are listed in Table 4 and Figure 4. The organic matter values ranged from 0.08 to
1.72 % with a mean value of 0.56%, obtaining the highest percentage from station 7 (NIOF;
western region). The common values for coastal sediments are from 0.32% to 1.76% (Diaz-de
alba, 2011)(463). Therefore, the present study showed relatively low organic matter content at
most of the locations. Low organic carbon values were observed which might be due to the
mixing processes and marine sedimentation at the sediment water interface, where the rate of
delivery, as well as the rate of degradation by microbial-mediated processes, can be high
(Canuel and Martens, 1993((464). The present work findings are agreeable with Revelle and
Shepard (1938)(465), who mentioned that the decrease in the organic carbon seaward is
explained by the increase of the grain size and the carbonate content.
El-Wakeel (2006)(285) mentioned that the relatively lower values of organic carbon in
the beach sediments of the western region from El-Dakhaila to Sidi Kirir can be the result of:
a) the position of this area further away from the riverine input or terrestrial discharge which
is regarded as the main contributer of the organic detritus; b) the remarkable increase of
carbonate; and c) waves and longshore and rip currents may remove the fine particles and the
associated organic carbon away from the beach.
6.2. Trace metal distribution
6.2.1. Total metals
With the rapid industrialization and economic development in coastal region trace
metals are continuing to be introduced to estuarine and coastal environment around the world
(Santos et al., 2005)(466). Various studies have demonstrated that sediments from coastal areas
greatly contaminated by trace metals. Therefore, the evaluation of metal distribution in
surface sediments is useful to assess pollution in the marine environment (Jayaprakash et al.,
2008)(467). These trace metals participate in various biogeochemical mechanisms that have
significant mobility, which affects the ecosystems through bioaccumulation and biomagnification processes and are potentially toxic for environment and human life (Ip et al.,
2007)(428). Metals such as Ni, Cd, Cr and Zn, etc. are used in contamination studies in marine
systems due to their relationship with anthropogenic activities (Burton et al., 2004)(468).
The range and the average values of trace metal (Fe, Mn, Cr, Ni, Zn, Pb, Cu, Co, Cd)
concentrations enter the aquatic marine environment from both natural and anthropogenic
sources were given in Table 5 and Figures 6, 7, and 8. Metal contents ranged in the following
intervals: Fe: 243.5-38045.05µgg-1; Mn: 17.3-1085.7µgg-1; Cr: 4.1-297.95µgg-1; Ni: 1.760.3µgg-1; Zn: 2.1-62.21µgg-1; Pb: 3.34-53.67µgg-1; Cu: 0.46-26.3µgg-1; Co: 0.4-26.4µgg-1;
Cd: 0.04-0.47µgg-1 dry weight. The average contents of the study area were Fe: 13255.7µgg-1;
Mn: 380.7µgg-1; Cr: 82.7µgg-1; Ni: 25.9µgg-1; Zn: 22.2µgg-1; Pb: 13.2µgg-1; Cu: 8.5µgg-1; Co:
8.2µgg-1; Cd: 0.22µgg-1 dry weight, allow arranging trace metals as the following: Fe> Mn>
Cr> Ni> Zn> Pb> Cu> Co> Cd. In order to give more information about the level of
pollution of the studied sediments, the measured metals concentrations were also compared
with those given in the literature by Taylor and Wedepohl (1961)(469), Taylor (1964)(470),
Bowen (1979)(471), Salomons and Forstner (1984)(181), Taylor and Mclennan (1995)(472),
Wedepohl (1995)(473).
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The distribution patterns of metal concentrations increases towards the Middle region
of the study area. Metals concentrations were found to be higher in the middle region of the
study area compared to the eastern and the western region. It has been observed that the
average of the total trace metals (Fe, Mn, Zn, Ni, Cr, Co) contents in the western region was
lower than the average of those metals in the middle and the eastern region. Whereas, the
average of the total Pb and Cd content in the western region was higher than the average of
the total Pb and Cd content of the middle and eastern region. The average of the total cupper
content in the western region was higher than the average of the total cupper content in the
eastern region and lower than the average of the cupper content in the middle region. On the
other hand, the average of the CaCO3 in the western region was higher than the average of
CaCO3 in the middle and the eastern region. This result may be explained as carbonate
usually contains insignificant amounts of trace metals and act mainly as dilutents. Under
certain circumstances, however, carbonates can fix contaminants such as cadmium and copper
(UNAP/Med POL, 2000)(474).
Total Iron (TFe)
Iron was the most abundant metal measured and there were large variations in the
levels of total Fe among sites. The average of total Fe levels in µgg-1 dry weight at the
studied regions were: the middle region (25516.22 ± 10090.1)>the eastern region (17696.95 ±
23015.70)>the western region (4646.03 ± 3459.61) as shown in Figure 23. Table 14 showed
that the concentration of Fe in the sediments along the Egyptian Mediterranean coast was
below the crustal average (Taylor, 1964)(470), Average shale (Turekian and Wedeohl,
1961)(469), mean sediment concentration (Salomons and Forstner, 1984)(181) and the average
continental crust (Wedepohl, 1995)(473) at all locations. While, the concentration of iron at
New Damietta (Stations 15; 38045.05µgg-1) and at Ras El-Barr (station 16; 37398.1µgg-1)
(the middle region) exceeds the upper crust concentration reported by Taylor and Mclennan
(1995)(472). The results of Fe (unpolluted area) were in the range from 3,000 to 40,820 µgg-1
(Santamaria-Fernandez et al., 2005; Svete et al., 2001)(475, 476), while polluted regions had
higher values in the range between 51,000 and 116,000 µgg-1 (Buykx et al., 2000; Sulivan and
Taylor, 2003)(477, 478) compared to these studies, the sediments along the Egyptian
Mediterranean coast are considered unpolluted by Fe. By comparing the obtained results with
previous results of other Mediterranean regions, the results were slightly higher than that
reported by Mostafa et al. (2004)(298) for the western harbour, Egypt (8819-36140µgg-1),
Legorburu and Canton (1991)(479) for the Pasajes harbour, Spain (3830-35500µgg-1),
Buccolieri et al. (2006)(160) for Taranta Gulf, Italy (26313-36098µgg-1), Khairy (2008)(226) for
Abu Qir Bay (1120-37100µgg-1) and lower than the range reported by Gargouri et al.
(2010)(480) for the coast of Safax (41011-50163µgg-1) and the range reported by Diaz de Alba
et al. (2011)(463) for Algeciras Bay, Spain (18285-42756µgg-1) (Table 15).
Total Manganese (TMn)
Distribution of Mn in the study area ranged from 17.2µgg-1 to 1085.72µgg-1 dry
weight with an average value of 380.71µgg-1 (Figure 23). The average Mn levels (µgg-1 dry
weight) measured for the three regions was in the following sequence: the middle region
(597.02±203.84)> eastern region (570.24±729.00)> the western region (208.60±163.08). The
concentration of Mn in the study area was below the crustal average reported by Taylor
(1964)(470) and the average shale reported by Turekian and Wedepohl (1961)(469) at all
locations except station 19 in Port Said at the eastern region (1085.72µgg-1). Stations 13
(Rashid East) and station 15 (New Damietta), at the Middle region and station 19 (Port Said)
in the eastern region exceeds the Mean sediment (Salomons and Forstner, 1984)(181). The
obtained results are higher than the results reported by Mostafa et al. (2004)(298) for the
western harbour, Egypt, Pasajes harbour, Spain reported by Legorburu and Canton (1991)(479),
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Nables harbour (Italy) reported by Adamo et al. (2005)(184), El-Mex Bay (Egypt) reported by
Abdallah (2007a)(292), Abu Qir bay reported by Khairy (2008)(226) and quite similar with the
results obtained by Diaz-de Alba et al. (2011)(463) for Algeciras bay, Spain and Themaikos
Gulf, Greece reported by Violintzis et al. (2009)(481) and lower than the results obtained by
Buccolieri et al. (2006)(160) for Taranto Gulf, Italy (Table 15).
From the obtained results, Fe and Mn display quite similar pattern of distribution.
This is in agreement with the results of Sarkar and Bhattacharya (2010)(482) who stated that
this might be due to the early diagenetic processes as well as the strong association to the
geochemical matrix between the two elements. The total Fe and Mn concentrations were low
at the western region of the study area, which is dominated by high CaCO3 and sand content,
whereas the middle and the eastern indicates high values of total Fe and Mn. It seems likely
that the enrichment results in the reduction of Fe in the sediment during the oxidation of
organic matter (Francois, 1988)(483). The iron and manganese can be converted to complex
hydroxyl compounds that may eventually precipitate (Riley and Chester, 1971)(31). It is well
established that iron and manganese oxides are excellent scavengers for trace metals (Tessier
et al., 1979)(15). This would lead the co-precipitation of other metals in the water column and
so increase the concentration of many metals in sediments. Nasr et al. (1990)(325) stated that
the occurrence of these two metals is mainly controlled by natural processes such as Nile
water supply, or even coastal erosion deposition.
Total Chromium (TCr)
Chromium and Ni displays quite similar pattern of distribution, these elements are
used as markers of metal industries (Loska et al., 2004)(484).
The mean Cr concentration in µgg-1 dry weight for the three regions were
(178.37±149.81; eastern region) > (146.53±84.69; Middle region) > (24.76±22.97; western
region) as represented in Figure 23. There was large variability in the levels of Cr among the
sites. The highest concentration (297.95µgg-1) of Cr was found at station 13 (Rashid east;
Middle region) and the lowest concentration (4.08µgg-1) of Cr was found at station 2
(Baghoush; Western region). Cr concentrations in the middle region and the eastern region
are in agreement with previous studies which reported high values of Cr in polluted
sediments. For example, it was reported that the levels of Cr in sediments from the Aegean
Sea (Greec) were in the range of 44.6-154.0µgg-1 (Aloupi and Angelidis, 2002)(485) and 53.0117.0 µgg-1 in sediments from the Gulf of Finland (Leivouri, 1998)(134). The concentration of
Cr at station 9 (Abu Qir Bay; western region), stations 12, 13, 14, 15, 16, 17, 18 (Rashid west,
Rashid east, Burullus, Ras El-Burr, El-Gamil west and El-Gamil east, respectively; middle
region) and stations 19, and 20 (Port Said and Rafah; eastern region) was higher than the
upper crust background (35µgg-1) (Taylor and Mclennan, 1995)(472) and the Average
continental crust background (35 µgg-1) (Wedepohl (1995)(473), which explained a new input
of toxic Cr to those stations. On the other hand, according to Taylor (1964)(470) and Bowen
(1979)(471), the concentration of Cr at stations 13, 14, 16, 17, 18, and 19 exceeds the
background concentration of the upper crust and the mean crust (Table 14). Cr concentrations
in the western region are in agreement with the results obtained by Alomary and Belhadj
(2007)(120) in sediments from the Algerian Mediterranean Sea sediments which were in the
range (2.6-18.9µgg-1), which considered unpolluted with Cr. The results were in agreement
with the results obtained by Diaz de Alba et al. (2011)(463) for Algeciras bay, Spain and higher
than reported by Gargouri et al. (2010)(480) for the Coast of Safax, Taranto Gulf, Italy
(Buccolieri et al. (2006)(160), Nables Harbour, Italy (Adamo et al., 2005)(184), El-Mex Bay,
Egypt reported by Abdallah et al. (2007a)(292) and lower than the range reported for the
western harbour reported by Mostafa et al. (2004)(298) (Table 15).
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Total Nickel (TNi)
Figure 24 showed that the average Ni levels (µgg-1 dry weight) measured for the three
regions were in the following sequence: (49.48±8.64; Middle region) > (37.61±13.88; eastern
region) > (8.81±5.72, western region). There was large variability in the level of Ni among
sites. The highest concentration of Ni (60.25µgg-1) was observed at station 17 (El-Gamil
west; Middle region) and the lowest concentration of Ni (1.65µgg-1) was found at station 2
(Baghoush; western region). These results are in agreement with those of Gargouri et al.
(2010)(480) for the coast of Sfax, Tunisia, which was in the range (7-55µgg-1) and Alomary and
Belhadj (2007)(120) for the Algerian Mediterranean Sea (0.8-54.9µgg-1) Table 15. Table 14
showed that stations 13, 16, 17 (Rashid east, Ras El-Burr, El-Gamil west; Middle region) are
considered polluted with Ni when compared to mean sediment (52µgg-1) reported by
Salomons and Forstner (1984)(181) and the average continental crust (56µgg-1) reported by
Wedepohl (1995)(473). On the other hand, all sediments samples considered as unpolluted
with Ni when compared with the crustal average (75µgg-1) (Taylor, 1964)(470) and the average
shale (68µgg-1) (Turekian and Wedepohl, 1961)(469).
Total Zinc (TZn)
The average Zn levels (µgg-1 dry weight) at the studied regions were in the following
sequence: the western region (23.43±20.07) > the middle region (21.90±6.81) > the eastern
region (16.33±19.40) as shown in Figure 24. There were slight variations in the levels of Zn
at the three regions. High concentration of Zn (62.21µgg-1) was observed at station 7 (NIOF;
the western region). Zn was observed to be lower than the crustal average reported by Taylor
(1964)(470), the upper crust (Taylor and Mclennan, 1995)(472), Mean crust (Bowen, 1979)(471),
Average shale (Turekian and Wedepohl, 1961)(469), Mean sediment (Salomons and Forstner,
1984)(181) and the average continental crust reported by Wedepohl (1995)(473) at all stations
(Table 14). The levels of Zn in the sediments are within the permissible levels which indicate
its normal concentration and reflect the background value in sediments. These results are in
agreement with the results reported for Algerian Mediterranean Sea sediments (5.3-48.7 µgg1
) (Alomoray and Belhadj, 2007)(120) and lower than the surface marine sediments of Sfax
coast, Tunisia (39-117µgg-1) reported by Gargouri et al. (2010)(480), the Pasajes harbour, Spain
(447-1390µgg-1) reported by Legorburu and Canton (1991)(479), Taranto Gulf, Italy (86.8129µgg-1) reported by Buccolieri et al. (2006)(160), Nables harbour (41-1196µgg-1), Italy
(Adamo et al., 2005)(184), Themaikos Gulf (84-537µgg-1), Greece (Violontzis, 2009)(481),
Mediterranean sea (29.4-509.3µgg-1), France (Fernex et al., 2001)(486) and Sardina, Italy (1983239µgg-1) reported by Caredda et al. (1999)(487) and higher than reported by Sabhi et al.
(2000)(488) (9.2-37.7µgg-1) for the Moroccan Mediterranean coast (Table 15).
Total Lead (TPb)
Data obtained at Figure 24 showed that the average Pb levels (µgg-1 dry weight)
measured for the three regions were in the following sequence: (western region; 18.93±13.73)
> (Middle region; 6.70±1.37) > (Eastern region; 5.01±2.35). The maximum concentration of
lead (53.67µgg-1) reported at station 5 (El-Mex; western region). The elevated concentrations
of Pb observed in EL-Mex Bay may be related to the release of Pb through the previous
discharge from the Chlor-Alkali plant, discharge of freshwater from EL-Umum Drain
carrying industrial, agricultural and sewage effluents and atmospheric deposition, where Pb is
released from the combustion of gasoline from motor vehicles (Masoud et al., 2007)(294).
Abdel-Moati, (2001)(489) pointed out that 44 % of the TPb enters EL-Mex Bay from industrial
wastes, 33 % from agricultural wastes discharged via EL-Umum Drain by the EL-Mex
pumping station and 23 % from sewage effluents. Lead enters the bay either in the dissolved
or particulate form (Saad et al., 2001)(490). An important source of lead in EL-Mex Bay is
from the Portland Cement Factory (in Wadi EL-Kamar). Lead content was generally high in
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the surface of the western harbor. There is a relationship between dispersed oil from ships
and lead enrichment (Saad et al., 2004)(91). Table 15 showed that stations 5, 6, 7, 8 at El-Mex
Bay, Western harbour, NIOF and Eastern harbor respectively, in the western region are
considered as polluted with Pb when compared to upper crust (20µgg-1) (Taylor and
Mclennan,1995)(472), average shale (20 µgg-1) (Turekian and Wedepohl, 1961)(469) and the
Mean sediment (19µgg-1) (Salomons and Forstner, 1984)(181). The level of Pb in sediments of
the Middle and the Eastern region were agreed with the levels which indicate its normal
concentrations and reflects the background value concentrations in sediments. Sadiq et al.
(2003)(491) reported that low concentrations of lead still might pose a threat to life in a marine
environment in comparison with other heavy metals.
Table 15 showed that the values obtained in this study are in agreement with the
results of Alomary and Belhadj for the coast of Algeria (2007)(120) and lower than the results
obtained by Gargouri et al. (2010)(480) for the coast of Sfax, (Caredda et al. (1999)(487) for
Sardina, Italy, Fernex et al. (2001)(486) for Mediterranean Sea, France, Feldstein et al.
(2003)(492) and higher than the results reported by Sabhi et al. (2000)(488) for the Moroccan
Mediterranean coast.
Total Cupper (TCu)
Anthropogenic sources of copper are primarily related to textile production, marine
antifouling agents, pipes and copper based fungicides or pesticides (PIO, 2005)(493).
The average concentration of Cu (µgg-1 dry weight) for the three regions were as
follows: the Middle region (9.86±4.47)> the Western region (7.93±7.36) > the Eastern region
(6.48±6.94) as represented in Figure 25. Cupper values ranged from 0.46 to 26.26µgg-1 with
an average of 8.46µgg-1. Table 14 showed that the concentration of Cu in the study area was
below the crustal average reported by Taylor (1964)(470) and the mean crust (50µgg-1) (Bowen,
1979)(471) and the average shale (45µgg-1) (Turekian and Wedepohl, 1961)(469) and the mean
sediment (33µgg-1) (Salomons and Forstner, 1984)(181) at all stations. On the other hand, Cu
concentrations is lower than the upper curst and the average continental crust (25µgg-1)
reported by Taylor and Mclennan (1995)(472) and (Wedepohl, 1995)(473) respectively, at all
stations except at station 7 (NIOF) in the western regon (26.26µgg-1). Therefore, the
investigated samples are considered unpolluted with Cu. These results are in agreement with
that obtained by Alomary and Elhadj (2007)(120) for the Algerian Mediterranean Sediments
and the results of Gargouri et al. (2010)(480) for Safx coast and Feldsterin (2003)(492) and lower
than the results obtained by Benamar et al. (1999)(494) for Algiers Bay, Algeria and higher
than that reported for the Moroccan Mediterranean coast (Sabhi et al., 2000)(488) (Table 15).
Total Cobalt (TCo)
As shown from data in Figure 25, the average Co concentration (µgg-1 dry weight)
measured for the three regions were in the following sequence: the middle region
(16.96±6.91)> the eastern region (9.50±10.71)> the western region (2.46±1.89). The highest
value (26.39µgg-1) found at station 16 (Ras El-Barr) and the lowest value (0.43µgg-1) found at
station 2 (Baghoush; western region) when compared to the Co background concentrations
(Table 14), the studied surface sediments were found not to be contaminated by Co.
However, slight contamination is to be considered in station 16 in Ras El-Barr. These results
are in agreement with the results of Adamo et al. (2005)(184) for Nables harbour, Italy and
lower than that results obtained by Khairy (2008)(226) for Abu Qir Bay, Egypt (Table 15).
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Total Cadmium (TCd)
Cadmium is presented relatively rare in the earth with the average concentration of
0.2µgg-1 in earth's crust (Mason and Moore, 1991)(495). In present study, the concentration of
Cd was in the range of 0.04-0.47 µgg-1 (dry weight), the lowest concentration (0.04µgg-1)
observed at station 15 (New Damietta; Middle region) and station 19 (Port Said; Eastern
region) and the highest concentration (0.47µgg-1) observed at station 6 (Western harbour;
western region). Mostafa et al. (2004)(298) stated that the presence of high concentrations of
Cd in the sediments of the western harbour appears to be directly related to urban and
industrial runoff discharged from the El-Mahmoudiya Canal. The mean Cd concentrations for
the three regions can be arranged in the following sequence: the western region (0.29±0.15
µgg-1) > the Middle region (0.12±0.07µgg-1) > the eastern region (0.11±0.10µgg-1). In former
studies of Cd concentration in marine sediments, the concentration of Cd in unpolluted areas
were within the range of 1.56-8.0 µgg-1 (Buykx et al., 2000; Santamaria-Fernandez et al.
2005; Svete et al. 2001)(475-477). The concentrations of Cd in the Mediterranean Sea sediments
ranged between 0.1 and 2.3 µgg-1 dry weight (UNEP, 1978)(496). Donazzolo et al. (1984)(497)
reported probable background cadmium concentrations, as calculated from core samples, of
1.2 µgg-1 dry weight. A probable background cadmium concentration should be in the range
of 0.1 to 2.5 µgg-1 dry weight (Sabhi et al., 2000)(488). Frignani and Giordani (1983)(498)
quote a value of 0.4 µgg-1 dry weight for Aegean Sea sediments. Thus the investigated
sediments are considered unpolluted by Cd. When compared with former studies on the other
Mediterranean regions (Table 15), the concentration of Cd of the present study was lower
than reported by Caredda et al. (1999)(487), Jdid et al. (1999)(499), Fernex et al. (2001)(486),
Feldstein et al. (2003)(492) and are in the range reported by Sabhi et al. (2000)(488).
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30000
Fe in µgg-1
25000
20000
15000
10000
5000
0
Western
Middle
Eastern
600
Mn in µgg-1
500
400
300
200
100
0
Western
Middle
Eastern
Western
Middle
Eastern
200
Cr in µgg-1
150
100
50
0
Figure 23: The average of total Fe, Mn and Cr levels in µgg-1 dry weight at the studied
regions
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50
Ni in µgg-1
40
30
20
10
0
Western
Middle
Eastern
Western
Middle
Eastern
Western
Middle
Eastern
25
Zn inµgg-1
20
15
10
5
0
20
Pb in µgg-1
15
10
5
0
Figure 24: The average of total Ni, Zn and Pb levels in µgg-1 dry weight at the studied
regions
-156-
10
Cu in µgg-1
8
6
4
2
0
Western
Middle
Eastern
Western
Middle
Eastern
Co in µgg-1
20
15
10
5
0
0.3
Cd in µgg-1
0.25
0.2
0.15
0.1
0.05
0
Western
Middle
Eastern
Figure 25: The average of total Cu, Co and Cd levels in µgg-1 dry weight at the studied
regions
-157-
Table 14. Average concentrations (µgg-1) of total trace metals in sediments along the Egyptian Mediterranean coastal area compared with
noncontaminated sediments
Fe
Mn
Zn
Cu
Ni
Cr
Co
Pb
Cd
Present study
243.48
17.25
2.05
0.46
1.65
4.08
0.43
3.34
0.04
Min. conc.
38045.05 1085.72 62.21
26.26
60.25 297.95 26.39
53.67
0.47
Max. conc.
13255.69 380.71
22.19
8.46
25.93 82.74
8.24
13.17
0.22
Average conc.
12911.36 305.38
15.84
6.22
20.96 90.18
8.40
11.90
0.15
SD (N=20)
a
56300
950
70
55
75
100
25
12.5
0.2
Crustal average
35000
600
71
25
20
35
10
20
0.098
Upper crustb
41000
950
75
50
100
14
0.11
Mean crustc
47000
850
95
45
68
90
19
20
0.3
Average shaled
e
41000
770
95
33
52
72
14
19
0.17
Mean sediment
43200
65
25
56
35
14.8
0.1
Average Continental crustf
a Taylor (1964)(470)
b Taylor and Mclennan (1995)(472)
c Bowen (1979)(471)
d Turekian and Wedepohl (1961)(469)
e Salomons and Forstner (1984)(181, 473)
f Wedepohl (1995)(473)
-158-
Table 15. Concentrations of trace metals (µgg-1) determined in sediments of the present study compared to other parts along the
Mediterranean coast.
Location
Cd
Cu
Cr
Co
Fe
Mn
Ni
Zn
Pb
The present study 0.04-0.47 0.46-26.3 4.1-297.95 0.43-26.39 243.5-38045.1 17.3-1086 1.7-60.25 2.05-62.21 3.34-53.67
Western harbour,
Egypt(298)
Western harbour,
Egypt(500)
Pasajes harbour,
Spain(479)
Coast of
Safax(480)
Taranto Gulf,
Italy(160)
Nables Harbour,
Italy(184)
El-Mex Bay,
Egypt(292)
Algeciras Bay,
Spain(463)
Mediterranean
Sea, Algeria(120)
Mediterranean
Sea, France(486)
Sardina, Italy(487)
0.61-2.44
39.207
33-649
NA
8819-36140
139-317
13-53
58.5-382
38-1070
0.07-0.64
0.3-19
NA
NA
5000-29000
15-40
NA
0.23-4.74
NA
1.2-6.64
25-3726
NA
NA
3830-35500
64-365
17-99
477-1390
45-346
5.5-7
13-29
41-82
NA
41011-50163
NA
7-55
39-117
18-88
NA
42.4-52.3
75.2-102.8
NA
26313-36098
552-2826
47.9-60.7
86.8-129
44.7-74.8
0.2-2.5
40-415
10.3-161.8
1.9-7.2
NA
95-535
NA
41-1196
37-314
4.3-11.9
11.5-52.2
27.6-114.9
NA
709-3063
132-592.3
NA
89.3-592.3
NA
0.1-22
5-25
30-251
5-22
18285-42756
235-967
19-144
33-117
12-39
0.1-2.3
1.1-10.4
2.6-18.9
NA
NA
NA
0.8-54.9
5.3-45.7
1.3-11.5
0.15-3.4
14.82.6
NA
NA
NA
NA
NA
29.4-509.3
20.1-393.6
0.21-13.4
2.77-51.3
NA
NA
NA
NA
0.9-51.3
198-3239
74-772
Abu Qir bay,
Egypt(226)
Themaikos Gulf,
Greece(481)
Moroccan Med.
Coast(488)
NA: Not available
0.08-3.54
0.95-127
4.66-186
0.727-22.8
1120-37100
58.6-582
2.05-40.5
5.77-717
2.71-129
0.3-8.4
32-130
21-470
NA
NA
590-890
63-130
84-537
38-190
0.03-0.25
1.2-6.7
NA
NA
NA
NA
NA
9.2-37.7
0.175-0.25
-159-
6.2.1.1. Assessment of sediment quality
Sediment quality guidelines (SQG)
Pollution of the natural environment by trace metals is a worldwide problem. Trace
metals from natural and anthropogenic sources continously enter the aquatic ecosystem where
they pose a serious threat because of their toxicity, long time persistence and bioaccumulaiton
(Chon et al., 2010)(501). Sediment is more conservative than water, as it accumulates historical
data on processes within water bodies and the effect of anthropogenic factors on these
processes. For these reasons, sediment quality parameters have been used as environmental
indicators and their ability to trace and monitor contamination sources is largely recognized.
Sediment show a high capacity to accumulate and eventually integrate the low concentrations
of trace metals usually found in water (Vallejuelo et al., 2010)(502).
Sediments can act as both a source and a sink for potential toxic compounds. In order
to predict adverse biological effects in contaminated sediments, numerous Sediment Quality
Guidelines (SQGs) have been developed over the past decade (McCready et al., 2006;
MacDonald et al., 2000)(201, 503), in order to protect aquatic organisms living in or near the
sediments from the toxic effects associated with sediment-bound contaminats. They include
sediment quality criteria, sediment quality objectives and sediment quality standards. These
guidelines are useful for the evaluation of spatial variations of sediment contamination, the
classification of the contamination state of the sediments, the design of monitioring programs,
interpretation of historical data, and for environmental assessments for future remedial
actions (MacDonald et al., 2000)(503).
Numerous sediment quality guidelines are used to protect aquatic biota from the
harmful and toxic effects related with sediment bound contaminants (McCready et al.,
2006)(201). These guidelines evaluate the degree to which the sediment-associated chemical
status might adversely affect aquatic organisms and are designed for the interpretation of
sediment quality. They are also used to rank and prioritize contaminated areas for further
investigation (Diaz-de Alba et al., 2011)(463).
SQGs are developed using a variety of approaches, such as effect-rang approach,
effect-level approach and apparent effect-threshold approach. The selection of the most
appropriate SQGs is not trivial; each derived numerical value may differ significantly based
on the derivation procedure, the objective of the calculation and the ability of the Guidelines
to match the selected geological background of the specific area (MacDonald et al., 2000)(503).
Based on the SQGs proposed by USEPA (Perin et al., 1997)(504), sediments were
categorized into three classes; nonpolluted, moderately polluted and heavily polluted. The
chemical contamination in the sediments was evaluated by comparison with the sediment
quality guidelines proposed by USEPA. These criteria are shown in Table 16. According to
US EPA, the concentration of Cd at all stations under investigation was belonged to
unpolluted sediments. Figure 26 showed that the concentration of Fe at stations (17 and 18)
in El-Gamil west and east was moderately polluted while stations (13, 14, 15, 16 and 19) in
Rashid east, Burullus, New Damietta, Ras El-Barr and Port Said respectively were highly
polluted. Figure 26 also showed that the concentration of Cr at stations (1, 7, 10, 11 and 20) in
El-Salloum, NIOF, Electric power station, Maadia and Rafah, respectively was moderately
polluted. Stations (9, 12, 13, 14, 15, 16, 17, 18 and 19) in Abu Qir Bay, Rashid west, Rashid
east, Burullus, New Damietta, Ras El-Barr, El-Gamil west, El-Gamil east and Port Said
respectively were heavily polluted. The concentration of Mn at stations (9, 10, 11, 17 and 18)
in Abu Qir Bay, Electric power station, Maadia, El-Gamil west and east respectively was
moderately polluted while at stations (13, 14, 15, 16 and 19) in Rashid east, Burullus, New
-160-
Damietta, Ras El-Barr and Port Said was heavily polluted. Finally from the obtained data in
Figure 27, it was observed that the concentration of Zn at all stations was belonged to
unpolluted sediments. Sediment at all stations was classified as not toxic with cupper except
station 7 in NIOF which was moderately polluted with Cu. While, Pb at all stations was
belonged to unpolluted except for station 5 in El-Mex which was moderately polluted. It was
noted that the concentration of Ni at stations (12, 14, 15, 18, 19 and 20) in Rashid west,
Burullus, New Damietta, El-Gamil east, Port Said and Rafah, respectively was belonged to
moderately polluted sediments. Also Figure 27 showed that Ni at stations (13, 16 and 17) in
Rashid east, Ras El-Barr and El-Gamil west, respectively was belonged to heavily polluted.
Another method used in this study to assess the impacts of trace metals in the
Egyptian Mediterranean coastal area was the Soil and Aquatic Sediment Guidelines and
Standards issued by the New York State Department of Environmental Conservation
(NYSDEC) (1999)(505). The total trace metal concentrations are compared with the two
benchmarks of lowest effect level (LEL) and severe effect level (SEL), which signifies
chronic, long term impacts of contamination to benthic organisms (Table 16). Macro benthic
communities are considered good indicators of ecosystem health as benthos shows the real
effects of pollution on the communities, being an integrator of the recent pollution history in
the sediment and of different kinds of pollutants, which can be synergistic (OcchipintiAmbrogi and Forni, 2004)(506).
If trace metal concentration in the sediment is lower than LEL, it would be tolerated
by the majority of benthic organisms. If the concentration is more than the LEL but lower
than the SEL concentration, the sediments is considered to be contaminated, but with
moderate impacts on benthic life. If SEL, is exceed; this indicates a clear disturbance to the
sediment-dwelling benthic communities and would be detrimental to the majority of them. If
all the heavy metal concentrations are higher than SEL it could severely impact health of this
biota (Graney and Erisen, 2004)(507).
From the obtained data at table 5, it was observed that Cd was less than the LEL at all
stations of the present study, which mean that the sediments would be tolerated by the
majority of benthic organisms. The concentration of Cr at stations (1, 9, 10, 11, 15, 16 and
20) in El-Salloum, Abu Qir Bay, Electric power station, Maadia, New Damietta, Ras El-Barr
and Rafah, respectively was higher than the LEL but lower than the SEL which mean that the
sediment is considered to be contaminated, but with moderate impacts on benthic life. While,
the concentration of Cr at stations (13, 14, 17, 18 and 19) in Rashid east, Busullus, El-Gamil
West and El-Gamil east and Port Said, respectively exceeded the SEL concentration, this
indicates a clear disturbance to the sediment-dwelling benthic communities and would be
detrimental to the majority of them
The concentration of Cu at most of the stations under investigation was lower than
LEL which means that it would be tolerated by the majority of benthic organisms. While, the
concentration of Cu at station 7 in NIOF and station 16 in Ras El-Barr were higher than the
LEL and lower than the SEL which mean that the sediment was considered to be
contaminated, but with moderate impacts on benthic life. Results showed that the
concentration of Pb at all stations of the Middle and the eastern region were lower than the
LEL which means that it would tolerated by the majority of benthic organisms. While, the
concentration of Pb at station 5 in El-mex Bay was higher than the LEL and lower than the
SEL which mean that the sediment was considered to be contaminated, but with moderate
impacts on benthic life. Stations (1, 6, 7 and 8) in El-Salloum, Western harbour, NIOF and
Eastern harbour was higher than the LEL but lower than which means that the sediment was
-161-
considered to be contaminated, but with moderate impacts on benthic life. It was observed
that Zn concentration was less than the LEL at all stations of the present study, which mean
that the sediments would be tolerated by the majority of benthic organisms.
SQGs are developed using a variety of approaches, such as effect rang approach,
effect –level approach and apparent effect threshold approach. The selection of the most
appropriate SQGs is based on the derivation procedure, the objective of the calculation and
the ability of the Guidelines to match the selected geological background of the specific area
(Macdonald et al., 2000)(503). The most widely used SQGs for marine sediment samples, have
been developed by the U.S National Oceanic and Atmospheric Administration (NOAA) and
they include sets of effect-range guidelines derived from a large series of chemical and
biological data collected from North American coastal regions that incorporate field and
laboratory data from many different methodologies, chemical and biological species
(McCready et al., 2006)(201).
The data set was compared with the sediment quality guidelines suggested by Long et
al. (1995)(199) on the basis of the potential to induce toxic effects in marine organisms. The
guidelines define two values ERL (effects range low) and ERM (effects range median). ERL
values indicate concentrations below which adverse effects rarely occur, and ERM values
represent concentrations above which effects frequently occur.
Chemical concentrations corresponding to the 10th and 50th percentiles of adverse
biological effects were called the Effects-rang-low (ERL) and Effects-range-median (ERM)
respectively. The NOAA guidelines provide two values for each chemical, classifying the
sediment either rarely (<ERL), occasionally (≥ERL and <ERM) or frequently associated with
adverse biological effects (Macdonald et al., 2000)(503).
Comparing results of the present study with ERL and ERM values (Figure 28 and 29),
it was observed that Cd, Cu, and Zn at all stations are below the ERL value (1.2, 34, and 150
µgg-1), respectively which indicate that these metals in sediments at all stations of the study
area are not likely to have adverse effects on animals that live in sediment. Only one station
(El-Mex) which had a Pb concentration > ERL, indicated that Pb at El-Mex will likely to has
effects on animals that live in this sediment. On the other hand, all the rest of the studied
station had a concentration of Pb below the ERL value which indicates that Pb in the study
area is not likely to have adverse effects on animals that live in sediments except station 5 in
El-Mex. On the other hand, Ni at stations 12, 14, 15, 18, 19, and 20 in (Rashid west,
Burullus, New Damietta, El-Gamil east, Port Said and Rafah) had a value over the ERL value
(36.384, 43.545, 44.305, 48.93, 47.415, and 27.79), respectively. This reflects that the adverse
effects on animals live at these stations are frequently occured. Stations 13, 16, and 17 in
(Rashid east, Ras El-Barr and El-Gamil west), respectively had concentration of Ni above the
ERM value (56.536, 56.413, and 60.246) which means that Ni probably has adverse effects
on animals live in this sediment.
Concentrations of (Zn, Cu, Ni, Pb and Cd) were evaluated in a screening level
ecological risk assessment, by comparing to numerical SQGs such as TEL and PEL. It is
interpreted that TEL as the concentrations, bellow which adverse biological effects rarely
occur. Hence, it is considered to provide a high level of protection for aquatic organisms.
Similarly, PEL interpreted as the concentrations above which adverse biological effects
frequently occur. Hence, it is considered to provide a lower level of protection for aquatic
organisms. When compared to the TEL-PEL SQGs, the concentrations of Cd and Zn are
lower than the TEL value at 100% of the sampling stations, while Pb and Cu showed values
-162-
lower than the TEL at 95% of the sampling stations. On the other hand, in case of Ni, 20% of
samples fall in the range between TEL and PEL at Abu Qir, Electric power station, Rashid
west and Rafah indicated that associated adverse biological effects may occasionally occur.
About 35% of samples are higher than PEL values at Rashid east, Ras El-Barr, New
Damietta, Burrllus, El-Gamil east and west and Port Said.
To assess metal concentration in sediments, Numerical Sediment Quality Guidelines
(SQG) were applied. SQG include a threshold effect concentration (TEC) and a probable
effect concentration (PEC). The values of consensus-based TECs and consensus-based PECs
(MacDonald et al., 2000)(503) are shown in Table 16.
If the metals in sediments are below the TEC, harmful effects are unlikely to be
observed. If the metals are above the PEC, harmful effects are likely to be observed. Most of
the TECs provide an accurate basis for predicting the absence of sediment toxicity
(MacDonald et al., 2000)(503). The concentrations of Cd and Cu at all sediment samples are
lower than the proposed TECs indicated that there are no harmful effects from these metals.
On the other hand, the concentrations of Ni exceed TECs at stations (12, 13, 14, 15, 19 and
20) in Rashid west, Rashid east, Burullus, New Damietta, Port Said and Rafah respectively,
while stations (16 and 17) in Ras El-Barr and El-Gamil west are above the PECs. At stations
(9, 12, 15, 16 and 20) in Abu Qir Bay, Rashid west, New damietta, Ras El-barr and Rafah,
respectively the concentrations of Cr are above the TECs and at stations (13, 14, 17, 18, 19) in
Rashid east, Burullus, El-Gamil west, El-Gamil east and Port Said, respectively the
concentration of Cr exceeded the PECs. Station 5 in El-Mex Bay showed concentration of Pb
exceeded the PECs. This indicated that these stations were in potential risk.
In order to obtain a more realistic measure of predicted toxicity that simply summing
the numbers of ERMs/PELs exceeded, mean ERM or PEL quotients (ERMQ PELQ) were
calculated as follows:
ERMQ or PELQ =  [Ci/(ERMi or PELi)] / n
Where Ci is the concentration of element i in sediments, ERMi, PELi the guidelines values for
the element i and n is the number of metals. Mean quotients are considered as useful tools for
reducing a large amount of contaminants into a single number. By calculating mean quotients
it is assumed that adverse effects to marine organisms caused by individual chemicals are
additional limitation is that this approach does not consider all the chemicals present in
sediments but only those include in the SQG list (Anders et al., 2005)(508). Mean quotients
can be used to identify, delineate and prioritize areas of potential concern with respect to
quality of sediments (Chapman and Mann, 1999)(509). ERMQ values of <0.1, 0.11-0.5, 0.51.5 and >1.5 related to 12%, 30%, 46% and 74% likehood, respectively, that sediments
present toxicity in amphipod survival bioassays. Similarly, PELQ values of <0.1, 0.11-1.5,
1.51-2.3 and >2.3 coincide with 10%, 25.5, 50% and 76% likehood of toxicity, respectively
(Long et al., 1995)(199). Consequently, four relative levels of priority (highly toxic, medium
toxic, slightly toxic and non toxic) have been proposed. The mean ERM and PEL quotients
calculated for the sampling sites (based on metals Cd, Cr, Cu, Ni, Pb and Zn) are shown in
Figure 30 and 31 and Table 17, only two stations (station 7; NIOF and station 9; Abu Qir
Bay) in the western region have ERLQ (0.11-1.5) and are categorized as slightly toxic and the
rest of the stations in this region are categorized as not toxic. While, all stations of the middle
and the eastern region are categorized as slightly toxic. Stations (2, 3 and 6) in Baghous, ElNobarreya and Western harbour, respectively in the western region have PELQ< 0.1 and are
categorized as not toxic while the other stations have PELq (0.11-1.5) and categorized as
slightly toxic.
-163-
Table 16: Comparison of the surfacial sediments from the Egyptian Mediterranean coast with Sediment quality guidelines
U S NOAA's
NYSDEC
US DOE
USEPA guidlines
Metal TEL
PEL ERL ERM LAL HAL LEL SEL TEC PEC
Present work
Not
Moderately
Heavily
polluted
polluted
polluted
0.68
4.21
1.2
9.6
0.04
9.6
1.2
9.6
0.99
4.98
NA
NA
>6
0.04-0.47
Cd
NA
NA
NA
NA
0.5
120
NA
NA
NA
NA
NA
NA
NA
0.43-26.39
Co
52.3
160
81
370
4
370
26
110 43.4
111
<25
25-75
>75
4.08-297.95
Cr
18.7
108
34
270
2
270
16
110 31.6
149
<25
25-50
>50
0.46-26.26
Cu
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
<17000 17000-25000
>25000
243.48-38045.05
Fe
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
<300
300-500
>500
17.25-1085.72
Mn
15.9
42.8
21
52
3
50
NA
NA
NA
NA
<20
20-50
>50
1.65-60.25
Ni
30.2
112
47
220
2
218
16
50
35.8
128
<40
40-60
>60
3.34-53.67
Pb
124
271
150
410
5
410
120
270
121
459
<90
90-200
>200
2.05-62.21
Zn
(505)
Lowest effect level (LEL)
Severe effect level (SEL)(505)
Effects rang low (ERL)(199)
Effects range median (ERM)(199)
Threshold effect concentration (TEC)(510)
Probable effect concentration (PEC)(510)
Low alert level (LAL)(511)
High Alert Level (HAL)(512)
Threshold effect level (TEL)(510)
Probable effect level (PEL)(510)
-164-
40000
35000
Fe (µgg-1)
30000
Heavily polluted
25000
20000
Moderately polluted
15000
10000
Not polluted
5000
0
1
2
3
4
5
6
7
8
9 10 11 12 13 14 15 16 17 18 19 20
1200
Mn (µgg1-)
1000
800
Highly polluted
600
Moderately polluted
400
200
Not polluted
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
350
Cr (µgg1-)
300
250
200
150
Highly polluted
100
50
Moderately polluted
Not polluted
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
Stations
Figure 26: Concentration of Fe, Mn and Cr, in sediments along the Egyptian
Mediterranean coastal area comparable to US SQG.
-165-
70
60
Highly polluted
Ni (µgg1-)
50
40
Moderately polluted
30
20
Not polluted
10
0
1
2
3 4
5 6
7
8 9 10 11 12 13 14 15 16 17 18 19 20
60
Heavily polluted
Pb (µgg1-)
50
40
Moderately polluted
30
20
Not polluted
10
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
Heavily polluted
60
Cu (µgg-1)
50
40
Moderately polluted
30
20
Not polluted
10
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
Stations
Figure 27: Concentration of Ni, Pb and Cu, in sediments along the Egyptian
Mediterranean coastal area comparable to US SQG.
-166-
Cr (µgg1-)
400
350
300
250
200
150
100
50
0
> ERM
ERM
ERL
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
70
60
> ERM
Ni (µgg1-)
50
ERM
40
30
20
ERL
10
0
Zn (µgg1-)
1
2
3 4
5 6
7
8 9 10 11 12 13 14 15 16 17 18 19 20
450
400
350
300
250
200
150
100
50
0
> ERM
ERM
ERL
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
Stations
Figure 28: Concentration of Cr, Ni and Zn, in sediments along the Egyptian
Mediterranean coastal area comparable to US NOAA's.
-167-
> ERM
250
Pb (µgg1-)
200
ERM
150
100
50
ERL
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
300
> ERM
Cu (µgg-1)
250
200
150
ERM
100
50
> ERM
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
Stations
Figure 29: Concentration of Pb and Cu in sediments along the Egyptian Mediterranean
coastal area comparable to US NOAA's.
-168-
Highly toxic
1.6
1.4
ERMq
1.2
1
Moderate toxic
0.8
0.6
0.4
Slightly toxic
0.2
Non toxic
0
1
2
3
4
5
6
7
8
9 10 11 12 13 14 15 16 17 18 19 20
S tation
Figure 30: Estimated mean ERMq of surface sediments along the Egyptian
Mediterranean coast
Highly toxic
2.5
PELq
2
Moderate toxic
1.5
1
Slightly toxic
0.5
Non toxic
0
1
2
3
4
5
6
7
8
9 10 11 12 13 14 15 16 17 18 19 20
S tation
Figure 31: Estimated mean PELq of surface sediments along the Egyptian
Mediterranean coast
-169-
Table 17: Estimated mean ERMq and PELq of surface sediments along the Egyptian
Mediterranean coastal area
Region
Western region
Middle region
Eastern region
Station
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
ERMq
0.08
0.01
0.05
0.07
0.10
0.04
0.11
0.05
0.12
0.09
0.07
0.17
0.34
0.21
0.20
0.26
0.32
0.23
0.31
0.13
PELq
0.13
0.02
0.08
0.12
0.18
0.09
0.21
0.11
0.14
0.14
0.11
0.20
0.31
0.26
0.27
0.33
0.33
0.29
0.28
0.15
The National Standard of China (NSC) GB18668-2002 (SEPA, 2002)(513) has defined
three grades of marine sediment, in which the content of some heavy metals is regarded as
parameters used to classify marine sediments quality (Table 18). According to this criterion,
the first class quality is suitable for mariculture, nature reserves, and endangered species
reserves, and leisure activities such as swimming; the second class quality can be used for
industry and tourism sites; and the third class can only be used for harbors. The contents of
Cd, Cu, Pb and Zn at all samples are lower than the upper limit of the first class criteria of
NSC GB 18668-2002.
In Hong Kong, a stricter criterion has been to distinguished sediment quality (Lau et
al., 1993)(514). The contents of Cd at station 6, 7 and 11 in the western harbour, NIOF and
Maadia, respectively at the western region to some extent, higher than the value which
regarded as the upper limit of the desired quality for fairly clean sediments but, lower than the
threshold values which are used to indicate moderately contaminated sediments (Table 18).
Also, the content of Cu in station 7 in NIOF at the western region and the content of total Ni
at station 20 in Rafah at the eastern region considered to some extent, higher than the value
which regarded as the upper limit of the desired quality for fairly clean sediments that is close
to background levels. The content of Ni in station 12 in Rashid west (middle region)
considered to some extent higher than the value which indicated that the sediment is
moderately contaminated but, lower than the threshold values which are used to indicate
heavily contaminated sediments, while stations 13, 14, 15, 16, 17, and 18 in Rashied east,
Burrullus, New Damietta, Ras El-Burr, El-Gamil west and El-Gamil east (middle region) and
station 20 in Rafah (eastern region) considered to higher than value which indicated heavily
polluted sediment. The content of Pb at station 5 in El-Mex (western harbour) considered to
some extent, higher than the value which regarded as the upper limit of the desired quality for
fairly clean sediments that is close to background levels but, lower than the threshold values
-170-
which are used to indicate moderately contaminated sediments. The contents of Zn in all
samples are lower than the values which are regarded as the upper limit of the desired quality
for fairly clean sediments. That is close to background levels.
Comparing the sediment of the present study with classification system from Hong
Kong environmental Protection Department (EPD) Classification system. The value of the
Mean Cupper, Zinc, and Lead concentrations showed to be calssified as uncontaminated
sediment (Class A). Wherease, Nickel and Cadmium represented (Class B) contaminations.
Sediments were considered as seriously contaminated (Class D) when comparing the mean
concentration of Chromium with the classification system adopted by the Hong Kong
Government (EPD,1992)(515) (Table 19).
Table 18: The metal threshold values of some different criteria used to distinguish
marine sediment quality (µgg1-)
Class I
Class II
Class III
Target
Trigger
Action
Cd
0.5
1.5
5
Cu
35
100
200
0.4
1.0
1.5
0.22
20
55
65
8.46
Ni
Pb
60
130
250
Zn
150
350
600
20
35
75
35
65
150
40
75
200
25.93 13.17 22.19
Reference
(SEPA, 2002)(513)
(SEPA, 2002)(513)
(SEPA, 2002)(513)
(Lau et al., 1993)(514)
(Lau et al., 1993)(514)
(Lau et al., 1993)(514)
The present study
Class I; suitable for mariculture, nature reserves, endangered species reserves and leisure activities like
swimming.
Class II; quality can be used for indursty and tourism activities
Class III; can only be used for harbours
Target; fairly clean sediment that is closed to background
Trigger; moderately contaminated sediment
Action; heavily polluted sediment
Table 19: Classification of sediments according to the background and classification
system from the Hong Kong environmental Protection Department (EPD, 1992)
Class
A
B
C
D
Degree of contamintion
Uncontaminated
Slightly contaminated
Moderately contaminated
Seriously contaminated
The present study
Cu
<10
10-41
55-64
>64
8.46
Zn
<70
70-150
150-200
>200
22.19
-171-
Cr
<25
25-50
50-80
>80
82.74
Ni
<15
15-35
35-40
>1.5
25.93
Cd
<0.1
0.1-1.0
1.0-1.5
>1.5
0.22
Pb
<25
25-65
65-75
>75
13.17
6.2.1.2. Estimation of Pollution impact
A number of calculation methods have been put forward for quantifying the degree of
metal enrichment in sediments. Various authors (e.g. Muller, 1969; Hakanson, 1980)(516, 517)
have proposed pollution impact scales (or ranges) to convert the calculated numerical results
into broad descriptive bands of pollution ranging from low to high intensity (Abrahim and
packer, 2008)(518). Five methods are discussed in the following sections along with proposed
modifications.
6.2.1.2.1. Enrichment factor (EF)
Natural weathering and human activities are known as metals sources for coastal
environment. However, the relative influence of natural and anthropogenic sources on the
geochemistry of marine sediments is not always clear, especially in the areas where the local
minerals contain high natural concentrations (Valdes et al., 2005)(519). Therefore, for a better
assessment of contamination process in the marine coastal environment, it is important to be
able to distinguish between natural and human related metal enrichments in sediments. Thus,
for a meaningful comparison of metals from different stations, normalizing procedures for
metals concentrations needed to be used. The normalization is often done based on a
conservative component in which the levels are unaffected by contaminant inputs (Alomary
and Belhadj, 2007)(120).
A common approach for estimating the anthropogenic impact on sediments is to
calculate a normalized enrichment factor (EF) for metal concentrations above uncontaminated
background levels (Salomons and Forstner, 1984; Dickinson et al., 1996)(181, 520). The EF
calculation seeks to reduce the metal variability associated with variations in mud/sand ratios,
and is a convenient tool for plotting geochemical trends across large geographic areas
(Abrahim and Parker, 2008)(518).
To evaluate the magnitude of contaminants in the environment, the enrichment factors
were computed relative to the abundance of species in source material to that found in the
Earth's curst (Atgin et al., 2000)(7) . Deely and Fergusson (1994)(521) proposed Fe as an
acceptable normalization element factor since they considered the Fe distribution was not
related to other heavy metals. Fe usually has a relatively high natural concentration, and is
therefore not expected to be substantially enriched from anthropogenic sources in estuarine
sediments (Niencheski et al., 1994)(522).
According to Forstner and Wittman (1983)(427), in case of Fe, particularly the redox
sensitive iron hydroxide and oxide under oxidation condition constitute significant sink of
trace metals in aquatic system. Even a low percentage of Fe(OH)3, in aquatic system, has a
controlling influence on trace metal distribution (Rath, 2005)(523). Therefore, Iron is taken as
a normalization element in determining enrichment factor (Chakravarty and Patgiri, 2009)(524).
EF represents the actual contamination level in the sediment and is a good tool to
differentiate the metal source between anthropogenic and naturally occurring (Valdes et al.,
2005)(519). According to Zhao et al. (2007)(525), Fe was used for the following reasons: (1) Fe
is associated with fine solid surfaces; (2) its geochemistry is similar to that of many trace
metals; and (3) its natural sediment concentration tends to be uniform.
A reference element is often the one characterized by low occurrence variability, such
as the most commonly used elements: Al, Fe, Ti, Si, Sr, K, etc., (Duzgoren-Aydin, 2007)(526).
The EF calculation is expressed below as the follwoing:
-172-
EF = [Cx/Cref]sample /[Cx/Cref]background
where, Cx is the concentration of the element of interest, Cref is the concentration of
reference element for normalization.
In the present study, the background values were adopted from Salomons and Forstner
(1984) , there values in µgg-1: 41000 for Fe, 0.17 for Cd, 19 for Co, 72 for Cr, 33 for Cu,
850 for Mn, 68 for Ni, 20 for Pb, 95 for Zn and the geochemical normalization was obtained
using Fe as the reference element.
(181)
EF values less than 5 are not considered significant, because such small enrichments
may arise from differences in the composition of local soil material and reference soil used in
EF calculations (Sezgin et al., 2003)(527). However, there is no accepted pollution ranking
system or categorization of degree of pollution on the enrichment ratio and/or factor
methodology. EF values were interpreted as suggested by Chen et al. (2007)(528) who
interpreted the value of EF into seven classes. EF<1 indicates no enrichment, EF< 3 is minor
enrichment, EF = 3-5 is moderate enrichment, EF = 5-10 is moderately severe enrichment,
EF= 10-25 is severe enrichment, EF= 25-50 is very severe enrichment and EF>50 is
extremely severe enrichment.
The computed metal EFs in the surface sediments of the Egyptian Mediterranean
coastal area represented in Table 20 showed that the EF of Cd ranges from 0.22 to 74.03, the
EF of Co ranged from 0.61 to 10.50, the EF of Cr ranged from 1.18 to 28.99, the EF of Cu
ranged from 0.32 to 6.78, the EF of Mn ranged from 0.86 to 8.04, the EF of Ni ranged from
0.62 to 26.99, the EF of Pb ranged from 0.40 to 114.22, the EF of Zn ranged from 0.29 to
14.77. The calculated EFs were found to fall in the following sequence:
Cd>Pb>Cr>Ni>Mn>Zn>Co>Cu. As shown in Figure 32 Cadmium showed EFs ranging from
minor enrichment (average: 1.49; Middle region) to sever enrichment (average: 18.34; Eastern
region) and very sever enrichment (average: 29.57; western region) indicating significant
participation of anthropogenic sources. On the other hand, lead showed EFs ranging from no
enrichment (average: 0.63; Middle region), minor enrichment (average: 2.74; eastern region)
to sever enrichment (average: 24.05; western region). The EFs values of Cr were (average:
3.81; middle region), (average: 4.13; Middle region) indicationg moderate enrichment,
(average: 16.88; eastern region) indicating sever enrichment. Nickel showed EFs ranged from
minor enrichment (average: 1.77; middle region) to moderate enrichment (average: 4.06;
western region) and (average: 8.25; eastern region). Manganese showed EFs ranged from
minor enrichment at the middle region (average: 1.29) and the eastern region (average: 1.88)
and moderate enrichment at the western region (average: 3.11). The EFs of Zinc were viewed
to be with no enrichment at the middle region (average: 0.39) and the eastern region (average:
0.59) to moderate enrichment at the western region (average: 3.74). The EFs for Co were of
minor pollution at all the studied regions in the following order: middle reigon<eastern
region<western region with an average 1.47, 2.01 and 2.22 respectively. Cupper showed EFs
with no enrichment at the middle region (average: 0.48) and the eastern region (average: 0.89)
and minor enrichment at the western region (average: 2.75).
Despite the high EF in the sediments of the western region at stations (2, 3, 4, 5, 6, 7
and 8) in Baghoush, Nobarreya, El-Dikhaila, El-Mex, Western harbour, NIOF and Eastern
harbour, respectively and station 20 in Rafah at the eastern region. There seems to be an
overestimation of pollution of metals in these sediments. This may results from the low
concentration of Fe in these sediments (243.48, 277.34, 3946.02, 3565.22, 1527.85, 4813.8,
3440.82 and 1422.39), respectively which may lead to higher EF of trace metals at these
stations. These results are in agreement with the results of Karabassi et al. (2008)(529) who
-173-
found that applying EFs for metals pollution detection may leads to incorrect results due to
naturally lower concentration of Al and Fe or higher concentration of trace metals in
sediments.
Table 20: Enrichment factor (EF) in the
Mediterranean coastal area
Region
Station
Pb
Zn
Ni
4.93
2.33
0.83
1
39.08
3.63
5.33
2
114.22
14.77
26.99
3
8.72
2.51
1.75
4
32.64
5.99
1.26
5
Western
28.70
1.78
1.90
6
region
13.41
5.58
1.43
7
15.34
2.51
0.62
8
1.78
0.60
1.33
9
3.93
0.66
2.13
10
1.86
0.79
1.12
11
Average 24.05
3.74
4.06
1.03
0.47
2.97
12
0.51
0.40
1.65
13
0.50
0.29
1.36
14
Middle
0.50
0.30
0.92
15
region
0.40
0.34
1.19
16
0.76
0.43
2.60
17
0.74
0.50
1.67
18
Average
0.63
0.39
1.77
0.42
0.38
1.10
19
Eastern
5.06
0.79
15.40
20
region
Average
2.74
0.59
8.25
All region Average 13.73
2.25
3.68
surface sediments of the Egyptian
Cd
9.66
73.30
55.65
21.15
10.89
74.03
22.09
26.85
3.99
14.24
13.43
29.57
3.55
0.82
2.22
0.22
0.32
1.97
1.32
1.49
0.31
30.69
15.50
18.34
Cu
2.40
2.33
3.85
3.05
2.56
3.36
6.78
3.95
0.62
0.56
0.77
2.75
0.44
0.32
0.49
0.39
0.57
0.50
0.65
0.48
0.42
1.37
0.89
1.77
Mn
0.86
5.71
3.31
2.37
1.56
8.04
1.98
2.11
2.44
3.32
2.45
3.11
1.42
1.61
1.14
1.15
1.03
1.48
1.18
1.29
1.70
2.05
1.88
2.35
Cr
2.51
9.55
10.79
2.36
3.70
2.64
3.01
1.56
4.24
2.91
2.17
4.13
4.96
6.27
2.65
1.18
1.56
7.28
2.77
3.81
4.77
28.99
16.88
5.29
Co
0.61
3.82
10.50
0.94
0.81
1.95
0.84
0.84
0.87
1.40
1.78
2.22
1.77
1.19
1.06
1.41
1.52
1.48
1.85
1.47
1.08
2.93
2.01
1.93
35.00
30.00
25.00
EF
20.00
15.00
10.00
5.00
0.00
Pb
Zn
Ni
western region
Cd
Cu
Middle region
Mn
Cr
Co
Eastern region
Figure 32: Mean Enrichment factor for surface sediments along the Egyptian
Mediterranean coastal area
-174-
6.2.1.2.2. Geoaccumulation Index (Igeo)
Another commonly tool used to evaluate the trace metal pollution in sediments of the
present study is the geoaccumulation index (Igeo). The geoaccumulation index (Igeo)
originally introduced by Müller (1969)(516), determine and define metal contamination in
sediments. The geoaccumulation index (Igeo) is defined by the following equation:
Igeo = log2 (Cn/1.5*Bn)
where Cn is the measured concentration of the examined metal (n) in the sediment and Bn is
the geochemical background concentration of the metal (n). The factor 1.5 is the background
matrix correction factor due to lithological variability. The background values of the metals
here are the same as those recorded in the enrichment factor calculation. Similar to metal
enrichment factor, geoaccumulation index can be used as a reference to estimate the extent of
metal pollution (Zhang et al., 2009)(530). Müller has distinguished seven classes of the
geoaccumulation index from Class 0 (Igeo=0) to Class 6 (Igeo>5). The Igeo is associated with
a qualitative scale of pollution intensity, samples may be classified as unpolluted (Igeo≤0),
unpolluted to moderately polluted (0≤Igeo≤1), moderately polluted (1≤Igeo≤2), moderate to
strongly polluted (2≤Igeo≤3), strongly polluted (3≤Igeo≤4), strongly to extremely polluted
(4≤Igeo≤5), and extremely polluted (Igeo≥5). The highest class (Class6) reflects at least 100fold enrichment above the background values (Table 21).
Table 21: Geoaccumulation Index Igeo
Igeo Value
≤0
0-1
1-2
2-3
3-4
4-5
>5
Igeo class
0
1
2
3
4
5
6
Pollution level
Unpolluted
Unpolluted to moderately polluted
Moderately polluted
Moderately polluted to highly polluted
Highly polluted
Highly polluted to very highly polluted
Very highly polluted
This classification is a methodological approach based on the geochemical data that
makes possible to map the study area and discriminate various sub-areas according to their
pollution degree. In addition it is possible to obtain a proper comparison between various
marine areas in terms of their heavy metal quality (Christophoridis et al., 2009)(531).
The calculated Igeo values are presented in Table 22 and the variations are shown
graphically in Figure 33. It is evident from the figure that the Igeo values for Zn, Mn, Ni, Fe,
Cu and Co fall in class "0" (unpolluted) at all the sampling locations indicating that there is no
pollution from these metals. The Igeo value for Pb fall in class "0" (unpolluted) at almost all
the sampling location except station 5 at El-Mex Bay and station 7 at NIOF which fall in class
"1" (unpolluted to moderately polluted). The Igeo value for Cd fall in class "0" (unpolluted) at
almost all locations except El-Salloum, El-Dikhaila, Western harbour, NIOF, Eastern harbour,
Electrical power stations and Maadia falls in class "1" (unpolluted to moderately polluted).
The Igeo value for Cr fall in class "0" (unpolluted) at all stations of the western region, while
stations 13 and 17 in (Rashid east, El-Gamil east) at the middle region and station 19 in Port
Said at the eastern region were classified as "moderately polluted" .
-175-
Table 22: Igeo values of trace metals in the surface marine sediments along the Egyptian Mediterranean coastal area
Region
Western
region
Middle region
Eastern region
All region
Stations
1
2
3
4
5
6
7
8
9
10
11
Average
12
13
14
15
16
17
18
Average
19
20
Average
Average
Pb
-0.88
-2.77
-1.03
-0.91
0.85
-0.56
0.00
-0.29
-1.67
-1.31
-2.12
-0.97
-2.70
-2.23
-2.37
-1.76
-2.11
-2.22
-1.93
-2.19
-2.17
-3.17
-2.67
-1.57
Zn
-1.88
-6.12
-3.91
-2.64
-1.53
-4.50
-1.20
-2.83
-3.16
-3.83
-3.28
-3.17
-3.76
-2.52
-3.07
-2.44
-2.26
-2.95
-2.43
-2.78
-2.25
-5.77
-4.01
-3.12
Ni
-3.75
-5.95
-3.43
-3.55
-4.16
-4.79
-3.55
-5.23
-2.40
-2.51
-3.16
-3.86
-1.49
-0.85
-1.23
-1.20
-0.85
-0.76
-1.06
-1.06
-1.11
-1.88
-1.49
-2.65
Cd
0.17
-1.78
-1.99
0.44
-0.66
0.88
0.79
0.59
-0.42
0.62
0.81
-0.05
-0.84
-1.47
-0.14
-2.87
-2.35
-0.78
-1.02
-1.35
-2.57
-0.49
-1.53
-0.65
-176-
Cu
-1.84
-6.76
-5.85
-2.35
-2.75
-3.58
-0.91
-2.18
-3.12
-4.05
-3.32
-3.34
-3.85
-2.81
-2.32
-2.04
-1.52
-2.76
-2.04
-2.48
-2.12
-4.98
-3.55
-3.06
Mn
-3.45
-5.61
-6.21
-2.86
-3.61
-2.47
-2.83
-3.23
-1.28
-1.63
-1.79
-3.18
-2.31
-0.64
-1.24
-0.63
-0.82
-1.33
-1.32
-1.18
-0.23
-4.54
-2.39
-2.40
Fe
-3.10
-7.98
-7.79
-3.96
-4.11
-5.33
-3.68
-4.16
-2.42
-3.22
-2.94
-4.43
-2.67
-1.19
-1.29
-0.69
-0.72
-1.75
-1.42
-1.39
-0.86
-5.43
-3.15
-3.24
Co
-3.81
-6.05
-4.40
-4.04
-4.41
-4.37
-3.93
-4.41
-2.62
-2.73
-2.10
-3.90
-1.84
-0.94
-1.21
-0.19
-0.11
-1.19
-0.53
-0.86
-0.74
-3.88
-2.31
-2.68
Cr
-1.77
-4.73
-4.36
-2.72
-2.22
-3.93
-2.09
-3.51
-0.34
-1.67
-1.82
-2.65
-0.36
1.46
0.11
-0.45
-0.08
1.11
0.05
0.26
1.40
-0.58
0.41
-1.32
Figure 33: Igeo values distribution of trace metals in the surface marine sediments along the
Egyptian Mediterranean coastal area
-177-
6.2.1.2.3. Contamination factor (Cf) and Degree of contamination (Cd)
Hakanson (1980)(517) proposed an overall indicator of contamination based on
integrating data for a series of seven specific heavy metals and the organic pollutant PCB.
This method is based on the calculation for each pollutant of a contamination factor (Cf).
Surfacial sediment samples are compared to a baseline pristine reference level, according to
the following equation:
Cf = Mx/Mb
Where, Mx and Mb respectively, refer to the concentration of a pollutant in the
contaminated sediments and the pre-industrial "baseline" sediments. According to Hakanson
(1980)(517) classification the contamination factor was classified into four groups: Cf< 1 low
contamination factor; 1 ≤ Cf < 3 moderate contamination factors; 3≤ Cf <6 considerable
contamination factors; Cf ≥ 6 very high contamination factor. Data at Table 24 showed that
the contamination factors values are low for Zn, Cu and Fe at all sampling sites. They are
between low and moderate for Ni, Cd, Mn and Co and between low and considerable
contamination for Cr.
The study by Hakanson (1980)(517) analysed seven specific heavy metals (As. Cd, Cu,
Cr, Hg, Pb, Zn) and the organic pollutants PCB and thus considers eight possible measures of
contamination. Hakanson's study also propsed that the numeric sum of the eight specific
contamination factors expressed the overall degree (Hakanson, 1980)(517) of sediment
contamination (Cd) using the following formula:
8
Cd =  Cif
i=1
The calculated Cd is therefore defined as the sum of the Cf for the eight pollutant
species specified by Hakanson (1980)(517).
The Hakanson (1980)(517) classification
terminology and calculation formula is based on and is restricted to the seven metals plus
PCB specified in Hakanson's study. Furthermore all eight species must be analysed in order
to calculate the correct Cd for the range of classes defined by Hakanson (1980)(517).
6.2.1.2.4. Modified degree of contamination (mCd)
As a result of the above limitations Abrahim (2005)(532) presented a modified and
generalized form of the Hakanson (1980)(517) equation for the calculation of the overall degree
of contamination at a given sampling site as follows:
i n
mCd 
c
i 1
i
f
n
where, n = number of analyzed elements and i = the element (or pollutant) and Cf =
contamination factor. Using this generalized formula to calculate the mCd allows the
incorporation of as many metals as the study may analyze with no upper limit. For the
classification and description of the modified degree of contamination (mCd) in estuarine
sediments the follwong gradation are proposed:
-178-
Table 23: The classification and description of the modified degree of contamination
(mCd)
mCd < 1.5
1.5 < mCd < 2
2 < mCd < 4
4 < mCd < 8
8 < mCd < 16
16 < mCd < 32
mCd ≥ 32
Nil to very low degree of contamination
Low degree of contamination
Moderate degree of contamination
High degree of contamination
Very high degree of contamination
Extremely high degree of contamination
Ultra high degree of contamination
In the Egyptian Mediterranean coastal sediments the revised Hakanson equation was
used to calculate the modified degree of contamination (mCd) for the nine analyzed elemetns
(Cd, Cu, Cr, Co, Fe, Mn, Pb, Zn and Ni). The results for each sampling site are presented in
Table 24 and illustrated in Figure 34. The calculated Cfs were found to fall in the following
sequence: Cd>Cr>Pb>Co>Ni>Mn>Fe>Cu>Zn. The Egyptian Mediterranean mCd are less
than 1.5 indicating zero to very low degree of contamination.
mCd values are calculated based only on the separate contamination factors, which
don't consider the normalizing factor of Fe, therefore it gives values not reproducible to every
environment with the same degree of contamination and not easily compared to other areas
studied in the past with different geological backgrounds (Christophoridis et al., 2009)(531).
6.2.1.2.5. Pollution Load Index
The extent of pollution by trace metals has been assessed by employing the method
based on Pollution load index (PLI) developed by Tomlinson et al. (1980)(533) as the following
equation:
Where CF = contamination factor and n= number of metals
PLI provides a simple, comparative means for assessing a site or estuarine quality: a
value of zero indicates perfection, a value of one indicates only baseline levels of pollutants
present and values above one would indicate progressive deterioration of the site and
estuarine quality (Tomlinson et al., 1980)(533). PLI calculated for total metals in surface
sediments of the Egyptian Mediterranean Sea from El-Salloum to Rafah are presented in
Table 24. PLI values of the analyzed samples ranged from 0.04 at station 2 (Baghoush;
western region) and 0.72 at station 19 (Port Said; eastern region) with an average 0.41.
Figure 35, shows the variation of PLI at different stations. In general, there is an increase of
PLI toward the eastern region. However, results of PLI were less than 1 at all the studied
locations indicating only baseline levels of pollutants present. The PLI can provide some
understanding to the public of the area about the quality of a component of their environment,
and indicates the trend spatially and temporarily. In addition, it also provides valuable
information to the decision makers on the pollution level of the area (Harikumar et al.,
2010)(534).
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Table 24: Modified degree of contamination (mCd) and contamination factors (Cf) for the surfacial sediments along the Egyptian
Mediterranean coastal area
Region
Station
Cf Pb
Cf Zn
Cf Ni
Cf Cd
Cf Cu
Cf Mn
Cf Fe
Cf Co
Cf Cr
Sum Cif mCd
0.86
0.41
0.15
1.69
0.42
0.15
0.17
0.14
0.44
4.43
0.49
1
0.23
0.02
0.03
0.44
0.01
0.03
0.01
0.03
0.06
0.86
0.10
2
0.77
0.10
0.18
0.38
0.03
0.02
0.01
0.10
0.07
1.66
0.18
3
0.84
0.24
0.17
2.04
0.29
0.23
0.10
0.12
0.23
4.25
0.47
4
2.84
0.58
0.11
0.95
0.22
0.14
0.09
0.10
0.32
5.34
0.59
5
Western region
1.07
0.07
0.07
2.76
0.13
0.30
0.04
0.10
0.10
4.62
0.51
6
1.57
0.65
0.17
2.59
0.80
0.23
0.12
0.13
0.35
6.62
0.74
7
1.29
0.21
0.05
2.25
0.33
0.18
0.08
0.10
0.13
4.62
0.51
8
0.50
0.17
0.37
1.12
0.17
0.68
0.28
0.33
1.19
4.81
0.53
9
0.63
0.11
0.34
2.30
0.09
0.54
0.16
0.31
0.47
4.95
0.55
10
0.36
0.15
0.22
2.63
0.15
0.48
0.20
0.47
0.42
5.09
0.57
11
0.24
0.11
0.70
0.84
0.10
0.33
0.24
0.57
1.17
4.30
0.48
12
0.34
0.26
1.09
0.54
0.21
1.06
0.66
1.06
4.14
9.36
1.04
13
Middle region
0.31
0.18
0.84
1.36
0.30
0.70
0.61
0.88
1.62
6.80
0.76
14
0.46
0.28
0.85
0.21
0.36
1.07
0.93
1.78
1.10
7.04
0.78
15
0.37
0.31
1.08
0.29
0.52
0.94
0.91
1.88
1.42
7.73
0.86
16
0.34
0.19
1.16
0.88
0.22
0.66
0.45
0.90
3.25
8.04
0.89
17
0.41
0.28
0.94
0.74
0.36
0.67
0.56
1.41
1.55
6.93
0.77
18
0.35
0.32
0.91
0.25
0.35
1.41
0.83
1.22
3.95
9.58
1.06
Eastern region
19
0.18
0.03
0.53
1.06
0.05
0.07
0.03
0.14
1.01
3.10
0.34
20
All region
Average
0.70
0.23
0.50
1.27
0.26
0.49
0.32
0.59
1.15
5.51
0.61
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PLI
0.34
0.04
0.08
0.29
0.29
0.18
0.43
0.23
0.43
0.34
0.36
0.35
0.69
0.61
0.63
0.71
0.62
0.66
0.72
0.15
0.41
2
mCd
1.5
1
0.5
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
Stations
Figure 34: modified degree of contamination (mCd) for sediments along the
Egyptian Mediterranean coast
Pollution Load Index
1.50
1.00
0.50
0.00
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
Stations
Figure 35: Pollution load index distribution at different stations along the
Egyptian Mediterranean coastal sediments.
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6.2.2. Leachable metals
Sediments are the important component of ecosystem in which toxic compounds
accumulate through complex physical and chemical adsorption mechanisms depending on the
properties of the adsorbed compounds and the nature of the sediment matrix (Leivouri,
1998)(134). In order to avoid the pollution of trace metals in marine environment, it is
necessary to establish the data and understand the mechanisms influencing the distribution of
trace metals in marine environment (Raj and Jayaprakash, 2008)(535).
The mobiltiy and bioavailabiltiy of heavy metals in the environment are
predominanted by their existing fractions. There are two methods to extract the active
fraction of heavy metals in sediments. One is to adopt the total content of the nonresidual
fractions using multiple sequential extraction procedure. The other is to choose a certain
selective reagent to extract the active fraction directly, named single step extraction
procedure. The selective reagents generally include dilute acids, chelating agents, neutral
salts, buffer reagents, and so on. The extraction procedure using dilute hydrochloric acid is
one of the most widely used methods in single reagent extraction, and it is considered to be a
rapid, effective means to assess the mobility of heavy metals in sediment and to evaluate
whether the sediment has been polluted by anthropogenic activities (Jayaprakash et al.,
2008)(467). After a series of environmental evolution, the acid leachable heavy metals in
sediment are easy to release into water column and bring "secondary pollution"
(Ayyamperumal et al., 2006)(536).
The leaching of metals provides an accurate measure of the bioavailable metals in any
aquatic environment which are often readily available to organisms affecting them directly.
Hence assessment of trace metal enrichment in sediments on the acid leachable (non-residual)
elements is of prime interest, as it often yields more data on the extent of trace metal
enrichment than the total sediments, which include the residual or non-residual fraction, and
so may mark the relationships sought (Sarkar and Bhattacharya, 2010)(482).
The leachable metal fraction is defined as the anthropogenic fraction of metals
involved with the sediment particles. The assessment of trace metals on the acid leachable
elements is of great interest as the results are more informative than total sediments, which
include the residual or non residual or non polluted fraction (Janki-Raman et al., 2007)(537).
The acid leachable fraction extracts almost the whole degree of elements as it is absorbed by
sediments depicting the contamination of an area (Janaki-Raman et al., 2007)(537). It is now
widely accepted that the use of total element concentrations in environmental impact
assessments does not represent an optimal approach (Sahuquilloet et al., 2003)(538).
The leachable metals fraction defined as that including exchangeable, carbonate
bound, iron and manganese oxide bound and organically bound fractions (Nriagu, 1990)(539).
Analysis of the leachable (labile) metal fraction of the sediment may be more useful, in terms
of discovering its biological significance and the new inputs, than analysis of the total metal
fraction (Lacerda et al., 1992)(540). However, sediment analysis is more indicative than water
analysis for evaluating the degree of contamination in the aquatic medium. Concentrations of
toxic elements are usually higher in sediments with less possibility of contamination of
samples during handling and processing, and the analytical methods are simpler (El-Nemr et
al., 2006)(437).
The importance of leaching studies has gained significant attention in recent decade
due to the assessment of metals present in different fractions in the sediment. The leaching of
metals provides an accurate data base of the bioavailable metals in any aquatic environment
which are often readily available to organisms affecting them directly. This bioavailable
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fraction is thus defined as the amount of metal that can be exchanged with biological
organisms and be incorporated into their structure (Vangronveld and Cunnihgham, 1998)(541).
The identification of the main binding sites of trace metals helps in the understanding
of geochemical processes and allows the prediction of potential remobilisation, along with the
determination of bioavailability and possibly the discrimination of sources (Hall et al.,
1996)(542). Leachability of metals from soils and sediments depends on several factors,
sediment characteristics, including pH, CaCO3 content, organic matter content (Sahuquilloet
al., 2003)(538).
The ratio of the trace metal concentrations in the labile fraction to the total
concentration of metals in the sediment is expressed as a percentage, known as the percentage
of extractability. The labile fraction is usually defined as exchangeable, carbonate bound,
iron and manganese oxide bound and organically bound fractions. The ratio between labile
and total trace metals may be referring to the anthropogenic input or the new input percentage
(Masoud et al., 2007)(294).
For investigating the bioavailability of metals, a one step extraction procedure was
applied. 1N HCl attack was selected because the extraction procedure is simple and its
replication is reliable. Many terms have been given to the leachable fraction (acid soluble,
labile, non-residual and non detrital). They are the unstable fraction of metals that with a
slight change in the environmental conditions (pH, Eh), may be released to the upper water
column thus named non-residual or labile. It is important to asses the availabily of heavy
metals in sediments because of their persistence, subsequent bioaccumulation and toxification
to biota (Tanner and Leong, 1999)(422). This one-step extraction process yields metals bound
as hydroxides, carbonates, sulfides, oxides and weakly bound to organic matter.
The concentrations of the selected nine trace metals of labile fractions in the surfacial
sediments of the Egyptian Mediterranean coastline are shown in Table 6. The surface layer
was chosen for this study, where this layer controls the exchange of metals between sediments
and water as well as constitutes a reserve of metals to which benthic organisms are exposed
(El-Nemr et al., 2006)(437). Visual inspection of Figures 36, 37 and 38 represented the
variations of trace metal content in labile fractions of the studied sediments. From the
obtained data, it is observed that the concentration of trace metals in labile fraction decreasing
in the order Fe>Mn>Pb>Zn>Ni>Cr>Co>Cu>Cd. The percentage of extractability which is the
ratio of leachable to total heavy metal concentrations were expressed in Table 25. This has
been used as a theoretical estimation of the relative importance of metals of anthropogenic
origin (Carrel et al., 1994)(543) and refers to the new input percentage (NIP) of heavy metals.
The NIP showed the following decreasing order: Mn>Pb>Cd>Zn>Cu>C>Fe>Cr>Ni.
Leachable Iron (LFe)
Iron is one of the most abundant metals in sediments because it is the common
element in the earth's crust (Wei et al., 2007)(13). In the western region of the present study,
the total Iron concentrations ranged from 243.48 to 11478.38µgg-1 and from 106.63 to
2100.42µgg-1 for labile fraction. While in the middle region, the total Iron concentration
ranged from 9653.67 to 38045.05µgg-1 and from 1494.21 to 6831.61µgg-1 for labile fraction.
Finally in the eastern region, the total Iron concentration ranged from 1422.39 to
33971.50µgg-1 and from 94.98 to 3667.31µgg-1 for labile fraction (Figure 36). % of LFe for
Fe fluctuated between 6.68% at station 20 in Rafah to 69.46% in station 2 in Baghoush.
The maximum % of LFe (69.46%) was found at station 2 (Baghoush) in the western
region and the minimum percentage (6.68%) of LFe was found at station 20 (Rafah) in the
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eastern region. It was also noted that Fe has a relatively low percentage of LFe in almost all
stations except some stations 2, 3, 4, 6, 7, and 8 in Baghoush, Nobareya, Dekhaila, El-Mex,
NIOF and the Eastern harbour, respectively in the western region which exceed (30%).
Results showed that there was a significant increase in the residual Fe from station 9 in Abu
Qir at the western region to station 20 in Rafah at the eastern region. The low percentage of
LFe in most cases (average: 22.83%) was agreeable with the findings of (Dassenakis et al.,
2003)(180) who found that the low percentage (<20%) of weakly bound iron may be due to the
increased contribution of the lattice-held fraction of the metal.
Leachable Manganese (LMn)
Manganese is the 11th element in terms of abundance in the earth crust (Anschutz et
al., 2005)(113). Mn in Earth's crust is 1600 µgg-1; in soils it is 61-1060 µgg-1. The common
aqueous species found in water is predominantly M2+ and M4+. Mangansese is essential for
plants and animal (WHO, 2004)(73). Natural levels of manganese in soil ranged from 40 to
900 µgg-1, with an estimated mean of 330 µgg-1 (Rope et al., 1988)(116). Accumulation of
manganese in soil usually occurs in the subsoil and not on the soil surface (WHO, 1981)(117).
From the obtained data at Figure 36 it was observed that, the concentration of
manganese ranged between 17.25 and 526.53 for total and from 10.83 to 261.85µgg-1 for
labile fractions at the western region. In the middle region, the total manganese
concentrations ranged between 257.30 and 822.60 µgg-1; and from 29.99 to 307.27µgg-1 for
labile fraction. While, in the esatern region, the total manganese concentration ranged
between 54.76 and 1085.72µgg-1 and from 30.20 to 171.53µgg-1 for labile fraction. The % of
LMn ranged between 19.94 to 91.67% at the western region, 3.66 to 46.21% at the middle
region and 15.80 to 55.15% at the eastern region. These results were in agreement with the
data reported by Dassenakis et al. (1995)(544), who found that high percentage of weakly
bound Mn (>50%) are probably due to the elevated association of Mn with carbonates that are
dissolved by the dilute HCl (Dassenakis et al., 1995)(544).
Leachable Lead (LPb)
Lead forms stable organic compounds. Tetraethyllead and tetramethyllead are used
extensively as fuel additives. Both are volatile and poorly soluble in water. Trialkyllead
compounds are formed in the environment by the breakdown of tetraalkyllead. These trialkyl
compounds are less volatile and more readily soluble in water. Lead is mined, most usually as
the sulfide, "galena". Lead is a non essential element for living organisms and is an
accumulative poison (Schulz and Zabel, 2000)(79). The average abundance in Earth's crust is
13 µgg-1; in natural soils background level ranges from 2.6-25 µgg-1; the common aqueous
species are hydroxides and carbonates of Pb2+ (Sharma and Pervez, 2003)(93). Lead is
reported to be in the range 15 to 50 µgg-1 for coastal and estuarine sediments around the world
(Denton et al., 1997)(48) with < 25 µgg-1 in clean coastal sediments. Hence, entry into the
aquatic environment occurs through releases (directly or through atmospheric deposition)
from the smelting and refining of lead, the burning of petroleum fuels containing lead
additives and, to a lesser extent, the smelting of other metals and the burning of coal and oil.
Metallic lead deriving from shotgun cartridges or used as fishing weights is lost in the
environment and often remains available to organisms (Schulz and Zabel, 2000)(79).
Total and labile fractions of the surfacial sediments along the Egyptian Mediterranen
coast represented Pb content fluctuated between 4.41 and 53.67 and from 1.78 to 51.35µgg-1
respectively, for the western region. It ranged from 4.61 to 8.83 in total fraction and from 0.77
to 3.98µgg-1 in labile fraction for the middle region. Finally in the eastern region, lead
concentrations ranged between 3.34 and 6.67 µgg-1 in total fraction, and from 0.54 to 1.56
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µgg-1 for labile fraction (Figure 36), which indicate a new input of lead to the western region
of the Egyptian Mediterranean coast.
Table 25 showed that Lead exhibited a high % of LPb in the western region ranged
between 25.72 and 95.68% (average: 60.95%). The high extractability percentage of Pb
reported at station 5 (El-Mex Bay) indicated a new input in El-Mex Bay, which was probably
due to the domestic, industrial effluents and the atmospheric deposition; many reported
studies have also indicated elevated levels of heavy metals in aquatic systems receiving
effluents from urban areas, domestic, and untreated sewage (El-Nemr, 2003)(545).
Leachable Zinc (LZn)
Sediments are known as major sinks for zinc in the aquatic environment, and residues
in excess of 3000 µgg-1 have been reported close to mines and smelters (Denton et al.,
2001)(66). The highest sedimentary zinc levels are found to be from enclosed harbors reaching
as high as 5700 µgg-1. This is mainly due to restricted water circulation and also particularly
prone to zinc contamination from a variety of localized sources including brass and
galvanized fittings on boats, wharves and piers; zinc-based anti-corrosion and anti-fouling
paints (Denton et al., 1997)(48). The average zinc content of the lithosphere is approximately
80 µgg-1 (Callender, 2003)(65) and depending on the local geology, sediments from
uncontaminated waters typically contain zinc concentration in the order of 5-50 µgg-1.
However, levels normally encountered in carbonate sediments from pristine reef waters, away
from coastal developments are usually less than 1 µgg-1 (Denton et al., 2001)(66).
Results at Figure 37 showed that, the total concentration of zinc ranged between 2.05
and 62.21 and from 0.11 to 49.48µgg-1 labile fractions at the western region. In the middle
region, the total zinc concentrations ranged between 10.49 and 29.82 µgg-1 and from 1.49 to
6.54µgg-1 for labile fraction. While, in the esatern region, the total zinc concentrations ranged
between 2.61 and 30.05 µgg-1 and from 0.15 to 3.63µgg-1 for labile fraction. Table 25 showed
that the % of LZn ranged between 1.16 and 89.27% (average: 43.35%) at the western region,
and from 8.11 to 24.85% (average: 15.46%) at the middle region, and from 5.83 to 12.07%
(average: 8.95%) at the eastern region, indication that about 45% of Zn detected in the
western region was a new input.
Leachable Nickel (LNi)
Nickel is a relatively mobile heavy metal. In natural waters, nickel is transported in
both particulate and dissolved forms. The pH, oxidation –reduction potential, ionic strength,
type, and concentration of organic and inorganic ligands (in particular, humic and fulvic
acids), and the presence of solid surfaces for adsorption (in particular, hydrous iron and
manganese oxides) can all affect the transport, fate, and biological availability of nickel in
fresh water and seawater (Snodgrass, 1980)(100). In some sediments under reducing conditions
and in the presence of sulphur, relatively insoluble nickel sulphide is formed (Ankley et al.,
1991)(228). Under aerobic conditions and pH less than 9, nickel forms compounds with
hydroxide, carbonate and sulphate. Naturally occurring organic ligands are sufficiently
soluble to maintain aqueous Ni2+ concentrations above 60 µg/L (Callahan et al., 1979)(546).
Most of the nickel in sediments and suspended solids is distributed among organic materials,
precipitated and co-precipitated as particle coatings, and crystalline particles (CEPA,
1994)(547). Typically, nickel residues in sediments can be up to 100 µgg-1 or higher but may
fall below 1 µgg-1 in some clean coastal waters (Denton et al., 1997)(48) with the average
concentration of nickel in the lithosphere of 55 µgg-1 (Callender, 2003)(65).
From the
obtained data at Figure 37, it was observed that, the concentration of nickel ranged between
1.65 and 19.36 and from 0.27 to 3.29µgg-1 for total and labile fractions espectively, at the
western region. In the middle region, total concentrations ranged between 36.38 and.60.25
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µgg-1 and from 2.70 to 11.16µgg-1 for labile fraction. While, in the eastern region, total nickel
concentration ranged between 27.79 and 47.42 µgg-1 and from 0.15 to 6.72 µgg-1 for labile
fraction. The % of LNi ranged between 2.86 and 28.83% (average: 15.61%) at the western
region, from 4.49 to 25.20% (average: 12.07%) at the middle region and from 0.55 to 14.17%
(average: 7.36%) at the eastern region (Table 25), which indicates that stations along the
western region have a relatively low input of Ni. This indicates that this element is tightly
bound to the sediment minerals.
Leachable Chromium (LCr)
Chromium is the 21st
most abundant element in Earth's crust with an average
concentration of 100µgg-1. Chromium compounds are found in the environment, due to
erosion of chromium containing rocks and can be distributed by volcanic eruptions. The
concentrations range in soil is between 1 and 3000 µgg-1, in sea water 5 to 800 µg/L, and in
rivers and lakes from 26µgL-1 to 5.2mgL-1. Chromium like zinc is one of the most abundant
heavy metals in the lithosphere with an average concentration of about 69µgg-1 (Callender,
2003)(65). Levels of chromium in marine sediments range from 2.4 at unpolluted sites to 749
µgg-1 at grossly contaminated sites. Under the US EPA sediment quality classification,
sediments with chromium levels of less than 25 µgg-1 are considered as non-polluted whereas
sediments with 25-75 µgg-1 and >75 µgg-1 are considered to be moderately polluted and
heavily polluted respectively (Denton et al., 1997)(48). Calcareous sediments of biogenic
origin, such as those found in coral reefs, are often lower and normally contain 3-5 µgg-1
(Denton et al., 2001)(66). From the obtained data in Table 25, the % of LCr ranged between
1.58 and 61.35% (average: 23.60%) at the western region, and from 0.42 to 3.56 % (average:
1.51%) at the middle region, and from 0.96 to 1.04 (average: 1.00%) at the eastern region,
indicating that this element is tightly bound to the sediment minerals.
Leachable Cobalt (LCo)
The average cobalt abundance in earth's crust is 29µgg-1; in soils it is 1.0-14 µgg-1.
(WHO, 2004)(73). Data at Figure 38 showed that cobalt content in total and labile fractions of
the surfacial sediments along the Egyptian Mediterranen coast were in the range (0.43 to 6.63
and 0.10 to 2.85 µgg-1, respectively) for the western region, and from 7.94 to 26.39µgg-1 for
total fraction and from 1.88 to 8.77 µgg-1 for the labile fraction in the middle region. Finally
in the eastern region, the total cobalt concentrations ranged between 1.93 and 17.07 µgg-1 and
from 0.34 to 7.87 µgg-1 for labile fraction, which indicate a new input of cobalt to the the
Egyptian Mediterranean coast.
In Table 25 the % of LCo rangrd between 5.34 and 43.01% (average: 24.96%) in the
western region, and from 16.94 to 35.22% (average: 24.92%) in the middle region, and from
17.69 to 45.93% (average: 31.81%) in the eastern region. Results showed that relatively low
% of LCo were found at most of the stations of the study area not exceeding 30%. This
indicates that this element is tightly bound to the sediment minerals.
Leachable Copper (LCu)
Copper is a moderately abundant heavy metal with mean concentration in the
lithosphere of about 39µgg-1. It is an essential trace element for the growth of most aquatic
organisms. However it becomes toxic to aquatic organisms at levels as low as 10 µgg-1
(Callender, 2003)(65). Heavily polluted sediments have been reported to exceed 200 µgg-1
(Denton et al., 1997)(48). Its mean concentration in soil is 18 µgg-1 (Shacklette, 1970)(548)
and from 4 to 28 µgg-1 in forage plants (Engel et al., 1964)(549). Uses of copper include
electrical wiring and electroplating, the production of alloys, copper piping, photography,
antifouling paints and pesticide formulations. Major industrial sources include mining,
smelting, refining and coal-burning industries. Certain of these anthropogenic sources may
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led to significant concentrations entering the aquatic environment (either directly via sewage
or industrial discharges or through atmospheric deposition) but copper will also enter the
aquatic environment through natural sources, e.g. from the weathering or the solution of
copper minerals (CCREM,1987)(103).
From the data obtained in Figure 38, the total cupper concentrations in the western
region, ranged between 0.46 and 26.26 µgg-1 and from 0.23 to 8.20 µgg-1 for labile fraction.
While, in the middle region, the total cupper concentration ranged between 3.44 and 17.27
µgg-1 and from 0.56 to 2.08 µgg-1 for labile fraction. Finally in the eastern region, the total
cupper concentrations ranged between 1.57 and 11.39 µgg-1 and from 0.27 to 1.79µgg-1 for
labile fraction, which indicate a new input of cupper to the Egyptian Mediterranean coastal
area.
The % of LCu ranged between 8.92 and 86.46% (average: 41.45%) at the western
region, from 10.58 to 17.34% (average: 13.48%) at the middle region, from 15.71 to 17.46%
(average: 16.59%) at the eastern region. The maximum % of LCu (86.46%) observed at
station 5 in El-Mex Bay. The antifouling paints used for ships and boats are regarded as one
of the important sources, which increased the level of copper in El-Mex Bay (Masoud et al.,
2007)(294). From the obtained data at Figure 38, it was observed that the relatively high % of
LCu (> 30%) was observed at stations 2, 3, 4, 5, 6, 7, 8, 10 (Baghoush, Nobarreya, Dekhaila,
El-Mex, Western harbor, NIOF, Eastern harbor and Power station).
Cu, although a micronutrient, can cause a toxic response in bacteria at certain
concentrations. Dissolved copper may exist as free cupric ions (Cu2+), inorganic complexes,
and complexes with various organic ligands (e.g. humic substances, phytoplankton
metabolites, proteins, etc.) (Wei et al., 2007)(13). Anthropogenic sources of cupper are
primarily related to textile production, marine anti-fouling agents, pipes and copper based
fungicides or pesticides (PIO, 2005)(550).
Leachable Cadmium (LCd)
The average concentration of cadmium in the lithosphere is ~0.1µgg-1 and it is
strongly chalcophilic (Callender, 2003)(65). Concentrations in pristine areas are <0.2 µgg-1
with levels exceeding 100 µgg-1 at severely contaminated sites (Naidu and Morrison,
1994)(551). Cadmium is widely distributed in the Earth's crust at an average concentration of
about 0.2 µgg-1 and is commonly found in association with zinc. Higher levels are present in
sedimentary rocks: marine phosphates often contain about 15µgg-1. Weathering and erosion
result in the transport by rivers of large quantities of cadmium to the world's oceans and this
represents a major flux of the global cadmium cycle; an annual gross input of 15,000 tonnes
has been estimated (Bryan,1971)(552).
The results at Table 25 showed that the % of LCd in the western region ranged
between 11.63 and 79.44% (average: 31.94%) and from 12.5 to 86.0% (average: 45.39%) for
the middle region and from 13.26 to 62.79% (average: 38.03%) for the eastern region, which
suggested a new input of Cd to the most of locations. Cd is contained in some phosphate
based fertilizers and petroleum. Such sources could constitute a major source of Cd that may
reach humans. In additions, sewage sludge from wastewater treatment may contain
significant quantities of Cd (El-Nemr et al., 2006)(437). These results are in agreement with
El-Nemr et al. (2007a)(61).
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Table 25: The relative percentage of leachable Metals from total metals in the sediments of the
Egyptian Mediterranean coastal area
Region
Station
1
2
3
4
5
6
Western
7
region
8
9
10
11
Min
Max
Average
12
13
14
15
Middle
16
Region
17
18
Min
Max
average
19
Eastern
20
region
Min
Max
Average
Min
All
Max
region Average
%LZn
13.53
25.59
1.16
49.88
89.27
67.93
28.00
73.50
25.08
52.99
49.89
1.16
89.27
43.35
15.77
8.74
15.34
24.85
21.47
8.11
13.92
8.11
24.85
15.46
12.07
5.83
5.83
12.07
8.95
1.159
89.272
30.145
%LCu %LPb %LCd %LCr %LCo %LNi
14.14
47.33
79.44
8.47
14.96
9.79
49.56
71.43
43.24
61.35
22.30
20.38
37.09
36.80
40.63
24.02
21.34
2.86
46.77
65.18
21.10
19.36
33.23
15.19
86.46
95.68
39.13
35.42
31.38
21.50
31.11
61.51
20.26
25.25
13.08
21.51
31.23
66.76
21.09
29.12
32.59
19.96
56.72
94.47
25.07
42.77
26.78
17.67
8.92
45.26
31.05
1.58
30.56
7.95
76.59
60.30
18.67
4.76
5.34
6.11
17.36
25.72
11.63
7.53
43.01
28.83
8.92
25.72
11.63
1.58
5.34
2.86
86.46
95.68
79.44
61.35
43.01
28.83
41.45
60.95
31.94
23.60
24.96
15.61
16.35
28.90
40.85
0.66
23.72
7.42
14.36
14.67
36.96
0.92
16.94
6.88
10.58
21.99
12.50
0.67
24.87
8.65
17.34
45.03
62.86
3.56
35.22
25.20
11.20
23.92
86.00
2.91
29.92
17.73
11.99
12.01
43.62
0.42
19.02
4.49
12.54
47.95
34.92
1.41
24.71
14.11
10.58
12.01
12.50
0.42
16.94
4.49
17.34
47.95
86.00
3.56
35.22
25.20
13.48
27.78
45.39
1.51
24.92
12.07
15.71
23.45
62.79
0.96
45.93
14.17
17.46
16.28
13.26
1.04
17.69
0.55
15.71
16.28
13.26
0.96
17.69
0.55
17.46
23.45
62.79
1.04
45.93
14.17
16.59
19.87
38.03
1.00
31.81
7.36
8.915
12.01
11.63
0.42
5.34
0.55
86.464 95.68
86.00
61.35
45.93
28.83
29.174 45.23
37.25
13.61
25.63
13.55
-188-
%LFe %LMn
14.78
47.38
69.46
74.98
38.45
62.80
39.69
85.02
26.56
91.67
40.97
70.48
32.59
52.81
41.20
84.33
9.08
49.73
9.88
19.94
26.17
55.74
9.08
19.94
69.46
91.67
31.71
63.17
15.48
46.21
7.49
3.66
8.85
20.49
17.96
37.35
14.66
33.26
9.08
25.73
16.83
28.02
7.49
3.66
17.96
46.21
12.91
27.82
10.80
15.80
6.68
55.15
6.68
15.80
10.80
55.15
8.74
35.47
6.68
3.66
69.46
91.67
22.83
48.03
40000
Fe µgg-1
30000
20000
10000
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
1200
Mn µgg-1
1000
800
600
400
200
0
Pb µgg-1
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
60
50
40
30
20
10
0
T Pb
LPb
1 2 3 4 5 6 7 8 9 10 11 12 1314 15 16 17 18 19 20
Stations
Figure 36: Distribution of total trace metals concentrations (Fe, Mn and Pb) and leachable
fractions in sediments along the Egyptian Mediterranean coast
-189-
Zn µgg-1
70
60
50
40
30
20
10
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
70
60
Ni µgg-1
50
40
30
20
10
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
T Cr
350
L Cr
Cr µgg-1
300
250
200
150
100
50
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
Stations
Figure 37: Distribution of total trace metals concentrations (Zn, Ni and Cr) and leachable
fractions in sediments along the Egyptian Mediterranean coast
-190-
30
Coµgg-1
25
20
15
10
5
0
Cu µgg-1
1 2 3 4 5 6 7 8 9 10 11 12 1314 15 16 17 18 19 20
30
25
20
15
10
5
0
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20
0.5
TCd
0.4
Cd µgg-1
LCd
0.3
0.2
0.1
0
1 2 3 4 5 6 7 8 9 10 11 1213 14 15 1617 18 19 20
Stations
Figure 38: Distribution of total trace metals concentrations (Co, Cu and Cd) and leachable
fractions in sediments along the Egyptian Mediterranean coast
-191-
6.2.3. Fractionation of trace metals
Measurements of total metal content are a poor indicator of metal bioavailability,
mobility or toxicity, when these properties basically depend on the chemical association of the
different components of the sample (Tuzen et al., 2004)(553). Aquatic sediments consist of
several different geochemical phases that act as reservoirs of trace metals. These phases
include carbonates, sulfides, organic matter, iron and manganese oxides, and clays, all of
which may occur in a variety of structural forms (Gao et al., 2010)(250). According to
Quevauviller (1998)(554) the determination of specific chemical forms, or the nature of
binding, is much more valuable than the determination of the total metal content, since the
toxic effects and the geochemical pathways of elements are determined mainly by their
mobile species. Fractionation of heavy metal contents may give indications of the source of
the metals. According to Tack and Verloo (1995)(555), exchangeable, carbonate, reducible and
oxidizable fractions are considered the fractions that are more available to aquatic biota where
the bioavailability decreases in the order of exchangeable > carbonate > reducible >
oxidizable.
Trace metals speciation studies are important because slight changes in metal
availability and in environmental conditions can cause these elements to be toxic to animals
and plants (Ure et al., 1993)(16). The sequential extraction procedure allowed determination of
the distribution of the trace metals between the different geochemical fractions, reflecting the
relative proportions of each metal transported by different chemical mechanisms (Passos et
al., 2011)(263). In the present work, fractionation of Cr, Fe, Mn, Ni, Pb and Zn into the
different chemical forms has been conducted in order to demonstrate in which forms these
metals are chemically chelated, as well as to differentiate between the residual metals (natural
background) and non-residual ones (man-made sources of pollution).
Iron (Fe)
Figure 39 and 40 showed that the examined Fe fractions are represented in the
following sequence: Residual> Fe-Mn oxides> organic form > carbonate> exchangeable.
This sequence corresponding to average percentages of 80.27% for residual, 12.24% for
oxides, 7.42 for organic, 0.07% for carbonate and 0.009% for exchangeable Fe. The
exchangeable fraction and the carbonate fraction represented only minor portions from the
total concentration (less than 1%) which considered very low, limiting its potential toxicity as
pollutant. These results are in agreement with the result of Wei et al. (2007)(13) who found
that a negligible amount of Fe was found in nonstable phase compared to the total
concentration, suggesting relatively low mobility of this element. Residual fraction
dominated the different phase of Fe in all the studied sites. The high average pecentage of
residual iron fraction (80.27%) reflects the natural background of iron in the investigated area.
The residual fraction corresponding to the metal fraction in the mineral crystalline matrix is
strongly retained and would not normally be released into solution under natural
environmental conditions (Morillo et al., 2007)(163). This indicates that iron is tightly bound to
the sediment minerals and it is not easily released to the environment. The high content of
both total Fe (13255.685 µgg-1 on average) and residual Fe (10388.84 µgg-1 on average) may
emphasize the strong influence of water discharge and sediment influx from the river Nile
into the Mediterranean Sea (Nasr et al., 1990)(325). The high percentage of Fe in the residual
fraction indicates that most of the iron exists as crystalline iron oxides (goethite, limonite,
magnetite, etc). Large amounts of Fe accumulate in the residual fraction, probably because
this is an element basically of natural origin (it is one of the most common elements in the
earth's crust) (Morillo et al., 2007)(163). These results were also in agreement with the results
of Usero et al. (1998)(169). The metal present in the residual fraction can be used as a baseline
data for the assessment of the degree of contamination of the system. The association
-192-
between metals and the residual fraction of uncontaminated soils is so strong that metal
association with non-residual fraction has been used as an indicator of anthropogenic
enrichment (Sutherland et al., 2000)(556).
Table 7 showed that the association of Fe with different fractions in the studied
regions was in order: western region, residual (3892.37µgg-1, 72.68%)> Fe-Mn oxides
(525.47µgg-1, 19.08%)> organic form (357.10µgg-1, 8.12%)> carbonate (2.0 µgg-1, 0.11%)>
exchangeable (0.20 µgg-1, 0.012%), the middle region, residual (19967.46 µgg-1, 90.41%)>
orgnic form (1294.69 µgg-1, 5.85%)> Fe-Mn oxides 825.74 µgg-1, 3.73%) > carbonate (2.16
µgg-1, 0.01%)> exchangeable (0.43µgg-1, 0.002%), the eastern region, residual (12594.33
µgg-1, 86.50%)> organic form (631.85µgg-1, 9.02%) > Fe-Mn oxides (475.78µgg-1, 4.39%) >
carbonate (3.39µgg-1, 0.07%) > exchangeable (0.27µgg-1, 0.01%).
Manganese (Mn)
Figure 41 showed that the examined Mn fractions are represented in the following
sequence: Residual > Fe-Mn oxides > carbonat > organic form > exchangeable. This
sequence corresponds to the average percentage of 44.22%, 43.08%, 8.83, 3.75%, and 0.12%,
respectively (Figure 42).
Table 8 showed that the association of Mn with different fractions in the studied
regions was in the order: western region, Fe-Mn oxides (95.04 µgg-1, 65.01%)> residual
(53.53 µgg-1, 23.06%)>, carbonate (17.46 µgg-1, 7.54%)> organic form (6.46 µgg-1, 4.31%)>
exchangeable (0.15µgg-1, 0.08%), middle region, residual (337.35µgg-1, 71.64%)> Fe-Mn
oxide (77.32µgg-1, 15.90%) > carbonate (43.58µgg-1, 8.99%) > organic form (16.00µgg-1,
3.29%) > exchangeable (0.73 µgg-1, 0.17%), eastern region, residual (309.97µgg-1, 64.55%),
> Fe-Mn oxides (30.96 µgg-1, 17.60%)> carbonates (54.24µgg-1, 15.43%) > organic form
(13.02µgg-1, 2.26%)> exchangeable (0.56µgg-1, 0.16%).
The data revealed that Mn fractionation was similar in the exchangeable fraction for
all the three regions (<1%). The concentration of Mn associated with carbonate was
relatively high in the sediments of the eastern region (15.43% on the average), may be due to
anthropogenic origin. Caplat et al. (2005)(557) found that manganese is bound to the
nonresidaul and the residual fractions of the sediments. Mn is mainly present in the acid
soluble fraction (carbonate form) and significant affinity for the residual fraction (up to 50%).
Also, it is noted from the results that the relatively high % of Mn in the western region
(65.01%) were found in the Fe-Mn oxides form. Relatively higher concentration of Mn
associated with the Fe–Mn oxides may caused by the adsorption of this metal by the Fe-Mn
colloids (Purushothaman and Chakranpani, 2007)(273). These results were in agreement with
the results of Faragallah and Khalil (2009)(326) they found that manganese is principally
present as amorphous oxides rather than carbonate. This indicated that Mn comes from
anthropogenic sources, which increases potential Mn mobility and bioavailability.
In general, Mn can be present as oxides, hydroxides or in association with iron oxides
and hydroxides. Mn ions can be adsorbed and / or partially ion exchanged on the surface of
MnO2 and they can be environmentally mobile in certain conditions (Weisz et al., 2000)(558).
Relatively low percentage of Mn was found to bind to organic matter and sulfides (4.31, 3.29,
and 2.26% for the western, the middle, and the eastern region respectively). This indicates
that organic matter is not the primary factor impacting the behavior of Mn in the study area.
Large amounts of Mn accumulate in the residual fraction in the sediments of the middle and
the eastern region of the Egyptian Mediterranean coastal area (71.64 and 64.55%,
-193-
respectively). The dominance of the residual fraction for Mn in the sediments may be
attributed to most of Mn comes from the parent material of geologic origin.
Zinc (Zn)
The different fraction species and the total Zn content extracted from the surfacial
sediments along the Egyptian Mediterranean coastal area are shown in Figure 43. This Figure
showed that the examined Zn fractions are represented in the following sequence: Residual >
Fe-Mn oxides > organic > carbonate form > exchangeable. Figure 44 showed that this
sequence corresponding to average percentages of 71.70% for residual, 18.24% for Fe-Mn
oxides, 5.79% for organic, 3.79% for carbonate and 0.47% for exchangeable Zn.
The association of Zn with different fractions in the studied regions was observed at
Table 9, in the order: western region, residual (11.44µgg-1, 59.20%)> Fe-Mn oxides (4.93
µgg-1, 27.75%), carbonate (1.02 µgg-1, 6.22%)> organic forms (1.42 µgg-1,
6.19%)>exchangeable (0.06µgg-1, 0.63%), middle region, residual (14.68µgg-1, 87.25%)>
organic form (1.10µgg-1, 6.14%)> Fe-Mn oxides (1.05 µgg-1, 6.12%)>carbonates (0.06µgg-1,
0.40%)>exchangeable (0.01 µgg-1, 0.10%), eastern region, residual (10.62µgg-1, 86.03%), >
Fe-Mn oxides (0.62µgg-1, 8.34%)> organic form (0.51µgg-1, 2.41%) > carbonate (0.08µgg-1,
2.33%)> exchangeable (0.03µgg-1, 0.89%).
The data revealed that the fractionation of Zn was similar in the exchangeable fraction
for all the three regions (<1%) except station 2 in Baghoush (3.84%; western region) and
station 20 in Rafah (1.61%; eastern region). The carbonate fraction was relatively low (less
than 10%) except at station 6 in the western region (14.89%). Zinc showed relatively high
percentage in the reducible (Fe-Mn oxides) fraction (27.72%) in the western region. The
relatively high percentage of Zn in this fraction is in agreement with the known ability of
amorphous Fe-Mn oxides to scavenge Zn from solution (Ramos et al., 1999)(559). This
relatively high percentage of Zn in the reducible fraction of the western region are a hazard
for the aquatic environment because Fe and Mn species can be reduced into the porewater
during early diagenesis, microbially mediated redox reactions (Canfield, 1989)(560).
Dissolution will also release the Zn associated with oxide phases to the porewater, possibly to
the overlying water column (Petersen et al., 1995) and to the benthic biota (Jones and Turki,
1997)(561). Abdallah (2007a)(292) stated that zinc mostly partitioning in the Fe-Mn oxides
reducible and in the residual fractions, wherease in the carbonate and organic matter fractions
slightly decreased with increasing concentrations in the sediments. Residual fraction is the
first most important fraction of Zinc in the western region with percent concentration
(average: 59.20%). While in the middle and the western regions, the residual form was the
predominant for Zn (87.25 and 86.03% respectively). Relatively low percentage of Zn was
found to bind to organic matter and sulfides (6.19, 6.14, and 2.41% for the western, the
middle, and the eastern region respectively). This indicates that organic matter is not the
primary factor impacting the behavior of Zinc in the study area.
Nickel (Ni)
The different fraction species and the total Ni content extracted from the surfacial
sediments along the Egyptian Mediterranean coastal area are shown in Figure 45. This Figure
showed that the examined Ni fractions are represented in the following sequence: Residual >
Fe-Mn oxides > organic form > exchangeable > carbonate. Figure 46 illlustrated that this
sequence corresponding to average percentages follows: 80.05% for residual, 10.32% for FeMn oxides, 6.91% for organic form, 1.91% for exchangeable and 0.81% for carbonate bound
Ni.
-194-
The association of Ni with different fractions in the studied regions was listed at Table
10 in the following order: western region, residual (5.64 µgg-1, 76.96%)> Fe-Mn oxides (0.59
µgg-1, 11.91%), organic form (0.45 µgg-1, 6.61%)> exchangeable (0.12 µgg-1, 3.19%)>
carbonate (0.07µgg-1, 1.35%), middle region, residual (30.85µgg-1, 81.75%)> Fe-Mn oxides
(3.45µgg-1, 9.47%) > organic form (3.00µgg-1, 8.26%) > exchangeable (0.14µgg-1, 0.36%) >
carbonates (0.06 µgg-1, 0.16%), eastern region, residual (26.05µgg-1, 91.12%), > Fe-Mn
oxides (1.46µgg-1, 4.56%)> organic form (1.30µgg-1, 3.87%) > exchangeable (0.11µgg-1,
0.34%)> carbonate (0.04µgg-1, 0.12%).
From the obtained data, relatively low percentage of Ni was found to bind to
exchangeable and carbonate forms for the western, the middle, and the eastern region. While
the bound to Fe-Mn oxides form showed relatively moderate percentage of Ni in the western
and the middle region (average: 11.91 and 9.47%, respectively). The high average pecentage
of residual Ni fraction (91.12%, eastern region; 81.75, middle region and 76.96, western
region) reflects the natural background of Nickel in the investigated area. These results were
inagreement with someother studies (e.g. Martin et al., 1998; Yuan et al., 2004)(562, 563).
Silicate and residual fraction of metals are generally much less toxic for organisms in aquatic
environment because this fraction is chemically stable and biologically inactive (Zakir and
Shikazono, 2011)(264). Francois (1988)(483) showed that Ni distribution was primarily
controlled by the mineralogical sediments more than its association with organic matter
accumulation. These results were in agreement with the results of Verlimirovic et al.
(2010)(564) who mentioned that the dominant proportions of nickel (59.1%) were found in the
residual fraction, which indicates low availabilty. The residual fraction of metals is most
strongly associated to the crystalline structures of the minerals, presenting small risk to the
environment, and this holds for Ni. This is in accordance with literature data (Kartal et al.,
2006)(565).
Chromium (Cr)
The different fraction species and the total Chromium content extracted from the study
area are represented in Figure 47. Results showed that the examined Cr fractions are
represented in the following sequence: Residual > Fe-Mn oxides > organic form > carbonate
> exchangeable. As shown in Figure 48 Cr percentages were: 86.17% for residual, 8.14% for
Fe-Mn oxides, 3.61% for organic form, 1.75% for carbonate and 0.33% for exchangeable Cr.
The average values and its relative percentage of the five fractions in the studied
regions (Table 11) can be arranged as the following: western region, residual (17.63 µgg-1,
76.63%)> Fe-Mn oxides (1.48 µgg-1, 14.34%), organic form (0.74 µgg-1, 5.37%)> carbonates
(0.23 µgg-1, 3.1%)> exchangeable (0.05µgg-1, 0.56%), middle region, residual (117.34µgg-1,
97.72%)> organic form (1.54µgg-1, 1.63%) > Fe-Mn oxides (0.52µgg-1, 0.52%) > carbonate
(0.10µgg-1, 0.09%) > exchangeable (0.04 µgg-1, 0.04%), eastern region, residual (125.73µgg-1,
98.19%), > organic form (1.07µgg-1, 0.90%)> Fe-Mn oxides (0.63µgg-1, 0.74%) > carbonates
(0.07µgg-1, 0.09%)> exchangeable (0.051µgg-1, 0.08%).
The obtained data revealed that the fractionation of Cr was similar in the exchangeable
fraction for all the three regions (<1%) except station 2 and 3 in (Baghoush; 2.79% and ElNobarreya, 1.04%, respectively) in the western region, which considered very low, limiting
its potential toxicity as pollutant, despite the total concentration for this metal being higher
than the background value at most to the studied stations. Relatively low percentage of Cr
was found to bind to carbonate (less than 1%) at the middle and the eastern region indicating
its low mobility. On the other hand station 8 in the western region showed relatively
moderate percentage of Cr bound to carbonate (12.17%). There was a relatively moderate
-195-
percentage (14.34%) of Cr found in the Fe-Mn oxides in the western region, while the middle
and the eastern region showed low percentage of Cr bounded to the Fe-Mn oxides (0.52 and
0.74% respectively). Relatively low percentage of Cr was found to bind to organic matter and
sulfides (5.37, 1.63, and 0.90% for the western, the middle, and the eastern region
respectively). This indicates that organic matter is not the primary factor impacting the
behavior of chromium in the study area. The high average pecentage of residual Cr fraction
(98.19%, eastern region; 97.72, middle region and 76.63, western region) reflects the natural
background of chromium in the investigated area. This indicates that this element is tightly
bound to the sediment minerals and that it is not easily released to the environment.
Lead (Pb)
Lead is highly toxic to the environment. In the areas influenced by extensive human
activities, a considerable proportion of Pb in the environment is of anthropogenic origin,
because it is an important additive that has been widely used in many products (Gao et al.,
2010)(250). The different fraction species and the total lead content extracted from the present
study are presented in Figure 49. The examined Pb fractions are represented in the following
sequence: Residual >Fe-Mn oxides > carbonate > organic form > exchangeable. This
sequence corresponding to average percentages of 48.7% for residual, 26.05% for Fe-Mn
oxides, 13.89% for carbonates, 8.44% for organic and 2.92% for exchangeable Pb as shown
in Figure 50.
The association of Pb with different fractions in the following order: western region,
residual (4.45 µgg-1, 35.21%)> Fe-Mn oxides (5.84 µgg-1, 33.80%), carbonate (4.72 µgg-1,
21.36%)> organic form (1.18 µgg-1, 7.79%)> exchangeable (0.22µgg-1, 4.72%), middle
region, residual (5.16µgg-1, 67.47%)> Fe-Mn oxides (1.07µgg-1, 15.18%) > organic form
(0.74 µgg-1, 9.47%) > exchangeable (0.30µgg-1, 4.12%) > carbonate (0.29 µgg-1, 3.76%),
eastern region, residual form (3.42µgg-1, 57.17%), > Fe-Mn oxides (1.18µgg-1, 21.47%)>
organic form (0.38µgg-1, 8.83%) > carbonates (0.42µgg-1, 8.28%)> exchangeable (0.29µgg-1,
4.70%).
From the obtained data, it is observed that the content of Pb measured in the
exchangeable fraction were (4.72, 4.12, and 4.70%) on the average for the western, middle
and the eastern region respectively. Lead showed appreciable concentration in the carbonate
fraction, with a relative abundance value of 21.36 % in the western region. Similar results
were reported for Pb in carbonate in surface sediment samples from Naha port, Japan (Noah
and Oomari, 2006)(566). Zhang et al. (1988)(567) attributed the high carbonate bounded lead in
Huanghe estuarine sediments to the similarity of its ionic radii to that of calcium, which
allows lead to substitute calcium in the carbonate crystal lattice to form a mixed carbonate
phase. Carbonate fraction is also termed to be the most bioavailable fraction to the aquatic
organisms since the elements are loosly bound to the sediments and can be released back to
the environment (Tack and Verloo, 1995)(555). A major portion of Pb is bound to the Fe/Mn
oxides with comparable amounts associated with the carbonate fraction. The fact that Pb can
form stable complexes with Fe and Mn dioxide (Ramos et al., 1994)(568) may be the reason
why reducible Pb is more abundant than other tow non-residual fractions of Pb (Gao et al.,
2010)(250). Lead was dominant in the reducible fraction, which has been proved to be sensitive
to anthropgenic inputs (Modak et al., 1992)(569), and it represented 33.80, 15.18, and 21.47%
of the total Pb for the western, middle and eastern region respectively. Amorphous Fe
oxyhydroxide phase is well recognized for its scavenging properties of metals in the
environment (Zakir and Shikazono, 2011)(264). This phase accumulated metals from the
aqueous system by the mechanism of adsorption and co-precipitation (Bordas and Bourg,
2001)(570). Fe and Mn hydrous oxides are important scavengers of Pb in sediments (Wong et
-196-
al., 2007)(571). Lead species are strongly sorbed to Fe/Mn oxides, which are reported to be
more important than association with clays and organic materials (Fergusson, 1990)(3). The
sorption of lead onto Fe/Mn oxides is not affected by aging, but does decrease with an
increase in Cl¯ ion concentration, due to the formation of lead chloride complexes. As a
result, lead is immobile in the aquatic environment and tends to accumulate in sediments
close to its point of entry (Morrison, 1996)(68). The metal in this fraction may be released if
the sediments is subjected to more reducing conditions (Singh et al., 2005)(266).
-197-
Concentration (µgg-1)
12000
10000
8000
6000
4000
2000
0
F1
F2
F3
F4
F5
Figure 39: The average concentration of Fe speciation µgg-1 in surfacial sediments from
the Egyptian Mediterranean coast
0.01%
0.07%
12.24%
7.42%
80.27%
F1
F2
F3
F4
F5
Figure 40: The relative percentage of Fe species in surfacial sediments from the
Egyptian Mediterranean coast
-198-
Concentration (µgg-1)
200
150
100
50
0
F1
F2
F3
F4
F5
Figure 41: The average concentration of Mn speciation µgg-1 in surfacial sediments from
the Egyptian Mediterranean coast
8.83%
0.12%
44.22%
43.08%
3.75%
F1
F2
F3
F4
F5
Figure 42: The relative percentage of Mn species in surfacial sediments from the
Egyptian Mediterranean coast
-199-
Concentration (µgg-1)
14
12
10
8
6
4
2
0
F1
F2
F3
F4
F5
Figure 43: The average concentration of Zn speciation µgg-1 in surfacial sediments from
the Egyptian Mediterranean coast
0.47%
3.79%
18.24%
5.79%
71.70%
F1
F2
F3
F4
F5
Figure 44: The relative percentage of Zn species in surfacial sediments from the
Egyptian Mediterranean coast
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Concentration (µgg-1)
20
15
10
5
0
F1
F2
F3
F4
F5
Figure 45: The average concentration of Ni speciation µgg-1 in surfacial sediments from
the Egyptian Mediterranean coast
1.91%
0.81%
10.32%
6.91%
80.05%
F1
F2
F3
F4
F5
Figure 46: The relative percentage of Ni species in surfacial sediments from the
Egyptian Mediterranean coast
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Concentration(µgg-1)
70
60
50
40
30
20
10
0
F1
F2
F3
F4
F5
Figure 47: The average concentration of Cr speciation µgg-1 in surfacial sediments from
the Egyptian Mediterranean coast
0.33%
1.75%
8.14%
3.61%
86.17%
F1
F2
F3
F4
F5
Figure 48: The relative percentage of Cr species in surfacial sediments from the
Egyptian Mediterranean coast
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Concentration (µgg-1)
5
4
3
2
1
0
F1
F2
F3
F4
F5
Figure 49: The average concentration of Pb speciation µgg-1 in surfacial sediments from
the Egyptian Mediterranean coast
2.92%
13.90%
48.70%
26.05%
8.44%
F1
F2
F3
F4
F5
Figure 50: The relative percentage of Pb species in surfacial sediments from the
Egyptian Mediterranean coast
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6.2.3.1. Risk assessment code (RAC)
The distribution of metal speciation associated with different geochemical fraction is a
critical parameter to assess the potential mobility and biovailability of heavy metals in
sediments (Wang et al., 2010)(258). The RAC considers the different binding strengths of the
metals in the various sediments fractions. It assesses the availability of metals in solution by
applying a scale to the percentage of metals in the exchangeable and carbonate sediment
fractions. This classification is considered by Perin et al. (1985)(572).
The RAC classification defines risk levels as zero, low, medium, high and very high,
depending on the percentage value (Jain, 2004)(275). Risk assessment code has been used to
assess environmental risks and estimate possible damage to benthic organisms caused by
contaminated sediments (Passos et al., 2011)(263).
Accordingly, if this value is < 1% there is no risk for the aquatic system, 1-10%
indicates low risk, 11-30 % medium risk, 31-50% high risk, and > 50% very high risk (Table
26) (Passos et al., 2011)(263). This classification scheme is presented as the following:
Table 26: Classification of Risk assessment code (RAC)
RAC
No risk
Low risk
Medium risk
High risk
Very high risk
Criterion (%)
<1
1-10
11-30
31-50
>50
The risk assessment code of (Fe, Mn, Zn, Cr, Ni and Pb) was evaluated in the
sediments of the Egyptian Mediterranean coastal area. From the present results (Table 27), it
can be seen that the sum of the exchangeable and carbonate-bound fraction of Cr, Fe, Mn, Ni,
Pb, and Zn was 0.04-12.58 (average: 2.08%), 0.01-0.34 (average: 0.06%), 0.87-18.96
(average: 8.95%), 0.14-10.88 (average: 2.72%), 4.32-55.43 (average: 16.82%), and 0.2615.27 (average: 4.26%), respectively. This suggests that Fe has no risk to local environment,
Cr, Mn, Ni, and Zn have posed low risk to local environment, Pb was at Medium risk level.
The percentage of bioavailable forms (exchangeable and carbonate fractions) in
sediments of the Egyptian Mediterranean coastal area were lower than (10%) except for Pb
reached to 16.82% on average. The potential bioavailable form (Fe/Mn oxide and organic
matter and sulfide fractions) were higher than 10%, the highest value reached to 46.83% for
Mn. When the environmental conditions are changed, the potential bioavailable form can
dissolve or degrade and release the metal. The potential ecological risk presented by this case
can not be neglected. Elevated levels of metals in the Fe/Mn oxide and organic matter and
sulfide fractions can reflect the pollution extent caused by human activities (Liu et al.,
2009)(573).
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Table 27: Risk assessment code (RAC) (%) and risk rank for trace metals in the sediments of the Egyptian Mediterranean coastal area
Station
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
Average
Cr
RAC
1.45
12.43
3.30
2.28
0.80
3.50
1.80
12.58
0.39
0.79
0.93
0.27
0.10
0.07
0.23
0.10
0.05
0.10
0.04
0.30
2.08
Rank
Low risk
Medium risk
Low risk
Low risk
No risk
Low risk
Low risk
Medium risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
Low risk
Fe
RAC
0.032
0.256
0.342
0.029
0.038
0.119
0.017
0.063
0.035
0.069
0.035
0.027
0.015
0.007
0.005
0.008
0.025
0.011
0.021
0.148
0.068
Mn
Rank
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
RAC
4.41
3.58
4.45
8.71
6.14
6.62
6.34
2.79
16.93
4.82
18.96
6.23
0.87
4.39
12.04
11.55
13.72
15.35
13.17
17.99
8.95
Rank
Low risk
Low risk
Low risk
Low risk
Low risk
Low risk
Low risk
Low risk
Medium risk
Low risk
Medium risk
Low risk
No risk
Low risk
Medium risk
Medium risk
Medium risk
Medium risk
Medium risk
Medium risk
Low risk
Ni
RAC
3.28
9.40
3.48
2.53
6.41
7.18
3.29
10.88
1.07
0.80
1.46
0.51
0.46
0.54
0.55
0.81
0.18
0.56
0.76
0.14
2.72
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Rank
Low risk
Low risk
Low risk
Low risk
Low risk
Low risk
Low risk
Low risk
Low risk
No risk
Low risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
Low risk
Pb
RAC
14.95
19.84
47.94
9.86
55.20
55.43
24.05
6.65
5.96
8.12
7.19
6.48
8.45
13.92
8.11
8.15
5.71
4.32
11.61
14.35
16.82
Rank
Medium risk
Medium risk
High risk
Low risk
High risk
High risk
Medium risk
Low risk
Low risk
Low risk
Low risk
Low risk
Low risk
Medium risk
Low risk
Low risk
Low risk
Low risk
Medium risk
Medium risk
Medium risk
Zn
RAC
2.72
11.10
1.62
5.54
6.50
15.27
4.52
9.01
4.77
9.58
4.77
0.93
0.54
0.46
0.31
0.35
0.60
0.26
0.48
5.95
4.26
Rank
Low risk
Medium risk
Medium risk
Low risk
Low risk
Medium risk
Low risk
Low risk
Low risk
Low risk
Low risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
No risk
Low risk
Low risk
6.2.4. Toxicity assessment of trace metals in sediments based on AVS, SEM
models
Acid volatile sulfide (AVS) is one of the major chemical components that control the
activities and availability of metals in the interstitial waters of sediments. Sulfide reacts with
several divalent transition metal cations (cadmium, copper, nickel, lead, and zinc) to form
highly insoluble compounds that are not bioavailable (USEPA, 2004)(444).
Metal
bioavailability refers to the ability of metals to be accumulated by animals and plants, which
is subjected to the ingestion habits of animals, e.g., through water (pore water and overlying
water) and/or through sediments (Chapman et al., 2002)(574). Although metals bound to AVS
cann't be released from sediment into water, they may be ingested by animals. The AVS
procedure extracts metals not only from the target sulfide phase, but also from other
nonsulfide phases that may be bioavailable (Feng et al., 2005)(575).
The parameters that determine the bioavailability/toxicity of trace metals (results of
SEM and AVS analysis) in sediment samples from the Egyptian Mediterranean Coast are
presented in Table 28. From the obtained data, it was observed that the concentrations of AVS
ranged from 0.015 µmoleg-1 dry weight at station 2 in Baghoush to 31.326 µmoleg-1 dry
weight at station 7 in NIOF (western region) (average: 1.847 µmoleg-1dry weight). This is in
agreement with literature values reported from marine and fresh water sediments in the US
and Canada, which were found to range from 0.1 to 75 µmoleg-1dry weight (Di Toro et al.,
1990; Ankley et al., 1991; Ankley et al., 1993; Casas and Crecelius, 1994; Carlson et al.,
1993; Allen et al., 1993)(22, 228, 424, 576-578). Relatively small AVS levels (<2µmoleg-1) have
been found at almost all the studied locations. Station 1 in salloum has AVS concentration
(3.51µmoleg-1) and station 7 in NIOF has an AVS level of more than 10 µmoleg-1 (31.326
µmoleg-1). Hence, according to Van den Hoop et al. (1997)(579) the sediment in station 7 in
NIOF are able to bind heavy metals in large quantities. The range and average concentrations
of SEM were 0.0001-0.0011 µmole/g (mean: 0.0003 µmole/g) for Cd, 0.002-0.046 µmole/g
(average: 0.0169 µmole/g) for Cu, 0.002-0.059 µmole/g (average: 0.0186 µmole/g) for Ni,
0.004-0.033 (average: 0.0118 µmole/g) for Pb and 0.002-0.174 (average: 0.0549 µmole/g) for
Zn.
The use of different relationships between AVS and SEM to establish mechanical
models such as the ratio of SEM and AVS (SEM/AVS), the difference between SEM and
AVS (SEM-AVS) or the organic carbon normalized difference between SEM and AVS
(SEM-AVS)/ foc to assess metal toxicity has been widely applied (Di Roro et al., 1990;
Burton et al., 2005;Yin et al., 2008)(22, 246, 247). The AVS-SEM models are generally based on
the hypothesis that when there are enough sulfides in the sediment, metals will react with the
sulfide to form insoluble compounds that are not bioavailable to benthic biota; however, if the
levels of sulfide are too low, the metals will be toxic toward the benthic biota (Yin et al.,
2010)(580).
It has been verified (Di Toro et al., 1990)(22) that divalent transition metals do not
begin to cause toxicity in sediment until the reservoir of sulfide is used up (i.e. until the molar
concentration of metals exceeds the molar concentration of sulfide), typically at relatively
high dry weight metal concentrations. This observation has led to a laboratory measurement
technique for calculating the differences between SEM concentration and AVS concentration
in field samples, to determine potential toxicity. To evaluate the potential effects of metals on
benthic species, the molar concentration of AVS ([AVS], µmole g-1) is compared to the sum
of SEM molar concentrations ([SEMi], µmole g-1) for five metals: cadmium, copper, nickel,
lead, and zinc; when [SEMi]/[AVS]<1, acute toxicity due to heavy metals is not probable.
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On the other hand, the sediment may be considered potentially toxic when this ratio is >1 (Di
Toro et al., 1990)(22).
From the obtained data at Table 28 and according to Di Toro et al. (1990)(22) the
sediment may be considered potentially toxic at stations 2, 6, 8, 10, 11, 12, 13, 14, 15, 16, 18,
and 19 in Baghoush, Western Harbour, power station, Maadia, Rashid West, Rashid East,
Burullus, New Damietta, Ras El-Barr, El-Gamil East, and Port Said respectively. Acute
toxicity due to heavy metals is not probable at stations 3, 4, 5, 7, 9, 17, and 20 in Nobarreya,
Dekhaila, El-Mex, NIOF, AbuQir, El-Gamil west, and Rafah respectively.
Di Toro et al. (1990)(22) proposed that a SEM/AVS>1 indicated toxicity, then on the
contrary, the pore water concentrations would be low and no toxicity to sediment dwelling
organisms would be observed under these conditions. However, this prediction was not
always accurate when the molar sum of the SEM was greater than that of the AVS, which
could be due to the different sensitivities of the benthic biota in the sediment to metals (Hare
et al., 1994)(581).
Burton et al. (2005)(246) conducted a long period field toxicity experiment and
concluded that sediments with a SEM to AVS ratio of 8.32 or greater would result in high
macroinvertebrate toxicity, the ratios between 2 and 8.32 are occasionally toxic, and the ratio
less than 2 are not toxic. Table 28 showed that according to Burton et al. (2005)(55) the
sediments at stations 6, 8, 16, 18, and 19 (Western Harbour, Eastern Harbour, Ras El-Burr,
El-Gamil East, and Port said), respectively are occasionally toxic while the other stations are
not toxic. SEM/AVS has been endorsed as the best technology for assessing the
bioavailability of five important toxic metals (Pb, Zn, Cu, Cd, and Ni) (Allen et al., 1993)(424).
Other relationships between SEM and AVS can be expressed as the molar difference
between SEM and AVS (SEM-AVS). According to USEPA (2004)(444), the assessment of
metal toxicity is based on the difference between the corresponding SEM and AVS molar
concentrations. In this approach each sample falls into one of three categories (tiers):
associated adverse effects on aquatic life are probable (Tier 1); associated adverse effects on
aquatic life are possible (Tier 2); or no indication of associated adverse effects (Tiers 3).
According to the USEPA evaluation, when [SEM]–[AVS] is greater than 5, the sampling
site is classified as Tier 1. If [SEM] – [AVS] is between zero and 5, the sampling site is
classified as Tier 2. If [SEM]–[AVS] less than zero, the sampling site is classified as Tier 3.
From this point of view, Table 29 Shows that the values of [SEM]-[AVS] calculated at sites
2, 6, 8, 10, 11, 12, 13, 14, 15, 16, 18, and 19 in Baghoush, Western Harbour, Eastern Harbour,
Power station, Maadia, Rashid west, Rashid East, Burullus, New Damietta, Ras El-Burr, ElGamil East, and Port Said are positive and less than 5 means associated adverse effects on
aquatic life are possible (tier 2). Whereas, (SEM-AVS) calculated at sites 1, 3, 4, 5, 7, 9, 17,
and 20 (El-Salloum, Nobarreya, Dekhaila, El-Mex, NIOF, El-Gamil West, and Rafah)
respectively, are negative reflecting no indication of associated adverse effects at these
sampling sites (tier 3). This result is consistent with the findings of SEM /AVS. The
obtained data concluded that the remobility and bioavailabilty of trace metals contained in
sediments of the Egyptian Mediterranean coastal area are low.
TOC and AVS are significant parameters for the evaluation of toxicity of sediments
since they play an important role in controlling the availability of inorganic and organic
contamination and the toxicity of sediments (McGrath et al., 2002)(245). EPA has developed a
recommended approach, Equilibrium Partitioning Sediment Benchmarks (ESBs) for
estimating metal toxicity based on the bioavailable metal fraction, which can be measured in
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pore water and /or predicted based on the relative sediment concentrations of AVS, SEM and
organic carbon (TOC). Both AVS and TOC are capable of sequestering and immobilizing a
range of metals in sediments (USEPA, 2005)(582). The benchmarks application document
advises that:
(i) any sediment in which [SEM-AVS]/foc < 130 μmol/g OC should pose a low risk of
adverse biological effects due to cadmium, copper, lead, nickel and zinc
(ii) any sediment in which 130 μmol/gOC < [SEM-AVS]/foc < 3000 μmol/g OC may have
adverse biological effects due to cadmium, copper, lead, nickel and zinc
(iii) in any sediment in which [SEM-AVS]/foc >3000 μmol/g OC adverse effects due to
cadmium, copper, lead, nickel and zinc may be expected
(iv) any sediment in which [AVS] > 0.1 μmol/g OC will not cause adverse biological effects
due to chromium or silver.
The results as shown in Table 30 revealed that [SEM-AVS]/foc < 130 μmol/g at all
sampling sites. This sediments should pose a low risk of adverse biological effects due to
cadmium, copper, lead, nickel and zinc at all stations according to (USEPA, 2005)(582). The
models (SEM/AVS, SEM-AVS, SEM-AVS/foc) are useful tools for the assessment of
sediment quality (Yin et al., 2010)(580). The concentration of AVS in sediment is a product of
the equilibrium between AVS generation and the loss due to oxidation or diffusion.
Consequently, the AVS contents in sediments may vary with those factors affecting the
supply of organic matter, the rate of SO42- reduction, and the redox status of the sediments
(Oehm et al., 1997)(583). Low positive correlation (r2= 0.39) between AVS and OC as shown
in Figure 51. Therefore, organic matter may not be the major controlling factor for AVS
content in the sediments because of the lower TOC concentrations in comparison with
freshwater sediments (Song and Muller, 1999)(236). High concentrations of SO42- (20-30 mM)
in marine sediments, compared with 0.2 mM in freshwater sediments, can be responsible for
most of the anoxic oxidation process (Blazer, 1989)(584). Thus, SO42- reduction is probably
more important for AVS formation than the supply of organic matter in marine sediments
(Fang et al., 2005)(26).
OC
0.012
0.01
R2 = 0.3919
0.008
0.006
0.004
0.002
0
0
10
20
30
40
AVS
Figure 51: The relationship between OC and AVS
Although this method can be used for the detection of pollution problems in the
polluted sediments, sediment classification on the basis of these analyses should not be final.
Relation SEM/AVS>1 is not showing the actual bioavailability and toxicity, because there
are bound metals in sediment in different forms with iron and manganese (Ankely et al.,
1996)(23). When SEM/AVS>1, other binding forms for metals in sediments should be
considered to assess the bioavailability of metals. The sequential extraction procedure must
be used as an additional tool for assessing the potential bioavailability and toxicity of metals
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in sediment. Although the role of AVS is to transforms the metals from biologically available
chemical species into insoluble sulfides, the sedimentary sulfide system is of highly dynamic
nature, metals that are associated with AVS may be released within sediments through storms,
dredging ativities, oxidation, etc. and may have adverse environmental impact (Feng et al.,
2005)(575). To have better understanding of SEM and metals associated with particular
geochemical phases, sequential extraction procedure was conducted.
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Table 28: The concentrations of AVS, SEM (µmoleg-1), and the ratio of SEM and AVS in surface sediments of the Egyptian
Mediterranean coastal area.
Station
AVS
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
3.505
0.015
0.036
0.26
0.335
0.032
31.326
0.058
0.576
0.123
0.116
0.049
0.052
0.044
0.11
0.043
0.071
0.038
0.029
0.119
SEM SEM/AVS
0.1513
0.0159
0.0162
0.1465
0.2408
0.0661
0.2014
0.2063
0.0892
0.1401
0.1342
0.0501
0.0640
0.0485
0.1329
0.1220
0.0448
0.0874
0.0792
0.0122
0.043
1.060
0.449
0.563
0.719
2.065
0.006
3.557
0.155
1.139
1.157
1.023
1.231
1.102
1.208
2.836
0.630
2.300
2.731
0.102
Toxicity according to Di Toro et al. (1990)(22)
acute toxicity due to heavy metals is not probable
the sediment may be considered potentially toxic
acute toxicity due to heavy metals is not probable
acute toxicity due to heavy metals is not probable
acute toxicity due to heavy metals is not probable
the sediment may be considered potentially toxic
acute toxicity due to heavy metals is not probable
the sediment may be considered potentially toxic
acute toxicity due to heavy metals is not probable
the sediment may be considered potentially toxic
the sediment may be considered potentially toxic
the sediment may be considered potentially toxic
the sediment may be considered potentially toxic
the sediment may be considered potentially toxic
the sediment may be considered potentially toxic
the sediment may be considered potentially toxic
acute toxicity due to heavy metals is not probable
the sediment may be considered potentially toxic
the sediment may be considered potentially toxic
acute toxicity due to heavy metals is not probable
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Toxicity according to
Burton et al. (2005)(55)
Not toxic
Not toxic
Not toxic
Not toxic
Not toxic
Occasionally toxic
Not toxic
Occasionally toxic
Not toxic
Not toxic
Not toxic
Not toxic
Not toxic
Not toxic
Not toxic
Occasionally toxic
Not toxic
Occasionally toxic
Occasionally toxic
Not toxic
Table 29: The concentrations of AVS, SEM (µmole/g), and the SEM-AVS in surface
sediments of the Egyptian Mediterranean coastal area.
Sample
AVS
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
3.505
0.015
0.036
0.26
0.335
0.032
31.326
0.058
0.576
0.123
0.116
0.049
0.052
0.044
0.11
0.043
0.071
0.038
0.029
0.119
SEM SEM-AVS
0.1513
0.0159
0.0162
0.1465
0.2408
0.0661
0.2014
0.2063
0.0892
0.1401
0.1342
0.0501
0.0640
0.0485
0.1329
0.1220
0.0448
0.0874
0.0792
0.0122
-3.354
0.001
-0.020
-0.114
-0.094
0.034
-31.125
0.148
-0.487
0.017
0.018
0.001
0.012
0.004
0.023
0.079
-0.026
0.049
0.050
-0.107
Toxicity according
to USEPA (2004)(444)
Tier 3
Tier 2
Tier 3
Tier 3
Tier 3
Tier 2
Tier 3
Tier 2
Tier 3
Tier 2
Tier 2
Tier 2
Tier 2
Tier 2
Tier 2
Tier 2
Tier 3
Tier 2
Tier 2
Tier 3
Tier 2; associated adverse effects on aquatic life are possible
Tier 3; no indication of associated adverse effects
Table 30: The concentrations of AVS, SEM (µmole/g), and the SEM-AVS/fOC in surface
sediments of the Egyptian Mediterranean coastal area.
Station AVS SEM OC/g
(SEM-AVS)/foc
3.505 0.1513 0.0027
-1242.113
1
0.015 0.0159 0.0032
0.282
2
0.036 0.0162 0.0072
-2.753
3
0.26
0.1465 0.0034
-33.385
4
0.335 0.2408 0.0031
-30.380
5
0.032
0.0661
0.0038
8.964
6
31.326 0.2014
0.01
-3112.458
7
0.058 0.2063 0.0056
26.482
8
0.576 0.0892 0.0022
-221.267
9
0.123 0.1401 0.0079
2.160
10
0.116 0.1342 0.0028
6.508
11
0.049 0.0501 0.0005
2.265
12
0.052 0.0640 0.0011
10.921
13
0.044 0.0485 0.0018
2.498
14
0.11
0.1329 0.0029
7.898
15
0.043 0.1220 0.0014
56.398
16
0.071 0.0448 0.0019
-13.812
17
0.038 0.0874
0.002
24.709
18
0.029 0.0792 0.0014
35.858
19
0.119 0.0122
0.002
-53.401
20
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6.3. Ecological risk assessment
The ecological risk assessment (REA) approach provides an adequate framework for
estimating the effects associated to the occurrence of contaminants in sediments. ERA is a
process for evaluating the likelihood that adverse ecological effects may occur or are
occurring as a result of organism exposure to one or more environmental stressors (USEPA,
1998)(330). A screening ERA is typically the initial phase of the process, aiming at identifying
stressors and receptors that pose potentially unacceptable risks and those which can be ruled
out (Duke and Taggart, 2000)(585). Accordingly, the effects in the evaluation phase are often
estimated by the application of the appropriate quality benchmarks, namely the suitable
environmental quality criteria. The term criteria refers to a definitive value, refined over time
based on region-specific effects and contaminant information, but unfortunately, these values
are not always available (Chapmand et al., 1999)(586). In these circumistances, values that are
more generic are selected in the screening ERA, generally adopted from other regions. In the
case of sediment contamination, several sets of sediment toxicity benchmarks, the sediment
quality guidelines (SQGs), are usually adopted (Chapman and Mann, 1999)(509).
The basic approach used to derive the risk assessment in the Egyptian Mediterranen
coast was shown in Chapter II (review; section 2.4). The ecological risk assessment is
generally conducted in a three step process: problem formulation, Analysis step and risk
characterization phase. The proceeding sections will discuss in details the formulation of the
ERA in the Egyptian Mediterranean Sea.
6.3.1. Problem formulation
In the problem fromulation, the purpose of the assessment is defined and a plane for
analyzing and characterizing the risks is determined, generating a conceptual model and
defining the endpoints (Gomez-Gutierrez et al., 2007)(587). In this study the conceptual model
is based on the hypothesis that industrial, urban and agricultural land based sources of
contamination have contributed to the release of trace metals into the Egyptian Mediterranean
coast, which may find their ultimate fate in the sediments. Thus, the assessment will consist
on the recognition of possible adverse effects of trace metals towards the marine ecosystem,
and particularly the benthic community, which is to be the most susceptible to the sediment
bound contaminants.
A full description of the Egyptian Mediterranean Sea is given in Chapter III
(Socioeconomic and environmental aspects).
Selected assessment and measurement
endpoints in the present study are given in Table 31. The assessment will consist on the
recognition of possible generic adverse effects of the investigated chemicals (trace metals)
towards the marine ecosystem, and particularly the benthic community, which is to be the
most susceptible to the sediment bound contaminants (Jones et al., 1999)(588).
In the present study, measurement endpoints depended on measuring the chemical
concentrations of the target pollutants in the sediments. For toxicity to occur, a contaminant
must be present, a receptor must be present, and there must be a complete exposure pathway
by which the receptor is exposed to the contaminant. The conceptual site model indentifies
where contaminant interactions with biota can occur. A conceptual model describes the fate
and transport of site contaminants through the physical system (sediment of the Egyptian
Meditrranean coast) and key receptor exposure pathways. The conceptual site model used in
the Ecological risk assessment of the Egyptian Mediterranean Sea is presented in Figure 52.
As shown in Figure 52, the selected exposure media was the Sea sediments. The ecological
receptors included in this risk assessment were the benthic organisms, although they were not
-212-
investigated directly. The selected exposture route was the direct contact of the benthic
organisms with the bay sediments.
Table 31: Assessment and measurement endpoints for the SERA of the Egyptian
Mediterranean coast.
Receptor Assessment endponts
Receptor
Measurement endpoint
type
- Trace metals concentrations
(site)
Sediments
Sediment quality
Abiotic
- Sediment quality guidelines for
protection of aquatic life
(collected from literature)
6.3.2. Analysis setup
The analysis part is designed to characterize the exposure, determining how likely is to
occur and, given this exposure, the potential and type of ecological effects that can be
expected (Gomez-Gutierrez et al., 2007)(587).
6.3.2.1. Exposure characterization
The assessment of the exposure was based on the concentrations of trace metals found
in the Egyptian Mediterranean sediments. Concentrations were reported in µgg-1 dry weight.
Only trace metals with available guidelines were evaluated in the SERA. The concentrations
of trace metals in sediment samples are given in Chapter V (Results; Table 5). The range and
average ± SD concentrations (µgg-1) were: 2.05-62.21 (22.19 ±15.84) for Zn, 0.46-26.26
(8.46 ±6.22) for Cu, 1.65-60.25 (25.93 ±20.96) for Ni, 4.08-297.95 (82.74 ±90.18) for Cr,
3.34-53.67 (13.17 ±11.90) for Pb, 0.04-0.47 (0.22 ±0.15) for Cd.
The maximum concentration for each trace metal was used in the SERA as a
conservative assumption.
The maximum concentrations were: (62.21µgg-1) for Zn,
-1
(26.26µgg ) for Cu, (60.25µgg-1) for Ni, (297.95µgg-1) for Cr, (53.67µgg-1) for Pb and
(0.47µgg-1) for Cd
6.3.2.2. Effect characterization
The effect characterization was based on several sets of sediment quality guidelines
(SQGs) reported in literature. These SQGs have been calculated using different theoretical
and empirical approaches, such as the equilibrium partitioning (EqP) method (Di Toro et al.,
1991; USEPA, 1994)(589, 590), the screening level concentration approach (Neff et al.,
1988)(591), the effects range approach including the effect range low (ERL) and the effect
range medium (ERM) (Long et al., 1995)(199), the effects level approach (ELA) including the
threshold (AET) approach including the apparent effect threshold low (AET-L) and the
apparent effects threshold high (AET-H) (Barrick et al., 1988)(592) and the logistic regression
models including T20, which represents the probability of occurrence of adverse ecological
effects in 20% of the investigated samples, T50 and T80 (USEPA/NOAA, 2005)(593). In most
of the developed sediment quality guidelines, more than one value is available, one
representing the low level of adverse ecological effects (ERL, ETL and T20) and the other
representing higher level of adverse ecological effects (ERM, PEL, T50 and AET-L). For
some sediment quality guidelines, values representing higher level of adverse ecological
effects are available (T80 and AET-H), where extreme adverse ecological effects could be
expected.
-213-
A full assessment of the impacts of trace metals in the Egyptian Mediterranean coastal
area using the different methods of SQGs were previously discussed in this chapter (section
6.3). As a result of the different approaches mentioned above, SQGs reported for different
pollutants exhibit a great variability, so that some authors have developed the consensus based
approach trying to harmonize the existing values (MacDonald et al., 2000)(503), which was
used in the present study for the risk calculations and for pointing out the presence of any hot
spots in the Egyptian Mediterranean Coast. Consensus based sediment quality values
developed for the investigated pollutants in the Egyptian Mediterranean coast sediments,
together with the sediment quality guidelines used in their calculation are given in table 30.
Because consensus based marine SQGs were not available for all the target metals, values
were calculated for all the target metals in the present study following the methodology
described by MacDonald et al. (2000)(503).
In the consensus approach (CA), consensus-based SQGs are derived from the existing
SQGs that have been established for the protection of sediment-dwelling organisms.
Derivation of numerical sediment quality guidelines using the CA involves several steps.
First, the SQGs for the protection of sediment-dwelling organisms are grouped into two
categories according to their original narrative intent, including threshold effect
concentrations (TECs) and probable effect concentrations (PECs). TEC group included ERL
(Long and Morgan, 1991; Ingersoll et al. 1996a, b)(594-596), TEL (Smith et al. 1996; Ingersoll
et al. 1996 a, b)(595-597) and T20 (USEPA/NOAA, 2005)(593). The TECs are intended to
identify concentrations below which harmful effects on sediment-dwelling organisms are
unlikely to be observed. The PECs are intended to identify concentrations above which
harmful effects on sediment-dwelling organisms are likely to be frequently or always
observed (MacDonald et al. 1996; Swartz, 1999)(510, 598). PEC group included PEL (Smith et
al., 1996; Ingersoll et al., 1996a, b)(595-597), ERM (Long and Morgan, 1991; Ingersoll et al.,
1996a)(594, 595), T50 (USEPA/NOAA, 2005)(593) and AET-L (Barrick et al., 1988)(592).
Following classification of the existing SQGs, consensus-based TECs were calculated
by determining the geometric mean of the SQGs that are included in this category. Likewise,
consensus-based PECs were calculated by determining the geometric mean of the PEC-type
values. The geometric mean, rather than the arithmetic mean, was calculated because it
provides an estimate of central tendency that is not unduly affected by outliers and because
the SQGs may not be normally distributed. Consensus-based TECs or PECs are calculated
only if three or more published SQGs are available for a chemical substance or group of
substances (MacDonald et al., 2000)(503).
Recent evaluations based on combining several sets of guidelines into one to yield
"consensus-based" guidelines have shown that such guidelines can substantially increase the
reliability, predictive ability, and level of confidence in using and applying the guidelines
(MacDonald et al., 2000; Ingersoll et al., 2000)(503, 599). The agreement of guidelines derived
from a variety of theoretical and empirical approaches helps to establish the validity of the
consensus-based values. Use of values from multiple guidelines that are similar for a
contaminant provides a weight-of evidence for relating to actual biological effects. A series of
papers were produced (Swartz, 1999; MacDonald et al., 2000)(503, 598) that addressed some of
the difficulties associated with the assessment of sediment quality conditions using various
numerical sediment quality guidelines. The results of these investigations demonstrated that
combining and integrating the effect levels from several sets of guidelines to result in
consensus based sediment quality guidelines provide a unifying synthesis of the existing
guidelines, reflect causal rather than correlative effects, and can account for the effects of
contaminant mixtures in sediment (Swartz, 1999)(598).
-214-
Aquatic receptors
Point
source
1- Nobarreya drain
2- Abu Qir Drain (ElAmiaa)
2- Lake Mariout (El-Mex
outfall)
3- Lake Idku outlet
4- Burullus outlet
5- Lake Manzala outlet
(El-Gamil outlet)
6- Rosetta (River Nile
Estuary)
7- Damietta (River Nile
Estuary)
Exposure
Media
fish
h
Benthic
s
Ingestion
Industrial, domestic and
agricultural effluents
Local
residents
I
Water
Direct
contact
X
Ingestion
Industrial effluents
discharged directly into the
sea
Sediments
Direct
contact
I
X
I
Amosheric input
Non
Point
source
Aquatic
food items
fish &
benthics
Shipping activities
Ingestion
X = Pathway is complete
I = Pathway is insignificant
Figure 52: Conceptual site model for the SLERA in the Egyptian Mediterranean Coast
-215-
X
Additionally, MacDonald et al. (2000)(503) have evaluated the consensus based effect
levels for reliability in predicting toxicity in sediments by using matching sediment chemistry
and toxicity data from field studies conducted throughout the United States. The results of
their evaluation showed that most of the consensus-based threshold effect concentrations
(TEC -lower effect level) and probable effect concentrations (PEC - upper effect level) for
individual contaminants provide an accurate basis for predicting the absence or presence,
respectively, of sediment toxicity.
The consensus-based SQGs as developed only involve effects to benthic
macroinvertebrate species. The guidelines do not consider the potential for bioaccumulation
in aquatic organisms and subsequent food chain transfers to humans or wildlife. Where
bioaccumulative compounds are involved, the consensus-based SQGs need to be used in
conjunction with other tools, such as bioaccumulation-based guidelines, bioaccumulation
studies, food chain modelling, and tissue residue guidelines to evaluate the direct toxicity and
upper food chain effects of these compounds.
6.3.3. Risk characterization
The risk characterization integrates the exposure and effect characterization to assess
whether chemical concentrations are sufficiently high to pose unacceptable risks to ecolocigal
receptors. It should be emphasized that this SLERA, where possible, incorporated
conservatism where uncertainities were apparent, which is typical for a screening analysis (i.e.
risks are likely to be overestimated rather than underestimated). This allows for chemicals
posing negligible risk to be confidently removed from further evaluations. The chemicals
indentified as being of potential concern (COPCs) have to be evaluated further in more
detailed site-specific assessement to further characterize the risks they pose (Khairy,
2008)(226).
In the present study, risk was characterized by compring the maximum concentration
of each metal with its corresponding sediment quality guidelines (Table 32). The selected
approach was the hazard quotient (HQ).
The detailed approach used to evaluate the investigated contaminants in the SLERA is
shown in Figure 53. Two HQs were calculated for each pollutant; TEC HQ, which was
calculated by dividing the maximum concentration of each pollutant by the calculated
consensus-based TEC (CBSQGTEC) and the PEC HQ, which was calculated by dividing the
maximum concentration of each pollutant by the calculated consensus-based PEC
(CBSQGPEC) (Figure 53). When TEC HQ < 1 for a given pollutant, rare adverse ecological
effects are expected to occur with respect to this pollutant, when PEC HQ > 1, frequent
adverse ecological effects are expected to occur and when TEC HQ > 1 > PEC HQ, adverse
ecological effects are possible but less frequent than the previous level (Figure 53).
-216-
Table 32: Summary of the effect concentration levels from different SQGs applied in the SERA of the Egyptian Mediterranean coast
Extreme effect
Threshold effect concentration
Probable effect concentration
Trace
Max. conc.
concentrations
metal present study
ERL
TEL
T20
TEC‫ذ‬
ERM
PEL
AET-L
T50
PECa
AET-H
T80
EECa
Cd
0.47
1.20
0.680
0.380
0.700
9.60
4.21
2.70
1.40
3.50
14.0
4.90
8.30
Cr
297.95
81.0
52.3
49.0
59.2
370
160
97.0
140
168
1101
410
672
Cu
26.26
34.0
18.7
32.0
27.3
270
108
390
94.0
181
1300
280
603
Pb
53.67
46.7
30.2
30.0
34.8
218
112
430
94.0
177
1200
300
600
Ni
60.25
20.9
15.9
15.0
17.1
51.6
42.8
110
47.0
58.1
371
150
236
Zn
62.21
150
124
94.0
121
410
271
460
240
333
3800
640
1560
a: Calculated consensus based SQGs for each group calculated according to MacDonald et al. (2000)(503)
ERL, ERM (Long et al., 1995)(199)
TEL, PEL (MacDonald et al., 1996)(510)
T20, T50 and T80 (USEPA/NOAA, 2005)(593)
AET-L, AET-H (Barrick et al., 1988)(592)
-217-
TEC HQ = concentration of the pollutant
CBSQGTEC
PEC HQ = concentration of the pollutant
CBSQGPEC
TEC HQ <1
TEC HQ >1> PEC HQ
Possible but less
frequent effects
Rare effects
PEC HQ >1
Frequent effects
Use labile fraction
Stop
TEC HQ <1
Stop
TEC HQ >1> PEC HQ
Contaminants of potential
concern (less concern)
PEC HQ >1
Contaminants of potential
concern (max. concern)
Figure 53: SLERA approach applied in the Egyptian Mediterranean coast
-218-
Trace metals were further evaluated by calculating both HQs (TEC HQ and PEC HQ)
based on the concentration of the labile fraction (the sum of the first 4 fractions obtained from
the sequential extraction scheme). Calculated HQs for the investigated trace metals in the
study area are shown in Figure 54. TEC HQ based on the total metal concentrations for Cd,
Cu and Zn was less than 1 indicating rare effects. While, TEC HQ for Cr, Pb and Ni were
higher than 1 based the total metals concentrations; indicating frequent adverse ecological
effects are expected to occur. At the same time, the TEC HQ for Pb based on the labile
fraction was in the range (TEC HQ > 1 > PEC HQ) indicating that adverse ecological effects
are possible but less frequent with respect to Pb. While the TEC HQ for Ni and Cr based on
the labile fraction concentration was TEC HQ < 1 indicating rare effects (Stop). Based on the
calculated values of PEC HQ (Figure 54), it can be recognized that the calculated PEC value
for Pb was (TEC HQ>1>PEC HQ). Thus it can be concluded that Pb considered as
contaminant of less potential concern.
Trace metals (TEC HQ)
6
5
4
Total
Labile
3
2
1
0
Cd
Cr
Cu
Pb
Ni
Zn
Trace metals (PEC HQ)
2
1.5
Total
Labile
1
0.5
0
Cd
Cr
Cu
Pb
Ni
Zn
Figure 54: TEC HQ and PEC HQ for trace metals in the Egyptian Mediterranean
coastal sediments
Evaluation of the adverse ecological risk usually includes the calculated HQ based on
the maximum concentrations in the first stage, followed by determining the number of
samples exceeding the used quality guidelines to determine contaminants of potential concern
in a given area. Maximum detected TEC HQ and PEC HQ, together with the percentage of
samples exceeding TEC and PEC level are given in Table 33. The maximum calculated PEC
HQ for Cr and Ni was represented about 15% of the total samples.
-219-
6.3.4. Area of special concern based on the conducted SLERA
One of the principle goals of conducting this screening level ERA was the
identification of sites of special concern in the study area due to the occurrence of the
different investigated pollutants. The application of the HQ methodology is appropriate for
the screening level assessment and may provide a good visualization tool or risk indicator for
sites. The HQ was estimated by means of the two approaches considered in the present study
(TEC HQ and PEC HQ). Calculated values of PEC HQ for each pollutant in the investigated
sediment samples are shown in Table 34. It can be observed that the highest risk for adverse
ecological effects (PEC HQ>1) from sediments was located at Rashid east, El-Gamil west and
Port Said for Chromium. Also, the results revealed highest risk for adverse ecological effects
due to Ni was found at Rashid east, Ras El-Barr and El-Gamil west.
-220-
Table 33: Summary of the information obtained from the risk characterization process for pollutants in sediments of the Egyptian
Mediterranean Coast.
Trace
metal
Cd
Cu
Cr
Zn
Ni
Pb
Metal type
Total
Labile
Total
Labile
Total
Labile
Total
Labile
Total
Labile
Total
Labile
Max. PEC
HQ
297.95
1.87
60.25
4.75
-
Location of the max.
PEC HQ
Rashid east
El-Gamil west
-
% of
exceedance
of PEC HQ
0%
0%
0%
0%
15%
0%
0%
0%
15%
0%
0%
0%
-221-
Max. TEC
HQ
Location of the max.
TEC HQ
297.95
1.87
60.25
4.75
53.67
44.29
Rashid east
El-Gamil west
El-Mex Bay
% of
exceedance
of TEC HQ
0%
0%
0%
0%
50%
0%
0%
0%
55%
0%
5%
5%
Table 34: The calculated PEC HQ of the Egyptian Mediterranean Coastal sediments.
Region
Western region
Middle region
Eastern region
Station
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
Cd
0.08
0.02
0.02
0.10
0.05
0.13
0.13
0.11
0.05
0.11
0.13
0.04
0.03
0.07
0.01
0.01
0.04
0.04
0.01
0.05
Cu
0.08
0.00
0.00
0.05
0.04
0.02
0.15
0.06
0.03
0.02
0.03
0.02
0.04
0.05
0.07
0.10
0.04
0.07
0.06
0.01
Cr
0.19
0.02
0.03
0.10
0.14
0.04
0.15
0.06
0.51
0.20
0.18
0.50
1.77
0.70
0.47
0.61
1.39
0.67
1.69
0.43
Zn
0.12
0.01
0.03
0.07
0.17
0.02
0.19
0.06
0.05
0.03
0.04
0.03
0.07
0.05
0.08
0.09
0.06
0.08
0.09
0.01
Ni
0.13
0.03
0.16
0.15
0.10
0.06
0.15
0.05
0.33
0.31
0.20
0.63
0.97
0.75
0.76
0.97
1.04
0.84
0.82
0.48
Pb
0.09
0.02
0.08
0.09
0.30
0.11
0.17
0.14
0.05
0.07
0.04
0.03
0.04
0.03
0.05
0.04
0.04
0.04
0.04
0.02
6.3.5. The Potential Ecological Risk Index method
Potential Ecological Risk Index Method is simple, relative shortcut and precise (Hong
et al., 2004)(600), not only reflects the single impact of heavy metals to ecological environment
but also takes into accounts the different background values of the geography (Xin et al.,
2008)(601) and combines environmental chemistry with biological toxicology and ecology
(Guo et al., 2010)(602).
In the present study, Hakanson's ecological risk method was used to evaluate the
potential ecological risk of metal contaminants in sediments. In 1980, Lars Hakanson's
reported an ecological risk index for aquatic pollution control; therefore, Hakanson's method
has been often used in ecological risk assessment as a diagnostic tool to penetrate one of
many possible avenues towards a potential ecological risk. The method is based on the
hypothesis that a sediment ecological risk index for toxic substances should account for the
following requirements:
(1)
the potential ecological risk index (RI) increases with the metal pollution
increase in sediments;
(2)
the ecological harms of different heavy metals in sediments have
cooperatively, and the potential ecological risk of the cooperative harm is
more serious, especially for Cu, Zn, Pb, Cd, and Cr; and
(3)
toxicity response of each heavy metal is different, and those metals whose
biologic toxicity are strong have larger proportion in RI
The index is calculated as the following equations:
Cif = CiD / CiR ;
Efi = Tif x Cif ;
CH =  Cif
RI =  Eif
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In which, Cif is the pollution coefficient of single metal; CiD is the measured
concentration of sample; CiR is the background concentration of sediments; CH is the polluted
coefficient of many metals; Eif is the potential ecological risk factor of single metal; Tif is the
biological toxicity factor of different metals; and RI is the potential ecological risk index of
many metals. The highest pre-industrial background concentration of global sediments is
selected as the reference concentration in this study, Cd= 1, Cr=90, Cu=50, Pb=70, and
Zn=175 µg g-1.
The formula reveals the hazards of trace metals on the human and aquatic ecosystem
and reflects the level of heavy metal toxicity and ecological sensitivity to the heavy metal
pollution. The standardized response coefficient for the toxicity of trace metals, which made
by Hakanson (1980)(517), was adopted to be evaluation criterion, the corresponding
coefficients based on its toxicity were: Cd=30, Cu= Pb=Ni = 5, Cr=2, Zn=1 (Qinna et al.,
2005)(603).
According to Hakanson's ecological risk index method, the trace metal polluted
elements in the present sediment samples are analyzed and evaluated. Only five polluted
elements (Cu, Pb, Zn, Cd, and Cr) are investigated, which are less than eight elements
required for Hakanson's method. Therefore, the integrated pollution degree (CH) is defined as
following: CH<5, low pollution; 5≤CH<10, middle-pollution; 10≤CH<20, high pollution; CH
≥20, very highe pollution.
As shown in Table 34, the results of pollution index and integrated pollution index
indicated low pollution for the studied metals (Cd, Cr, Cu, Pb and Zn) in all the sampling
sites.
Table 35: Pollution index, integrated pollution index and pollution grade in sediments
along the Egyptian Mediterranean coastal area
Station
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
Cd
0.29
0.07
0.06
0.35
0.16
0.47
0.44
0.38
0.19
0.39
0.45
0.14
0.09
0.23
0.04
0.05
0.15
0.13
0.04
0.18
Pollution index Cif
Cr
Cu
Pb
0.35
0.28
0.23
0.05
0.01
0.06
0.06
0.02
0.21
0.18
0.19
0.23
0.26
0.15
0.77
0.08
0.08
0.29
0.28
0.53
0.43
0.10
0.22
0.35
0.95
0.11
0.14
0.38
0.06
0.17
0.34
0.10
0.10
0.93
0.07
0.07
3.31
0.14
0.09
1.30
0.20
0.08
0.88
0.24
0.13
1.14
0.35
0.10
2.6
0.15
0.09
1.24
0.24
0.11
3.16
0.23
0.10
0.80
0.03
0.05
Zn
0.22
0.01
0.05
0.13
0.28
0.04
0.36
0.11
0.09
0.06
0.08
0.06
0.14
0.10
0.15
0.17
0.11
0.15
0.17
0.01
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Integrated pollution
Integrated
index CH
pollution grade GG
1.37
Low pollution
0.20
Low pollution
0.40
Low pollution
1.08
Low pollution
1.62
Low pollution
0.96
Low pollution
2.03
Low pollution
1.17
Low pollution
1.48
Low pollution
1.06
Low pollution
1.07
Low pollution
1.27
Low pollution
3.78
Low pollution
1.91
Low pollution
1.43
Low pollution
1.80
Low pollution
3.09
Low pollution
1.87
Low pollution
3.70
Low pollution
1.08
Low pollution
The pollution degree and potential ecological risk are analyzed by the use of
potential ecological risk index method (Puente et al., 2008)(604). The evaluated standards for
the analyzing potential ecological risk factor (Eif) and potential ecological RI are given in
Table 36.
Table 36:.Relation between RI and pollution levels
Potential ecological risk factor Eif
Threshold range of single metal risk factor grade
Potential ecological risk index RI
Threshold range of five metals RI grade
< 30
30-60
60-120
120-240
>240
< 110
110-220
220-440
> 440
I Low
II Middle
III appreciable
IV high
V serious
A Low
B middle
C appreciable
D High
The evaluation of potential risk of trace metal associated within the sediments along
the Egyptian Mediterranean coastal area are summarized in Table 37. The potential ecological
risk factor (Eif) of Cd, Cr, Cu, Pb, and Zn were lower than 30, which indicated slight potential
ecological risk of all five metals in the studied stations. According to the evaluating standards,
all the stations of the study area had RI ranged from 2.68 to 18.91 (<110), implicating a low
ecological risk levels.
Table 37: Evaluation on potential risk of trace metal pollution in sediments along the
Egyptian Mediterranean costal area
Station
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
Cd
8.61
2.22
1.92
10.38
4.83
14.07
13.23
11.49
5.70
11.73
13.41
4.26
2.76
6.96
1.05
1.50
4.47
3.78
1.29
5.43
Potential ecological risk factor Eif
Cr
Cu
Pb
0.70
1.39
1.17
0.09
0.05
0.32
0.12
0.09
1.05
0.36
0.97
1.14
0.52
0.73
3.85
0.16
0.41
1.45
0.56
2.63
2.14
0.21
1.09
1.75
1.90
0.57
0.68
0.75
0.30
0.86
0.68
0.49
0.49
1.87
0.34
0.33
6.62
0.71
0.46
2.60
0.99
0.41
1.76
1.20
0.63
2.27
1.73
0.50
5.19
0.73
0.46
2.49
1.20
0.56
6.32
1.14
0.48
1.61
0.16
0.24
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Zn
0.22
0.01
0.05
0.13
0.28
0.04
0.36
0.11
0.09
0.06
0.08
0.06
0.14
0.10
0.15
0.17
0.11
0.15
0.17
0.01
RI
12.09
2.68
3.23
12.98
10.21
16.13
18.91
14.66
8.94
13.70
15.16
6.86
10.69
11.06
4.79
6.17
10.96
8.18
9.39
7.45
Risk
grade
Low
Low
Low
Low
Low
Low
Low
Low
Low
Low
Low
Low
Low
Low
Low
Low
Low
Low
Low
Low
6.4. Multivariate Statistical analyses
To assess the dynamics of trace metals and identify the contributing sources of
bioavailable trace metals in the surfacial sediments along the Egyptian Mediterranean coastal
area, multivariate statistical analyses such as factor and cluster analyses were carried out.
6.4.1. Application of principle component analysis (PCA)
Principal component analysis (PCA), a multivariate statistical technique, is generally
employed to reduce the dimensionality of a dataset while attempting to preserve the
relationships present in the original data. The PCA enables a reductions in data and
description of a given multidimensional system by means of a small number of new variables
(Loska and Wiechula, 2003)(605). Principal component analysis is one of the multivariate
statistical analytical tools used to assess metal behaviour in sediments (Zhou et al., 2004)(606),
and it is applied to detect the hidden structure and association of elements in the data set, in an
attempt to explain the influence of latent factors on the data distribution (Simeonov et al.,
2000)(607). The Eigenvalues and Eigenvectors were computed for the standardized data using
a specialized statistical software package (SPSS for windows version 17).
The results of the principal component analysis (PCA) on the data matrix obtained
from total trace metals analysis of surfacial sediments along the Egyptian Mediterranean
coastal area are shown in Table 38. Three principal components were obtained with Eigen
values greater than 1, together explaining more than 80.49% of total variance of the data, and
the variances of Factor1, Factor2 and Factor3 are 46.020%, 21.269% and 13.202%
respectively. Apparently the result of PCA corresponds well with the correlation coefficients.
Factor 1 has strong loadings of Ni, Fe, Co, Mn and Cr; correlation coefficients among this
group of elements exceed 0.7 (0.953, 0.920, 0.905, 0.888 and 0.831, respectively). Cobalt, Ni
and Cr belong to the siderophile elements, and are main rocks forming elements. It is easy for
them to enter into iron magnesium silicate minerals, because of their similar ionic radius.
This element association is considered to represent the lithology of the study area, and a
natural input, i.e., they are derived from terrigenous detrital material transported by surface
runoff.
Factor 2 indicated significant loadings in favour of Zn and Cu (0.875 and 0.858
respectively). These metals were believed to have the same sources. The impact of trace
metals occurs over a finite period of time, and many have been effectively retained in the
sediments near their sources, rather than resuspended and distributed uniformly throughout
the study area.
Factor 3, is strongly loaded with the textural parameters (sand and Mud), which are
inversely related to each other. This factor contributes 13.202% of the total variance and may
be termed as "Textural Factor".
Principal component analysis was used on the data matrix obtained from leachable
trace metals analysis of surfacial sediments along the Egyptian Mediterranean Coastal area to
quatify relationships between variables under simultaneous condition of their interactions and
the results were reported in Table 39. In this way, the number of variables under investigation
was reduced and interelement associations might be assessed in detail (Bakac, 2000)(608).
Results showed that two components (with Eigen values >1) were extracted describing
approximately 80.655% of the common variance. To interpreat a group of variables to be
associated with a particular factor, loadings greater than 0.7 were considered. The first factor
describes 46.216% of the common variance and is positively loaded by LPb, although
negative loading of LCo, LNi, LFe and LMn is also observed. The second factor, describing
-225-
34.439% of the common variance of the data set, is positively loaded by LCr and negatively
loaded by LCd.
The PCA of the metals in the five binding fractions were applied. In the case of
exchangeable fraction (F1), three components have been extracted, they accounted for
75.554% of the variance in the original data. Factor 1 has high negative association of Cr
(-0.850) and Zn (-0.758), while Factor 2 were negatively associated with Mn (-0.596) and
Factor 3 was positively associated with Ni (0.726) and negatively associated with Fe (-0.304)
(Table 40).
The principal component analysis of carbonate Fraction (F2) was shown in Table 41.
Two components have been extracted, together they accounted for 65.111% of the variability
in the original data. The factor 1 was positively associated with Zn (0.831) and Pb (0.774)
and negatively associated with Mn (-0.692) and Fe (-0.650), while the Factor 2 was
negatively associated with Cr (-0.732).
The PCA was applied on the results of trace metals associated with the Fe/Mn oxide
fraction (F3). Two components have Eignevalues were greater than 1, accounted for a
cumulative variance of 86.143% (Table 42). Factor 1 represent 51.194% of the data variance
was associated with Zn (0.986), Cr (0.927) and Pb (0.920). Factor 2 represented 34.949% of
the variance was associated with Ni (0.901) and Fe (0.883) and negatively associated with Zn
(-0.012), Cr (-0.094) and Pb (-0.226).
The plot of loading of the components in the organic fraction (F4) gives a distribution
of the metals into two components (Table 43). Factor 1 was associated with Fe (0.970), Ni
(0.952), Mn (0.948) and Cr (0.907). Factor 2 of this fraction was positively associated with
Pb (0.903) and Zn (0.822) and negatively associated with Fe (-0.082), Ni (-0.194), Mn
(-0.025) and Cr (-0.042).
The application of PCA on the data of trace metals binding to the residual fraction
(F5) (Table 44) explaining only the first two components (82.981% of the variance). The
factor 1 of this fraction represented 63.104% of the variance and was positively associated
with Mn (0.952), Cr (0.916), Fe (0.914) and Ni (0.873), while factor 2 (19.878% of the
variance) was positively associated with Zn (0.719) and negatively associated with Mn, Cr
and Ni (-0.121, -0.274, -0.362, respectively).
6.4.2. Cluster analysis
Cluster analysis was used in order to clarify the pattern of metal distributions in the
sediments of the selected area. Classification was performed using (MINITAB 14 for
windows).
Cluster analysis can be used in combination with other statistical techniques in order
to assess epidemiological relationships between variables and outcomes. Cluster analysis
refers to a set of techniques designed to classify observations so that members of the resulting
groups are similar to each other and distinct from the other groups. Hierachical clustering,
which successively joins the most similar observations, is the most common approach.
Because groups are simply based on their similarity to each other, hierarchical cluster analysis
can be useful when abundant data are available and clear hydrogeologic models have not yet
been developed. While other multivariate techniques, such as factor analysis or principal
component analysis, provide more insight into the underlying structure of a data set, the use
of these techniques might require further analyses to identify distinct groups. Cluster
-226-
analysis, on the other hand, may be thought of as a useful way of objectively organizing a
large data set into groups on the basis of a given set of characteristics. This can ultimately
assist in the recognition of potentially meaningful patterns. The set of characteristics chosen
for inclusion in the cluster analysis is assumed to include the important distinguishing
characteristics of the entities that are being clustered. It is recommended that the simplest
characteristics applicable to the data set be chosen because this is likely to ease interpretation
of the final results (Swanson et al., 2001)(609).
Cluster analysis provides a number of methods to obtain an insight into data sets, and
to extract relevant information from them. This analysis is the most important tool used for
interpreting multivariate data containing objects and features, and sometimes also properties.
Cluster analysis generically refers to different multivariate methods designed to creat
homogeneous sets of objects called cluster (Debska and Wianowska, 2002)(610). The basic
principal upon which all clusters analyses are based is very simple. All of them attempt to
group samples or objects into groups of similar objects called clusters. Objects are placed
into different clusters such that members of any other cluster are more similar to each other in
some way than that they are to members of any cluster. The major problem associated with
cluster analysis is that the techniques always produce clusters even in circumistances where
there are no natural groupings in the data. The analysis actually imposes a cluster structure on
the data. The success of the method will depend entirely on knowing whether the clusters
produced are real ones or simply artifacts of the method (Hagan and Fellowes, 2003)(611).
Figure 55 depicts a dendogram drived by group average methods and Squared Euclidean of
twenty sites based on all the parameters (Trace metals: Fe, Mn, Zn, Cu, Pb, Cr, Co, Ni;
Organic matter, Carbonate, Mud and Sand) in surfacial sediments along the Egyptian
Mediterranean coastal area. The cluster analysis results indicate six clusters: (1) Pb-Zn-Cu;
(2) TOM-CaCO3; (3) Cd; (4) Ni-Mn-Fe-Co-Cr; (5) Mud; (6) Sand in terms of similarities.
This indicates that Ni, Mn, Fe, Co, and Cr appear to have originated mainly from natural
sources. In addition, Pb, Zn and Cu seem to drive partly from sources other than Ni, Mn, Fe,
Co and Cr.
Cluster analysis allows the grouping of sampling sites on the basis of the similarities
of trace metals (Sundaray et al., 2011)(612). This analysis for metals concentrations is rendered
as dendogram (Figure 56), where 20 sampling sites of the surfacial sediments along the
Egyptian Mediterranean Coastal area are clustered into two groups depending upon the
enrichment of metals. The sites in groups have similar characteristics features and
anthropogenic/natural background source types. The first major grouping is formed by the
most stations from the western region (1, 4, 9, 11, 8, 7, 5, and 10), the Middle region (12, 13,
17, 14, 18, 15, and 16) and station 19 in the eastern region. The most relevant in this study is
the second group, formed by the four stations 20, 2, 3, and 6.
-227-
Table 38: Factor loadings on elements in surfacial sediments samples along the Egyptian
Mediterranean coastal area (n=20)
Element
Factor 1
Factor 2
Factor 3
Ni
0.953
0.036
0.034
Fe
0.920
0.268
0.136
Co
0.905
0.188
0.151
Mn
0.888
0.155
0.181
Cr
0.831
0.059
0.005
CaCO3
-0.824
-0.009
0.267
Cd
-0.649
0.114
0.072
TOM
-0.625
0.169
0.382
Pb
-0.583
0.415
0.436
Zn
-0.055
0.875
0.366
Cu
0.114
0.858
0.329
Sand
0.202
-0.675
0.702
Mud
-0.202
0.675
-0.702
Eigenvalue
5.983
2.765
1.716
% variance explained
46.020
21.269
13.202
Cumulative % variance
46.020
67.290
80.491
Table 39: Factor loadings on leachable metals in surfacial sediments samples along the
Egyptian Mediterranean coastal area (n=20)
Element
Factor 1
Factor 2
L Pb
0.801
0.522
LCo
-0.799
0.557
LNi
-0.767
0.609
LFe
-0.696
0.684
LZn
0.675
0.661
LCu
0.651
0.636
LCd
0.478
-0.065
LCr
0.641
0.705
LMn
-0.537
0.578
Eigenvalue
4.159
3.099
% variance explained
46.216
34.439
Cumulative % variance
46.216
80.655
-228-
Table 40: Factor loadings of variable (5 metals) using PCA technique of the data set in
the Exchangeable fraction of the surfacial sediments along the Egyptian Mediterranean
coastal area.
Element
Cr
Zn
Pb
Fe
Mn
Ni
Eigenvalue
% variance explained
Cumulative % variance
Factor 1
-0.850
-0.758
0.610
0.468
0.479
0.510
2.377
39.622
39.622
Factor 2
0.193
0.275
0.501
0.640
-0.596
0.103
1.139
18.977
58.598
Factor 3
0.163
0.486
0.239
-0.304
0.277
0.726
1.017
16.955
75.553
Table 41: Factor loadings of variable (5 metals) using PCA technique of the data set in
the Carbonate fraction of the surfacial sediments along the Egyptian Mediterranean
coastal area.
Element
Zn
Pb
Mn
Fe
Cr
Ni
Eigenvalue
% variance explained
Cumulative % variance
Factor 1
0.831
0.774
-0.692
-0.650
0.418
0.470
2.587
43.122
43.122
Factor 2
0.03
0.408
0.404
0.108
-0.732
0.664
1.319
21.989
65.112
Table 42: Factor loadings of variable (5 metals) using PCA technique of the data set in
the Fe/MN oxides fraction of the surfacial sediments along the Egyptian Mediterranean
coastal area.
Element
Zn
Cr
Pb
Ni
Fe
Mn
Eigenvalue
% variance explained
Cumulative % variance
Factor 1
0.986
0.927
0.920
-0.314
0.360
0.407
2.587
43.122
43.122
-229-
Factor 2
-0.012
-0.094
-0.226
0.901
0.883
0.668
1.319
21.989
65.112
Table 43: Factor loadings of variable (5 metals) using PCA technique of the data set in
the Organic fraction of the surfacial sediments along the Egyptian Mediterranean
coastal area.
Element
Fe
Ni
Mn
Cr
Pb
Zn
Eigenvalue
% variance explained
Cumulative % variance
Factor 1
0.970
0.952
0.948
0.907
0.023
0.370
3.704
61.740
61.740
Factor 2
-0.082
-0.194
-0.025
-0.042
0.903
0.822
1.538
25.634
87.373
Table 44: Factor loadings of variable (5 metals) using PCA technique of the data set in
the Residual fraction of the surfacial sediments along the Egyptian Mediterranean
coastal area.
Element
Mn
Cr
Fe
Ni
Zn
Pb
Eigenvalue
% variance explained
Cumulative % variance
Factor 1
0.952
0.916
0.914
0.873
0.443
0.497
3.786
63.104
63.104
-230-
Factor 2
-0.121
-0.274
0.032
-0.362
0.917
0.674
1.193
19.878
82.981
Similarity
59.78
73.19
86.59
100.00
Pb
Zn
Cu
TOM CaCO3
Cd
Ni
Mn
Fe
Co
Cr
Mud
Sand
Variables
Figure 55: Cluster analysis of the selected trace metals in the surfacial sediments
along the Egyptian Mediterranean coastal area
Similarity
99.94
99.96
99.98
100.00
1 12 13 19 17 14 18 15 16 4 9 11 8
Variables
7
5 10 20 2
3
6
Figure 56: Dendogram showing station groups formed by group averaging cluster
analysis of elemental concentrations
-231-
6.4.3. Correlation matrix
Pearson Correlation (PC) matrix for analysed sediment parameters was calculated to
see if some of the parameters were interrelated with each other and the results are presented in
Table 45. Examination of the matrix also provides clues about the carrier substances and the
chemical association of trace metals in the study area (Forstner and Witlmann, 1983)(613). The
correlation matrixes of the metal contents, the sand fraction, calcium carbonate (CaCO3) and
organic matter content (OM) in the surface sediments of the study area are given in Table 45.
The correlation matrixes show the differences in the inter-relationships between sediment
properties in the study area, which may relate to either compositional differences or process
variability between the areas. The data presented in Figure 57 showed that strong positive
correlation exists between Fe and (Mn, Co, Cr, and Ni) (r = 0.92, 0.96, 0.69, and 0.85
respectively). It means that these metals tend to accumulate together. The significantly
positive correlation with Fe indicates that the elements were derived from similar sources and
also moving together (Bhuiyan et al., 2009)(614). There is a negative correlation between Fe
and Pb (r = -0.3918). This result is in agreement with the results of Sohrabi et al. (2010)(615).
This result suggested that Pb was not strongly controlled by natural weathering processing
(Zhang et al., 2009)(530). The strong positive correlation exists between Zn and (Cu and Pb) (r
= 0.80 and 0.63 respectively) Figures (58 and 59 respectively). Also, Cr shows strong positive
correlation with (Ni, Fe, Mn, and Co), where (r = 0.83, 0.69, 0.79 and 0.60, respectively)
Figures (62, 57, 60 and 61 respectively). The data show that strong positive correlation exists
between Ni and (Fe, Mn, Cr, and Co) (r =0.85, 0.78, 0.83 and 0.86, respectively) Figures (57
and 62). It means that Ni strongly coupled with these metals. On the other hand it shows high
negative correlation with TCO3 (r = -0.81) indicating that Ni does not found in carbonate
form, and negative correlation with Pb and Cd (r = -0.48 and -0.59).
Carbonate gives negative correlation with most of the studied metals except Pb and Cd
(r = 0.56 and 0.39) and no correlation with Zn, and cupper (r= 0.10 and 0.007). There were
positive correlation between TOM and Cd (r= 0.49) whereas Pb had low association with total
organic matter with low positive correlation (r= 0.373). This results are in agreement with
Sohrabi et al. (2010)(615). Organic matter shows moderate positive correlation with Cd where
(r = 0.49). On the other hand it shows negative correlation with Ni, Fe, Mn, Co, and Cr (r= 0.54, -0.46, -0.40, -0.45 and -0.51). This negative correlation may indicate that these metals
are not associated with organic matter.
-232-
Table 45: Correlation matrix between heavy metals, CaCO3 and OM in sediments (n=20) along the Egyptian Mediterranean coast from
Salloum to Rafah (n= 20)
Zn
Ni
Pb
Cd
Fe
Cu
Mn
Co
Cr
CaCO3
OM
Zn
1
Ni
-0.0079
1
Pb
0.6621*
-0.5335
1
Cd
0.0620
-0.5954
0.3227
Fe
0.1850 0.8515* -0.3918 -0.5549
0.1710
0.2673
1
Mn
0.1074 0.7839* -0.4026 -0.4166 0.9198*
0.2652
1
Co
0.1159 0.8644* -0.4253 -0.5660 0.9627*
0.3426
0.8414
Cr
0.0580 0.8316* -0.4092 -0.4983 0.6919*
0.0925 0.7892* 0.6011*
CaCO3
0.1011
-0.808
0.5666
0.3911
-0.683
0.0067
-0.675
-0.686
-0.647
1
OM
0.2122
-0.537
0.3732
0.4907
-0.455
0.2452
-0.403
-0.446
-0.505
0.5803
Note:
* is significant at P< 0.05
-233-
0.1410
1
0.3887
Cu
0.8029*
1
1
1
1
Manganese
1200
1000
800
600
400
200
0
R = 0.9195
Cobalt
0
10000
20000
Iron
30
25
20
15
10
5
0
30000
40000
R = 0.9627
0
10000
20000
Iron
30000
40000
Chromium
400
300
200
100
R = 0.6918
0
0
10000
20000
Iron
30000
40000
40000
Iron
30000
20000
10000
R = 0.8512
0
0
20
40
Nickel
60
80
Figure 57: Relationship between Fe and Mn, Co, Cr, Ni
-234-
Cupper
30
25
20
15
10
5
0
R = 0.8028
0
20
40
Zinc
60
80
Lead
Figure 58: Relationship between Zn and Cu
60
50
40
30
20
10
0
R= 0.6621
0
20
40
Zinc
60
80
1200
1000
800
600
400
200
0
R = 0.7892
0
100
200
Chromium
300
400
Figure 60: Relashionship between Mn and Cr
Cobalt
Manganese
Figure 59: Relationship between Zn and Pb
30
25
20
15
10
5
0
R = 0.6011
0
100
200
300
400
Chromium
Figure 61: Relationship between Co and Cr
-235-
Chromium
400
300
200
100
R = 0.8316
0
Manganese
0
20
40
Nickel
60
1200
1000
800
600
400
200
0
80
R = 0.7839
0
20
40
60
80
Nickel
30
Cobalt
25
20
15
10
5
R = 0.8644
0
0
10
20
30
40
50
60
Nickel
Figure 62: Relationship between Ni and Cr, Mn, Co
-236-
70
Statistical analyses based on acid leachable elements were effectively utilized for the
interpretation of geochemical processes of heavy metals (Jonathan et al., 2004; Selvaraj et al.,
2004; Jayaprakash et al., 2008)(467, 616, 617). The correlation coefficients among acid leachable
trace metals and sediment properties in the Egyptian Mediterranean coastal area are listed in
Table 46. From the obtained results iron showed postitive correlation with Co (r= 0.94), and
Ni (r=0.98), indicating that acid leachable Co and Ni are mainly combined with Fe
oxyhydroxides but not Mn oxyhydroxides. The acid leachabel contents of Ni, Fe, Mn, and Co
are negatively correlated with OM (r= -0.36, -0.29, -0.26, and -0.40, respectively), on the
other hand Zn, Pb, Cd, and Cr, showed low positive correlation with OM (r= 0.15, 0.29, 0.18,
and 0.37, respectively), which indicates that the organic matter is not the important factors
controlling the potential perniciousness of trace metals in the sediment. Acid leachable Ni,
Fe, Mn and Co are negatively correlated with CaCO3, (r=-0.56, -0.48, -0.32, and -0.58), which
indicates that these metals are hardly combined with carbonates. Acid leachable Zn, Pb, Cd,
Cu, and Cr have a relatively low positive correlation with CaCO3 (r=0.33, 0.53, 0.28, 0.49,
and 0.46, respectively) indicating that these metals may be combined with CaCO3.
The behaviour of the six trace metals (Fe, Mn, Zn, Ni, Cr and Pb) in the sedimentary
five fractions could be known by performing correlation matrix (Tables 47-52). As shown in
Table 47 high correlation between total Fe and both organic bounded Fe (r=0.861) and
residual Fe (r=0.978). Moderate correlation was found between Fe bound to oxides and Total
Fe and Organic iron (r=0.611 and 0.750 respectively). It is clear from Table 48 that total Mn
correlates significantly with exchangeable Mn fraction (r=0.701), Mn bound to carbonate
fraction (r=0.769), and Residual Mn (r=0.936). Significant correlation was obsereved
between exchangeable Mn and Mn bound to carbonate (r=0.845) and exchangeable Mn and
Mn bound to organic fraction (r=0.799). Mn bound to carbonate was found to be significantly
related to Mn bound to organic fraction (r=0.830). The results of correlation coefficient
matrix between chemical characteristics and different species of Zn metal (Table 49) showed
that significant correlation observed between total Zn and Carbonate Zn (r=0.619), Zn bound
to oxides (r= 0.652), Organic Zn (r=0.831) and residual Zn (r=0.915). Zn bound to carbonate
fraction was observed to be significantly correlated with Zn bounded to Fe/Mn oxides fraction
(r=0.937) and Zn bound to organic fraction (r=0.743). There was significant correlation
between Zn bounded to Fe/Mn oxides and Zn bound to organic fraction (r=0.791). Finally Zn
bounded to organic was significantly correlated with the residual Zn (r=0.640). Table 50
showed that significant correlation found between Total Ni and Ni bound to Fe/Mn oxides
(r=0.757), Ni bounded to organic fraction (r=0.712), and residual Ni (r=0.9680. Also, Ni
bounded to Fe-Mn oxides was significantly correlated with Ni bounded to organic fraction
(r=0.978). Total Cr was significantly correlated with residual Cr (r=0.993) (Table 51).
Finally, Table 52 showed that total Pb was significantly correlated with both carbonate
fraction (r=0.891) and Fe/Mn oxides fraction (r=0.963). Pb bounded to Carbonate fraction
was significantly correlated with Fe/Mn oxides (r=0.805), while Pb bounded to Fe/Mn was
significantly correlated to Pb bounded to organic fraction (r= 0.716). Bird et al. (2005)(618)
have suggested that anthropogenically sourced metals preferentially partitioned to the non
residual phase of the sediment, and that the residual phase generally reflects background
geochemical conditions.
-237-
Table 46: Correlation matrix between leachable trace metals, CaCO3 and OM in sediments from Salloum to Rafah (n=20)
Zn
Ni
Pb
Cd
Fe
Cu
Mn
Co
Cr
CaCO3
Zn
1
Ni
-0.1195
1
Pb
0.9535**
-0.3022
1
Cd
0.1546
-0.3666
0.2539
1
Fe
-0.0521
0.9809**
-0.2240
-0.2686
1
Cu
0.7515*
-0.1187
0.7717*
0.3065
0.0116
1
Mn
0.0252
0.6682*
-0.1123
-0.1466
0.7126*
0.0007
1
Co
-0.1699
0.9716**
-0.3509
-0.4012
0.9443**
-0.1774
0.6623*
1
Cr
0.8659*
-0.0448
0.8354*
0.2105
0.0370
0.8764*
-0.0248
-0.0975
1
CaCO3
0.3337
-0.568
0.5278
0.2481
-0.487
0.4921
-0.322
-0.586
0.4594
1
OM
0.1529
-0.36
0.2881
0.1867
-0.294
0.5335
-0.26
-0.404
0.3704
0.5803
Note:
* is significant at P< 0.05
** is significant at P< 0.01
-238-
OM
1
Table 47: Correlation coefficient matrix between chemical characteristics and different
species of Iron
T Fe
F1.
F2
F3
F4
F5
%TOM CaCO3
T Fe
1.000
F1
0.300
1.000
F2
0.308
0.150
1.000
F3
0.611*
-0.152
-0.059
1.000
F4
0.861**
0.047
0.176
0.750*
1.000
F5
0.978**
0.304
0.382
0.572
0.870**
1.000
TOM
-0.455
-0.326
-0.137
0.086
-0.370
-0.475
1.000
CaCO3
-0.683
-0.370
-0.532
-0.113
-0.579
-0.725
0.580
1.000
Table 48: Correlation coefficient matrix between chemical characteristics and different
species of Manganese.
T Mn
F1
F2
F3
F4
F5
%TOM CaCO3
T Mn
1.000
F1
0.701*
1.000
F2
0.769*
0.845**
1.000
F3
0.203
0.230
0.474
1.000
F4
0.769*
0.799*
0.830**
0.556
1.000
F5
0.936**
0.621
0.585
-0.125
0.575
1.000
TOM
-0.403
-0.480
-0.362
0.012
-0.273
-0.413
1.000
CaCO3
-0.675
-0.664
-0.564
0.038
-0.433
-0.702
0.580
1.000
Table 49: Correlation coefficient matrix between chemical characteristics and different
species of Zinc.
T Zn
F1
F2
F3
F4
F5
%TOM CaCO3
T Zn
1.000
F1
0.362
1.000
F2
0.619*
0.245
1.000
F3
0.652*
0.192
0.937**
1.000
F4
0.831**
-0.030
0.743*
0.791*
1.000
F5
0.915**
0.396
0.333
0.336
0.640*
1.000
TOM
0.212
0.160
0.555
0.472
0.313
0.157
1.000
CaCO3
0.101
0.212
0.571
0.594
0.254
-0.142
0.580
Note:
* is significant at P< 0.05
** is significant at P< 0.01
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1.000
Table 50: Correlation coefficient matrix between chemical characteristics and different
species of Nickel.
T Ni
F1
F2
F3
F4
F5
%TOM CaCO3
T Ni
1.000
F1
0.165
1.000
F2
-0.181
0.060
1.000
F3
0.757*
0.412
-0.105
1.000
F4
0.712*
0.404
-0.076
0.978**
1.000
F5
0.968**
0.040
-0.204
0.577
0.523
1.000
TOM
-0.537
-0.006
0.143
-0.368
-0.374
-0.536
1.000
CaCO3
-0.808
0.113
0.273
-0.569
-0.595
-0.793
0.580
1.000
Table 51: Correlation coefficient matrix between chemical characteristics and different
species of Chromium.
T Cr
F1
F2
F3
F4
F5
%TOM CaCO3
T Cr
1.000
F1
-0.251
1.000
F2
-0.379
-0.001
1.000
F3
-0.350
-0.107
0.558
1.000
F4
0.371
-0.232
-0.331
-0.104
1.000
F5
0.993**
-0.264
-0.400
-0.383
0.408
1.000
TOM
-0.505
-0.154
0.401
0.557
-0.148
-0.529
1.000
CaCO3
-0.647
0.050
0.538
0.665
-0.574
-0.677
0.580
1.000
Table 52: Correlation coefficient matrix between chemical characteristics and different
species of Lead.
TPb
F1
F2
F3
F4
F5
%TOM CaCO3
TPb
1.000
F1
0.033
1.000
F2
0.891**
0.130
1.000
F3
0.963**
-0.011
0.805**
1.000
F4
0.655*
-0.220
0.386
0.716*
1.000
F5
-0.052
0.022
-0.326
-0.110
0.173
1.000
TOM
0.373
0.001
0.091
0.444
0.321
0.060
1.000
CaCO3
0.567
-0.259
0.426
0.651
0.371
-0.355
0.580
Note:
* is significant at P< 0.05
** is significant at P< 0.01
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1.000
Chapter VII
Summary
The coastal zone of Egypt on the Mediterranean Sea extends from Rafah in the eastern
region to El-Salloum in the western region for over 1200 km. It hosts a number of important
residential and economic centers, like the cities of Alexandria, Matruh, Rosetta, Damietta,
Port Said and Al-Arish. It also hosts five large lakes, namely Bardawil, Manzala, Burullus,
Edku and Maryut which represents about 25% in area of the total wetland of the
Mediterranean. Many activities are known in the coastal areas including fishing, industrial,
tourism, trading and agricultural, oil and gas production, and transportation.
The main objective was to study the concentration and distribution of some trace
metals in the surfacial sediments along the Egyptian Mediterranean coast and to evaluate the
contamination degree and the potential ecological risk for the selected metals.
To achieve this objective, twenty surfacial sediment samples were collected from
selected stations along the coastal area from El-Salloum to Rafah during July 2010. The
study area was divided into three regions as follows: western region, located in the western
area of the coast represented by 11 stations from El-Salloum to El-Maadia. The middle
region, located in the Delta area of the coast represented by 7 stations from Rashid west to ElGamil west and the eastern region, represented by 2 stations from Port Said to Rafah.
The samples were collected and placed into sealed polyethylene bags. Separate
samples were collected at the same time for the measurement of SEM/AVS. Samples were
collected in such a manner that exposure to atmospheric O2 is minimized, or sulfides will be
oxidized.
Several ecological assessment methods were applied (total and leachable
concentration, sequential extraction, and AVS/SEM) to assess the trace metals (Fe, Mn, Cr,
Ni, Zn, Pb, Cu, Co and Cd) pollution and potential biotoxicity in sediments from the Egyptian
Mediterraean coast. The interrelationship and corresponding relationships of different
evaluation methods were also discussed. The obtained results can be summarized as follows:
A- Geochemical analysis
Grain size analysis was performed on twenty sediment samples collected from the
study area. Results showed that fine sand and very fine sand were the dominated fractions of
all sediment samples. The range and average of sand were 86.85 to 100% (93.43%). The
highest value of sand content (100%) was observed at Nobareya, El-Mex, Western harbour,
Rashid west and Rafah. While the lowest value (86.85%) was observed at El-Salloum.
The range and average of total organic matter (TOM) in the surface sediments of the
study area were found to vary from 0.08 to 1.72 (0.56 ± 0.43%). The highest value was
observed at NIOF (western region) and the lowest value was observed at Rashid west (middle
region).
The Egyptian Mediterranean coast contains a wide variety of sediments. Coastal
sediments are mainly composed of two principal types: carbonate and quartze dominant sand.
The concentration of total carbonate in the sediment samples of the investigated area ranged
between 2.85 to 95.57%, with an average (37.78 ± 39.65%). The study area was divided into
three regions, which revealed differences in the distribution of CaCO3. The western region,
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where sediments at El-Salloum are characterized by relatively medium CaCO3 percentage
(34.97%) while sediments from Baghoush to the Eastern harbour (west of Alexandria) are
characterized by high percentage of CaCO3 ranged from (67.24%) at El-Mex Bay to a
maximum percentage (95.57%) at the western harbour. On the other hand the region east of
Alexandria characterized by relatively low % of CaCO3 (14.69, 15.69 and 22.11%) at
electrical power station, Maadia and Abu Qir, respectively. There was a relatively low
percentage of CaCO3 in the Middle region ranged from 2.85% at Rashid west to 6.08% at
New Damietta and the Eastern region 6.33% at Rafah and 7.32% at Port Said.
B- Trace metal Distribution
I - Total trace metals distribution
The distribution of the total trace metals (Fe, Mn, Cr, Ni, Zn, Pb, Cu, Co and Cd) of
sediments along the Egyptian Mediterranean coastal area was studied. The range and average
± SD concentrations (µgg-1) were 243.48–38045.05 (13255.69± 12911.36) for Fe, 17.251085.72 (380.71 ±305.38) for Mn, 4.08-297.95 (82.74 ±90.18) for Cr, 1.65-60.25 (25.93
±20.96) for Ni, 2.05-62.21 (22.19 ±15.84) for Zn, 3.34-53.67 (13.17 ±11.90) for Pb, 0.4626.26 (8.46 ±6.22) for Cu, 0.43-26.39 (8.24 ±8.40) for Co, 0.04-0.47 (0.22 ±0.15) for Cd.
The level of trace metals can be arranged as the following:
Fe> Mn> Cr> Ni> Zn> Pb> Cu> Co> Cd.
The lowest value of Fe, Zn, Cu, Ni, Cr and Co (243.48, 2.05, 0.46, 1.65, 82.74 and
0.43 µgg-1, respectively) observed at Baghoush, while Mn, Pb and Cd showed a minimum
concentration value (17.25, 3.34 and 0.04µgg-1) at El-Nobarreya, Rafah and Port Said,
respectively. On the other hand, the highest content of Fe, Mn, Ni, Cr, Co, Pb and Cd; 38045,
1085.72, 60.25, 297.95, 26.39, 53.67 and 0.47 µgg-1, respectively observed at New Damietta,
Port Said, El-Gamil west, Rashid east, Ras El-Barr, El-Mex and Western harbor, respectively.
The concentration of Zn (62.21µgg-1) and Cu (26.26µgg-1) was found to be high at NIOF.
The distribution of Fe, Mn, Ni and Co, among the three regions of the study area was similar
and arranged as the following: Middle region > Eastern region > Western region. While
the distribution of Zn, Pb and Cd are also similar and arranged as the following: Western
region > Middle region> Eastern region. Cupper was distributed among the three regions
of the study area as the following: Middle region > Western region > Eastern reigon. The
distribution of Chromium was different than the other metals and its distribution among the
different regions of the study area as the following: Eastern region> Middle region>
Western region
Assessment of pollution impact
1- Sediment quality guidelines
-
The magnitude and ecological relevance of metal pollution in the surfacial sediments
along the Egyptian Mediterranean coastal area was investigated by applying different
sediment quality approaches.
-
Based on the SQGs proposed by USEPA, sediments were categorized into three
classes: unpolluted, moderately polluted and heavily polluted.
-
Accordingly, the concentration of Cd and Zn at all stations under investigation was
belonged to unpolluted sediments.
-
The concentration of Cr at El-Salloum, NIOF, Electric power station, Maadia and
Rafah was moderately polluted, while the stations at Abu Qir, Rashid west, Rashid
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east, Burullus, New Damietta, Ras El-Barr, El-Gamil west and Port Said were heavily
polluted.
-
Sediment at all stations was classified as unpolluted with cupper except NIOF station
which was moderately polluted.
-
The concentration of Fe at El-Gamil stations was moderately polluted, while Rashid
east, Burullus, New Damietta, Ras El-Barr and Port Said were highly polluted. Pb at
all stations was belonged to unpolluted except El-Mex Bay which was moderately
polluted.
-
It was noted that the concentration of Ni at Rashid west, Burullus, New Damietta, ElGamil east, Port Said and Rafah was belonged to moderately polluted sediments,
while Ni at Rashid east, Ras El-Barr and El-Gamil west was belonged to heavily
polluted.
-
The concentration of Mn at the Eastern harbour, Electric power station, El-Gamil east
and El-Gamil west was moderately polluted, while Rashid east, Burullus, New
Damietta, Ras El-Barr and Port Said was heavily polluted.
2- Enrichment factor
-
The computed metal EF range in the surface sediments of the Egyptian Mediterranean
coastal area showed that the EF of trace metals (Cd, Co, Cr, Cu, Mn, Ni, Pb, Zn) were:
(0.22 to 74.03), (0.61 to 10.50), (1.18 to 28.99), (0.32 to 6.78), (0.86 to 8.04), (0.62 to
26.99), (0.40 to 114.22), and (0.29 to 14.77), respectively.
-
The calculated EF was found
Cd>Pb>Cr>Ni>Mn>Zn>Co>Cu.
-
Despite the high EF for almost all the studied metals in the sediments of the western
region at Baghoush, Nobarreya, El-Dikhaila, El-Mex, Western harbour, NIOF and the
Eastern harbour and Rafah Station at the eastern region. There seems to be not even
moderate pollution of metals in these sediments. This leads to incorrect results due to
the naturally lower concentration of Fe at these stations (243.48, 277.34, 3946.02,
3565.22, 1527.85, 4813.8, 3440.82 and 1422.39 respectively) which may lead to
higher EF of trace metals. For this reason we used Igeo index too.
to
fall
in
the
following
sequence:
3- Geoaccumulation load Index
Results showed that the Igeo values for Zn, Mn, Ni, Fe, Cu and Co fall in class "zero"
(unpolluted) at all the sampling locations indicating that there is no pollution from these
metals. The Igeo value for Pb fall in class " zero" (unpolluted) at almost all the sampling
location except El-Mex Bay and NIOF which fall in class "1" (unpolluted to moderately
polluted). The Igeo value for Cd fall in class "zero" (unpolluted) at almost all locations except
El-Salloum, El-Dikhaila, Western harbour, NIOF, Eastern harbour, Electrical power stations
and Maadia falls in class "1" (unpolluted to moderately polluted). The Igeo value for Cr fall in
class "zero" (unpolluted) at all stations of the western region except stations 13 and 17 in
(Rashid east, El-Gamil east) at the middle region and station 19 in Port Said at the eastern
region were classified as "moderately polluted".
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4- Contamination factor and modified degree of contamination:
-
The calculated CF was found to fall in the following sequence: Cd (1.27) >Cr (1.15)
>Pb (0.70)>Co (0.59) >Ni (0.50)>Mn (0.49)>Fe (0.32) >Cu (0.26)>Zn (0.23).
-
The contamination factors values are low for Zn, Cu and Fe at all sampling sites and
between low and moderate for Ni, Cd, Mn and Co and finally, between low and
considerable contamination for Cr.
-
Modified degree of contamination (mDc) was found to be less than 1.5 indicating
sediments along the Egyptian Mediterranean coast classified from zero to very low
contamination.
5- Pollution load index
- The Results of PLI were less than 1 at all the studied locations indicating only baseline
levels of pollutants present.
-
PLI values of the analyzed samples ranged from 0.04 at Baghoush and 0.72 at Port
Said with an average 0.61. In general, there is an increase of PLI toward the eastern
region.
II- Leachable trace metals
The range and average concentrations (µgg-1) were: 94-6831 (1930±1777) for Fe,
10.83-307.27 (126±81) for Mn, 0.55-8.21 (2.61±2.01) for Cr, 0.15-11.16 (3.04±3.23) for Ni,
0.11-49.48 (7.40±10.94) for Zn, 0.54-51.35 (8.13±11.97) for Pb, 0.22-8.20 (2.19±2.29) for
Cu, 0.09-8.77 (2.34±2.82) for Co, 0.02-0.23 (0.06±0.05) for Cd.
Leachable Iron: Results showed that the % of LFe ranged between (6.68 to 69.46%) with an
average (22.83%). The maximum % of LFe (69.46%) was found at Baghoush in the western
region which has high % of calcium carbonate (94.45%) and the minimum percentage
(6.68%) of LFe was found at Rafah in the eastern region. The low percentage of LFe in most
cases (average: 22.832%) may be due to the increased contribution of the lattice-held fraction
of the metal. This reflects low bioavailability of Fe.
Leachable Manganese: Data showed that the range and average of LMn % was (3.66-91.67;
48.03). The highest % of LMn (91.67%) was observed at El-Mex Bay (western region) while
the lowest % of LMn (3.66 %) was observed at Rashid east (Middle region). The high
percentage of weakly bound Mn (>50%) observed at almost all stations of the western region
are probably due to the elevated association of Mn with carbonate that are dissolved by the
dilute HCl. The relatively low % of LMn in the Middle (27.82%) and the eastern region
(35.47%) indicated that manganese is tightly bound to the sediment minerals.
Leachable Zinc: The % of LZn ranged between 1.159 and 89.27 (average: 30.15%). The
lowest value observed at Nobarreya while the highest value observed at El-Mex Bay. There
was a relatively high percentage of Zn (43.35%) associated with the non residual fraction
(leachable) at the western region. While the Middle and the Eastern region showed relatively
low % of leachable Zn (15.46 and 8.95%, respectively) reflecting the natural background of
this element in the Middle and the Eastern region.
Leachable Copper: Cupper exhibited LCu ranged between 8.92 and 86.46% (average:
29.17%). The lowest % of LCu was observed at Abu Qir Bay while the highest value found
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at El-Mex Bay. The antifouling paints used for ships and boats are regarded as one of the
important sources, which increased the level of copper in El-Mex Bay. Relatively high % of
LCu (> 30%) was observed at Baghoush, Nobarreya, Dekhaila, El-Mex, Western harbor,
NIOF, Eastern harbor and Electric power station in the western region. While, relatively low
percentage of the non residual Cu observed at stations of the Middle and the Eastern region
13.48% and 16.59%, respectively.
Leachable Lead: The LPb % ranged from 12.01 to 95.68% (average: 45.23%). The high %
found at El-Mex Bay, while the lowest value found at El-Gamil east. The high extractability
percentage of Pb reported at El-Mex Bay indicated a new input in El-Mex Bay, which was
probably due to the domestic, industrial effluents and the atmospheric deposition. Relatively
low percentage (<30%) of Pb (19.87% and 27.78) was extracted at the non residual fraction
(leachable) in the Eastern and the Middle region.
Leachable Cadmium: The results showed that LCd % ranged between 11.63 and 86%
(average: 37.25%). The lowest value observed at Maadia while the highest % observed at
Ras El-Barr. Cadmium showed relatively high % of LCd in the Middle region (average:
45.39%). Cd is contained in some phosphate based fertilizers and petroleum. Such sources
could constitute a major source of Cd that may reach humans. In additions, sewage sludge
from wastewater treatment may contain significant quantities of Cd. Relatively high
percentage of Cd was observed in the residual form in the Western and the Eastern region
(>60%), reflecting the natural background of this element in these regions.
Leachable Chromium: The LCr % ranged between 0.42 and 61.35% (average: 13.61%).
The highest value observed at Baghoush while the lowest value observed at El-Gamil west.
The western region showed relatively high percentage of LCr (61.35, 35.42 and 42.27%) at
Baghoush, El-Mex Bay and the Eastern harbour, respectively. On the other hand, it exhibited
a relatively low % of LCr ranged from 0.42 to 3.56 % (average: 1.51%) at the middle region
and from 0.96 to 1.04 (average: 1.00%) at the eastern region, reflecting the low Cr mobility in
the study area.
Leachable Cobalt: The results showed that the value of LCo ranged between 5.34 and
45.935 (average: 25.63%). The highest value found at the Electric power station and the
lowest value observed at Port Said. Relatively low % of LCo was found at most of the
stations not exceeding 30%.
Leachable Nickel: The % of LNi ranged between 0.55 and 28.83% (average: 13.55%) at the
western region, and from 4.49 to 25.20% (average: 12.07%) at the middle region and from
0.55 to 14.17% (average: 7.36%) at the eastern region. The highest value found at Maadia
while the lowest value found at Rafah. The relatively low % of LNi (<20%) at most of
locations reflects the natural background of Ni in the investigated area.
III- Fractionation of trace metals
In the present work, fractionation of Fe, Mn, Cr, Ni, Zn and Pb into the different
chemical forms (Exchangeable, Carbonate, Fe-Mn oxides, Organic and Residual fraction) has
been conducted in order to demonstrate in which forms these metals are chemically chelated,
as well as to differentiate between the residual metals (natural background) and non residual
ones (man made sources of pollution).
Exchangeable fraction
The exchangeable metal in sediments is labile, highly toxic and the most bioavailable.
The results showed that the exchangeable bound fraction for trace metals (Fe, Mn, Zn, Ni, Cr
-245-
and Pb) in surfacial sediments along the Egyptian Mediterranean coast was comparatively
low, (ND-0.04%), (0.01-0.32%), (ND -3.84%), (0.09-9.35%), (0.01-2.79%) and (0.246.64%), respectively suggesting lower pollution risk and the relatively low mobility of these
elemests.
Carbonate fraction
The range and average percentage of carbonate fraction were: 0.82-54.51 (13.90%) for
Pb; 0.87-18.82 (8.83%) for Mn; 0.2-14.89 (3.79%) for Zn and 0.03-12.17 (1.75%) for Cr and
only small fraction 0.54-5.30 (0.07%) and 0.02-2.64 (0.81%) for Fe and Ni, respectively.
Lead showed appreciable concentration (14.30, 14.93, 45.63, 54.40, 54.51 and 22.60%) in the
carbonate fraction at (El-Salloum, Bagoush, Nobarreya, El-Mex Bay, Western harbour and
NIOF, respectively) in the western region which characterized by high CaCO3 contents
(34.79, 94.45, 95.49, 67.24, 95.57 and 91.64%). This reveals that calcite has a strong affinity
for Pb. This also may be attributed to the similarity of its ionic radii to that of calcium in the
carbonate crystal lattic to form a mixed carbonate phase. On the other hand, manganese was
found in relatively moderate amounts (>10%) (18.82, 11.91, 11.35, 13.53, 15.11, 13.04 and
17.81%) at Maadia, new Damietta, Ras El-Barr, El-Gamil west, El-Gamil east, Port Sail and
Rafah, respectively. This may be attributed to anthropogenic inputs.
Fe/Mn oxide fraction
The range and average percentage of Fe/Mn oxide fraction were: 2.29-87.31 (43.08%)
for Mn; 4.92-49.83 (12.32%) for Pb; 1.22-67.26 (18.24%) 1.75-25.38 (10.32%) for Ni; 1.8240.45 (12.24%) for Fe; 0.22-37.34 (8.14%) for Cr. More than 65% of Mn was found to be
associated with Fe-Mn oxides in the western region suggesting, that hydrous Fe-Mn oxides
may play a major role in controlling the fate and transport of Mn in this region. Moderate
amounts of Pb were found binding to Fe-Mn oxide (average: 33.80, 21.47 and 15.18%) in the
the study area which has been proved to be sensitive to anthropogenic inputs. Finally,
moderately amounts (average: 27.75%) of Zn bounded to the Fe-Mn fraction. was found in
the western region.
Organic fraction
The range and average percentage
organic fraction were: 2.34-13.85 (7.42%)
(5.79%) for Zn; 0.22-22.97% (6.91%) for
(4.52%) for Pb were bounded to organic
benthic biota.
of percentage of trace metals bounded to the
for Fe; 0.91-6.72 (3.75%) for Mn; 0.03-13.01
Ni; 0.48-17.98 (3.61%) for Cr and 2.68-17.25
matter and sulfide, contributed low toxicity to
Residual fraction
Residual phase of metals are generally much less toxic for organisms in aquatic
environment. The results of residual fraction in the surfacial sediments along the Egyptian
Mediterranena coast was dominanted. The range and average percentage of the residual
fraction were: 39.14-99.20 (86.17%) for Cr; 51-35-95.52 (80.27%) for Fe 50.92-97.90
(80.05%) for Ni and 9.32-97.14 (71.70%) for Zn. This reflects the natural background of Ni,
Fe, Cr and Zn in the investigated area. The highest concentration of residual fraction was
reported for Mn (71.64 and 64.55%, respectively) and for Pb(67.47 and 57.17%) in the
Middle and the Eastern region. This indicates that Mn and Pb are tightly bound to the
sediment minerals in the Middle and the eastern region and are not easily released to the
environment.
Risk assessment code
- Risk assessment code has been used to assess environmental risks and estimate
possible damage to benthic organisms caused by contaminated sediments.
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-
-
From the present study, The range and average values of the exchangeable and
carbonate-bound fractions of Cr, Fe, Mn, Ni, Pb, and Zn were: 0.04-12.58 (average:
2.08%)for Cr; 0.005-0.342 (average: 0.068%) for Fe; 0.087-18.96 (average: 8.95%)
for Mn; 0.14-10.88 (average: 2.72%) for Ni; 4.32-55.43 (average: 16.82%) for Pb and
0.26-15.27 (average: 4.26%) for Zn.
This suggests that Fe has no risk to local environment; Cr, Mn, Ni, and Zn have posed
low risk to local environment while Pb was at Medium risk level.
V- Acid volatile sulfides and simultaneously extractable metals
-
Acid volatile sulfide (AVS) is one of the major chemical components that control the
activities and availability of metals in the interstitial waters of sediments. Sulfide
reacts with several divalent transition metal cations (Cd, Cu, Ni, Pb and Zn) to form
highly insoluble compounds that are not bioavailable.
-
The average concentration of SEM in the three regions can be arranged as follows:
western region (0.13±0.07µmoleg-1) > Middle region (0.08±0.04µmoleg-1)> eastern
region (0.05±0.05µmoleg-1).
-
The average value of AVS in the different regions can be arranged as follows: western
region (3.31±9.35µmoleg-1)> Eastern region (0.07±0.06µmoleg-1)> Middle region
(0.06±0.03µmoleg-1).
-
The use of different relationships between AVS and SEM to establish mechanical
model such as the ratio of SEM and AVS (SEM/AVS), the difference between
SEM and AVS (SEM-AVS) or the organic carbon normalized difference between
SEM and AVS (SEM-AVS/fOC) to assess metal toxicity has been applied in this
study.
-
According to SEM/AVS model: Western Harbour, Eastern Harbour, Ras El-Barr, ElGamil East, and Port Said are occasionally toxic while the other stations are not toxic.
-
According to SEM-AVS model: stations at El-Salloum, Nobarreya, El-Dikhaila, ElMex, NIOF, Abu Qir, El-Gamil east and Rafah showed no indication of associated
adverse effects, while at the western harbour, eastern harbour, electrical power station,
Maadia, Rashid, east and Rashid west, Burullus, New Damietta, Ras El-Barr, ElGamil east and Port Said associated adverse effects on aquatic life are possible.
-
Organic carbon normalized difference (SEM-AVS/fOC) showed that sediments
should pose a low risk of adverse biological effects due to Cd, Cu, Pb, Ni and Zn at all
sampling stations
C- The Screening level Ecological Risk Assessment:
The output of the SLERA (based on the consensus approach) revealed that:
- TEC HQ based on the total metal concentrations for Cd, Cu and Zn was less than 1
indicating rare effects. While, TEC HQ for Cr, Pb and Ni were higher than 1 based on the
total metals concentrations; indicating frequent adverse ecological effects are expected to
occur.
- TEC HQ for Pb based on the labile fraction was in the range (TEC HQ > 1 > PEC HQ)
indicating that adverse ecological effects are possible but less frequent with respect to Pb.
While the TEC HQ for Ni and Cr based on the labile fraction concentration was TEC HQ < 1
indicating rare effects (Stop).
-247-
- Based on the calculated values of PEC HQ, it can be recognized that the calculated PEC
value for Pb was (TEC HQ>1>PEC HQ). Thus it can be concluded that Pb considered as
contaminant of less potential concern.
D- The potential ecological risk
The potential ecological risk indices (Eif) of Cd, Cr, Cu, Pb, and Zn were lower than 30,
which indicated slight potential ecological risk of all five metals in the studied stations.
According to the evaluating standards, all the stations of the study area had low ecological
risk levels (RI <110).
E- Multivariate Statistical analyses
-
To assess the dynamics of trace metals and identify the contributing sources of
bioavailable trace metals in the surfacial sediments along the Egyptian Mediterranean
coastal area, multivariate statistical analyses such as principal component, cluster
analysis and correlation matrix coefficient were carried out.
-
Data showed that strong positive correlation exists between Fe and (Mn, Co, Cr, and
Ni) (r = 0.92, 0.96, 0.69, and 0.85, respectively). It means that these metals tend to
accumulate together. The significantly positive correlation with Fe indicates that the
elements were derived from similar sources and also moving together.
-
There is a negative correlation between Fe and Pb (r = -0.3918) suggested that Pb was
not strongly controlled by natural weathering processing.
-
There was a strong positive correlation exists between Zn and (Cu and Pb) (r = 0.80
and 0.63 respectively). These metals were believed to have the same sources which
differ from the sources of Fe, Co, Ni, Cr, and Mn.
-
The obtained results are in agreement with the results of the cluster analysis and the
principle component analysis.
-248-
Conclusion
-
Information of this study reported for the first time for the surfacial sediments of the
Egyptian Mediterranean Sea, constitute a baseline of metals speciation, AVS and SEM in
sediments and a reference for future studies on the changes of labile and residual metal
fractions over time.
-
Sediments are important hosts for trace metals and such should be included in environmental
monitoring programs.
-
The distribution of Fe, Mn, Ni and Co, among the three regions of the study area was similar
and arranged as the following: Middle region> Eastern region> Western region. While, the
distribution of Zn, Pb, Cd are also similar and arranged as the following: Western region>
Middle region> Eastern region. Copper was distributed among the three regions of the study
area as the following: Middle region> Western region> Eastern reigon. The distribution of
chromium was different from the other elements and its distribution among the different
regions of the study area as the follwoing: Eastern region> Middle region> Western region.
-
Different metal assessment indexes were used and their applicability to interpreat the
pollution status was discussed. The potential ecological risk indices of Cd, Cr, Cu, Pb and
Zn, indicated slight potential ecological risk of these metals. The calculated modified
degree of contamination was less than 1.5 indicating zero to very low contamination.
-
The application of EF as a tool for estimating the anthropogenic impact on sediments is not
an appropriate tool where a relatively low natural Fe concentration exists as it will gives
incorrect higher values of EF.
-
Application of the geochemical accumulation index may lead to more realistic results in
comparison with EF in sediments of the study area as low Fe content of sediments in almost
all stations of the western region and station 20 (Rafah) in the eastern region.
-
Relationships between metal speciation and total metals concentrations reflected the stability
of the metals, it can be arranged in the order of: Cr> Fe> Ni> Zn> Pb> Mn.
-
Results of single acid extraction method (Leachable) was in agreement with the results of
sequential extraction procedure for (Fe, Cr, Mn, Zn, Pb and Ni):
-
The single extraction procedure using 1 N HCl (leachable metals) was an easy, simple,
cheap and quick procedure that provides an accurate data base of the availability of metals in
any aquatic environment which are often readily available to organisms affecting them
directly. On the other hand, it does not differentiate between the non residual fractions
(exchangeable, carbonate, Fe/Mn oxides, Organic) and residual fraction, as it can be
obtained from the results of sequential extraction procedure.
-
Sequential extraction procedure should be used to better differentiate the form of Fe, Mn,
Cr, Zn, Ni and Pb. These fractions play an important role in controlling the fate and
transport of metals in the sediment.
-
Determination of the total concentration of a metal in sediment is not a particularly useful
indicator of sediment toxicity, as it does not distinguish between the natural and
anthropogenic components of the metal. This is supported by the finding obtained by
comparing the results with SQG proposed by USEPA and the obtained results of sequential
extraction procedure.
-
According to SQG proposed by USEPA the concentration of Cr, Fe, Ni was moderately
polluted to highly polluted. While according to RAC, these metals posed no risk to local
environmet.
-249-
-
The acid volatile sulfide (AVS) levels in the Egyptian Mediterranean Sea sediments revealed
wide variation. It varied from 0.015 to 31.33 µmoleg-1, which is in agreement with the
values reported for samples from the US and Canada.
-
The values of SEM/AVS at the western Harbour, Eastern Harbour, Ras El-Burr, El-Gamil
East, and Port Said are occasionally toxic while the other stations are not toxic.
-
According to SEM-AVS model: stations at El-Salloum, Nobarreya, El-Dikhaila, El-Mex,
NIOF, Abu Qir, El-Gamil east and Rafah showed no indication of associated adverse effects,
while associated adverse effects on aquatic life are possible at the western harbour, eastern
harbour, electrical power station, Maadia, Rashid, east and Rashid west, Burullus, New
Damietta, Ras El-Barr, El-Gamil east and Port Said.
-
Results showed that [SEM-AVS]/fOC < 130 µmoleg-1 at all the sampling sites revealing that
sediments posed a low risk of adverse biological effects due to Cd, Cu, Pb, Ni and Zn at all
stations.
-
The output of the SLERA (based on the consensus approach) revealed that rare negative
effects are possible due to Cd, Cu, Cr, Ni and Zn. And lead was considered as a contamint of
less potential concern.
Recommendations
-
Information of this study constitutes a baseline of metals speciation, AVS and SEM in
sediments and should be used as a reference for future studies on the changes of labile and
residual metal fractions over time.
-
The application of EF as a tool for estimating the anthropogenic impact on sediments is not
an appropriate tool where a relatively low natural Fe concentration exists as it will gives
incorrect higher values of EF.
-
Application of the geochemical accumulation index (Igeo) may lead to more realistic results
in comparison with EF in sediments of lower natural Fe content.
-
AVS/SEM models and sequential extraction procedure accompanied with measuring the
total metal concentration should be included when evaluating the environmental risk of trace
metals in sediments.
-
The capabilities of national institutions to carry out marine pollution monitoring and
research should be supported, developed and when necessary established to formulate and
apply pollution control and abatement measures.
-
Commitment with the international agreement and the environmental laws.
-
Establishment of a Management Information System that stores all previous information and
data which will help in future development, management and restoration of the Egyptian
Mediterranean coastal zone
-
Monitoring systems for trace metals in fish, water and sediments in the study area.
-250-
References
1.
2.
3.
4.
5.
6.
7.
8.
9.
10.
11.
12.
13.
14.
15.
16.
17.
Burges, M.R., and Scott, K.J. (1992). The significance of in-place contaminated marine
sediments on the water column: process and effects. In: Sediment toxicity assessment,
Burton, G.A. (Ed.). Lewis Publishers, London.
Angelidis, M.O. (2005). Exchange of Pollutants (Trace Elements) at the Sediment
Boundary. Environmental Chemistry, 5: 319-341.
Fergusson, J.E. (1990). The Heavy Elements. In: Chemistry, Environmental Impact and
Health Effects. Pergamon Press, Oxford, England.
Robson, A.J. and Neal. C. (1997). A summary of regional water quality for Eastern UK
Rivers, Science of Total Environment, 194-195: 15-37.
Gonzalez-Macias, C.G., Schifter, I., Lluch-Cota, D.B., Rodriguez, L. and Vanzquez.
S.H. (2006). Distribution, Enrichment and Accumulation of Heavy Metals In Coastal
Sediments Of Salina Cruz Bay, Me´ Xico. Environmental Monitoring Assessment, 118:
211-230.
Davies, A.G. (1978). Pollution studies with marine plankton; Part II. Heavy metals.
Advances in Marine Biology, 15: 381–508.
Atgin, R.S., El-Agha, O., Zarasiz, A., Kocatas, A., Parlak, H. and Tuncel, G. (2000).
Investigation of the sediment pollution in Izmir Bay: Trace elements. Spectrochemica
Acta, Part B, 55: 1151-1164.
Mucha, A.P., Vasconcelos, M.T.S.D. and Bordalo, A.A. (2003). Macrobenthic
community in the Douro Estuary: Relations with trace metals and natural sediment
characteristics. Environmental Pollution, 121: 169-180.
Tam, N. F. Y. and Wong, Y. S. (1995). Retention and distribution of heavy metal in
mangrove soils receiving wastewater. Environmental Pollution, 94(3): 283–291.
Moore, J.N., Brook, E.J. and Johns, C. (1989). Grain size partitioning of metals in
contaminated coarse-grained floodplain sediment, Clark Fork River, Montana, U.S.A.
Environmental Geology Water Research, 14 (20): 107-115.
Everaat, J. M. and Fischer, C. V. (1992). The distribution of heavy metals (Cu, Zn, Cd
and Pb) in the fine fractions of surface sediments of the North Sea. Netherlands Journal
of Sea Research, 29 (4): 323–331.
Al-Abdali, F., Massoud, M. S. and Al-Ghadban, A. N. (1996). Bottom sediments of the
Arabian Gulf-III. Trace metal contents as indicators of pollution and implications for the
effect and fate of the Kuwait Oil Slick. Environmental Pollution, 93: 285–301.
Wei, C., Guang-ren, Q., Thye, L.T. and Peng-cheong, C. (2007). Speciation of heavy
metals in surface sediments from Suzhou Creek. Journal of Shanghai University (English
Edition), 11 (4): 415-425.
Davidson, C.M., Thomas, R.P., McVey, S.E., Perala, R., Littlejohn, D. and Ure, A.M.
(2004). Evaluation of a sequential extraction procedure for the speciation of heavy
metals in sediments. Analytica Chimica Acta, 291: 277-286.
Tessier, A., Campbell, P.G.C. and Biason, M. (1979). Sequential extraction procedure
for speciation of particulate trace metals. Journal of Analytical Chemistry, 51 (7): 844851.
Ure, A.M., Quevauviller, P., Muntau, H. and Griepink, B. (1993). Speciation of heavy
metals in solids and harmonization of extraction techniques undertaken under the
auspices of the BCR of the Commission of the European Communities. International
Journal of Environmental Analytical Chemistry, 51(1-4): 135–151.
Rauret, G., L´opez-S´anchez, J.F., Sauquillo, A., Rubio, R., Davidson, C., Ure, A. and
Quevauviller, P. (1999). Improvement of the BCR three step sequential extraction
-251-
18.
19.
20.
21.
22.
23.
24.
25.
26.
27.
28.
29.
30.
31.
32.
33.
34.
procedure prior to the certification of new sediment and soil reference materials. Journal
of Environmental Monitoring, 1 (1): 57–61.
Rauret, G. (1998). Extraction procedures for the determination of heavy metals in
contaminated soil and sediment. Talanta, 46: 449–455.
Quevauviller P. (1998). Operationally defined extraction procedures for soil and
sediments analysis I. Standardization. Trends in Analytical Chemistry, 17 (5): 289–298.
Filgueiras, A.V., Lavilla., I. and Bendicho, C. (2002). Chemical sequential extraction for
metal partitioning in environmental solid samples. Journal of Environmental Monitoring,
4 (6): 823-857.
Davidson, C.M., Hursthouse, A.S., Tognarelli, D.M. and Ure, A.M. (2004). Should acid
ammonium oxalate replace hydroxylammonium chloride in step 2 of the revised BCR
sequential extraction protocol for soil and sediment? Analytica Chimica Acta, 508 (2):
193–199.
Di-Toro, D.M., Mahony, J.D., Hansen, J.D., Scott, K., Hicks, M.B., Mayr, S.M. and
Redmond, M.S. (1990). Toxicity of cadmium in sediments: The role of acid volatile
sulfide. Environmental Toxicology and Chemistry, 9 (12): 1487–1502.
Ankley, G.T., Di-Toro, D.M., Hansen, D.J and Berry, W.J. (1996). Technical basis and
proposal for deriving sediment quality criteria for metals. Environmental Toxicology and
Chemistry, 15 (12): 2056–2066.
O'Day, P.A., Carroll, S.A., Randall, S., Martinelli, R.E., Anderson, S.L., Jelinski, J. and
Knezovich, J.P. (2000). Metal speciation and bioavailability in contaminated estuary
sediments, Alameda Naval Air Station, California. Environmental Science and
Technology, 34 (17): 3665–3673.
Peltier, E., Dahl, A.L. and Gaillard, J.F. (2005). Metal Speciation in anoxic sediments:
when sulfides can be construed as oxides. Environmental Science and Technology, 39
(1): 311-316.
Fang, T., Li, X.D. and Zhang, G. (2005). Acid volatile sulfide and simultaneously
extracted metals in the sediment cores of the Pearl River Estuary, South China.
Ecotoxicology and Environmental Safety, 61 (3): 420–431.
Campbell, P.G.C., Chapman, P.M. and Hale, B.A. (2006). Risk assessment of metals in
the environment. Issues in Environmental Science and Technology, 22: 102-131.
Adu, O.K. (2010). An assessment of heavy metal contamination of sediments and tissues
of the clam Galatea Paradoxa (Born 1778) in the voltal estuary, Chana, M.Sc. In
fisheries science and watershed management. Kuame Nkrumah University of science
and technology, Faculty of Renewable Natural Resources, Kumasi.
Duffus, J.H. (2002). “Heavy Metals” A meaningless term. Pure and Applied Chemistry,
74: 793- 807.
GESAMP. (2001). IMO/FAO/UNESCO/WHO/IAEA/UNDP joint group of experts on
the scientific aspects of marine pollution (GESAMP): Reports and studies no. 71,
Protecting the oceans from land-based activities, land base sources and activities
affecting the quality and uses of the marine, coastal and associated fresh water
environment.
Riley, J.P. and Chester, R. (1971). Introduction to Marine Chemistry, Academic Press,
London, pp. 88-92.
Vernet, J.P. (1991). Heavy Metals in the Environment. Elsevier Science Publishers B.V,
New York, USA.
Dahrazma, D. (2005). Removal of heavy metals from sediment using Rhamnolipid,
Ph.D. Thesis, Concoria University, Montreal, Quebec, Canada.
Sedgwick, J.S. (2005). Contamination and remediation of Soil at a Former Orchard Site,
Hamilton, North Island, Thesis, The University of Waikato, New Zealand.
-252-
35.
36.
37.
38.
39.
40.
41.
42.
43.
44.
45.
46.
47.
48.
49.
50.
51.
52.
53.
54.
Furness, R.W. and Rainbow, P.S. (2000). Heavy Metals in the Marine Environment,
CRC Press, Inc, Florida, United States of America, pp. 1-3.
Calmano, W., Ahlf, W., Baabe, H. and Fostner, U. (1988). In: Heavy metals in the
hygrological cycle. Astuc, M and Lester, J. N. (Eds.), Selper Ltd., London.
Salbu, B. and Steinnes, E. (1995). Trace elements in natural waters. CRC. Press,
London.
USEPA (Unites State Environmental Protection Agency) (1999). Integrated Risk
Information System (IRIS). National Center for Environmental Assessment, Office of
Research and Development, Washington, DC.
Drever, J.I. (1997). The geochemistry of natural waters (3rd ed.), Prentice-Hall, Sydney.
Manahan, S.E. (2000). Environmental Chemistry. Seventh Edition, Lewis Publishers,
CRC Press LLC, pp. 898.
Wood, J.M. and Wang, H.K. (1983). Microbial resistance to heavy metals.
Environmental Science & Technology, 17 (2): 582A-590A.
Burton, J.D. and Statham, P.J. (2000). Trace Metals in Seawater. In: Heavy Metals in the
Marine Environment. Furness, R.W. and Rainbow, P.S. (Eds.), CRC Press, Inc, Florida,
United States of America, pp. 5-25.
Plant, J.A., Kinniburgh, D.G., Smedley, P.L., Fordyce, F.M. and Klinck, B.A. (2003).
Arsenic and Selenium. In: Treatise on Geochemistry. Holland, H. D. and Turekian, K. K.
(Eds.), Volume 9: Environmental Geochemistry, Pergamon, Oxford, pp.17-66.
Güven, D.E. and Akinci, G. (2008). Heavy metals partitioning in the sediments of Izmir
Inner Bay. Journal of Environmental Sciences, 20 (4): 413–418.
Munsgaard, N.C. and Parry, D.L. (2002). Metal arsenic and Land isotopes in near
pristine estuarine and marine coastal sediments from north Australia. Marine Fresh
Water Research, 53: 719-729.
Owens, R.E., Balls, P.W. and Price, N.B. (1997). Physicochemical processes and their
effects on the composition of suspended particulate material in estuaries: Implication for
monitoring and modeling. Marine Pollution Bulletin, 34: 51-60.
Onyari, J.M., Muohi, A.W., Omondi, J.G and Mavuti, K.M. (2003). Heavy metal
distribution in surface sediments from Mtwapa and Shirazi Creek, Kenya Coast. Bulletin
of Environmental Contamination and Toxicology, 70 (6): 1220-1227.
Denton, G.R.W., Wood, H.R., Concepcion, L.P., Siegrist, H.G., Eflin, V.S., Narcis, D.K.
and Pangelinan, G.T. (1997). Analysis of In-Place Contaminants in Marine Sediments
from Four Harbor Locations on Guam: A Pilot Study, Water and Environmental
Research Institute of the Western Pacific, Technical Report No. 87, University of Guam,
Mangilao, Guam.
Kabata-Pendia, A. and Pendia, H. (2000). Trace elements in soil and plants. Boca Raton,
LA: CRC Press.
Morel, F.M.M. and Hering, J.G. (1993). Principles and applications of aquatic chemistry,
New York: John Wiley & Sons.
Adams, W.J., Conard, B., Ethier, G., Brix, K.V., Paquin, P.R. and Di Toro. D.M, (2000).
The challenges of hazard identification and classification of insoluble metals and metal
substances for the aquatic environment. Human and Ecological Risk Assessment, 6 (6):
1019-1038.
Ip-Carman, C.M., Li, X.D., Zhang, G., Wai, O.W.H. and Li, Y.S. (2007). Trace metal
distribution in sediments of the Pearl River estuary and the surrounding coastal area,
south china. Environmental Pollution, 147(2): 311-323.
Morillo, J., Usero, J. and Gracia, I. (2002). Partitioning of metals in sediments from the
Odiel River (Spain). Environment International, 28: 263-271.
Rashed, M.N. (2004). Biomarkers as indicators for water pollution in rivers, seas and
oceans. Environment International, 27: 27-33.
-253-
55.
56.
57.
58.
59.
60.
61.
62.
63.
64.
65.
66.
67.
68.
69.
70.
71.
72.
Burton, E.D. (2005). Distribution and Partitioning of Trace Metals and Tributyltin in
Estuarine Sediments. Ph D. Thesis, Griffith University.
Csuros, M., Csuros, C. and Szabo, L.G. (2002). Environmental sampling and analysis for
metals. Lewis Publishers, CRC Press Company.
Makepeace, D.K., Smith, D.W. and Stanley, J.S. (1995). Urban storm water quality.
Summary of contaminant data. Critical Review in Environmental Science and
Technology, 25 (2): 93-139.
Teasdale, P.R., Apte, S.C., Ford, P.W., Batley, G.E. and Koehnken, L. (2003).
Geochemical cycling and speciation of copper in waters and sediments of Macquarie
Harbour, Western Tasmania. Estuarine Coastal and Shelf Science, 57 (3): 475-487.
WHO/UNEP. (1995). Health risks from marine pollution in the Mediterranean
Implication for policy makers. EUR/HFA-TARGET20, PART I, p. 255.
Dean, J.G., Bosqui, F.L. and Lanouette, K.H. (1972). Removing heavy metals from
wastewater. Environmental Science and Technology, 6: 518-524.
El-Nemr, A.M., Sikaily, A.E. and Khaled, A. (2007a). Total and leachable heavy metals
in muddy and sandy
sediments of Egyptian coast along Mediterranean Sea.
Environmental Monitoring Aessessment, 129: 151-168.
Fleischer, M., Sarofhn, A.F., Fassett, D.W., Hammond, P., Shacldette, H.T., Nisber, I.C.
and Epstein, S. (1974). Environmental impact of cadmium. A Review by the panel on
hazardous trace sub-stances. In: Environ. Health Perspect. Goyer, R. and Falk, H.
(Eds). Exp. Issue No. 7. Health, Education and Welfare. Nat. Inst. Environ. Health
Sci., p. 253.
Doyle, J.J. and Pfander, W.H. (1974). Acceptability by lambs of cadmium in feed and
water. Nutrition Reports International, 9 (4):273.
Finkelman, R.B. (2005). Sources and Health Effects of Metals and Trace Elements in our
Environment: An Overview, In: Metal Contaminants in New Zealand. Moore, T.A.,
Black, A., Centeno, J.A., Harding, J.S. and Trumm, D.A. (Eds.). Resolutions Press,
Christchurch, New Zealand, pp. 25-46.
Callender, E. (2003). Heavy Metals in the Environment--Historical Trends. In: Treatise
on Geochemistry. Holland, H.D. and Turekian, K.K. (Eds.). Volume 9: Environmental
Geochemistry, Pergamon, Oxford, pp. 67-105.
Denton, G.R.W., Bearden, B.G., Concepcion, L.P., Siegrist, H.G., Vann, D.T. and H.R.,
Wood. (2001). Contaminant Assessment of Surface Sediments from Tanapag Lagoon,
Saipan, Water and Environmental Research Institute of the Western Pacific, Technical
Report No. 93, University of Guam, Mangilao, Guam.
Morley, N.H., Burton, J.D., Tankere, S.P. and Martin, J.M. (1997). Distribution and
behavior of some dissolved trace metals in the western Mediterranean Sea. Deep-Sea
Research, 44: 675-691.
Morrison, R.J., Gangaiya, P. and Koshy, K. (1996). Contaminated Soils in the South
Pacific Islands. In: Contaminants and the Soil Environment in the Australasia- Pacific
Region. Naidu, R., Kookana, R.S., Oliver, D.P., Rogers, S. (Eds.). Kluwer Academic
Publishers, Amsterdam, pp. 659-675.
Owen, R.B. and Sandhu, N. (2000). Heavy metal accumulation and anthropogenic
impacts on Tolo Harbour, Hong Kong. Marine Pollution Bulletin, 40 (2): 174-180.
WHO. (1995). Environmental Health Criteria No 165, Lead, inorganic. IPCS, World
Health Organisation, Geneva.
Shen, G.T., Boyle, E.A. and Lea, D.W. (1987). Cadmium in corals as a tracer of
historical upwelling and industrial fallout. Nature, 328: 794–796.
Simpson, W.R. (1981). A critical review of cadmium in the marine environment.
Progress in Oceanography, 10 (1): 1-70.
-254-
73.
74.
75.
76.
77.
78.
79.
80.
81.
82.
83.
84.
85.
86.
87.
88.
89.
90.
91.
92.
93.
94.
WHO. (2004). Guidelines for Drinking Water Quality (3rd ed.), ISBN: 92-4-154638-7,
516.
Jumbe, A.S. and Nandini, N. (2009). Heavy metals analysis and Sediment quality values
in Urban Lakes. American Journal of Environmental Sciences, 5 (6): 678-687.
Miller, W.J. (1971). Cadmium nutrition and metabolism in ruminants: Relationship to
concentrations in tissues and products. Feedstuffs, July 17: 24.
Ansari, T.M., Marr, I.L. and Tariq, N. (2004). Heavy Metals in Marine Pollution
Perspective- A Mini Review, 4 (1): 1-20.
Burger, J., Gaines, K.F., Boring, C.S., Stephens, J. W.L., Snodgrass, Dixon, C., M.,
McMahon, Shukla, S., Shukla, T. and Gochfeldw, M. (2002). Metal levels in fish from
the Savannah river: potential hazards to fish and other receptors. Environmental
Research A., 89 (1): 85-97.
Abbasi, S.A., Abassi, N. and Soni, R. (1998). Heavy metals in the Environment (1st ed).
Mittal Publications, ISBN: 81-7099-657-0, pp: 314.
Schulz., H. and Zabel, M. (2000). Marine Geochemistry, Spingler-Verlag, Berlin,
Germany, pp. 455.
Muniz, P., Danulat, E., Yannicelli, B., Garcya-Alonso, J., Medina, G.G and Bycego, M.
(2004). Assessment of contamination by heavy metals and petroleum hydrocarbons in
sediments of Montevideo Harbour (Uruguay). Environment International, 29: 1019–
1028.
Cameron, R.E. (1992). "Guide to site and soil description for hazardous waste sites
characterization" 1: Metals, EPA/600/4-91/029.
Mohammed, N.I. (1991). Biological and biochemical studies of some Siganid fishes
from the Mediterranean waters off Alexandria, M.Sc. Thesis. Faculty of Science,
Biological Oceanography, Alexandria University.
Turnlund, J.R. (1998). Human Whole body copper metabolism. American Journal of
Clinical Nutrition, 67 (Suppl) (960S-964S).
LIanos, R.M. and Mercer, J.F.B. (2002). The molecular basis of copper homestasis and
copper-related disorders. DNA and cell Biology, 21: 259-270.
Gratten, J.P., Huxley, S.I. and Pyatt, F.B. (2003). Modern Bedouin exposure to copper
contamination: An imperial legacy? Ecotoxicology and Environmental Safety, 55: 108115.
Barceloux, D.G. (1999). Cobalt. Clinical Toxicology, 37 (2): 201–216.
Erdogan, M. (2009). Monitoring and statistical assessment of heavy metal pollution in
sediments along Izmir Bay using ICP-MS. Ph D. Thesis, Izmir Institute of Technology,
Izmir
Smith, I.C. and Carson, B.L. (1981). Trace metals in the environment. Ann Arbor, MI,
Ann Arbor Science Publishers.
Greenwood, N.N. and Earnshaw, A. (1984). Chemistry of the Environment, Pergamon
Press Ltd. ISBN 008020576.
Moore, J.W. and Ramamoorthy, S. (1984). Heavy metals in Natural waters, Springer
Verlag.
Saad, M.A.H., Beltagy, A.I., Fahmy, M.A. and Mahmoud, W.M. (2004). Lead speciation
in the sediments of the heavily polluted western harbor of Alexandria. Water, Air, and
Soil Pollution, Focus. 4: 375-384.
WHO. (1995). Human exposure to lead. Geneva: World Health Organization.
Sharma, R. and Prevez, S. (2003). Enrichment and exposure of particulate lead in a
traffic environment in India. Environmental Geochemistry and Health, 25 (3): 297-306.
USCDC (United State Centers for Disease Control) (1991). Preventing lead poisoning
in young children, US Government Printing Office, Atlanta, GA.
-255-
95.
96.
97.
98.
99.
100.
101.
102.
103.
104.
105.
106.
107.
108.
109.
110.
111.
112.
Needleman, H.L. (1987). Low level lead exposure and children's intelligence: A
quantitative and critical review of modern studies. Proc. 6th Int. Conf. on Heavy Metals
in the Environment, New Orleans. Volume 1. CEP Consultants Ltd., Edinburg. pp.1-8.
Anglin-Brown, B., Aromou-Brown, A. and Lalor, G.C. (1995). Heavy metal pollution in
Jamaica I: survey of cadmium, lead and zinc concentrations in the Kintyre and hope flat
districts. Environmental Geochemistry and Health, 17 (2): 51-56.
WHO. (1991). Environmental Health Criteria 124, LINDANE, World Health
Organization A report published jointly by the United Nations Environment Program,
the International Labor Organization and the World Health Organization, Genevea, p.
214.
USEPA (U.S. Environmental Protection Agency). (2001). Nickel compounds.
Technology Transfer Network Air Toxics Website.
Kakela, R., Kakela, A. and Hyvarinen, H. (1999). Effects of nickel chloride on
reproduction of the rat and possible antagonistic role of selenium, Comparative
Biochemistry and Physiolog part C: Pharmacology, Toxicology and Endocrinology,
123 (1): 27-37.
Snodgrass, W.J. (1980). Distribution and behaviour of Nickel in the aquatic
Environment. In: Nickel in the Environment. Nriagu, J.O. (Ed.), John Wiley and Sons,
New York, NY, 203-274.
Eaton, A.D. and Franson, M.A.H. (2005). Standard methods for the examination of
water and wastewater (20th edn.). American Public Health Association, Washington,
ISBN: 0875530478.
Mance, G. (1987). Pollution Threat of Heavy Metals in Aquatic Environments, Elsevier
Applied Science Publishers Ltd, New York, USA.
CCREM (Canadian Council of Resource and Environmental Ministry) (1987). Canadian
Water Quality Guidelines, Environmental, Ottawa, Canada.
Al-Ouran, N. (2005). Environmental Assessment, Documentation and Spatial Modeling
of Heavy Metal Pollution along the Jordan Gulf of Aqaba Using Coral Reefs as
Environmental Indicator. Ph.D. Thesis, Department of Hydrogeology and Environment,
Julius Maximilians University of Würzburg.
WHO. (2001). Environmental Health Criteria No. 221: Zinc, World Health Organization,
Geneva.
Franca, S., Vinagre, I. and Cacador, H.N. (2005). Cabral heavy metal concentrations in
sediment, benthic invertebrates and fish in three salt marsh areas subjected to different
pollution loads in the Tagus Esturay (Portugal). Marine Pollution Bulletin, 50: 993-1018.
Choi, D. and Koh, J.Y. (1998). Zinc and brain injury. Annual review. Neuroscience, 21:
347-375.
Smith, J.C., Zeller, J.A., Brown, E.D. and Ong, S.G. (1976). Elevated plasma Zinc: a
heritable anomaly. Science, 193: 496.
Hotz, C., Lowe, N.M., Araya, M. and Brown, K.H. (2003). Assessment of the trace
element status of individuals and populations: The example of zinc and copper. The
American society of nutritional science, 133 (5): 1563-1560.
Moolenaar, S.W. (1998). Sustainable Management of Heavy Metals in Agro-ecosystems.
Landbouw University, Wageningen.
ANZECC and ARMCANZ (2000). Australian and New Zealand guidelines for fresh and
marine water quality, Australian and New Zealand Environment and Conservation
Council/Agriculture and Resource Management Council of Australia and New Zealand,
Canberra.
Klaassen, C.F. (1996). Casarett & Doull's Toxicology - The basic science of poisons. (5th
ed.) McGrawHill.
-256-
113.
114.
115.
116.
117.
118.
119.
120.
121.
122.
123.
124.
125.
126.
127.
128.
129.
130.
131.
Anschutz, P., Dedieu, K., Desmazes, F. and Chaillou, G. (2005). Speciation, oxidation
state and reactivity of particulate manganese in marine sediments. Chemical
Geochemistry, 218 (3-4): 265-279.
Chester, R., Thomas, A., Lin, F.J., Basaham, A.S. and Jacimto, G. (1988). The solid state
speciation of copper in surface water particulates and oceanic sediments. Marine
Chemistry, 24: 261-292.
NAS (National Academy of Sciences) (1973). Medical and biological effects of
environmental pollutants: manganese. National Academy Press, Washington, DC.
Rope, S.K., Arthur, W.J., Craig, T H. and Craig, E.H. (1988). Nutrient and trace
elements in soil and desert vegetation of southern Idaho. Environmental Monitoring
Assessment, 10: 1–24.
WHO. (1981). Environmental health criteria 17: Manganese. Geneva, World Health
Organization.
Preston, A., Jefferies, D.F., Dutton, J.W.W., Harvey, B.R. and Steele, A.K. (1972).
British Isles coastal waters: The concentrations of selected heavy metals in sea water,
suspended matter and biological indicators-a pilot survey. Environmental Pollution, 3:
69-82.
Greaney, K.M. (2005). An assessment of heavy metal contamination in the marine
sediments of Las Perlas Archipelago, Gulf of Panama, M.Sc.Thesis, Heriot-watt
University, Edinburgh.
Alomary, A.A. and Belhadj, S. (2007). Determination of heavy metals (Cd, Cr, Cu, Fe,
Ni, Pb, Zn) by ICP-OES and their speciation in Algerian Mediterranean Sea Sediment
after a five stage sequential extraction procedure. Environmental Monitoring
Aessessment, 135 (1-3): 265-280.
Institute of Medicine. (2001). Food and Nutrition Board. Dietary Reference Intakes for
Vitamin A, Vitamin K, Arsenic, Boron, Chromium, Copper, Iodine, Iron, Manganese,
Molybdenum, Nickel, Silicon, Vanadium and Zinc. National Academy Press,
Washington, DC:
Dallman, P.R. (1986). Biochemical basis for the manifestations of iron deficiency.
Annual Review of Nutrition, 6: 13-40.
Andrews, N.C. (1999). Andrews NC. Disorders of iron metabolism. N. Engl. J. Med.,
341:1986-95.
Corbett, J.V. (1995). Accidental poisoning with iron supplements. MCN. American
Journal of Maternal and Child Nursing, 20: 234.
Alabaster, J.S. and Lloyd, R. (1980). Water Quality Criteria for Fish (2nd ed.). London:
Butterworths. p.546.
Alloway, B.J. and Ayres, D.C. (1998). Chemical principles of environmental pollution.
Pub. Stanley Thorens Ltd.
Murray, C.V. and Murray, L. (1973). Absorption-desorption equilibria of some
radionucleotides in sediment freshwater and sediment-seawater systems. In: Proceedings
of the Symposium on Interaction of Radioactive Contaminants with the Constituents of
the Marine Environment. International Atomic Energy Commission. pp. 105-124.
Renfro, W.C. (1973). Transfer of Zn from sediments by marine polycheate worm.
Marine Biology, 21: 305- 316.
Batley, G.E. (1989). Trace Metal Speciation: Analytical Methods and Problems. CRC
Press, Boca Raton, Florida.
Mackeviciene, G., Striupkuviene, N. and Berlinskas, G. (2002). Accumulation of Heavy
Metals and Radionuclides in Bottom Sediments of Monitoring Streams in Lithuania.
Ekologija, 2: 69-74.
Forstner, U. and Witlmann, G.T.W. (1979). Metal pollution in the aquatic Environment,
Berlin, Springer-Verlag, Heidelberg, New York.
-257-
132.
133.
134.
135.
136.
137.
138.
139.
140.
141.
142.
143.
144.
145.
146.
147.
148.
149.
Elder, J.F. (1988). Metal Biogeochemistry in Surface-Water Systems. A Review of
Principles and Concepts. U.S. Geological Survey Circular 1013.
Zeri, C. and Voutsinou-Taliadouri, F. (2003). Processes affecting the distribution of
dissolved trace metals in the North Aegean Sea (Eastern Mediterranean). Continental
Shelf Research, (23): 919-934.
Leivouri, M. (1998). Heavy metal contamination in surface sediment in the Gulf of
Finland and comparison with the Gulf of Bothnia. Chemosphere, 36: 43-59.
Sullivan, A.B. and Drever, J.I. (2001). Geochemistry of suspended particles in a mineaffected mountain stream. Applied Geochemistry, 16: 1663-1676.
Sokolowski, M., Wolowicz, M. and Hummel, H. (2001). Distribution of dissolved and
labile particulate trace metals in the overlying bottom water in the Vistula River plume
(southern Baltic Sea). Marine Pollution Bulletin, 42: 967-980.
Nguyen, H.L., Leermakers, M., Osan, J., Torok, S. and Baeyens, W. (2005). Heavy
metals in Lake Balaton: water column, suspended matter, sediment and biota. Science of
the Total Environment, 340: 213-230.
Leivuori, M., Joksas, K., Seisuma, Z., Kulikova, I., Petersell, V., Larsen, B., Pedersen,
B. and Floderus, S. (2000). Distribution of heavy metals in sediments of the Gulf of
Riga, Baltic Sea. Boreal Environment Research, 5 (2): 165-185.
Sherrell, R.M. and Boyle, E.A. (1992). The trace metal composition of suspended
particles in the oceanic water column near Bermuda. Earth and Planetary Science
Letters, 111: 155-174.
Bishop, J.K.B. and Fleisher, M.Q. (1987). Particulate manganese dynamics in Gulf
Stream warm-core rings and surrounding waters of the N.W. Atlantic. Geochimica et
Cosmochimica Acta, 51: 2807-2827.
Aguilea, A. and Amils, R. (2005). Tolerance to cadmium in chlamydomonas sp.
(Chlorophyta) strains isolated from an extreme acidic environment, the Tinto River (SW,
Spain). Aquatic Toxicology, 75 (4): 316-329.
Warnken, J., Dunn, R.J.K. and Teasdale, P.R. (2004). Investigation of recreational boats
as a source of copper at anchorage sites using time-integrated diffusive gradients in thin
film and sediment measurements. Marine Pollution Bulletin, 49: 833-843.
Baeyenes, W., Leemakers, M., De-Gieter, M., Naguyen, H.L., Parmentier, K,
Panutrakul, S. and Elskens, M. (2005). Overview of trace metal contamination in the
scheldt esturay and effect of regulatory measures. Hydrobiolgia, 540: 141-154.
Hirose, K. (2006). Chemical speciation of trace metals in seawater: a review. Analytical
Sciences, 22 (8): 1055-1063.
Apte, S.C., Batley, K.C., Bowles, G.E., Brown, P.L., Creighton, N., Hales, L.T., Hyne,
R.V., Julli, M., Markich, S.J., Pablo, F., Rogers, N.J., Stauber, J.L. and Wild, K. (2005).
A comparison of copper speciation measurements with the toxic response of three
sensitive fresh water organisms. Envionmental Chemistry, 2: 320-330.
Smith, D.S., Bell, R.A. and Kramer, J.R. (2002). Metal speciation in natural waters with
emphasis on reduced sulfur groups as strong metal binding sites. Comparative
Biochemistry and Physiology Part C: Toxicology & Pharmacology, 133 (1-2): 65-74.
Forstner, U. (1979a). Chapter E: Metal transfer between solid and aqueous phases. In:
Metal Pollution in the Aquatic environment. Forstner U and Wittmann GT. (Eds.).
Springer Verlag, Berlin, Heidelberg, New York, pp. 197-269.
Forstner, U. (1979b). Chapter G: Metal concentrations in river, lake and ocean waters In:
Metal pollution in the aquatic environment. Fostner U and Wittmann GT (Eds.).
Springer-velage, Berlin Headelberg, New York, pp. 71-107.
Nelson, A. and Donkin, P. (1985). Processes of bioaccumulation: the importance of
chemical speciation. Marine Pollution, 15: 164-169.
-258-
150.
151.
152.
153.
154.
155.
156.
157.
158.
159.
160.
161.
162.
163.
164.
165.
166.
167.
Simkiss, K. and Taylor, M.G. (1989). Metal fluxes across the membranes of aquatic
organisms. Aquatic Sciences, 1: 173-188.
Goldberg, E.D., Hodge, V., Koide, V., Griffin, M., Gramble, E., Bricker, D.P., Matisoff,
G., Holderen, G.R.J. and Braun, R. (1976). Metal pollution in Tokyo as recorded in
sediments of the Palace Moot. Geochemical Journal, 10: 165-174.
Turner, R.R. and Southworth, G.R. (1999). Mercury Contaminated industrial and mining
sites in North America: an overview with selected case studies. In: Mercury
contaminates sites characteristics, risk assessment and remediation. Ebinghaus R.,
Turner, R.R., De lacerada, L.D., Vasillev, O. and Salomans, W. (Eds.) Springre pub.
Verlg Berlin Beidlberg, pp. 89-112.
FAO (Food and Agriculture Organization of the United Nations). (1976). Manual of
methods in aquatic environment research. Part 2 Guidelines for the use of biological
accumulators in marine pollution monitoring. Portman J. E. (Eds).
Culler, W.R. and Reimer, K.J. (1989). Arsenic speciation in the environment: Chemical
Reviews, 89: 713-764.
Puyate, Y.T., Rim-Rukeh, A. and Awatefe, J.K. (2007). Metal Pollution Assessment and
Particle Size Distribution of Bottom Sediment of Orogodo River, Agbor, Delta State,
Nigeria. Journal of Applied Sciences, 3 (12): 2056-2061.
Basaham, A.S. and El-Sayed, M.A. (1998). Distribution and phase association of some
major and trace elements in the Arabian Gulf sediments. Estuarine Coastal and Shelf
Science, 46: 185-194.
Forstner, U. and Wittman, G.T. (1983). Metal Pollution in the Aquatic Environment,
Springer Verlag, Berlin, Heidelberg, 486.
Howari, F.M. (2005). Distribution of Heavy Metal Concentrations in Surface Sediments
in Dubai Creeks, United Arab Emirates. Annali di Chimica, 95 (9-10): 667-675.
Marchand, C., Lallier-verges, E., Batlzer, F., Alberic, P., Cossa, D. and Bailiff, P.
(2006). Heavy metals distribution in mangrove sediments along the mobile coastline of
French Guiana. Marine Chemistry, 98: 1-17.
Buccolieri, A., Buccolieri, G., Cardellicchio, N., Dell'Atii, A., Leo, A.D. and Maci, A.
(2006). Heavy metals in marine sediments of Taranto Gulf (Ionian Sea, Southern Italy).
Marine Chemistry, 99 (1-4): 227-235.
Allen, H.E. (1995). Metal Contaminated Aquatic Sediments. Michigan, Ann Arbor
Press.
Topcuoglu, S., Kirbasoglu, C. and Gungor, N. (2002). Heavy metals in organisms and
sediments from Turkish coast of the Black Sea. Environment International, 27 (7): 521526.
Morillo, J., Usero, J. and Gracia, I. (2007). Potential Mobility of Metals in Polluted
Coastal Sediments in Two Bays of Southern Spain. Journal of Coastal Research, 32 (2):
352-361.
Berner, R.A. (1981). A new geochemical classification of sedimentary environments.
Journal of Sedimentary research, 51: 359-369.
Hong, J., Calmano, W. and Forstner, U. (1995). Interstitial waters. In: Trace elements in
natural water. Salbu, B. and Steinnes, E. (Eds.), CRC Press, London.
Berry, W.J., Hansen, D.J., Mahony, J.D., Robson, D.L., Di-Toro, D.M., Shipley, B.P.,
Rogers, B., Corbin, J.M. and Boothman, W.S. (1996). Predicting the toxicity of metal
spiked laboratory sediment using acid volatile sulfide and interstitial water
normalizations. Environmental Toxicology and Chemistry, 15 (12): 2067-2079.
Tessier, A., Carignan, R. and Belzile, N. (1994). Processes occurig at the sedimentwater interface: emphasis on the trace elements. In: Chemical and biological regulation
of aquatic systems. Buffle, J and De Vitre, R.R. (Eds.), Lewis Publishers, London.
-259-
168.
169.
170.
171.
172.
173.
174.
175.
176.
177.
178.
179.
180.
181.
182.
183.
Song, Y.H., Peng, J.F., Yuan, P., Cui, X.Y. and Qiu, G.L. (2009). The remediation of
heavy metals contaminated sediment. Journal of Hazardous Materials, 161: 633-640.
Useo, J., Gamero, M., Morillo, J. and Gracia, I. (1998). Comparative study of three
sequential extraction procedures for metals in marine sediments. Environment
International, 24: 487-496.
Kersten, M. and Forstner, U. (1987). Effect of sample pretreatment on the reliability of
solid speciation data of heavy metals implications for the study of early diagenetic
processes. Marine Chemistry, 22: 299-312.
Davies, O.A. and Abowei, J.F.N. (2009). Sediment quality of lower reaches of Okpoka
Creek, Niger Delta, Nigeria. European Journal of Scientific Research, 26 (3): 437-442.
Olubunmi, F.E. (2010). Evaluation of the Status of Heavy Metal Pollution of Sediment
of Agbabu Bitumen Deposit Area, Nigeria. European Journal of Scientific Research, 41
(3): 373-382.
Adeyemo, O.K., Adedokun, O.A., Yusuf, R.K. and Adeleye, E.A. (2008). Seasonal
changes in physicao-chemical parameters and nutrient load of river sediment in Ibadan
city, Nigeria. Global NEST Journal, 10 (3): 326–336.
Suciu, I., Cosma, C., Todica, M., Bolboaca, S.D. and Jantschi, L. (2008). Analysis of
soil heavy metal pollution and pattern in Central Transylvania. International Journal of
Molecular Science, 9 (4): 434-453.
Praveena, S.M., Radojevic, M., Abdullah, M.H. and Avis, A.Z. (2007). Factor-cluster
analysis and enrichment study of mangrove sediments – An example from Mengkabong
Sabah. The Malysian Journal of Analytical Sciences, 2 (2): 421 - 430.
Akinmosin, A., Osinowo, O. and Oladunjoye, M.A. (2009). Radiogenic components of
the Nigeria Tarsand Deposits. Earth Science Research, 13 (1): 64 - 73.
Ajayi, S.O. and Osibanjo, O. (1981). Pollution studies in Nigerian Rivers. Water quality
of some Nigerian rivers. Environmental Pollution (Series B), 2: 87 - 95.
Khan, A.G., Kuek, T.M., Chaudhry, C. and Hayes, W. (2000). Role of plants,
mycorrhizae and phytochelators in heavy metal contaminated land remediation.
Chemosphere, 41: 197-207.
Calace, N., Ciardullo, S., Petronio, B.M., Pietrantonio, M., Abbondanzi, F., Campisi, T.,
and Cardellicchio, N. (2005). Influence of chemical parameters (heavy metals, organic
matter, sulphur and nitrogen) on toxicity of sediments from the Mar Piccolo (Taranto,
Ionian Sea, Italy). Microchemical Journal, 79 (1-2): 243-248.
Dassenakis, M., Adrianos, H., Depiazi, G., Konstantas, A., Karabela, M., Sakellari, A.
and Scoullos, M. (2003). The use of various methods for the study of metal pollution in
marine sediments, the case of Euvoikos Gulf, Greece. Applied Geochemistry, 18: 781794.
Salomonos, W. and Forstner, U. (1984). Metals in the hydrocycle, Springer Publishing
Co. Berlin, Heidelberg, p. 349.
Saad, M.A.H. and Badr, N.B. (2003). Copper speciation in the sediments of the polluted
Eastern Harbor of Alexandria, Egypt. In: Proceeding of the Third International
Conference on Environmental and Health, Chennai, India, 15-17 December. Martin, J.,
Bunch, V., Madha, S. and Vasantha K (Eds.), Chennai: Department of Geography,
University of Madras and Faculty of Environmetal Studies, York University. pp. 417426.
Sarkar, S.K., Franciskovic-Bilinski, S., Bhattacharya, A., Saha, M., Bilinski, H. (2004).
Levels of elements in the surfacial estuarine sediments of the Hugli River, northeast
India and their environmental implications, Environment International, 30 (8): 10891098.
-260-
184.
185.
186.
187.
188.
189.
190.
191.
192.
193.
194.
195.
196.
197.
198.
199.
200.
201.
202.
Adamo, P., Arienzo, M., Imperato, M., Naimo, D., Nardi, G. and Stanzione, D. (2005).
Distribution and partition of heavy metals in surface and sub-surface sediments of
Naples city port. Chemosphere, 61(6): 800–809.
UNEP (1999). Guidelines for Municipal Solid Waste Management: Planning in Small
Islands Developing States in the Pacific Region, South Pacific Regional Environment
Programme (SPREP), Apia, Western Samoa.
Chandra, S. (2002). Investigations into the Lami municipal dump as a source of heavy
metal contamination, M.Sc.Thesis, University of the South Pacific (USP).
Gangaiya, P. (1994). Land-based Pollution Sources in Kiribati: A Case Study, SPREP
Reports and Studies Series No. 80, South Pacific Regional Environment Programme,
Apia, Western Samoa.
Forstner, U. (1990). Inorganic Sediment Chemistry and Elemental Speciation, In:
Sediments: Chemistry and Toxicity of In-Place Pollutants. Baudo, R., Giesy, J. and
Muntau, H. (Eds.), Lewis Publishers, Inc, pp. 61 - 105.
Binning, K. and Baird, D. (2001). Survey of heavy metals in the sediments of the
Swartkops River Estuary. Water South Africa, 27 (4): 461-465.
Burden, F.K., McKelvie, I., Forstner, U. and. Guenther, A (2002). Environmental
Monitoring Handbook. eBook, McGraw-Hill Company.
Malins, D.C. (1984). Chemical pollutants in sediments and diseases of bottom-dwelling
fish in Puget Sound, Washington. Environmental Science and Technology, 18: 705-713.
Wang, X.C., Feng, H. and. Ma, H.Q. (2007). Assessment of Metal Contamination in
Surface Sediments of Jiaozhou Bay, Qingdao, China. Clean soil, Air, Water, 35 (1): 62 –
70.
USEPA (US Environmental Protection Agency) (1998). EPA’s Contaminated Sediment
Management Strategy. 823-R–98–001.
Begum, A., HariKrishna, S. and Khan, I. (2009). Analysis of Heavy metals in Water,
Sediments and Fish samples of Madivala Lakes of Bangalore, Karnataka. International
Journal of Chemtech Research, 1 (2): 245-249.
Begum, A., HariKrishna, S. and Khan, I. (2008). Chemical Composition of Rain water in
South Bangalore, Karnataka, Rasayan. Journal of Chemistry, 1 (4): 774-781.
Gangaiya, P., Tabudravu, J., South, R. and Sotheeswaran, S. (2001). Heavy metal
contamination of the Lami coastal environment, Fiji'. South Pacific. Journal of Natural
Science, 19: 24 - 29.
Langston, W.J. (2000). Toxic Effects of Metals and the Incidence of Metal Pollution in
Marine Ecosystems. In: Heavy Metals in the Marine Environment. Furness, R.W and
Rainbow, P. S. (Eds.), Florida, United States of America, pp.101-122.
Williamson, B., Lewis, Mills, G. and Vant, B. (2003). Contaminants on the coast, In:
The New Zealand Coast. Goff, J. R., Nichol, S. L. and Rouse, H. L. (Eds.), Dunmore
Press, Palmerston North, New Zealand, pp. 237-259.
Long, E.R., McDonald, D., Smith, S.L. and Calder, F.D. (1995). Incidence of adverse
biological effects within ranges of chemical concentrations in marine and estuarine
sediments. Environmental Management, 19 (1): 81-97.
Casado-Martinez, M.C., Buceta, J.L, Belzunce, M.J. and DelValls, T.A. (2006). Using
sediment quality guidelines for dredged material management in commercial ports from
Spain. Environment International, 32 (3): 388-396.
McCready, S., Birch, G.F., and Long, E.R. (2006). Metallic and organic contaminants in
sediments of Sydney Harbour, Australia and vicinity-a chemical dataset for evaluating
sediment quality guidelines. Environment International, 32 (4): 455-465.
Guerra-Garcia, J.M. and Garcia-Gomez, J.C. (2005). Assessing pollution levels in
sediments of a harbour with two opposing entrances. Environmental implications.
Journal of Environmental Management, 77 (1): 1-11.
-261-
203.
204.
205.
206.
207.
208.
209.
210.
211.
212.
213.
214.
215.
216.
217.
218.
219.
220.
Red fern, F.M. (2006). Heavy Metal Contamination from Landfills in Coastal Marine
Sediments: Kiribati and New Zealand, M.Sc. Thesis, The University of Waikato.
Williamson, R.B. and Wilcock, R.J. (1994). The distribution and fate of contaminants in
Estuarine Sediments: Recommendations for environmental monitoring and assessment.
Technical Publication. 47, New Zealand.
Peng, J., Song, Y., Yuan, P., Cui, X. and Qiu, G. (2009). The remediation of heavy
metals contaminated sediment. Journal of Hazardous Materials, 161: 633-640.
Luoma, S.N. (2000). Processes Affecting Metal Concentrations in Estuarine and Coastal
Marine Sediments. In: Heavy Metals in the Marine Environment. Furness, R.W. and
Rainbow, P.S. (Eds.), CRC Press, Inc, Florida, United States of America, pp. 51-66.
Williamson, R.B., Blom, A., Hume, T.M., Glasby, G.P. and Larcombe, M. (1992).
Heavy Metals in Manukau Harbour Sediments, Water Quality Center Publication 23,
Hamilton, New Zealand.
Krumgalz, B.S., Fainshtein, G. and Cohen, A. (1992). Grain size effect on anthropogenic
trace metals and organic matter distribution in marine sediments, The Science of the
Total Environment, Elsevier Science Publishers, 116: 15 - 30.
Loomb, C.A.M. (2001). Muddy Sedimentation in a Sheltered Estuarine Marine,
Westpark Marina, Auckland, New Zealand, M.Sc.Thesis, The University of Waikato.
Qian, J., Shan, X.Q., Wang, Z.J. and Qiang, T. (1996). Distribution and plant availability
of heavy metals in different particle size fractions of soil. The Science of the Total
Environment, 187: 131-141.
Fan, W., Wang, W.X., Chen, J., Li, X. and Yen, Y.F. (2002). Cu, Ni and Pb speciation in
surface sediments from a contaminated bay of northern China. Marine Pollution Bulletin,
44: 816–832.
Sturm TW, Amirtharajah, A. and Tiller, C.L. (2002). Mobilization and fate of inorganic
contaminants due to resuspension of cohesive sediment. Final Project Report submitted
to the HSRC/S & SW, February 1998. http://w ww.hsrc.org/hsrc/html/ssw/yr5mobil.html.
Tabata, M., Ghaffar, A. and Nishimoto, J. (2009). Accumulation Of Metals In Sediments
Of Ariake Bay, Japan. Electronic Journal of Environmental, Agricultural and Food
chemistry, 8 (10): 937-949.
Taulis, M.E. (2005). Metal Contaminants in Leachate from Sanitary Landfills. In: Metal
Contaminants in New Zealand. Moore, T.A., Black, A., Centeno, J.A., Harding, J.S. and
Trumm, D.A. (Eds.), Resolutionz Press, Christchurch, New Zealand, pp. 173-190.
Gundersen, P. and Steinnes, E. (2003). Influence of pH and TOC concentration on Cu,
Zn, Cd, and Al speciation in rivers. Water Research, 37: 307-318.
Uta, K. and Jens, W. (2006). Long-term effects of the Aznalcollar mine spill heavy metal
content and mobility in soils and sediments of the Guadiamar river valley (SW Spain).
Science of Total Environment, 367: 855–871.
Valerie, C., Rudy, S. and Jisika, V. (2004). Assessment of acid neutralizing capacity and
potential mobilization of trace metals from land disposed dredged sediments. Science of
Total Environment, 333: 233-247.
Sekaly, A.L.R., Mandal, R., Nouri, M.H., Murimboh, J., Chakrabart, C.L., Back, M.H.,
Gregoire, D.C. and Schroeder, W.H. (1999). Effect of metal/fulvic acid mole ratios on
the binding of Ni (II), Pb (II), Cu (II), Cd (II), and Al (III) by two well characterized
fulvic acids in aqueous model solutions. Analytic Chimica Acta, 402: 211-221.
Salomonos, W. and Stigliani, W.M. (1995). Biogeodynamics of pollutants in soils and
sediments: Risk assessment of Delayed and Non linear responses. Environmental
Science, Springer-Verlag, New York: 331-343.
Hansen, D.J., Berry, W.J., Mahony, J.D., Boothman, W.S., Di Toro, D.M., Robson,
D.L., Ankley, G.T., Ma, D. and Pesch, C.E. (1996). Predicting the toxicity of metal
-262-
221.
222.
223.
224.
225.
226.
227.
228.
229.
230.
231.
232.
233.
234.
235.
236.
contaminated field sediments using interstitial concentrations of metals and acid volatile
sulfide normalizations. Environmental Toxicology and Chemistry, 15 (12): 2080-2094.
Calmano, W., Hong, J.F. and Forstner, U. (1993). Binding and mobilization of heavy
metal in contaminated sediment affected by the pH and redox poptential. Water Science
and Technology, 28 (8-9): 223-235.
Tneofanis, Z.U., Astrid, S., Lidia, G. and Calmano, W.G. (2001). Contaminants in
sediments remobilization and demobilization. Science of the Total Environment, 266:
195–202.
Emile, M.L. and Jonathan, R.P. (1998). The effect of salinity on binding of Cd, Cr, Cu
and Zn to dissolved organic matter. Chemosphere, 37: 861–874.
Garnier, J.M., Ciffroy, P. and Lakhdar, B. (2006). Implications of short and long term
(30 days) sorption on the desorption kinetic of trace metals (Cd, Zn, Co, Mn, Fe, Ag, Cs)
associated with river suspended matter. Science of the Total Environment, 366: 350–
360.
Loring, D.H., Rantala, R.T.T., Morris, A.W. and Asmund, G. (1985). Non-detriral and
detrital particulate metal transport in estuarine systems. In: Planteary Ecology.
Caldawell, D.E., Brierley, J.A. and. Brierley, C.L. (Eds.). Van Nostrand Reinhold
Company, N. Y., pp. 187-197.
Khairy, M. (2008). Risk Posed by Anthropogenically Driven Toxic Pollutants in Abu
Qir Bay, East Alexandria, Egypt. Ph.D. Thesis, Faculty of Life Sciences, Technical
University of Carolo-Wilhelmina of Braunschweig, Germany.
Prica, M., Dalmacija, B., Roncevic, S., Krcmar, D. and Becelic, M. (2008). A
comparison of sediment quality results with acid volatile sulfide (AVS) and
simultaneously extracted metals (SEM) ratio in Vojvodina (Serbia) sediments. Science
of The Total Environment, 389: 235–244.
Ankley, G.T., Phipps, G.P., Leonard, E.N., Benoit, D.A., Mattson, V.R., Kosian, P.A.,
Cotter, A.M. Dierkes, J.R., Hansen, D.J. and Mahony, J.D. (1991). Acid volatile
sulphide as a factor mediating cadmium and Nickel bioavailability in contaminated
sediments. Environmental Toxicology and Chemistry, 10 (10): 1299-1307.
Chen, J., Dong, L. and Deng, B. (1989). A Study on heavy metal partitioning in
sediments from Poyang Lake in China. Hydrobiologia, 176-177(1): 159-170.
Sposito, G., Lund, J. and Chang, A.C. (1982). Trace metal chemistry in arid zone field
soils amended with sewage sludge. In: Fractionation of Ni, Cu, Zn, Cd and Pb in solid
phases. Journal of the American Society and Soil Science, 46: 260-264.
Wette, B., Bles, N.N. and Montiel, A. (1983). Study of the different methods of
speciation of heavy metals in the sediments: II Applications. Environmental Technology,
4: 223-238.
Clevenger, T.E. (1990). Use of sequential extraction to evaluate the heavy metals in
mining wastewater. Water, Air and Soil Pollution, 50: 241-255.
Howard, J.L. and Vandenbrink, W.J. (1999). Sequential extraction analysis of heavy
metals in sediments of variable composition using nitrilotriacetic acid to counteract
resorption. Environmental Pollution, 106: 285-292.
Kersten, M. and Forstner, U. (1986). Chemical fractionation of heavy metals in anoxic
estuarine and coastal sediments. Water Science and Technology, 18: 121-130.
Li, X., Shen, Z. Wai, O. and Li, Y. (2001). Chemical forms of Pb, Zn and Cu in the
sediments profiles of the Pearl River estuary. Marine Pollution Bulletin, 42: 215-223.
Song, Y.S. and Muller, G. (1999). Sediment-water interaction in anoxic freshwater
sediments mobility of heavy metals and nutrients (lecture notes in earth sciences),
Berlin, Heidelberg, New York, Springer-Verlag, pp. 14-16.
-263-
237.
238.
239.
240.
241.
242.
243.
244.
245.
246.
247.
248.
249.
250.
251.
252.
253.
254.
Griethuysen, C.V., Gillissen, F. and Koelmans, A.A. (2002). Measuring acid volatile
sulphide in floodplain lake sediment: effect of reaction time, sample size and aeration.
Chemosphere, 47: 395.
Simpson, S.T., Apte, S.C. and Batlekny, G.E. (2000). Effect of short term resuspension
events on the oxidation of cadmium, lead, and zinc sulphide phases in anoxic estuarine
sediments. Environmental Science and Technology, 34 (21): 4533-4537.
Hansen, W., Berry, W. Boothman, Di Toro, D.M. and Mahony, J.D. (1998). U.S
Environment Protection Agency, National Health Effects Environmental Research
Laboratory: Atlantic Ecology Division, 27 tarzwell Drive, Narragansett, RI 02882.
Daskalakis, K.D. and O'Connor, T.P. (1995). Distribution of chemical concentrations in
US coastal and estuarine sediment. Marine Environmental Research, 40 (4): 381-398.
Yu, K.C., Tsai, L.J., Chen, S.T. and Ho, S.T. (2001). Chemical binding of heavy metals
in anoxic river sediments. Water Research, 17: 4086-4094.
Argeses, E., Ramieri, E., Bettiol, C., Pavoni, B., Chiozzootto, E. and Sfriso, A. (1997).
Pollutant Exchange at the Water/Sediment Interface in the Venice Canals. Water, Air &
Soil Pollution, 99 (1-4): 255-263.
Forstner, U. (1989). Contaminated sediments: Lectures on environmental aspects of
particle associated chemicals in aquatic systems, Berlin, Heidelberg, New York,
Springer-Verlag, 157.
Di Toro, D.M., Hansen, D.J., McGrath, J.M. and Berry, W.J. (1992). Acid volatile
sulfide predicts the acute toxicity of cadmium and nickel in sediments. Environmental
Science and Technology, 26: 96-101.
McGrath, J.A., Paquin, P.R. and Di Toro, D.M. (2002). Use of the SEM and AVS
approach in predicting metal toxicity in sediments, Fact sheet on Environmental risk
assessment. The International Council on Mining and Metals (ICMM).
Burton, G.A., Nguyen, J.L., Janssen, C., Baudo, R., McWilliam, R. and Bossuyt, B.
(2005). Field validation of sediment zinc toxicity. Environmental Toxicology and
Chemistry, 24: 541-553.
Yin, H.B., Fan, C.X., Ding, S.M., Zhang, L. and Li, B. (2008). Acid volatile sulfides and
simultaneously extracted metals in a metal polluted area of Taihu Lake, China. Bulletin
of Environmental Contamination and Toxicology, 80 (4): 351-355.
Gomes, M.V.T., Costa, A.S., Garcia, C.A.B., Passos, E.A. and Alves, J.P.H. (2010).
Concentrations and geochemical associations of Pb and Zn in sediments of the river São
Francisco impacted by wastes from industrial zinc production Quimica Nova, 33 (10):
2088-2092.
Martinez-Sanchez, M.J., Navarro, M.C., Perez-Sirvent, C., Marimon, J., Vidal, J.,
Garcia-Lorenzo, M.L. and Bech, J. (2008). Assessment of the mobility of metals in a
mining-impacted coastal area (Spain, Western Mediterranean). Journal of Geochemical
Exploration, 96: 171-182.
Gao, X., Chen, C.T.A., Wang, G., Xue, Q., Tang, C. and Chen, S. (2010). Environmental
status of Daya Bay surface sediments inferred from a sequential extraction technique.
Estuarine Coastal Shelf Science, 86: 369-378.
Cornelis, R., Crews, H., Caruso, J. and Heumann, K. (2005). Handbook of elemental
speciation II. Species in the environment, food, medicine and occupational health.
Chichester, John Wiley & Sons.
WHO (World Heath Organization). (2006). Environmental Health Criteria 234.
Elemental Speciation in Human Health Risk Assessment.
Verloo, M., Kiekens, L. and Cottenie, A. (1980). Distribution patterns of essential and
nonessential trace elements in the soil-soil solution system. Pedologie, 30 (2): 163–175.
Adriano., D.C. (1986). Trace Elements in Terrestrial Ecosystems. Springer Verlag, New
York. p.533
-264-
255.
256.
257.
258.
259.
260.
261.
262.
263.
264.
265.
266.
267.
268.
269.
270.
271.
272.
273.
Lindsay, W.L. (1979). Chemical Equilibria in Soils, John Wiley, New York, p. 449.
Hlavay, J., Prohaska, T., Weisz, M., Wenzel, W.W. and Stingeder, G.J. (2004).
Determination of Trace Elements Bound to Soils and Sediment Fractions. Pure Applied
Chemistry, 76 (2): 415-442.
Martin, J.M., Nirel, P. and Thomas, A.J. (1987). Sequential extraction techniques:
promises and problems. Marine Chemistry, 22: 313.
Wang, S., Jia, Y., Wang, S., Wang, X., Wang, H., Zhao, Z. and Liu, B. (2010).
Fractionation of heavy metals in shallow marine sediments from Jinzhou Bay, China.
Journal of Environmental Sciences, 22 (1): 23-31.
Horowitz, A.J. (1991). A primer on trace metal – sediment chemistry. 2nd edition, Lewis
publisher, Chelsea (Michigan), p.136.
Larner, B.I., Paler, A.S., Seen, A.J. and Townsend, A.T. (2008). A comparison of an
optimized sequential extraction procedure and dilute leaching of elements in anoxic
sediments, including the effects of oxidation on sediment metal partitioning. Analytica
Chimica Acta, 608: (2) 147-157.
Gleyzes, C., Tellier, S. and Astruc, M. (2002). Fractionation studies of trace elements in
contaminated soils and sediments: a review of sequential extraction procedures Trends
Analytical Chemistry, 21 (6-7): 451-467.
Kestern, M. and Foِrstner, U. (1995). Speciation of trace metals in sediments and
combustion waste. In: Chemical speciation in the environment. Ure, A.M and Davidson,
C.M. (Eds.), Blackie Academic & Professional, London,
Passos, E.A., Alves, J.P.H., Garcia, C.A.B. and Costa, A.C.S. (2011). Metal
fractionation in the sediments for Sergipe River, Northeast, Brazil. Journal of Brazilian
Chememical Society, 22 (5):828-835.
Zakir, H.M. and Shikazono, N. (2011). Environmental mobility and geochemical
partitioning of Fe, Mn, Co, Ni and Mo in sediments of an urban river. Environmental
Chemistry & Ecotoxicology, 3 (5): 116-126.
Zakir, H.M., Shikazono, N. and Otomo, K. (2008). Geochemical distribution of trace
metals and assessment of anthropogenic pollution in sediments of Old Nakagawa River,
Tokyo, Japan. American Journal of Environmental Science, 4 (6): 654-665.
Singh, K.V., Singh, P.K. and Mohan, D. (2005). Status of heavy metals in water and bed
sediments of river Gomti a tributary of the Ganga River, India. Environmental
Monitoring Assessment, 105 (1-3): 43-67.
Pardo, R., Barrado, E., Perez, L. and Vega, M. (1990). Determination and association of
heavy metals in sediments of the Pisucraga, river. Water Research, 24 (3): 373-379.
Krishna, R.R., Xu, C.Y. and Subraja, C. (2001). Assessment of electrokinetic removal of
heavy metals from soils by sequential extraction analysis. Journal of Hazardous
Materials, 84: 279-296.
Zerbe, J., Sobezynski, T., Elbanowska, H. and Siepak, J. (1999). Speciation of heavy
metals in bottom sediments of lakes. Journal of Environmental Studies, 8 (5): 331-339.
Tokalioglu, S., Kartal, S. and Elci, L. (2000). Determination of heavy metals and their
speciation in lake sediments by flame atomic absorption spectrometry after a four stage
sequential extraction procedure. Analytica Chimica Acta, 413 (1-2): 33-40.
Turner, A. (2000). Trace metal contamination in sediments from UK estuaries: An
empirical evaluation of the role of hydrous iron and manganese oxides. Estuarine Coastal
and Shelf Science, 50: 355-371.
Root, R.A., Dixit, S., Campbell, K.M., Jew, A.D., Hering, J.G. and O'Day, P.A. (2007).
Arsenic sequestration by sorption processes in high iron sediments. Geochimica et
Cosmochimica Acta, 71: 5782-5803.
Purushothaman, P. and Chakrapani, G.J. (2007). Heavy metals fractionations in Ganga
river sediments, India. Environmental Monitoring Assessment, 132 (1-3): 475-489.
-265-
274.
275.
276.
277.
278.
279.
280.
281.
282.
283.
284.
285.
286.
287.
288.
289.
290.
291.
Sharmin, S., Zakir, H.M. and Shikazono, N. (2010). Fractionation profile and mobility
pattern of trace metals in sediments of Nomi River, Tokyo, Japan. Journal of Soil
Science and Environmental Management, 1 (1): 1-14.
Jain, C.K. (2004). Metal fractionation study on bed sediments of river Yamuna, India.
Water Research, 38: 569-578.
Rubio, R., Lopez-Sanchez, J.F. and Rauret, G. (1991). La especiacio`n so` lida de trazas
de metales en sedimentos. Aplicacio`n a sedimentos muy contaminados. Analytica De
Quimica, 87 (5): 599–605.
Castro, H., Aguilera, P.A.k, Martinez, J.L. and Carrique, E.L. (1999). Differentiation of
clams from fishing areas an approximation to coatal quality assessment. Environmental
Monitoring Aessessment, 54: 229-237.
David, C.P. (2003). Heavy metal concentrations in growth bands of corals: a record of
mine tailings input through time. Marine Pollution Bulletin, 46 (2): 187-196.
Fan, W.H., Wang, W.X. and Chen, J.S. (2002). Geochemistry of Cd, Cr, and Zn in
highly contaminated sediments and its influences on assimilation by marine bivalves.
Environmental Science & Technology. 36: 5164-5171.
UNEP (1996). Workshop on policies for Sustainable Development of Mediterranean
coastal areas, Santorini Island, 26–27 April 1996. Papers by a group of experts. MAP
Technical Reports Series, 114,UNEP, Athens.
El-Nemr, A., Said, T.O, Khaled, A., El-Sikaily, A. and Abdallah, A.M.A. (2007b). The
distribution and sources of polycyclic aromatic hydrocarbons in surface sediments along
the Egyptian Mediterranean Coast. Environmental Monitoring Aessessment, 124: 343359.
MESAEP (Mediterranean Scientific Association of Environmental Protection). (2003).
In: Proceedings 11th International Symposium on Environmental Pollution and its Impact
on Life in the Mediterranean Region, Antalya, Turkey, 4-8th October, 2003.
EEA (1999). State and pressures of the marine and coastal Mediterranean environment.
Environmental assessment Series No, 5. Copenhgen: European Environment Agency.
Fowler, S.W. (1986). PCBs and the environment: the Mediterranean marine ecosystem.
In: Waid JS, ed. PCBs and the environment, vol.III. Boca, Florida: CRC Press.
El-Wakeel, S.K., Dogheim, M., Khadr, A. and Sayed, M.K.E. (2006). Study of coastal
sediments, west of Alexandria (El-Agami - Sidi Kreir). Egyptian Journal of Aquatic
Research, 32 (1): 105-123.
UNEP. (2002). Revision of pollution hot spots in the Mediterranean, 28-29 January.
First Meeting of the ad-hoc Technical Committee to select pollution hot spots for the
preparation of pre-investment studies within the GEF Project, UNEP, Athens.
EEAA (Egyptian Environmental Affairs Agency). (2009a). Alexandria integrated coastal
zone management project. Environmental and Social Impact Assessment, p.113.
Nessim, R.B., Bassiouny, A.R., Zaki, H.R., Moawad, M.N. and Kandeel, K.M. (2010).
Environmental studies at El-Mex Region (Alexandria-Egypt) during 2007-2008. World
Applied Sciences Journal, 9 (7): 779-787.
Ahdy, H.H.H. and Khaled, A. (2009). Heavy metals contamination in sediments of the
western part of Egyptian Mediterranean Sea. Australian Journal of Basic and Applied
Science, 3 (4): 3330-3336.
El-Sammak, A.A. and Aboul-Kassem, T.A. (1999). Metal pollution in the sediments of
Alexandria Region, Southeastern Mediterranean, Egypt. Bulletin of Environmental
Contamination and Toxicology, 63 (2): 263-270.
Rifaat, A.E. (1982). Bottom sediments of the western Harbour of Alexandria and its
neighborhood. M.Sc. Thesis, Faculty of Science, Alexandria University. p.114.
-266-
292.
293.
294.
295.
296.
297.
298.
299.
300.
301.
302.
303.
304.
305.
306.
307.
308.
Abdallah, M.A.M. (2007a). Speciation of trace metals in coastal sediments of El-Mex
bay South Mediterranean Sea west of Alexandria (Egypt). Environmental Monitoring
Aessessment, 132: 111-123.
El-Gindy, A., Aboul-Dahab, O. and Halim, Y. (1986). Preliminary estimation of water
and trace metals balances in El-Mex Bay, west of Alexandria. Rapp. Com. Ins. Mer.
Medit., 30 (2): 127.
Masoud, M.S., El-Samra, M.I. and El-Sadawy, M.M. (2007). Heavy-metal distribution
and risk assessment of sediment and fish from El-Mex Bay, Alexandria, Egypt.
Chemistry and Ecology, 23 (3): 201–216.
Abdallah, M.A.M. (2008). Trace metals behavior in Mediterranean climate coastal bay:
El-Mex bay, Egypt and its coastal environment. Global Journal of Environmental
Research, 2 (1): 23-29.
Dorgham, M.M., E.Abdel-Aziz, N., El-Deeb, K.Z. and Okbah, M.A. (2004).
Eutrophication problems in the Western Harbour of Alexandria, Egypt. Oceanologia, 46
(1): 25-44.
Butzer, K.W. (1960). On the Pleistocene shorelines of Arab's Gulf, Egypt. The Journal
of Geology, 68 (6): 626-637.
Mostafa, A.R. and Barakat, A.O. (2004). An overview of Metal Pollution in the Western
Harbour of Alexandria, Egypt. Soil & Sediment Contamination, 13: 299-311.
Salem, A. and Sharkawi, F. (1981). Industrial waste released to the aquatic environment
in Mex area. Aquat. Pollt. Project, EGY/73/058/A/10/13.
Saad, M.A.H. and Hemeda, E.I. (1992). Effect of Pollution on the Western Harbor of
Alexandria. I. Environmental Characteristics. Science of Total Environment, Suppl.:
737-753.
El-Mamoney, M.H. and Mohamed, L.A. (2006). Concentration of some heavy metals in
short cores in the Nearshore sediments of Alexandria, Egypt. Qatar Univiversity Science
Journal, 26: 101-112.
Abdallah, M.A.M. and Abdallah, A.M.A. (2008). Biomonitoring study of heavy metals
in biota and sediments in the south eastern coast of Mediterranean sea, Egypt.
Environmental Monitoring Aessessment, 146: 139-145.
Abdallah, M.A.M. (2007b). Accumulation and distribution of heavy metals in surface
sediments of a semi enclosed basin in the southeastern Mediterranean Sea, Egypt.
Mediterranean Marine Science, 8 (1): 31-40.
El-Rayis, O.A., Hemedan, E.I., Ismael, A.M. and Jammo, K. (2003). Approaches to
environmental restoration of a polluted harbour with submerged archaeology: the
Alexandria case study. Marine Pollution Bulletin, 47: 193-197.
El-Rayis, O.A. and Abdallah, M.A.M. (2006). Contribution of nutrients and some trace
metals from a huge Egyptian drain to the SE-Mediterranean Sea, west of Alexandria.
Mediterranean Marine Science, 7 (1): 79-86.
Nasr, S. M., El-Eskandarani, M. and El-Gamal, S.M.M. (2011). Geochemical and
granulometric normalization for heavy metals in bottom sediments of eastern harbor,
Alexandria. The XIX International conference on Marine geology, The Russian
Academy of Science, P.P. Shirshov Institute of Oceanography, Mosco, November 14-18
Saad, M.A.H. and Hassan, E.M. (2002). Heavy metals in the Rosetta estuary of the Nile
and the adjoining Mediterranean waters: evidence of removal of dissolved heavy metals
from waters as a result of possible binding to suspended matter. Hydrobiologia, 469:
131–147.
Shereet, S.M. (2009). Pollution of Petroleum hydrocarbon in the new Damietta Harbour,
Egypt. Emirates Journal of Engineering Research, 14 (1): 65-71.
-267-
309.
310.
311.
312.
313.
314.
315.
316.
317.
318.
319.
320.
321.
322.
323.
324.
325.
Said, T.O. and Hamed, M.A. (2001). Determination of organo chlorine compounds in
coastal water of the Gulf of Suez and river Nile, Egypt. The Journal of Egyptian
Academic Society for Environmental Development, 2 (2): 125- 139.
Said, T.O. and Hamed, M.A. (2006). Determination of Persistent Organic Pollutants in
Water of New Damietta Harbor, Egypt. Egyptian Journal of Aquatic Research, 32 (1):
235-245.
Faragallah, H.M., Askar, A.I., Okbah, M.A. and Moustafa, H.M. (2009). Physicochemical characteristics of the open Mediterranean Sea water far about 60 Km from
Damietta harbor, Egypt. Journal of Ecology and The Natural Environment, 1 (5): 106119.
Shata, M.A. and Deghedy, E.E.M. (2004). Sedimentological and petrographical
characteristics of subsurface sediments in El-Dekheila harbor, Alexandria, Egypt.
Egyptian Journal of Aquatic Research, 30 (A): 142-149.
Fahmy, M.F., Tayel, F.T. and Sheriadah, M.A. (1997). Spatial and seasonal variations of
dissolved trace metals in two contaminated basins of the coastal Mediterranean Sea,
Alexandria, Egypt. Bulletin of Faculty of Science, Alexandria University, 37 (2): 187198.
Abdalla, R., F.A. Zaghloul, Y.A. Hassan, and Moustafa, H.M. (1995). Some water
quality characteristics of El-Dekhaila harbour, Alexandria, Egypt. Bulletin of National
Institute of Oceanography and Fisheries, A.R.E. 21(1): p. 85-102.
Faragallah, H.M. (2004). Chemical fractioantion of phosphorus and some heavy metals
in suspended matter in water column and sediments in Abu-kir Bay, Egypt. Ph.D.
Thesis, Faculty of Science, Alexandria University, p. 347.
Aboul Dahab, O. (1992). Mercury pollution in Abu Kir Bay East of Alexandria, Egypt.
Qatar Univiversity Science Journal, 12: 222-227.
El-Deeb, M.K.Z. (1977). Hydrography and Chemistry of Abu Kir Bay. M.Sc. Thesis,
Faculty of Science, Alexandria University.
Abdel-Moati, A.R. (1991). An input/output flux for lead in a coastal bay of Alexandria
region. Water, Air and Soil Pollution, 59: 261-269.
Saad, M.A.H. and Badr, N.B.E. (2005). Dissolved Oxygen and Organic Matter in a
Harbor and a Bay in Egypt receiving Organic Pollutants from different sources,
CM2005/Z: 13.
Khalil, M.K., Radwan, A.M. and El-Moselhy, K.M. (2007). Distribution of phosphorus
fractions and some of heavy metals in surface sediments of Burullus lagoon and adjacent
Mediterranean Sea. Egyptian Journal of Aquatic Research, 33 (1): 277-289.
Elsammak, A.A. and El-Sabrouti, M.A. (1995). Assessment of metals pollution in the
sediments of Lake Burullus. S.E. Mediterranean, Egypt. Qatar University Science
Journal, 15 (2): 451-455.
Fathi, A.A. and Abdelzahar, H.M.A. (2005). Environmental assessment and monitoring
of lake Burullus, Egypt with special reference to sediment and water quality, Proc. 2nd
Saudi Sci, Conf., Fac Sci, KAU, 15-17 March 2004, part 1, pp 149-166 A.d./1425 A.H.
Saeed, S.M. and Shaker, I.M. (2008). Assessment of Heavy Metals Pollution in Water
and Sediments and Their Effect on Oreochromis Niloticus in The Northern Delta Lakes,
Egypt, 8th International Symposium on Tilapia in quaculture: pp. 475-490.
El-Tokhi, M., Abdelgawad, E. and Lotfy, M.M. (2008). Impact of heavy metals and
petroleum hydrocarbons contamination of the East Port Said Port area, Egypt. Journal of
Applied Sciences Research, 4 (12): 1788-1798.
Nasr, S.M., Sokkary, I.H.E., Fakhry, A.A. and Shibiny, I.A.R. (1990). Fractionation of
some heavy metals in the coastal belt sediments of Alexandria, Regional symp. Environ.
Stud. (UNARC), Alexandria: 415-436.
-268-
326.
327.
328.
329.
330.
331.
332.
333.
334.
335.
336.
337.
338.
339.
340.
341.
342.
343.
Faragallah, H.M. and Khalil, M.K. (2009). Chemical fractionation of Copper and
Manganese in the sediment of Alexandria two bays, Egypt. Global Journal of
Environmental Research, 3 (1): 61-67.
El-Anany, W.I. (2008). Trace metals speciation in the sediments of Burullus Lake.
M.Sci. Thesis, Institute of Graduate Studies and Research, Alexandria University.
Okbah, M.A., ElKhair, E.M.A. and Moustafa, H.M. (2011). Simultaneously extracted
metals / acid volatile sulfide in surface sediments from Damietta Port and the offshore
area. Egyptian Journal of Aquatic Research, 37 (1): 31-40.
Andresen, M.C. and Draggan, S. (2007). "Ecological risk assessment" In: Encyclopedia
of Earth. Cutler J. (Eds), Cleveland (Washington, D.C.: Environmental Information
Coalition, National Council for Science and the Environment), First Published in the
Encyclopedia of Earth August 31, 2007; Last revised Date August 31, 2007; Retrieved
July 1, 2011. http://www.eoearth.org/article/Ecological_risk_assessment.
USEPA. (1998). Guidelines for Ecological Risk Assessment. Published on May 14,
1998, Federal Register 63 (93):26846-26924). Risk Assessment Forum. U.S.
Environmental Protection Agency, Washington, DC.
Frihy, O.E., Iskander, M.M. and Badr, A.A. (2004). Effects of shoreline and bedrock
irregularities on the morphodynamics of the Alexandria coast littoral cell, Egypt. GeoMarine Letters, 24 (4): 195-211.
Morcos, S., Tongring, N., Halim, Y., El-Abbadi, M. and Awad, H. (2003). Toward
integrated management of Alexandria’s coastal heritage. The United Nations
Educational Scientific and Cultural Organization.
EEAA (Egyptian Environmental Affairs Agency). (2001). National Environmental
Action Plane.
Susmita, D. (2009). Sea level rise and storm surges; A comparative analysis of impacts
in developing countries, Policy Research Working Paper 4901 (Washington, DC: World
Bank Development Research Group Environment and Energy Team).
EEAA (Egyptian Environmental Affairs Agency). (1999). The Arab Republic of Egypt:
Initial National Communication on Climate Change. Prepared for the United Nations
Framework Convention on Climate Change UNFCCC (Cairo: EEAA, 1999).
Dewidar, K.M.A. (1997). Impact analysis of environmental changes in the coastal zone
by remote sensing techniques (case study: El-Alamein to port said coast). Ph.D. Thesis,
Environmental Studies. Institute of Graduate Studies and Research, University of
Alexandria.
Hamed, A.A. (1983). Atmospheric circulation features over the Southeastern part of the
Mediterranean Sea in relation to weather conditions and wind waves along the Egyptian
coast, Ph.D. Thesis, Faculty of Science. Alexandria University, Alexandria, Egypt, pp.
150.
Shata, A.A.S. (2000). Site potential of Egypt for wind energy applications. M.Sc.
Thesis, Physics department, Faculty of Science, Zagazig University, Egypt.
Kamel, M.S. (2010). Geostrophic current patterns off the Egyptian Mediterranean coast,
Oceanologia, 52 (2): 299-310.
Nafaa, M.G. and Frihy, O.E. (1993). Beach and nearshore features along the dissipative
coastline of the Nile Delta, Egypt. Journal of Coastal Research, 2: 423-433.
Stanley, D.J. and Warne, A.G. (1993). Nile Delta: recent geological evolution and
human impact. Science, 260: 628-634.
El-Raey, M., Dewidar, K. and El-Hattab, M. (1999). Adaptation to the impact of sea
level rise in Egypt. Mitigation and Adaptation Strategies for Global Change, 4 (3-4):
343-361.
Stanley, D.J. (1988). Subsiding in the northeastern Nile Delta, rapid rates, possible
causes and consequences. Science, 240: 497–500.
-269-
344.
345.
346.
347.
348.
349.
350.
351.
352.
353.
354.
355.
356.
357.
358.
359.
360.
361.
362.
363.
364.
Frihy, O.E. (2001). The necessity of environmental impact assessment (EIA) in
implementing coastal projects: lessons learned from the Egyptian Mediterranean Coast.
Ocean & Coastal Management, 44: 489-516.
Abdallah, A.M., El-Din, S.H.S. and Shereet, S.M. (2006). Analysis of wave observations
and wave transformations in Abu-Qir bay, Egypt. Egyptian Journal of aquatic
research, 32 (1): p. 22-33.
Shata, A.S.A. and Hanitsch, R. (2006). Evaluation of wind energy potential and
electricity generation on the coast of Mediterranean Sea in Egypt. Renewable Energy,
31: 1183-1202.
Salem, A.I. (1995). Characteristics of surface wind speed and direction over Egypt. Solar
Energy for Sustainable Development. Romania, 4 (2): 56-63.
Hogben, N., Dacunha, N.M.C. and Olliver, G.F. (1986). Global wave statistics, Unwin
Brothers.
Wobber, F.J. (1990). Longshore current analysis and littoral drift near Burullus Outlets,
Egypt, Coastal research Institute.
Delf Hydraulics. (1992). Vulnerability assessment to accelerated sea level rise, case
study Egypt. Final Report, 1992. Prepared for Commission of European Communities
in cooperation with the Coastal Research Institute, Alexandria.
Fishawy, N.M.E. and Fanos, A.M. (1989). Prediction of sea level rise by 2100, Nile
delta coast. INQUA, Com. Quat. Shoreline Newsl., 11: 43-47.
Manohar, M. (1976). Dynamic factors affecting the Nile Delta Coast, Proceedigns
UNESCO Seminar on Nile Delta Sedimentology, Alexandria, pp. 104-129.
Fishawy, N.M.E. and Khafaga, A.A. (1991). Sea level characteristics along the Nile
Delta coast. Internat. Union Quaternay Res., INQUA, Spain, 13:29-37.
Said, M.A. and Eid, F.M. (1990). On the formation of the intermediate water masses off
the Egyptian Mediterranean coast. Archives of Hydrobiology, 120 (1): 111-122.
Lascaratos, A., Roether, W., Nittis, K. and Klein, B. (1999). Recent changes in deep
water formation and spreading in the Eastern Mediterranean Sea: a review. Progress in
Oceanography, 44: 5–36.
Tomczak, M. and Godfrey, J.S. (1994). Regional Oceanography: an Introduction.
Pergamon, New York.
Dowidar, N.M. (1988). Productivity of the South-eastern Mediterranean. In: Natural and
Man-Made Hazards. El-Sabh M.I. and Murty T.S. (Eds.), pp. 477–498.
EEAA/EIMP. (2011). Arabic annual report on water quality data from the coastal waters
of
the
Mediterranean
Sea
Year
2010.
http://
www.
eeaa.
gov.
eg/eimp/coastalwater_reports.html.
Hilmy, M.E. (1951). Beach sands of the Mediterranean coast of Egypt. Sedimentary
Petology, 21 (2): 109–120.
Anwar, Y.M., Askary, M.A.E. and Nasr, S.M. (1984). Arab’s Bay oolitic carbonate
sediments: bathymetric, granulometric and chemical studies. Neues Jahrbuch Fur¨
Geologie und Palaontologie¨ Monatshefte, 10, 594–610.
Anwar, Y.M., Askary, M.A.E. and Nasr, S.M. (1981). Petrography and origin of the
oolitic carbonate sediments of Arab’s Bay, western part of the continental shelf of Egypt.
Neues Jahrbuch Fur¨ Geologie und Palaontologie¨ Monatshefte, 65–75.
Nasr, S.M (1978). Bottom sediments in the Arabs Bay along the Mediterranean coast of
Egypt. M.Sc. Thesis, faculty of Science. Department of Geology, Alexandria University.
Frihy, O.E. (1994). Discrimination of accreted and eroded coasts using heavy mineral
compositions of the Nile Delta beach sands. Egypt. Sedimentology, 41 (5): 905-912.
Frihy, O.E., Lotfy, M.F. and Komar, P.D. (1997). Spatial variations in heavy minerals
and patterns of sediments sorting along the Nile Delta, Egypt Sediment. Sedimentary
Geology, 97(1-2): 33–41.
-270-
365.
365.
367.
368.
369.
370.
371.
372.
373.
374.
375.
376.
377.
378.
379.
380.
381.
382.
Saleh, I., El-Gamal, A. Nasr, S. and Naim, M.M. (2004). Spatial and temporal variations
of uranium and thorium series along the Egyptian Mediterranean coast. In: International
Conference on Isotopes in Environmental Studies Aquatic Forum 2004, Monte-Carlo,
Monaco 25–29 October 2004, IAEA-CN-118/113P, pp. 550–551.
El-Khatib, A.M. and El-Khier, A.A.A. (1988). Regional study of black radioactivity.
Isotopenproxis, 24: 333–336.
El-Gamal, A., Nasr, S. and Eltaher, A. (2007). Study of the spatial distribution of natural
radioactivity in the upper Egypt Nile sediments. Radiattion Measurements, 42 (3): 457–
465.
El-Gamal, A., Rashed, M. and Ghatass, Z. (2010). Radiological responses of different
types of Egyptian Mediterranean coastal sediments. Radiation Physics and Chemistry.
79: 831-838.
Dorgham, M.M. (2011). Eutrophication problem in Egypt. In: Eutrophication: Causes,
consequences and control. Ansari, A.A., Gill, S.S., Lanza, G.R. and Rast, W. (Eds).
Springer Science & Media B.V.
Sestini, G. (1990). Impacts of global climate change in the Mediterranean region:
Responses and policy options. In: Changing Climate and the Coast. Titus, J.G. (Ed.).
Volume 2: Western Africa, the Americas, the Mediterranean Basin, and the Rest of
Europe, pp. 115–125.
Barakat, A.O., Mostafa, A., Wade, T.L., Sweet, S.T. and El-Sayed, N.B. (2011).
Distribution and characteristics of PAHs in sediments from the Mediterranean coastal
environment of Egypt. Marine Pollution Bulletin, 62: 1969-1978.
El-Raey, M. (1999). Impact of Climate Change on Egypt. Environmental Software and
Services. http://www.ess.co.at/GAIA/CASES/EGY/impact.html (accessed July 14,
2009).
EEAA (Egyptian Environmental Affairs Agency). (2006). Egypt State of the
Environment, Report 2005. p.88.
DG Environment European Commission. (2006). Support to DG Environment for the
development of the Mediterranean De-pollution Initiative: Horizon 2020, Review of
ongoing and completed activities, December, 2006.
Beltagy, A.I. (2009). The impact of human activities in Alexandria and Port Said
Governorates on the coastal marine environmental along the Egyptian Mediterranean
coast., Impact of large coastal Mediterranean cities on marine ecosystems-Alexandria,
Egypt 10-12, February.
Abdrabo, M.A., Farag, M., Okab, A.B. and Khalifa, M. (2002). Economic Valuation of
an Environmental damage using hedonic pricing: case study Wadi El-Qamar,
Alexandria. Journal of the Faculty of Commerce for Scientific Research, March, 42 (1):
45-53.
CAMPS (1996). Central Agency for Public Mobilization and statistics, Population
Census Alexandria and Cairo Governorate.
CAMPS (1986). Central Agency for Public Mobilization and statistics, Population
Census Alexandria and Cairo Governorate.
CAMPS (2004). Central Agency for Public Mobilization and statistics, Population
Census Alexandria and Cairo Governorate.
Oczkowski, A.J., Nixon, S.W., Granger, S.L., El-Sayed, A.M. and Mckinney, R.A.
(2009). Anthropogenic enhancement of Egypt's Mediterranean fishery. Proceedings of
the National Academy of Sciences of the United States of America, 106 (5): 1364-1367.
Agramala, (2004). Development and climate change in Egypt: Focus on coastal
resources and the Nile. OECD.
Sestini, G. (1992). Implications of climate changes for the Nile Delta Chapter 14. In:
Climatic changes and the Mediterranean. Teftic, L., Milliman, J.D. and Sestinin, G.
-271-
383.
384.
385.
386.
387.
388.
389.
390.
391.
392.
393.
394.
395.
396.
397.
398.
399.
400.
(Eds.), Publication Edward Arnadd, a division of Heider and Stroughton, 1992, pp. 535601.
Rabenau, V. (1994). Sustainable Development Strategies for the northwest coast of
Egypt. Qasr Rural Development Project. Egyptian Environmental Affairs Agency, Cairo.
Iskander, M.M., Abo Zed A.I., El-Sayed, W.R. and Fanos, A.M. (2008). Existing marina
coastal problems, western Mediterranean coast, Egypt. Emirates Journal for Engineering
Research, 13 (3): 27-35.
El-Raey, M. (2010). Impacts and Implications of climate change for the coastal zones of
Egypt. In: Coastal zones and climate change. Michel, D. and Pandya, A. (Eds). The
Henry L. Stimson Center, Washington, DC. pp. 31-50.
Saad, M.A.H., Beltagy, A.I. and Mahmoud, W.M. (2003). Total dissolved and
particulate lead in the Western Harbor of Alexandria, a Mediterranean basin under stress.
Marine Pollution Bulletin, 47: 52–58.
Abdelmoneim, M.A. and Shata, M.A. (1993). Study of some heavy metals in Abu Qir
Bay under the effect of industrial effluents. Bulletin of the High Institute of Public
Health, xxiii (3): 675-692.
Ezzat, A.A., El-Rayis, O.A. and El-Nady, F.E. (1985). Bioaccumulation of some heavy
metals in coastal marine animals, in vicinity of Alexandria, Egypt I-Surveying. Journal
of the Faculty of Marine Science. 4 1405 H -1985: 157-166.
IDSC (1995). Information Description of Egypt. Decision support Centre, M.C. Periodic
Bulletin produced by the Cabinet Information & Decision support Center of Port Said
Governorate.
EEAA (Egyptian Environmental Affairs Agency). (2009b). National Circumstances.
Egypt Second National Communication on Climate Change.
UNDP (2009). United Nations Development Programme, Coastal research Institute and
shore protection Agency. Country: Egypt. Project document.
Mehanna, S.F. (2007). Stock Assessment and Management of the Egyptian Sole Solea
aegyptiaca Chabanaud, 1927 (Osteichthyes: Soleidae), in the Southeastern
Mediterranean, Egypt. Turkish Journal of Zoology, 31 (4): 379-388.
El-Haweet, A. and Gaber, M. (2011). Pelagic longline fishery for Albacore, Thunnus
alalunga in the Mediterranean sea, Egypt. General fisheries commission for the
Mediterranean. Marseille, France, 7-11 February 2011.
CAPMS (Central Agency for Public Mobilisation and Statistics). (1994). Statistical
Yearbook, 1994 Cairo, Egypt, # 71-12413/94.
Mohammed, W.E., El-Kaffas, S.M. and El-Raey, M. (2000). Development of a
Classification Scheme for Land Cover / Land Use Mapping in a Semi Arid Coastal
Area. Proceedings of the 2nd International Conference on Earth Observation and
Environmental Information, 11-14 November, 2000, Cairo, Egypt.
El-Kaffas, S.M., El-Aref, S.A. and El-Raey, M. (2000). Remote Sensing of the
Environmental Changes of New Burg El Arab, Egypt. Proceedings of the 2nd
International Conference on Earth Observation and Environmental Information, 11-14
November, 2000, Cairo, Egypt.
El-Bastawisy, M., Abdel-wahab, H. and Ali, R.H. (2006). Integrated socio-economic
development for accelerating regional role of Port-Said in tourism development of
Egypt. 42nd ISoCaRP Congress, p.16
Saad, M.A.H. (1990). State of the Egyptian delta lakes, with particular reference to
pollution problems. Proc. Reg. Symp. Environ. Std., (UNARC), Alexandria, 275-292.
Saad, M.A.H. (1974). Calcareous deposits of the brakish water lakes in Egypt, Rapp.
Com. Ins. Mer. Medit., 22 (6): 53.
Meininger, P.L. and Mullie, W.C. (1981). The significant of Egyptian wetlands for
wintering waterbirds. Holy Land Conservation Fund, New York.
-272-
401.
402.
403.
404.
405.
406.
407.
408.
409.
410.
411.
412.
413.
414.
415.
416.
417.
418.
419.
Alyses, A.A. and Soliman, I.A. (1993). The use of experimental tramel nets for studying
the distribution of fish population in Lake Edku (Egypt). Journal of Applied Science, 8
(11): 19-46.
GAFRD (General Authority for Fishery Resources Development). (2006). Year-Book of
fishery statistics in Egypt (1990-2005), Cairo.
EMDB (Egyptian Maritime Data Bank Alexandria). (2003). Egyptian Maritime Data
Bank Alexandria, Egypt.
APA (Alexandria Port Authority). (2005). Annual Statistical Book.
EL-Rayis, O.A. and Hinckley, D. (1999). Water and sediment quality of Mediterranean
Sea coastal area containing underwater archaeology of ancient Alexandria, Egypt.
Proceedings of the 9th international conference on environmental protection is a must,
Alexandria, Egypt, 4–6 May.
Abdallah, R.R., Zaghloul, F.A., Hassan, Y.A. and Moustafa, H.M. (1995). Some water
qulatiy characteristics of El-Dekhela Harbour, Alexandria, Egypt. Bulletin of National
Institute of Oceanography and Fisheries, 21 (1): 85-102.
EMDB (Egyptian Maritime Data Bank Alexandria). (2005). Analytical Report. Vol. No.
11.
Clarke, M., Campbell, A.C., Hameid, W.S. and Ghoneim, S. (2000). Preliminary report
on the status of marine turtle nesting populations on the Mediterranean coast of Egypt.
Biological Conservation, 94 (3): 363-371.
El-Rayis, O. (2005). Impact of man’s activities on a closed fishing-lake, lake Mariut in
Egypt, as a case study. Mitigation and Adaptation Strategies for Global Change, 10:
145-157.
Kamal, H.S. and Magdy, T.K. (2005). Socioeconomic Features. In: Lake Burullus
(Burullus Protected Area). Kamal, H.S and Magdy, T.K (Eds). Cairo: Publication of
ational Biodiversity Unit, p. 578.
Abdel-Moati, M.A.R. and EL-Sammak, A.A. (1997). Man-made Impact on the
Geochemistry of the Nile Delta Lakes. A Study of Metals Concentrations in Sediments.
Water, Air and Soil Pollution, 97: 413-429.
Soliman, A.S., Wang, X., Stanley, J.D., El-Ghawalby, N., Bondy, M.L., Ezzat, F.,
Soultan, A., Abdel-Wahab, M. Fathy, O., Ebidi, G., Abdel-Karim, N., Anh, D. K, Levin,
B., Hamilton, S.R. and. Abbruzzese, J.L. (2006). Geographical Clustering of Pancreatic
Cancers in the Northeast Nile Delta, Region of Egypt. Archives of Environmental
Contamination and Toxicology, 51, 142–148.
Siegel, F.R., Slaboda, M.L. and Stanley, D.J. (1994). Metal pollution loading, Manzala
Lagoon, Nile Delta, Egypt: Implications for aquaculture. Environmental Geology, 23:
89-98.
Said, M.A. and Rajkovic, B. (1996). A study of water circulation along the Egyptian
Mediterranean coast using a three dimensional numerical model. International Journal of
Environmental Studies, 50 (3): 223-235.
SMASRT (2005). Sustainable management of scarce resources in the coastal zone.
D06.2: Case Study Report (EGYPT) Abu Qir Bay, Egypt. May, 2005.
Loring, D.H. and Rantala, R.T.T. (1992). Manual for the geochemical analyses of marine
sediments and suspended particulate matter. Earth Science Reviews, 32: 235-283.
Yan, C., Li, O., Zhang, X. and Li, G. (2010). Mobility and ecological risk assessment of
heavy metals in surface sediments of Xiamen Bay and its adjacent areas, China.
Environmental Earth Science, 60 (7): 1469-1479.
Folk, R.L. and W.C. Ward. (1957). Brazos River bar: a study in the significance of grain
size parameters. Journal of Sedimentary Petrology, 27 (1): 3-26.
Black., C.A. (1965). Methods of soil analysis, part 2, “Chemical and Microbiological
properties, American society of Agronomy Inc., Madison, Wisc.
-273-
420.
421.
422.
423.
424.
425.
426.
427.
428.
429.
430.
431.
432.
433.
434.
435.
436.
437.
438.
Baruah, T.C. and Barthakur, H.P. (1997). A Textbook of Soil Analysis, Vikas Publishing
House, Pvt. Ltd.
Oregioni, B. and Aston, S.R. (1984). The determination of selected trace metals in
marine sediments by flame atomic absorption spectrophotometry. IAEA Monaco
Laboratory Internal Report.
Tanner, P.A. and Leong, L.S. (1999). Metals in a marine sediment core from near MA
WAN, Hong Kong. Water Air Soil Pollution, 121: 309-325.
Kersten, M. and Fostner, U. (1985). Trace metal partitioning in suspended matter with
special reference to pollution in the southeastern North Sea. In: Transport of carbon and
minerals in major world rivers. Degans, E.T., Kempe, S. and Herrera, R. (Eds), Pt. 3
Mitt. Geol. Palaont. Inst. Univ. Hamburg, 58, 631-645.
Allen, H.S., Fu, G. and Deng, B. (1993). Analysis of acid volatile sulfide and
simultaneously extracted metals (SEM) for estimation of potential toxicity in aquatic
sediments. Environmental Toxicology and Chemistry, 12 (8): 1441-1453.
Nessim, R.B., Mamdouh, S., Masoud, A. and Maximous, N. (2005). Water
characteristics of Alexandria hot spots. Egyptian Journal of Aquatic Research, 31: 2537.
Smith, S.R. (1994). Effect of soil pH on availability to crops of metals in sewage sludgetreated soil, in nickel, copper and zinc uptake and toxicity to ryegrass.
Environmental Pollution, 85: 321-327.
Forstner, K. and Witman, G.T.W. (1983). Metal pollution in the aquatic environment.
Springer-Verlage, Berlin, Heindelberg, New York, Tokyo.
IP, C., Li, X.D., Zhang, G., Wai, O.W.H. and Li, Y.S. (2007). Trace metal distribution in
sediments of the pearl river estuary and the surrounding coastal area, south China.
Environmental Pollution, 147 (2): 311-323.
Shepard, F.P. and Moore, D.G. (1954). Sedimentary environments differentiated by
coarse fraction studies. Bulletin of American Association of Petrololeum Geology, 38
(8): 1792-1802.
Abdulla, M.I., Dunlop, H.M. and Gardner, D. (1973). Chemical and hydrographic
observations in the Bristol channel during April and June 1971. Journal of Marine
Biological Association of the United Kingdom, 36: 509-517.
Millero, F.J. (1996). Chemical oceanography, second Edition, Marine Science Series,
CRC. Press. Inc., New York, USA.
Masoud, M.S., Elewa, A.A., Ali, A.E. and Mohamed, E.A. (2005). Distribution of some
metal concentrations in water and sediments of lake Edku, Egypt. Bulletin of Chemists
and Technologists of Macedonia, 24 (1): 21-34.
Hayward, J. (1969). Studies on the growth of Rhaeodactylum. V. The relationship to
iron, manganese and zinc. Journal of Marine Biological Association of the United
Kingdom, 49: 439-446.
Gratten, J.P., Huxley, S.I. and Pyatt, F.B. (2003). Modern Bedouin exposure to copper
contamination: An imperial legacy. Ecotoxicology and Environmental Safety, 55 (1):
108-115.
WHO. (1998). Copper: Environmental Health Criteria Monograph 200, International
program on chemical safety (IPCS).
Nriagu, J.O. (1980). Global cycle and properties of nickel. In: Nickel in the
Environment. Nriagu, J.O. (Ed), Wiley, New York, pp. 1–26.
EL-Nemr, A., Khaled, A. and Elsikaily, A. (2006). Distribution and statistical analysis of
leachable and Total Heavy metals in the sediments of the Suez gulf. Environmental
Monitoring and Assessment, 118:89–112.
Beltagy, A.I. (1973). The geochemistry of some recent marine sediments from the Gulf
of St.Lowrence, Ph.D. Thesis, Mc.Grill University, p.203.
-274-
439.
440.
441.
442.
443.
444.
445.
446.
447.
448.
449.
450.
451.
452.
453.
454.
455.
456.
Huzinger, O. (1980). The handbook of environmental chemistry. The natural
environment and the biogeochemical cycles. Springer-Verlage, New York, I, p. 258.
Davidson, C.M., Thomas, R.P., McVey, S.E., Perala, R., Littlejohn, D. and Ure, A.M.
(1994). Evaluation of a sequential extraction procedure for the speciation of heavy
metals in sediments. Analytica Chimica Acta, 291: 277-286.
Bernhard, T. and Neff, J. (2001). Metals bioavailability in the Navy's tiered ecological
risk assessment process. Naval Facilities engineering command. Issue papers, 1-15,
(http://web.ead.anl.gov/ecorisk/issue).
Glosinska, G., Sobczynski, T., Boszke, L., Bierla, K. and Siepak, J. (2005). Fractionation
of some heavy metals in bottom sediments from the middle Odra river
(Germany/Poland).
Izquierdo, C., Usero, J. and Gracia, I. (1997). Speciation of heavy metals in sediments
from salt marshed on the southern Atlantic coast of Spain. Marine Pollution, 34: 123128.
USEPA (2004). The incidence and severity of sediment contamination in surface waters
of the United States: National Sediment Quality Survey, EPA 823-r-04-007, 2nd ed.
Washington, DC: Environmental Protection Agency, Office of Water.
Edenborn, H.M. (2005). Rapid detection of bioavailable heavy metals in sediment pore
waters using acid volatile sulfide gel probes. Environmental Geology, 47: 660-669.
Usero, J., Morillo, J. and El-Bakouri, H. (2008). A general integrated ecotoxicological
method for marine sediment quality assessment: Application to sediments from littoral
ecosystems on Southern Spain's Atlantic Coast. Marine Pollution Bulletin, 56: 20272036.
Silva, S., Re, A., Pestana, P., Rodrigues, A. and Quintino, V. (2004). Sediment
disturbance off the Tagus Estuary, Western Portugal: chronic contamination, sewage
outfall operation and runoff events. Marine Pollution Bulletin, 49: 154-162.
Ridgway, J. and Shimmield, G. (2002). Estuaries as repositories of historical
contamination and their impact on shelf seas. Estuarine, Coastal and Shelf Science, 55:
903-928.
Stigliani, W.M. (1991). Chemical time bomb; definition, concepts, and examples,
International institute for applied system analysis, Laxenburg, Austria, 23, 16.
Szefer, P., Szefer, K., Glasby, G.P., Pempkowiak, J. and Kaliszan, R. (1996). Heavy
metal pollution in surfacial sediments from the southern Baltic Sea of Poland, Journal of
Environmental Science and Health, 31 (10): 2723-2754.
Zhang, X.P., Xueming, W. and Yang, M. (2002). The background concentrations of 13
soil trace elements and their relationships to parent materials and vegetation in Xizang
(Tibet), China. Journal of Asian Earth Sciences, 21: 167-174.
Ewais, T.A., Grant, A. and Fattah, A.T.A. (2000). The role of surface coatings on
sediments in sediment: water partitioning of trace elements and radionuclides. Journal of
Environmental Radioactivity, 49: 55-64.
Stevenson, A.G. (2001). Metal concentrations in marine sediments around Scotlans: a
baseline for environmental studies. Continental Shelf Research, 21: 879-897.
Rosales, H.L., Cundyb, A.B. and Manjarrez, J.L. (2003). Heavy metals in sediment cores
from a tropical estuary affected by anthropogenic discharges: Coatzacoalcos estuary,
Mexico. Estuarine, Coastal and Shelf Science, 58: 117-126.
Fishawi, N.M.E., Lotfy, M.F., Sestini, G. and Shawki, A. (1976). Grain size of the Nile
Delta beach sands. Proc. UNESCO Seminar on the Nile Delta Sedimentology, 2: 79-94.
Korfali, S.I. and Davies, B.E. (2004). Speciation of metals in sediment and water in a
river underline by limestone: role of carbonate species for purification capacity of rivers.
Advances in Environmental Research, 8: 599-612.
-275-
457.
458.
459.
460.
461.
462.
463.
464.
465.
466.
467.
468.
469.
470.
471.
472.
473.
474.
475.
476.
Selim, A.A. (1974). Origin and lithofication of the Pleiostocene carbonates of the Salum
area, western coastal plain of Egypt. Journal of Sedimentary Petrology, 44 (1): 70-78.
Ottosen, L.M. and Villumsen, A. (2006). High Cu and Cd pollution in sediments from
Sisimiut, Greenland. Adsorption to organic matter and fine particles. Environmental
Chemistry Letters, 4 (4): 195-199.
Grathwohl, P. (1990). Influence of organic matter from soils and sediments from various
of the sorption of some chlorinated aliphatic hydrocarbons. Environmental Science &
Technology, 24 (11): 1687-1693.
Draz, S.C. (1983). The Texture and Chemistry of the Nile Sediments in the Rosetta
Branch, M.Sc. Thesis, Faculty of Science, Alexandria University, Egypt.
Peltier, E.F., Webb, S.M. and Gaillard, J.F. (2003). Zinc and lead sequestration in an
impacted wetland system. Advances in Environmental Research, 8: 103-122.
Seralathan, P., Meenakshikutty, N.R., Asaref, K.V. and Padmalal, D. (1993). Sediment
and organic carbon distribution in the Cochin harbor area. Indian Journal of Marine
Sciences, 22: 252-255.
Dias de Alba, M., Galindo-Riano, M.D., Casanuueva-Marenco, M.J., Garcia-Vargas, M.
and Kosore, C.M. (2011). Assessment of the metal pollution, potential toxicity and
speciation of sediment from Algeciras Bay (South of Spain) using chemometric tools.
Journal of Hazardous Materials, 190 (3): 177-187.
Canuel, E.A. and Martens, C.S. (1993). Seasonal variations in the sources and alteration
of organic matter associated with recently-deposited sediments. Organic Geochemistry,
20: 563-577.
Revelle, R. and Shepard, F.P. (1938). Sediments off California coast In: Recent Marine
Sediments. Amer. Assoc. Petrol. Geo. Tulsa, Oklahoma, 245-282.
Santos, I.R., Silva-Filho, E.V., Schaefer, E.G.R., Albuquerque-Filho, M.R. and Campos,
L.S. (2005). Heavy metal contamination in coastal sediments and soils near the Brazilian
Antarctic Station, King George Island. Marine Pollution Bulletin, 50: 185-194.
Jayaprakash, M., Jonathan, M.P., Srinivasalu, S., Muthuraj, S., Ram-Mohan, V. and
Rajeshwara-Rao, N. (2008). Acid leachable trace metals in sediments from an
industrialized region (Ennore Creek) of Chennai city, SE coast of India: an approach
towards regular monitoring. Estuarine, Coastal and Shelf Science, 76: 692-703.
Burton, E., Phillips, I. and Hawker, D. (2004). Trace metals and nutrients in bottom
sediments of the south port Broadwater, Australia. Mar. Pollut. Bull., 48: 378-384.
Turekian, K.K. and Wedepohl, K.H. (1961). Distribution of the elements in some major
units of the earth's crust. Geological Society of America, Bulletin, 72 (2): 175-192.
Taylor, S.R. (1964). Abundance of chemical elements in the continental crust, a new
table. Geochimica et Cosmochimica Acta, 28: 1278-1785.
Bowen, H.J.M. (1979). Environmental chemistry of the elements. Academic Press,
London.
Taylor, S.R. and Mclennan, S.M. (1995). The geochemical evaluation of the continental
crust. Review of Geophysics, 33 (2): 241-265.
Wedepohl, K.H. (1995). The composition of the continental crust. Geochim.
Cosmocheim Acta, 59: 1217-1232.
UNAP/MED POL (2000). Guidelines for the management of dredged material. MAP
technical reports series No. 129, UNAP, Athens 2000.
Santamaria-Fernandez, R., Cave, M.R. and Hill, S.J. (2005). Trace metal distribution in
the Arosa estuary (N.W. Spain): The application of a recently developed sequential
extraction procedure for metal partitioning. Analytica Chimica Acta, 557 (1-2): 344-352.
Sevete, P., Milacic, R. and Pihlar, B. (2001). Partitioning of Zn, Pb and Cd in river
sediments from a lead and zinc mining area using the BCR three step sequential
extraction procedure. Journal of Environmental Monitoring, 3: 586-590.
-276-
477.
478.
479.
480.
481.
482.
483.
484.
485.
486.
487.
488.
489.
490.
491.
492.
Buykx, S.E.J., Bleijenberg, M., Van Den Hoop, M.A.G.T. and Gustavloch, J.P. (2000).
The effect of oxidation and acidification on the specification of heavy metals in sulfide
rich freshwater sediments using a sequential extraction procedure. Journal of
Environmental Monitoring, 2: 23-27.
Sulivan, P. and Taylor, K. (2003). Sediment and porewater geochemistry in a metal
contaminated estuary, Dulas Bay, Anglesey. Environmental Geochemistry and Health,
25: 115-122.
Legoburu, I. and Canton, L. (1991). Heavy metal concentration in sediments from
Pasajes Harbour. Spain. Marine Pollution Bulletin, 22: 207-209.
Gargouri, D., Azri, C., Serbaji, M.M., Jedoui, Y. and Montacer, M. (2010). Heavy metal
concentrations in the surface marine sediments of Sfax Coast, Tunisia. Environmental
Monitoring Aessessment, 175(1-4):519-530.
Violintzis, C., Arditsoglou, A. and Voutsa, D. (2009). Elemental composition of
suspended particulate matter and sediments in the coastal environment of Thermaikos
Bay, Greece: Delineating the impact of inland waters and wastewaters. Journal of
Hazardous Materials, 166 (2-3): 1250-1260.
Sarkar, S.K. and Bhattacharya, B. (2010). Bioavailability of trace metals in sediment
cores from sunderban mangrove wetland, India: An urgent need for bioremediation. 19
World Congress of Soil Science, Soil Solutions for a Changing World, 1– 6 August,
Brisbane, Australia. Published on DVD. pp. 31-33.
Francois, R. (1988). A study on the regulation of the concentrations of some trace metals
(Rb, Sr, Zn, Pb, Cu, V, Cr, Ni, Mn and Mo) in Saanich inlet sediments, British
Columbia, Canada. Marine Geology, 83 (1-4): 285-308.
Loska, K., Wiechula, D. and Korus, I. (2004). Metal contamination of farming soil
affected by industry. Environment International, 30 (2): 159-165.
Aloupi, M. and Angelidis, M.O. (2001). Normalization to lithium for the assessment of
metal contamination in coastal sediments cores from the Aegean Sea, Greece. Marine
Environmental Research, 52: 1-12.
Femex, F.E., Migon, C. and Chisholm, J.R.M. (2001). Entrapment of pollutants in
Mediterranean sediments and biogeochemical indicators of their impact. Hydrobiologia,
450 (1-3): 31-46.
Caredda, A.M., Cristini, A., Ferrara, C., Lobina, M.F. and Baroli, M. (1999).
Distribution of heavy metals in the Piscinas beach sediments (SW Sardina, Italy).
Environmental Geology, 38(2): 91-100.
Sabhi, Y., Chaoui, M., El-Quessar, S., Bakkas, S. and Ramdani, M. (2000).
Identification of the northern Moroccan hot spots and contamination baseline of coastal
sediments by heavy metals. Bulletin de l'Institute Scientifique, Rabat, (22):59-69.
Abdel-Moati, M.A.R. (2001). Long-term impact of landbased activities on the coastal
waters of Alexandria, Egypt. Proceedings of the Fifth International Conference on the
Mediterranean Coastal Environment, MEDCOAST 01, E. Ozhan (Editor), Tunisia, Vol.
23-27, pp: 463-474.
Saad, M.A.H., Abdel Moati, A.A. and Younis, W.A. (2001). Impact of landbased
sources of pollution on the levels of lead and cadmium in Mex-Bay, a Mediterranean
Basin under stress. Fresenius Environmental Bulletin, 10 (6): 561-565.
Sadiq, R., Husain, T., Bose, N. and Veitch, B. (2003). Distribution of heavy metals in
sediment pore water due to offshore discharges; an ecological risk assessment.
Environmental Modeling & Software, 18 (5): 451-461.
Feldstein, T., Kashman, Y., Abelson, A., Fishelson, L., Mokady, O. and Bresler, V.
(2003). Marine molluscs in environmental monitoring. III. Trace metals and organic
pollutants in animal tissue and sediments. Helgoland Marine Research, 57 (3-4): 212219.
-277-
493.
494.
495.
496.
497.
498.
499.
500.
501.
502.
503.
504.
505.
506.
507.
508.
509.
510.
PIO (Pesticide Information Office). (2005). PI-66, Florida cooperative Extension Office,
Institute of food and Agriculture Sciences, University of Florida. September 2005.
Benamar, M.A., Toumert, I., Tobbeche, S., Tchantchane, A. and Chalabi, A. (1999).
Assessment of the state of pollution by heavy metals in the surfacial sediments Algiers
Bay. Applied Radiation and Isotopes, 50: 975-980.
Mason, B. and Moore, C.B. (1991). Principles of geochemistry. New Delhi: Basic
Books.
UNEP (1978). Preliminary study on the state of Intergovernmental Review Meeting of
Mediterranean Coastal States on the Mediterranean Action Plan. UNEP/IG.11/INF, 4,
209p.
Donazzolo, R., Merlin, H.O., Vitturi, M.L. and Pavoni, B. (1984). Heavy metal content
and lithological properties of recent sediments in the northern Adriatic. Marine Pollution
Bulletin, 15: 93-101.
Frignani, M. and Giordani, P. (1983). Metalli pesanti in sedimenti antichi e recenti dei
mari Italiani. Atti del 5 Congresso della Associazione Italiana di Oceanologia e
Limnologia. Stressa, 19-22 May 1982, pp. 525-534.
Jdid, E.A., Blazy, P., Kamoun, S., Guedria, A., Marouf, B. and Kitane, S. (1999).
Environmental impact of mining activity on the pollution of the Medjerda river, north
west Tunisia. Bulletin of Engineering Geology and the Environment, 57: 273-280.
El-Sayed, M., Wakeel, S.K.E. and Rifaat, A. (1988). Factor analysis of sediments in the
Alexandria western Harbour, Egypt. Oceanology Acta, 11: 1-11.
Chon, H.S., Ohandja, D.G. and Voulvoulis, N. (2010). Implementation of U.E. water
framework directive: source assessment of metallic substances at catchment levels.
Journal of Environmental Monitoring and Assessment, 12 (1): 36-47.
De-Vallejuelo, S.F., Arana, G., Diego, A.D. and Mandariaga, J.M. (2010). Risk
assessment of trace elements in sediments: The case of the estuary of the NerbioiIbaizabal River (Basque Country). Journal of Hazardous Materials, 181: 565-573.
MacDonald, D.D., Ingersoll, C.G. and Berger, T.A. (2000). Development and evaluation
of consensus-based sediment quality guidelines for freshwater ecosystem. Archives of
Environmental Contamination Toxicology, 39 (1): 20-31.
Perin, G., Fabris, R., Manente, S., Wagener, A.R., Hamacher, C. and Scotto, S. (1997).
A five years study on the heavy metal pollution of Guanabara bay sediments (Rio De
Janeiro, Brazil) and evaluation of the metal bioavailability by means of geochemical
speciation. Water Research, 31: 3017-3028.
NYSDEC (New York State Department of Environmental Conservation). (1999).
Technical guidance for screening contaminated sediments.
Occhipinti-Ambrogi, A. and Forni, G. (2004). Biotic indices. In: Mediterranean marine
benthos: a manual of methods for its sampling and study. Gambi, M.C. and Dappiano,
M. (Eds). Biological Marine Mediterranea, 11(Suppl. 1): 545-572.
Graney, J.R. and Eriksen, T.M. (2004). Metals in pond sediments as archives of
anthropogenic activities: a study in response to health concerns. Applied Geochemistry,
19 (7): 117-118.
Anders, B., Reischmann, T., Poller, U. and Kostopoulos, D. (2005). Age and origin of
granitic rocks of the eastern Vardar Zone, Greece: new constraints on the evolution of
the Internal Hellenides. Journal of Geological Society, 162: 857–870.
Chapman, P.M. and Mann, G.S. (1999). Sediment quality values (SQVs) and ecological
risk assessment. Marine Pollution Bulletin, 38 (5): 339–344.
MacDonald, D.D., Carr, R.S., Calder, F.D., Long, E.R. and Ingersoll, C.G. (1996).
Development and evaluation of sediment quality guidelines for Florida coastal waters.
Ecotoxicology, 5: 253-278.
-278-
511.
512.
513.
514.
515.
516.
517.
518.
519.
520.
521.
522.
523.
524.
525.
526.
527.
528.
529.
Manheim, F.T. and Commeau, J.A. (1981). Chemical analyses (trace metals), in Poppe,
J.L., Data File Atlantic Magin Coring Project (AMCOR) of the U.S. Geological Survey:
U.S. Geological Survey Open-File Report 81-239, 1981, 244-51.
USEPA. (1996). The National Sediment Quality Survey: A report to Congress on the
extent and severity of sediment contamination in surface waters of the United States:
U.S. Environmental Protection Agency, Office of Science and Technology, Washington
D.C., Report No. EPA-823-D-96-002, Chapter-Paginated, 1996, 9 chapters (200 p.).
SEPA (State Environmental Protection Administration of China). (2002). Marine
Sediment Quality (GB 18668-2002). Standards Press of China, Beijing.
Lau, M.M., Roothman, R.C. and Bradley, G.C. (1993). A Strategy for the management
of contaminated dredged sediment in Hong Kong. Journal of Environmental
Management, 38: 99-114.
EPD (1992). Marine water quality in Hong Kong. Environment Protectection
Department, Hong Kong Government Printer, Hong Kong.
Muller, G. (1969). Index of geoaccumulation in sediments of the Rhine River,
Geojournal, 2 (3): 108-118.
Hakanson, L. (1980). An Ecological Risk Index for Aquatic Pollution control. a
sediment-Ecological Approach. Water Research, 14: 975-1001.
Abrahim, G.M.S. and Parker, R.J. (2008). Assessment of heavy metal enrichment factors
and the degree of contamination in marine sediments from Tamaki Estuary, Auckland,
New Zealand. Environmental Monitoring Aessessment, 136: 227-238.
Valdes, J., Vargas, G., Sifeddine, A., Ortlieb, L. and Guinez, M. (2005). Distribution and
enrichment evaluation of heavy metals in Mejillones Bay (23 S), Northern Chile
geochemical and statistical approach. Marine Pollution Bulletin, 50: 1558-1568.
Dickinson, W.W., G.B. Unbar, and H. McLeod, 1996. Heavy metal history from cores in
Wellington Harbour, New Zealand, Environmental Geology, 27: 59-69.
Deely, J.M. and Fergusson, J.E. (1994). Heavy metal and organic matter concentration
and distributions in dated sediments of a small estuary adjacent to a small urban area,
The Science of the Total Environment, 153: 97-111.
Nizncheski, L.F., Windom, H.L. and Smith, R. (1994). Distribution of particulate trace
metal in patos Lagoon Estuary (Brazil). Marine Pollution Bulletin, 28: 96-102.
Rath, P., Panda, U.C., Bhatta, D. and Sahoo, B.N. (2005). Environmental Quantification
of Heavy Metals in the Sediments of the Brahami and Nandira Rivers, Orissa. Journal of
Geological Society of India, 65 (4): 487-492.
Chakravarty, M. and Patgiri, A.D. (2009). Metal Pollution Assessment in Sediments of
the Dikrong River, N.E. India. Journal of Human Ecology, 27 (1): 63-67.
Zhao, F., Cong, Z., Sun, H. and Ren, D. (2007). The geochemistry of rare earth elements
(REE) in acid mine drainage from the Sitai coal mine, Shanxi Province, North China,
International Journal of Coal Geology, 70 (1-3): 184-192.
Duzgoren-Aydin, N.S. (2007). Sources and characteristics of lead pollution in the urban
environment of Guangzhou. Science of the Total Environment, 385: 182-195.
Sezgin, N., Ozcan, J.K., Demir, G., Nemlioglu, S. and Bayat, C. (2003). Determination
of heavy metal concentrations in street dusts in Istanbul E-5 highway. Environment
International, 29 (7): 979-985.
Chen, C.W., Kao, C.M., Chen, C.F. and Dong, C.D. (2007). Distribution and
accumulation of heavy metals in sediments of Kaoshiung Harbor, Taiwan. Chemosphere,
66(8): 1431-1440.
Karbassi, A., Saeedi, M. and Amirnejad, R. (2008). Historical changes of heavy metals
content and sequential extraction in a sediment core from the Gorgan Bay, Southeastern
Caspian Sea. Indian Journal of Marine Science, 37 (3): 267-272.
-279-
530.
531.
532.
533.
534.
535.
536.
537.
538.
539.
540.
541.
542.
543.
544.
545.
546.
Zhang, W.G., Feng, H., Chang, J, Qu, J.G. and Yu, L.Z. (2009). Heavy metal
contamination in surface sediments of Yangtze River intertidal zone: an assessment from
different indexes. Environmental Pollution, 157: 1533–1543.
Christophoridis, C., Dedepsidis, D. and Fytianos, K. (2009). Occurrence and distribution
of selected heavy metals in the surface sediments of Thermaikos Gulf, N. Greece.
Assessment using pollution indicators. Journal of Hazardous Materials, 168 (2-3): 10821091.
Abrahim, G.M.S. (2005). Holocene sediments of Tamaki Fstuary: Characterization and
impact of recent human activity on an urban esturary in Auckland, New Zealand. Ph.D
Thesis, University of Auckland, Auckland, New Zealand, p. 361.
Tomlinson, D.C., Wilson, J.G., Harris, C.R. and Jeffrey, D.W. (1980). Problems in the
assessment of heavy-metal levels in estuaries and the formation of a pollution index.
Helgoland Meeresunter Suchungen, 33 (1-4): 566-575.
Harikumar, P.S. and Jish, T.S. (2010). Distribution pattern of trace meal pollutants in the
sediments of an urban wetland in the southwest coast of India. International Journal of
Engineering Science and Technology, 2 (5): 840-850.
Raj, S.M. and Jayaprakash, M. (2008). Distribution and enrichment of trace metals in
marine sediments of Bay of Bengal, off Ennore, south-east coast of India. Environmental
Geology, 56 (1): 207-217.
Ayyamperumal, T., Jonathan, M.P., Srinivasalu, S., Armstrong-Altrin, J. and RamMohan, V. (2006). Assessment of acid leachable trace metals in sediment cores from
River Uppanar, Cuddalor, southeast coast of India. Environmental pollution, 143 (1): 3445.
Janki-Raman, D., Jonathan, M.P., Srinivasalu, S., Altrin, J.S.A., Mohan, S.P. and RamMoham, V. (2007). Trace metal enrichments in core sediments in Muthupet mangroves,
SE coast of India: Application of acid leachable technique. Environmental Pollution,
145: 245-257.
Sahuquillo, A., Rigol, A. and Rauret, G. (2003). Overview of the use of
leaching/extraction tests for risk assessment of trace metals in contaminated soils and
sediments. Trends in Analytical Chemistry, 22: 152–159.
Nriagu, J.O. (1990). Global metals pollution: Poisoning the biosphere. Environment, 32:
7-33.
Lacerda, L.D., Fernandez, M.A., Calazans, C.F. and Tanizaki, K.F. (1992).
Bioavailability of heavy metals in sediments of two coastal lagoons in Rio de Janeiro,
Brazil. Hydrobiology, 228: 65-70.
Vangronveld, J. and Cunnihgham, S.D. (1998). Metal contaminated soils: in-situ
inactivation and Phytoremediation. Springer Verlag, Berlin.
Hall, G.E.M., Vaive. J.E., Beer. R. and Hoashi, M. (1996). Selective leaches revisited,
with emphasis on the amorphous Fe oxyhydroxide phase extraction. Journal of
Geochemical Exploration, 56 (1): 59-78.
Carral, E., Villares, R., Puente, X. and Carballeira, A. (1994). Characterization of
sediment metal pollution in Galician Estuaries (NW Spain), in Proceedings of the 6th
International Conference of Environmental Contamination, Delphi, Greece, October,
1994, pp. 281-283.
Dassenakis, M., Degaita, A. and Scoullos, M. (1995). Trace metals in sediments of a
Mediterranean esturay affected by human activites: Acheloos river estuary, Greece.
Science of the Total Environment, 168: 19-13.
El-Nemr, A. (2003). Assessment of heavy metal pollution in surface muddy sediments of
Lake Burullus, southern Mediterranean, Egypt. Egypt J. Aquat. Biol. Fish, 7: 67-90.
Callahan, M.A., Slimak, M.W., Gabel, N.W., May, I.P., Fowler, C.F., Freed, J.R.,
Jennings, R.L., Durfee, R.L., Whitmore, B., Maestri, B. and Mabey, W.R. (1979).
-280-
547.
548.
549.
550.
551.
552.
553.
554.
555.
556.
557.
558.
559.
560.
561.
562.
563.
Nickel, in water-related Environmental fate of 129 Priority pollutant, I: Introduction and
Technical Background, Metals and Inorganics, Pesticides and PCBs, US. Environmental
Protection Agency Report EPA-440/4-79-029a, U.S. EPA, Office of water and waste
Management, Washington, DC.
CEPA. (1994). Canadian Environmental Protection Act, Priority substances list
assessment report: Nickel and its compounds.
Shacklette, H.T. (1970). A.U.S. Geological Survey study of elements in soils and other
surficial materials in the United States. In Trace Subs. Environ. Health. Univ. Missouri.,
4: 35.
Engel, R.W., Hardison, W.A., Miller, R.F., Pierce, N.O. and Huber, J.T. (1964). Effect
of copper intake on concentration in body tissue and on growth, reproduction and
production in dairy cattle. Jouranl of Animal Science, 23: 1160.
PIO. (2005). Pesticide Information Office, PI-66, florida cooperative Extension Office,
Institute of Food and Agricultural Sciences, University of Florida. September 2005.
Naidu, S. and Morrison, R.J. (1994). Contamination of Suva Harbour, Fiji'. Marine
Pollution Bulletin, 29: 126-130.
Bryan, G.W. (1971). Heavy metals in the marine environment. In: Furness R.W., and
Rainbow P.S., (Eds.), CRC Press Inc., Florida, USA, p. 256.
Tuzen, M., Sari, H. and Soylak, M. (2004). Microwave and wet digestion procedures for
atomic absorption spectrophotometric determination of trace metals contents of sediment
samples. Analytical Letters, 37 (9): 1925-1936.
Quevauviller, P. (1998). Operationally defined extraction procedures for soil and
sediment analysis. I. Standardization, Trac-Trend. Analytica Chemica Acta, 17: 289298.
Tack, F.M. and Verloo, M.G. (1995). Chemical speciation and fractionation in soil and
sediment heavy metal analysis: a review. International Journal of Environmental
Analytical Chemistry, 59 (2-4): 225-238.
Sutherland, R.A., Tack, F.M.G., Tolosa, C.A. and Verloo, M.G. (2000). Operationally
defined metal fractions in road deposited sediment, Honolulu, Hawaii. Journal of
Environmental Quality, 29 (5): 1431-1439.
Caplat, C., Texier, H., Barillier, D. and Lelievre, C. (2005). Heavy metals mobility in
harbour contaminated sediments: The case of Porten-Bessin. Marine Pollution Bulletin,
50 (1): 504-511.
Weisz, M., Polyak, K. and Hlavay, J. (2000). Fractionation of elements in sediment
samples collected in rivers and harbors at Lake Balaton and its catchment area.
Microchemical Journal, 67: 207-217.
Ramos, L., Gonzalez, M.J. and Hernandez, L.M. (1999). Sequential extraction of copper,
lead, cadmium, and zinc in sediments from Ebro river (Spain): relationship with levels
detected in earthworms. Bulletin of Environmental Contamination and Toxicology, 62
(3): 301-308.
Canfield, D.E. (1989). Reactive iron in marine sediments, Geochimca et Cosmochimica
Acta, 53: 619-632.
Jones, B. and Turki, A. (1997). Distribution and speciation of heavy metals in surficial
sediments from Tees Estuary, northeast England. Marine Pollution Bulletin, 34 (10):
768–779.
Martin, R., Sanchez, D.M. and Gutierrez, A.M. (1998). Sequential extraction of U, Th,
Ce, La and some heavy metals in sediments from Ortigas River, Spain. Talanta, 46:
1115-1121.
Yuan, C., Shi, J., He, N., Liu, J., Liang, L. and Jiang, G. (2004). Speciation of heavy
metals in marine sediments from the East China Sea by ICP-MS with sequential
extraction. Environment International, 30: 769-783.
-281-
564.
565.
566.
567.
568.
569.
570.
571.
572.
573.
574.
575.
576.
577.
578.
579.
580.
581.
Velimirovic, M.B., Prica, M.D., Dalmacija, B.D., Roncevic, S.D., Dalmacija, M.B.,
Becelic, M.D. and Trickovic, J.S. (2011). Characterization, Availability, and Risk
Assessment of the Metals in Sediment after aging. Water Air Soil Pollution, 214 (1-4)
219-229
Kartal, S., Aydin, Z. and Tokalioglu, S. (2006). Fractionation of metals in street
sediments samples by suing the BCR sequential extraction procedure and multivariance
statistical elucidation of the data. Journal of Hazardous Materials, 132: 80-89.
Noah, N.M. and Oomori, T. (2006). Evaluation of heavy metal pollution on the coastal
marine environments of Okinaws Island, Japan. Bulletin of Faculty of Science.
University of Ryukyus, 81: 93-104.
Zhang, J., Huang, W.W. and Martin, J.M. (1988). Trace metal distribution in Huanghe
(Yellow River) estuarine sediments. Estuarine Coastal Shelf Science, 26 (5): 499-516.
Ramos, L., Hernandez, L.M. and Gonzalez, M.J. (1994). Sequential fractionation of
copper, lead, cadmium and zinc in the soil from or near Donana National Park. Journal
of Environmental Quality, 23 (1): 50-57.
Modak, D.P., Singh, K.P., Chandra, H. and Ray, P.K. (1992). Mobile and bound forms
of the lower Ganges. Water Research, 26 (11): 1541-1548.
Bordas, F. and Bourg, A. (2001). Effects of solid/liquid ratio on the remobilization of
Cu, Pb, Cd, and Zn from polluted river sediment. Water Air Soil Pollution, 128: 391400.
Wong, C.S.C., Wu, S.C., Duzgoren-Aydin, N.S., Aydin, A. and Wong, M.H. (2007).
Trace metals contamination of sediments in an e-waste processing village in China.
Environmental Pollution, 145: 434-442.
Perin, G., Craboledda, L., Lucchese, M., Cirillo, R., Dotta, L., Zanette, M.L. and Orio,
A.A. (1985). In: Heavy metal in the environment: Lekkas. T. D. (Ed.), CEP Consultants:
Edinburg, 1985. vol. 2, pp. 454-456.
Liu, J., Zhang, B., Cao, J., Cao, Z. and Domagalski, J. (2009). Ecological risk of heavy
metals in sediments of the Luan River source water. Ecotoxicology, 18 (6): 748-758.
Chapman, P.M., Wang, F.Y., Germano, J.D. and Batley, G. (2002). Pore water testing
and analysis: the good, the bad, and the ugly. Marine Pollution Bulletin, 44: 359–366.
Feng, T., Li, X. and Zhang, G. (2005). Acid volatile sulfide and simultaneously extracted
metals in the sediment cores of the Pearl River Estuary, south China. Ecotoxicology and
Environmental Safety, 61: 420-431.
Ankley, G.T., Mattson, V.R., Leonard, E.N., West, C.W. and Bennett, J.L. (1993).
Predicting the acute toxicity of copper in freshwater sediments: evaluation of the role of
acid volatile sulfide. Environmental Toxicology and Chemistry, 12 (2): 315-320.
Casas, A.M. and Crecelius, E.A. (1994). Relationship between acid volatile sulfide and
the toxicity of zinc, lead and copper in marine sediments. Environmental Toxicology and
Chemistry, 13 (3): 529-536.
Carlson, A.R., Phipps, G.L., Mattson, V.R., Kosian, P.A. and Cotter, A.M. (1991). The
role of acid volatile sulfide in determining cadmium bioavailability and toxicity in
freshwater sediments. Environmental Toxicology and Chemistry, 10: 1309-1319.
Van den Hoop, M.A.G.T., den. Hollander, H.A. and Kerdijk, H.N. (1997). Spatial and
seasonal variations of acid volatile sulphide (AVS) and simultaneously extracted metals
(SEM) in Dutch marine and freshwater sediments. Chemosphere, 35 (10): 2307-2316.
Yin, H., Deng, J., Shao, S., Gao, F., Fao, J. and Fan, C. (2011). Distribution
characteristics and toxicity assessment of heavy metals in the sediments of Lake Chaohu,
China. Environmental Monitoring Aessessment, 179 (1-4): 431-442.
Hare, L., Carignan, R. and Huerta-Diza, M.A. (1994). A field study of metal toxicity and
accumulation by benthic invertebrates; implication for the acid volatile sulfide (AVS)
model. Limnology and Oceanography, 39: 1653-1668.
-282-
582.
583.
584.
585.
586.
587.
588.
589.
590.
591.
592.
593.
594.
595.
USEPA. (2005). Procedure for the Derivation of Equilibrium Partitioning Sediment
Benchmarks (ESBs) for the Protection of Benthic Organisms: Metal Mixtures
(Cadmium, Copper, Lead, Nickel, Silver and Zinc). U.S. EPA. 2005, EPA-600-R-02011. Office of Research and Development. Washington, DC 20460.
Oehm, N.J., Luben, T.J. and Ostrofsky, M.J. (1997). Spatial distribution of acid volatile
sulfur in the sediments of Canadohta Lake, PA., Hydrobiology, 345: 79-85.
Blazer, W. (1989). Chemische reaktionen und transportptozesse in oberflahennahen
sedimenten borealer und polarer Meeresgebiete, Habilitationsschrift Universitat Kiel
312S.
Duke, L.D. and Taggart, M. (2000). Uncertainty factors in screening ecological risk
assessment. Environmental Toxicology and Chemistry, 19: 1668-1680.
Chapman, P.M., Allard, P.J. and Vigers, G.A. (1999). Development of sediment quality
values for Hong Kong special administrative region: a possible model for other
jurisdictions. Marine Pollution Bulletin, 38: 161-169.
Gomez-Gutierrez, A., Garnacho, E., Bayona, J.M. and Albaiges, J. (2007). Screening
ecological risk assessment of persistent organic pollutants in Mediterranean Sea
sediments. Environment International, 33: 867-876.
Jones, D.S., Barnththouse, L.W., Suter, I.G.W., Efroymson, R.A., Field, J.M. and
Beauchamp, J.J. (1999). Ecological risk assessment in a large river-reservoir: 3. Benthic
invertebrates. Environmental Toxicology and Chemistry, 18: 599-609.
Di-Toro, D.M., Zarba, C., Hansen, D.J., Berry, W., Swartz, R.C., Cowan, C.E., Pavlou,
S.P. and Allen, H.E. (1991). Technical basis for establishing sediment quality criteria for
nonionic organic chemicals using equilibrium partitioning. Environmental Toxicology
and Chemistry, 10: 1541-1583.
USEPA. (1994). Technical basis for deriving sediment quality criteria for nonionic
organic contaminants for the protection of benthic organisms by using equilibrium
partitioning. EPA 822-R-93-011. U.S. Environmental Protection Agency, Office of
Water, Washington, DC.
Neff, J.M., Cornaby, B.W., Vaga, R.M., Gulbransen, T.C., Scanlon, J.A. and Bean, D.J.
(1988). An evaluation of the screening level concentration approach for validation of
sediment quality criteria for freshwater and saltwater ecosystems. In Aquatic Toxicology
and Hazard Assessment: 10th Volume, ASTM STP 971. Adams, W. J., Chapman, G. A.
and Landis, W. G. (Eds.), American Society for Testing and Materials, Philadelphia, pp.
115–27
Barrick, R., Becker, S., Brown, L., Beller, H. and Pastorok, R. (1988). Sediment quality
values refinement: Update and evaluation of Puget Sound AET. Bellevue: PTI
Environmental Services. Prepared for U.S. Environmental Protection Agency. Region
10-Office of Puget Sound.
USEPA/NOAA (United States Environmental Protection Agency/National Oceanic and
Atmospheric Administration). (2005). Predicting toxicity to Amphipods from sediment
chemistry. Washington, DC: National Center for Environmental Assessment. Office of
Research and Development.
Long, E.R. and Morgan, L.G. (1991). The potential for biological effects of sedimentSorbed contaminants tested in the national status and trends program, NOAA Technical
Memorandum NOS OMA 52, National Oceanic and Atmospheric Administration.
Ingersoll, C.G., Haverland, P.S., Brunson, E.L., Canfield, T.J., Dwyer, F.J., Henke, C.E.,
Kemble, N.E. and Mount, D.R. (1996a). Calculation and evaluation of sediment effect
concentrations for the amphipod Hyalella azteca and the midge Chironomus riparius.
Assessment and Remediation of Contaminated Sediments (ARCS) Program. U.S. EPA
Great Lakes National Program Office. Region 5. EPA 905-R96-008.
-283-
596.
597.
598.
599.
600.
601.
602.
603.
604.
605.
606.
607.
608.
609.
610.
611.
612.
Ingersoll, C.G., Haverland, P.S., Brunson, E.L., Canfield, T.J., Dwyer, F.J., Henke, C.E.,
Kemble, N.E., Mount, D.R. and Fox, R.G. (1996b). Calculation and evaluation of
sediment effect concentrations for the amphipod Hyalella azteca and the midge
Chironomus riparius. Journal of Great Lakes Research, 22: 602-623.
Smith, S.L., MacDonald, D.D., Keenleyside, K.A., Ingersoll, C.G. and Field, J. (1996).
A preliminary evaluation of sediment quality assessment values for freshwater
ecosystems. Journal of Great Lakes Research, 22: 624-638.
Swartz, R.C., Schultz, D.W., Demitt, T.H., Distsworth, G.R. and Lamberson, J.O.
(1999). Toxicity of fluoranthene in sediment to partitioning approach to sediment quality
criteria. Environmental Toxicology and Chemistry, 9: 1071-1080.
Ingersoll, C.G., MacDonald, D.D., Wang, N., Crane, J.L., Field, L.J., Haverland, P.S.,
Kemble, N.E., Lindskoog, R.A., Severn, C. and Smorong, D.E. (2000). Prediction of
toxicity using consensus-based freshwater sediment quality guidelines. U.S. EPA Great
Lakes National Program Office. EPA-905/R-00/007.
Hong, H., Yajuan, Y. and Xiaodong, W. (2004). Pollution of Heavy Metals in Surface
Sediments from Huaihe River (Jiangsu Section) and its Assessment of Potential
Ecological Risk. Environmental Pollution Control, 26 (3): 115-118.
Xin, Z., Taofa, Z. and Xifer, Y. (2005). Study on assessment methods of heavy metal
pollution in river sediments. Journal of Hefei Univiversity Technology, Nature Science,
28 (11): 1419-1423.
Guo, W., Liu, X., Liu, Z. and Li, G. (2010). Pollution and potential ecological risk
evaluation of heavy metals in the sediments around Dangjiang harbor, Tianjin.
Procededia Environmental Science, 2: 729-736.
Qinna, Z., Qixin, X. and Kai, Y. (2005). Application of Potential Ecological Risk Index
in Soil pollution of Typical Pollution Industries. Journal of East China Normal
University, Natural Science, 2005 (1): 110-116.
Puente, A., Juanes, J.A. and Garcia, A. (2008). Ecological assessment of soft bottom
benthic communities in northern Spanish estuaries. Ecological Indicators, 8 (4): 373388.
Loska, K. and Wiechula, D. (2003). Application of principal component analysis for the
estimation of source of heavy metal contamination in surface sediments from the Rybnik
Reservoir. Chemosphere, 51: 723-733.
Zhou, H.Y., Peng, X.T. and Pan, J.M. (2004). Distribution, source and enrichment of
some chemical elements in sediments of the Pearl River Estuary, China. Continental
Shelf Research, 24 (16): 1857-1875.
Simeonov, V., Massart, D.L., Andreev, G. and Tsakovski, S. (2000). Assessment of
metal pollution based on multivariate statistical modeling of "hot spot" sediments from
the Black sea. Chemosphere, 41: 1411-1417.
Bakac, M. (2000). Factor analysis applied to a geochemical study of suspended
sediments from the Gediz River, Western Turkey. Environmental Geochemistry and
Health, 22 (2): 93-111.
Swanson, S.K., Bahr, J.M., Schwar, M.T. and Potter, K.W. (2001). Two way cluster
analysis of geochemical data to costrain spring source waters. Chemical Geology, 179
(1-4): 73-91.
Debska, B. and Wianowska, E. (2002). Property modeling application of cluster analysis
method in prediction of polymer properties, Polymer Testing, 21: 43-47.
Hagan, P. and Fellowes, D. (20030. Multivariate statistical methods in battery research.
Journal of Power Sources, 122: 77-84.
Sundaray, S.K., Nayak, B.B., Lin, S. and Bhatta, D. (2011). Geochemical speciation and
risk assessment of heavy metals in the river estuarine sediments- A case study: Mahanadi
Basin, India. Journal of Hazardous Materials, 186: 1837-1846.
-284-
613.
614.
615.
616.
617.
618.
Forstner, U. and Witlmann, G.T.W. (1979b). Metal pollution in the aquatic
Environment, Berlin, Springer-Verlag, Heidelberg, New York.
Bhuiyan, M.A, Parver, L., Isalm, M., Dampare, S. and Suzuki, S. (2009). Heavy metal
pollution of coal mine affected agricultural soils in the northern part of Bangladesh.
Journal of Hazardous Materials, 173 (1-2): 384-392.
Sohrabi, T., Ismail, A. and Nabavi, M.B. (2010). Distribution and Normalization of
some metals in surface sediments from south Capsican Sea. Bulletin of Environmental
Contamination and Toxicology, 85 (5): 502-508.
Jonathan, M.P., Ram-Mohan, V. and Srinivasalu, S. (2004). Geochemical variations of
major and trace elements in recent sediments, off the gulf of Manner, southeast coast of
India. Environmental Geology, 45: 466-480.
Selvaraj, K., Ram-Mohan, V. and Szefer, P. (2004). Evaluation of metal contamination
in coastal sediments of the Bay of Bengal, India: Geochemical and statistical approaches.
Marine Pollution Bulletin, 49: 174-185.
Bird, G., Brewer, P.A., Macklin, M.G., Serban, M., Balteanu, D. and Driga, B. (2005).
Heavy metal contamination in the Aries river catchment, western Romania: implications
for development of the Rosia Montana gold deposit. Journal of Geochemical
Exploration, 86 (1): 26–48.
-285-
-286-