Assessing Alternatives for Disposal of Reject Brine from Inland Desalination Plants

ii

Assessing Alternatives for Disposal of Reject Brine

from Inland Desalination Plants

By

Mohamed Ezzat AbdelMohsen Mahmoud Ammar

B.Sc. in Civil Engineering, Cairo University, 2010

A Thesis Submitted to the

Faculty of Engineering at Cairo University

In Partial Fulfillment of the

Requirements for the Degree of

MASTER OF SCIENCE

In

IRRIGATION AND HYDRAULICS

ENGINEERING

FACULTY OF ENGINEERING – CAIRO UNIVERSITY

GIZA – EGYPT

2012

iii



By







MASTER OF SCIENCE

In


ENGINEERING

Supervised by

Dr. Ahmed Emam Ahmed Hassan

Professor

Irrigation and Hydraulics Department

Faculty of Engineering

Cairo University

Dr. Hesham Bekhit Mohamed Bekhit

Associate Professor

Irrigation and Hydraulics Department

Faculty of Engineering

Cairo University


GIZA – EGYPT

2012

iv



By







MASTER OF SCIENCE

In


ENGINEERING

Approved by the

Examining Committee:

_____________________________

Prof. Dr. Ahmed Emam Ahmed Hassan, Thesis Main Advisor

_____________________________

Prof. Dr. Ahmad Wagdy Abdel Dayem, Member

_____________________________

Prof. Dr. Ahmad Ali Hassan, Member


GIZA – EGYPT

2012

�� !"� �� "�� "��#�$�� %�"�� "��%��"�� "��"��#�&�� %�!��"��'��(�"��%�� )�� "�� %��"��"��#�� %�� "��"� �"��%�� "�� "� ��"��%��#�*��"�� "� "��"�� "� �+�� "� �� "� "��"�� "��"��%�� #�,�� "�� -� ��.� � �� #�*�� (�"�� (�"��%�� "�� #�*��"��!�"�� "��!/� %�� -� ��.� �#�0�� %� "��1�� "� �� "�� "� �� #�2�� (�"�� !�� (�"��%�� +��#�*�� 1��/3$2430� ��/*5$/.�"�� "�� "��.��0�*�"�� %�� #��

�� !�� "#��$��%��"#�� & �� "�'�� '�� ($)��*+$�� '�� ,*�� -�� .�� &�� / ��$�� &��

�� !" �#��"$��! �� % ��$�# ��&��$$�"��!�'�!$��(��$��)�� $��&�$��$��!��"�# �� # ��!�$ ��$� �� $��$$��$�� &�� $�� !� ��! ��!�� !��$�� #��$$� !�� ! ��$�# ��*��#��$��! �� !��!��&��%$*�+�# !�'��!�%$*�' ��#�, ��-�#�� !��$�-��#��#�� $ ��!�� $��"$�� &��&�$�� $�"��!�� !�� # �� $� �" $� ��#�� # �! �� $�.��$�*�� $ ��!� �$� ��&�� #��&�#��-�#��&�� $�/�"*�/00��+�! �1�� -�#��#�� $�2�&��-�#�� $�%��!�1��&��#��" $��$�� $�&�$�� $��$��!�� *�3�� !�� %$*�1��# !�4��$�/�!��+��#5��"��!�� !��$�� ! !�#��!�!��*�+��-��!�� #��"$��! ��%$*�6�� !�6��$�/�!��-�� $�7 ��#��" $��&�� 8��! ��-�&�$�� !�$��"�� &� �!��( ��$��8��!�/�"*�9�� $�4�"��&�1�%2+3/:*��!�� %$*�+�!��,�0�$��-�%$*�8�# ��+�! �;��!-��!�%$*�+�# !�2�"!��&�$�� $��!�� "�&�$�#��&��$ ��$ $��!��!�*��%$*�1��# !�4��$�&�$��$��"�# �!�$��"�� $�� $��" ��&�#��$ � �$��$ $��!��"�� # ��$�"$ ��&�� $�*��!�� %$*�1��# !�+��&�$��$��!��!�� *�+�!�&��-��$��#��&$� �!�� $ ��#��! -�' "�0�-�+��-�8�#�-�+��$�&-�'��#��-�1��$-�<��!��"�� :��!*� �1��# !�/*�+##�$-�=>?=*�

�� !��"�� "��#�$��%� ��!�� %�%�%�&��'(��')��%�*�&�'+',(-.�$/0'10�,2�03��%�4�&�'+',(-.��56'-0��'+�,17��'0./7/2/83��%�9�&�.'+�+��(8,1�:,0�/1��%�*�4�;��#�$��<��!�� !��= ��<�*�%�&�'+,2�1,0�/1�#�+0/(3�,17��/2>0�/1��<�*�*�&�.'��'+,2�1,0�/1��"?'(�'1-'��1��83?0��%*�*�4�&�'+,2�1,0�/1�,17��>0>('��'�'2/?@'10��1��83?0��%A�*�9�&��'(��')�/B��(,-C�+.��,0'(��1��83?0��%D�*�9�%�&�(,-C�+.��,0'(�='(+>+��,2�1'��,0'(��%E�*�9�*�&�/>(-'+�/B��(,-C�+.��,0'(��%E�*�9�4�&$('0(',0@'10�/B��(,-C�+.��,0'(�$(�/(�0/��'+,2�1,0�/1��%<�*�9�9�&�'+,2�1,0�/1�/B��(,-C�+.��(/>17),0'(��1��83?0��%<�*�9�;�&�1'(83��/1+>@?0�/1�/B��(,-C�+.��,0'(��'+,2�1,0�/1�F��G�,17��3?�-,2��'-/�'(3��,0'+��*%�*�9�A�&$(/52'@+��++/-�,0'7�)�0.��(,-C�+.��,0'(��'+,2�1,0�/1� **�*�;�&�'��')�/B��'+,2�1,0�/1��'-.1/2/8�'+��*4�*�;�%�&�2'@'10+�B/(��'+,2�1,0�/1�$(/-'++��*4�*�;�*�&�3?'+�/B��'+,2�1,0�/1�$2,10+��*9�*�;�4�&�'@5(,1'��'+,2�1,0�/1�$(/-'++'+��*;�*�;�9�&�.'(@,2��'+,2�1,0�/1�$(/-'++'+��*E�*�;�;�& /1��"-.,18'��'+,2�1,0�/1�$(/-'++'+� *<�*�;�A�&#35(�7��'+,2�1,0�/1�$(/-'++'+� *<�*�;�D�&�1'(83��/1+>@?0�/1�/B��BB'('10��'+,2�1,0�/1��'-.1/2/8�'+��4H�*�;�E�&�'1'),52'��1'(83�B/(��'+,2�1,0�/1��4%�*�A�&�'6'-0��(�1'��+?/+,2��'0./7+��44�*�A�%�&��,?/(,0�/1�$/17+��4;�*�A�*�&��+?/+,2��1��>1�-�?,2��')'(,8'��3+0'@+��4<�*�A�4�&�,20�$(/7>-0�/1��9H�*�A�9�&�''?��'22� 16'-0�/1��94�

�� !��"�"!��#��#��! ��$��!%��&'��&%��$��(��')�"�(��*��&%��$�$��"�(��*��&%�&+�� ,�&%-��.�"�(*&(� ��)(��/��$��0��&**�%1��&- �/��%�� ,�&%-��.�"�(*&(� ��)(��/��2�$��/ ��&%��&'� ��$��3��%'�� 4� �%��$�5��++��&+�� %��)�&+��0�*&��&%��(��5�$��( �(��%'�"�(�((�&%��2�$�2��%(��0��)��%� )(�(�+&��.��++��&+�� %��)�&%��.��&%'��+&�/�%��5��$��6��7�"�/�%(�&%(�+&��.��0�*&��&%��&%'(�+&��+��"�(*&(� ��5��$��&(��&+��.��0�*&��&%��&%'(��8��5��!��#��8#��!��5��.�/�� &'� �&+��"�#�3��&'��.�/�� &'� �&+��$"��&'��2��$��.�/�� &'� �&+��3��&'��2��! ��2��9��8�"3��#�3��"��"�#��2��"�(��*��&%�&+��')��1�&%��25��&%��*�� &7��&'� ��2��$��/�� &7��&'� �"�0� &*/�%��22��$�� %'� �� %��22��$��9�&%'7��#�0� (��%'��&0�/�%��66��$�$��9�&%'7��:� ��)��6��$��+��)'�� &*��(��62��$��"&/��%��*�� "�(��;��&%�+&��.��1�&%� ��&'� ��6��$��"&/��%��*�� "�(��;��&%�+&��.��#&�� &'� ��4&%'��)��%'�!%�� &%'��&%(��%'�!/* �/�%��&%�&+��"�#�3��!%<��&%�3� ��)'�� &%'��0��)��5��&'� �� -��&%��5��(��/��'��/��(��((�((/�%��&+�� -��&%��$��%(*&��&'� ��5��4�(��(��%��&��!=��2�!�>��!��!��8#��"�"!��8��!��2��9�&%'7�� &7��( �(�&+��.��4�(�,��(�� -��'��&'� ��$6��( �(�&+��.��(��!%<��&%��%��&��$��

�� !��"��!�#�� $��%�� &��!�#�� $��%�� '��!�#�� $��%�� "�� (��%� ) ��% ��*�� +��!�� +��,��-��*��(��.�� /��/��) ��0)�� %� ) ��1� ��-��/�#� )$��% ��2*%�!��!3!4��.��55*�67��04� ��04��*4/��!�055!4/*��04��+0��+��!�,0�8��.�"��$$��.�"�'�� '�"�� $$�� +��, �9��!+!�!4�!��

�� !��"��#��$�%��&'()%��*��+��+�� !�'��,��!��-��%��.�/�0�� 1�0��0��2�� #��0�� "�� .��'��345��+556�&�"/7�+5568*��5��+�+� ��8��9��1�0��!��-��%�&�"/7�+5568*��+�(� 1�0��!��-��%��$�%��&'� �.�� .��:��'��7�+55+*��(��+�;� 1�0��0��-��-��0��$�%��&$�<#� �=7�+5�5*��;��+�4� >��1�0��!��-��0��$�%��3?5<+554�&$�<#� �=7�+5�5*��;��+�6� 1�0��8��.��@��:A��.��#%0��'0��$�%��&:��*��+5��+�B� C�%�$��'��0��.��1�0��#%0��'�&0��.��D�'�'8��<8�0� �0%0��'*�'� �.�� .��&E��0��3?5*��+;��+�?� !�'��,��8��!��-��%��.�/�0�� 1�0��0�.��>��'�� @�'8��>�-��%��>��'��>�-��%��-�� 0�@$17�@#�� @F!��@�'8��>�-��%��-�� 0�G97�$1�� $1G��0�G�.��-��!��9��&��0�'� ��*�!��-��%��.�8��>�-��0��&�"/7�+5568*��+4��+�3� >��@�-��0'��.��90'�0�0�� G�,��0��90'�0�0�&G9*��-�00�0��+6��+��5�>%��-��:��'��.��$��-�� %0�0�@�'8��0��+B��+��G��.�:��-�8��%��.�1�..��1�0��>�-��0��'0��.�#��!��-��"��&��-��-�'��--�00� D�@��-�7�+5�+*��(5��

�� !�� "�#�� $%#�!&�''()�(��(�� !�� *� �� "�� $"+��&�'',)�(��-�.�"��#./*�!0�� -'��,�� ! �� "�� !�1��2��0*0��!0�� $.��&�'')�-��(�� "��#�� / �� #��,��(� /�� "�� 1��.�� .�*3 �� ,��((� #��0�� .�*3 �� ,4��(-� "��#��#�� #��$2� ��&�'�')�,5��(,� �� *#�� 6�� ,7��(4� �� .�� 6�� 4��(5� "��/ �� 6��3�� #�� .�*3 �� 4-��(7� # �� # �� #��"��$8 ��&�'',)�45��(9� #�� 6�� $3�� &�'��)�49��(�'�� 06�� $�)�� !��0��&�$�)�� .��6�� !��&�$�)�� #��&��$�)�� :��; ��< ��5��(�� 6�� $�)�� !��.��6�� &�$�)�� !��&�.��6�� #�� 6�� !��&��$�)�� !��&�.��6�� &�#�� 6�� 5-�

�� !"�� #��$� ��%&��'��(��$)��*�'�� )��+�*�$��+�'��%%��,��"�� )��+�*�$��+�'��%-��.�!'� ��"��/ � ��+�'��%0��&��$� ��1�2*�$��./3��4�5��*�'��465 ��!#��-7��%�!'� ��8��"��/ � ��+�'��-��-"��8��'��(��$)�� 9��:��*��*�'�� '��-7��0��"��1�2�+�'��-,��.�� $�� :�'��8�2��"�� )9��2��8� ��)�/:�6��0&��.�� !'� �� '�� "�� 4;��6� ��7�75��0%��.�� "��'8�+��<''��"�� )9�� 7��.�, "��(��8�+��<''��"�� )9��7��.�. "��(��8�+��8�2��"�� )9��7,��.�& ��$�� :�'��<''��"�� )9��7.��.�% ��$�� :�'��8�2��"�� )9��7&��.�- ��/��!"��:�'��<''��"�� )9�� 4��=�� !��$��00-5�� 7%��.�0 ��/��!"��:�'��8�2��"�� )9�� 4��=�� !��$��00-5�� 7-� �

�� !��"#�� $��"��%�&'��%() $��)��(�� !��"!�� *��$��"��+�, ��+��,&$��)��)-��.��%�� !��"#�� !�� *��$��"��/��%��%��0��1��$��2��2!��3+��!��%��!��4��*��%��!��&��%�-��5�+��67��%��%��7��%��!��48��67��!�+�-��48��(��&��!��-��-��!��%�0��!�+0��9�+��%��-��+��:��1��4��!��%��:��-��#��4��%��70��-��#��+��+��0��-��;��48��((��<��!��-��-��!��%�0��+0��!�+0��9�+��%��-��+��:��1��4��!��%��:��-��#��4��%��70��-��#��+��+��0��-��%��48�� '��,��+��0��%��+��=��0��70��=��#��70��-��;��48��(2��'��+��0��%��+��=��0��70��=��!��70��-��%��48��(2��( -��#��#��!��70��0��;��%��8��(� �

�� !��"��!�� #��$ ��%�&�� !��"��!��#��# '�(��)��!��'�(��*��&��*�!��!�+��$,-��'�(��.!��/��&��#$��0 '�(��)��!��'�(��*��&��*�!��!�+��$,-��'�(�� .��1�� 2�� *��!/��&��%�&��.��*�� 3�� *��!/��&��##��, )�!��+��'�(��2�!!��*��4��##�� '�(��)��!��*��'�(��*��&��*�!��!�+��$,-��'�(��.!��/��&��##��5 '�(��)��!��*��*��&��*�!��!�+��$,-��'�(�� .��1�� 2�� *��!/��&��%�&��.��*�� 3�� *��!/��&��#,��6 '�(��)��!�� '�(��*��&��*�!��!�+��$,-��'�(��.!��/��&��#��7 '�(��)��!�� '�(��*��&��*�!��!�+��$,-��'�(�� .��1�� 2�� *��!/��&��%�&��.��*�� 3�� *��!/��&��#5��8'�(��)��!��-��!�-%�-��&��*�!��!�+��$,-��'�(��.!��/��&��#6��'�(��)��!��-��!�-%�-��1��&��*�!��.!�+��.!��*��,8+��&�&��+��'�(��9��$,-��'�(��#7� �

�� !��"�� #��$��%��%��&��'��()��$��(��*��+��,��&��(-��$��(��*��./��0��'��+��+��$��'�� &��1��+*�� . ��.��$��2��3��)��!��&��$��45��+�6 /� 7�� .��#%��$��!��&��+%��.8��%��%��$��!��!��#/��8%��$��&��9��+��!�� #��:%��&��$��4��+��3��7��.;��;%��&��+��"�� 1��9��+��!�� .;�� /%��<��+$��$��1=��&��+&��'��+��>�+��4%��+��+��+��"��3+��7��#0�� %��<��+$��,��$��1=��&��+&��'��+��>�+��4%��+��+��+��"��3+��7��#0�� $��1=��&��+&��'��+��>�+��4%��+��+��+��"��3+��7��#.�� 0%��$��!��!��1��&��+��)��?��#8� �

�� !�"��#��$$$�%��"��&'(��$$$)��$�� !��*��+��,��-��.� /��0��,��*��-��1� 2��3*�&��345��,��*��-��-��67�&��*��6�8�0��9��:��;� <�7��=��#0��9�>��&'��/��#��5��%?��0��(��$$�)��$��@� >��&'��!��#�&��*��%?��#�!�6��(��$$�)��:� ��>��&'��%0��9��9��?��#�!�6��(��$$�)��A� �67�&��<��&��&��6�B��#07��9��>��*&��0��%��(��$$�)��$��67�&��!�&��6�=��#��9��>��*&��0��%��(��$$�)��B�00��2��0��<�&�#��>��&'��%��(��$$�)��.��.�� "��2��0��9��<�&��9��+��!4�#��-.��.�� "��2��0��9��<�&��9��!�C�&��!4�#��-.��.�.� =��#��9��0�?��2�#��%�AA�)�B��7��"��+��!��D#0��6�%E)��!#��%'0F��6)��;.��.�1� 8�&��9��G��+��"�&��*��;1��.�-� ?��<��7��7��!��<��0��<��7��!��9��G��+��"�&��*��;1�

�� !��"��#��!��#�� $��% ��&&��'�� '��(�� &��)"��*��!�"��&& ��+ (��,(��-#��(��.��!�� '��/��&+�� #��'��0��1��)��#��'!��)��2''��#��!� 3!��&+4��4 #��'��0��1��)��#��'!��)��5�0��#��!� 3!��&+��& �!��"��#��/�6��*��&4��+ ��#��5��'��(��0��&�� #��'*��/�6��'��1'��&��

1

CHAPTER ONE

INTRODUCTION

1.1. Overview

Owing to the rapid growth in population, the development expansion in all

aspects in the past decades, and the fixed share Egypt gets from the Nile water

according to the 1959 treaty (i.e., 55.5 billion m3), it is expected that Egypt will

rely more on other supplements to the Nile River water to meet the needs of

different sectors. The combination of continued rapid population growth and

severely constrained fresh water resources confronts Egypt with great

challenges in the pursuit of sustainable development. The total population of

Egypt increased from 22 million in 1950 to 82 million today, and is likely to

increase to above 92 million by 2025, which means a severe drop in the per

capita water share to reach 600 m3/year compared to 1,000 m

3/year described

as the international standard of water scarcity (Figure 1.1). Although vast

quantities of groundwater exist in the deserts of Egypt, most of these are non-

renewable and are stored at great distances below land surfaces. In some

places, large amounts of brackish groundwater exist and can be utilized. Where

freshwater availability is limited, desalination of brackish groundwater can be

used as an alternative supply, especially given the lower desalination cost

compared to seawater. Desalination of brackish groundwater in Egypt has a

great potential with respect to the availability of the resource.

2

Figure 1.1. Population Growth and Per Capita Water Share in Egypt (m3/year)

1.2. Research Potentiality

Desalination of seawater is becoming a reliable and cost effective mean of

providing fresh water especially in arid regions like Egypt. However, the

utilization of desalinated brackish groundwater is not as common, as it has to

go in parallel with designing a disposal system in a way that protects the

environment and be cost effective. Moreover, there is a lack of the information

and knowledge about the environmental impacts of the desalination byproduct

(reject brine) on the soil and groundwater. A need thus exits for assessing the

possible disposal options (i.e., evaporation ponds, well injection in deep

aquifers) from the perspectives of the technical viability, effect on the

environment as well as the economic feasibility.

Brackish groundwater desalination holds promise as a water supply

strategy. It offers opportunities such as providing a viable water resource where

other supply options are not readily available. It can also free up pressure on

freshwater resources that are of vital importance to the environment. However,

a number of important issues should be addressed when considering this as a

0

20

40

60

80

100

120

140

0

1000

2000

3000

4000

5000

6000

7000

1800

1900

1950

1961

1965

1970

1975

1979

1980

1985

1990

1994

1998

2000

2005

2012

2025

2050

Pop

ula

tion

(m

illi

on

)

Per C

ap

ita W

ater sh

are

Year

Water Share Water poverty line Population

3

water supply option. Atop of these issues are the cost, energy requirements, and

disposal of the reject water or brine resulting from the desalination process.

This research focuses on addressing this latter issue. In particular, it aims at

identifying the alternatives of land disposal of reject brine resulting from inland

desalination plants and studying the impacts of disposal on the environment.

The main objective is to assess the different options of reject brine disposal

including evaporation ponds, and deep well injection into deep aquifers. It is

expected that this research will deliver an assessment of the feasibility of each

of the possible disposal options and the long-term impacts on the environment.

1.3. Research Objectives and Methodology

This research will investigate the issue of brackish groundwater

desalination and the associated problems. The main problem that we will focus

on is the disposal of the desalination process by-product (i.e., the reject brine

water). The overall objective is to assess the technical and economic feasibility

and the environmental impacts of different options for inland disposal of the

reject water. To achieve this overall objective, the following specific objectives

are considered and are linked to the selected case study:

Assess the feasibility of each option in terms of sustainability of the

resource, cost of disposal system implementation, and short- and long-

term environmental impacts.

Study the effect of salinity of the reject brine on the evaporation rates

which reflects on the dimensions of the required evaporation ponds.

Evaluate the time-varying effects on groundwater quality in the injection

domain present in the chosen study region.

Suggest the most appropriate disposal system for the desalination

byproduct for the chosen site.

4

Suggest a development scheme for the regional area encompassing the

study region utilizing the brackish groundwater as a source of feed water

for desalination.

The abovementioned research objectives are achieved using the suitable

modeling tools and software such as the utilization of the MODFLOW code,

MT3DMS as well as using SEAWAT for variable density flow simulation. In

addition, routing the reject water into evaporation ponds is addressed using the

appropriate tools (e.g., water and salt balance, rainfall rates, evaporation rates,

and infiltration rates). Previous work and studies-however rare-concerned with

the inland disposal of reject brine and injection into deep saline aquifers are

compiled and integrally used for building a concrete research base for the

current research purpose.

The proposed methodology generally consists of the following tasks:

An extensive background on the desalination technologies and the

typical recovery rates, energy consumption, feasibility, and

environmental impacts.

Literature review of the relevant studies dealing with inland desalination

of brackish groundwater and means of brine disposal

Case study data collection, analysis and compilation of studies related to

the Upper Cretaceous and the Lower Cretaceous aquifers in Central

Sinai where the studied desalination plant is located.

Setting/developing conceptual processes for the different disposal

techniques and converting these conceptual processes into mathematical

models.

Developing a numerical model for the groundwater system taking into

consideration the effect of salinity in simulation as variable density

groundwater flow.

5

Using the developed models to assess the environmental impacts of the

different options on both the short- and the long-term.

1.4. Thesis Organization

Following this chapter which includes the introduction, this thesis contains

six other chapters. Chapter two gives a detailed background on the evolution of

desalination worldwide, in Middle East, and the desalination experience in

Egypt with a brief description for the desalination technologies, typical

desalination energy consumption, problems and environmental concerns

associated with desalination and especially stressing on brackish groundwater

desalination, sources of renewable energy for desalinating water, and the

potential remote areas for development and desalination in Egypt with

identification of the potential saline aquifer systems. Also, this chapter presents

a literature review on the disposal options of reject brine from inland

desalination plants. Design criteria, researches, case studies, and assumptions

made by researchers previously addressing the disposal options of the brine are

discussed in this chapter as well.

The description of the selected case study is presented in Chapter three.

This chapter also included an assessment of the evaporation pond disposal

option. Water and salt balance are utilized for the evaluation of the pond

performance as well as an approximate estimation of the costs.

Chapter four exhibits a brief background on the used software in this study,

GMS (Groundwater Modeling System) developed by Aquaveo. It also presents

a clarification for the codes used to develop the groundwater flow and transport

models to simulate the movement and dispersion of the injected brine into the

aquifer system through MODFLOW for the groundwater flow pattern and

MT3DMS for the transport simulation, which are then coupled with SEAWAT

to account for the variable density in the flow simulation. The chapter includes

an explanation for the mathematical equation used to solve these models.

6

In chapter five, the developed regional flow model is discussed through

describing the model domain and the conceptual model. Model calibration is

performed to estimate the values of the hydraulic conductivity of the model

layers. A local model is extracted from the regional model for local study of the

Al-Monbateh region, with boundary conditions based on the calibrated regional

model. Different cases of injection and sources of uncertainty are considered in

addition to the description of a base case scenario that is benchmarked as a

comparison model.

Chapter six includes the results of the base case scenario of the developed

regional model and the results of the different injection scenarios for the local

model. The results are presented for a time frame of 25 years and a comparison

is made between each case and the base case scenario. It also describes a

suggestion for the potential development areas for the regional study area

utilizing the Lower Cretaceous aquifer in Central Sinai based on the depth to

the aquifer, the salinity of the groundwater, the thickness of the water bearing

formation and the topography of the area. Production and injection well fields

are proposed upon the allowable well extraction rates and minimum

environmental risk of injection for a study period of 25 years. The developed

regional groundwater model is utilized to simulate the proposed future

groundwater extractions and to predict the aquifer response to the different

extraction scenarios and probable changes in groundwater quality over the

foreseen period of exploitation, to assess the optimum, economic and

sustainable groundwater extraction plan.

Chapter seven abridges a summary to the performed work in addition to the

conclusions derived from the analysis and results presented in the study as well

as the recommendations for future work that can complement the current study.

It should be stressed, however, that the conclusions made are specific to the

case study and the scale of the desalination plant considered. Such conclusions

will likely change as the scale of brine production changes and this should be

7

taken into consideration in interpreting the results and conclusions presented

herein.

9

CHAPTER TWO

BACKGROUND AND LITERATURE REVIEW

2.1. Desalination History and Evolution

About 470 million people live in areas with severe water shortages (e.g.,

northern China, northern Africa and the Middle East as well as the western

United States, parts of Mexico and northern India). By 2025, the number of

people living in water stressed regions is expected to reach 3 billion (Cosgrove

and Rijsberman, 2000). This dramatic increase raises the flag lead to a critical

need of more potable water for human uses, which has put more emphasis to

non-conventional water sources (i.e. desalination of seawater and/or brackish

water).

Desalination technologies and their applications have evolved dramatically

over the past half century. The use of desalination processes was very limited

to activities were distilled water was required until the early 1960s. There have

been efforts to identify unconventional water supplements to the traditional

water supplies in order to fulfill the needs of the rapid growth of population and

the development expansion in all aspects in the past three decades and due to

the increased expenses, unavailability or the controversies associated with the

use of traditional sources. These efforts yielded an exponential increase in the

desalination capacity both globally and nationally to overcome the water

scarcity problem faced by many societies as well as to provide fresh water to

localities experiencing rapid population growth with the decreasing or fixation

traditional water supplies as in Egypt.

A total desalination capacity of about 26 million m3/d was installed or

contracted worldwide by the end of 1999, counting only those plants with a

10

capacity more than 100 m3/day (Wangnick, 2000). Global desalination water

production capacity has been increasing exponentially since 1960 to value of

59.5 million m3/d in 2009 with an increase of 6.6 million m

3/d in the last year

according to the 22nd

GWI/IDA Worldwide Desalting Plant Inventory.

Middle East region has the largest share of desalination, where the leaders

of desalination are found there, followed by North America then Europe. Table

2.1 shows the Desalination in the world’s regions (Wangnick, 2000). Figure 2.1

shows the cumulative capacity of the desalination plants in the United States

and Worldwide until 2006 whereas Figure 2.2 shows the global desalination

capacities by countries (GWI, 2006b).

Table 2.1 Desalination in the World’s Regions until 2000 (Wagnick, 2000)

World’s region Desalination in 2000

Total capacity million

m3/d (%)

Seawater

million m3/d (%)

Australia & Pacific Islands 0.1 (0.4) Negligible

Asia 3.2 (13.3) 1.2 (8.5)

The Middle East 11.3 (47.1) 9.5 (67.4)

Africa 1.2 (5.0) 0.8 (5.7)

Europe 3.1 (12.9) 1.7 (12.1)

North America 4.3 (17.9) 0.3 (2.1)

Central America & Carribean 0.6 (2.5) 0.5 (3.5)

South America 0.2 (0.9) < 0.1 (0.7)

TOTAL: 24.0 (100) 14.1 (100)

Figure 2.1. Cumulative capacity of installed desalination plants in the United

States and Worldwide from 1950 to 2006 (GWI, 2006b).

11

Figure 2.2. Global online desalination capacity (GWI, 2006b)

12

More than 120 countries are now using desalination of seawater for

domestic uses. More than 90% of water of the Gulf countries (Oman, Qatar,

Bahrain, Emirates, Saudi Arabia and Kuwait) is from desalination. The cost of

desalination, especially reverse osmosis (RO) has reached a competitive level.

For instance, in 1948 the desalination cost was over US$1/m3, but now on

average it is about $0.50/m3 (El-Kady and El-Shibini, 2001).

2.2. The Desalination Experience in Egypt

In Egypt, the integrated water policy is based on three main fundamentals

namely: increasing the Nile discharges from the sources, enhancing the water

efficiency and preventing pollution, and finally using non-conventional water

sources (El-Kady and El-Shibini, 2000). One of the most promising non-

conventional water sources is desalting water especially with the current low

price of desalination and the continuous decline in desalination cost.

The desalination experience in Egypt is relatively new compared to other

countries. It began in Helwan (south Cairo) with a large distillation pond for

domestic uses, and then moved to electrodialysis (ED) (described in details in

section 2.5.2) processes in remote areas in the mid-1970s where the real

Egyptian experience began when a number of ED plants were installed in

remote areas, mainly for military and exploration camps, industries, hotels and

resorts. In contrast, recently the reverse osmosis (RO) desalination technology

became more widely spread and more common due to its cost effectiveness. It

was important to start considering more non-conventional water sources as the

conventional sources became already exhausted (El-Kady and El-Shibini,

2001).

Research efforts have moved the cost from being expensive to competitive

allowing the feasibility of desalination in obtaining a reliable source of water.

During the period 1975 to 1982, three different models of ED plants were

installed in Egypt, and their capacities differed from 50 to 1000 m3/d, and with

13

salinity levels between 2000 to 1000 parts per million of feed water (El- Sadek,

2010).

Desalination increased notably in Egypt, where the total installed capacity

has grown to some 228,900 m3/d on 2012 (Moawad). Figure 2.1 shows the

evolution of desalination in Egypt during the past five decades and until 2000.

Most of the plants treat seawater, however lately a growing number of

installations use brackish water and the capacity of these installations ranges

between 500 to 10,000 m3/d (Allam et al., 2002). Nowadays, the amount of

desalinated water in Egypt is in order of 84 million m3/year. Figure 2.2 through

Figure 2.4 show the desalination installation capacities in Egypt.

Figure 2.3. Desalination Capacity in Egypt (modified after Allam et al., 2002)

0

50000

100000

150000

200000

250000

1960 1970 1980 1990 2000 2010 2020

Cap

aci

ty, m

3/d

Years

14

Figure 2.4. Desalination Installation Capacities in Egypt (El-Sadek, 2010)

Figure 2.5. The Desalination Capacities in Egypt in 1980-2005 (El-Sadek,

2010)

The development of the Red Sea zone led to an increase in water demands

to meet the needs of the tourist, industrial and urban settlements. Table 2.2

shows the Red Sea modern desalination units. The desalination plant sizes for

such application are relatively small due to the nature of the coastline and the

dispersed locations of dwellings (Khalil, 2004).

15

Table 2.2. Red Sea Modern Desalination Units

Capacity, m3/d Technology Activity Location

3 x 1,500 VC* Tourism Abu Soma Bay

4,000 RO Hotel Movenpick, Sharm El-Sheikh

8,500 RO Tourism Hashish Bay

4,000 RO Tourism Hurghada

25 x 500 RO Various Isolated sites

30 x 100 - 200 RO Various Isolated sites

Desalination units can be categorized into two types based on their

ownership: first are government-owned units; and second are the private-

sector-owned units (Abou Rayan et al., 2001). Table 2.3 presents the

government-owned units and the technology used whereas table 2.4 presents

the private sector-owned units. As shown, the major supply of desalinated

water is from the private sector mostly owned by hotels.

Table 2.3. Governmental Desalination Units in Sinai

Place Taba Taba Nuweiba Dahab

Sharm

El-

Sheikh

Sharm

El-

Sheikh

Nuweiba

System ROa MVC

b ED

c RO VCD

d RO MED

e

Start date 1986 1996 1958 1995 1996 1998 1999

Total area m2 50,000 42,000 23,600 30,000 30,000 30,000

Capacity, m3/day 600 2,000 300 500 500 4,000 2,000

Feed water salinity,

ppm 48,000 48,000 2,400 44,000 44,000 44,000 45,000

Product salinity, ppm 450 30 500 500 30 500 50

Power consumption,

Kw/m3

13.5 9 4.3 8.5 9 6.5

Total coast/m3 , in LE 6.21 6.64 2.78 7.51 4.75 6.43 NA

aReverse osmosis,

bMechanical vapor compression,

cElectrodialysis

dThermal vapor compression,

eMultiple effect desalination

Table 2.4. Private-Sector-Owned Units in Sinai

Location Owner Technology Capacity

(m3/day)

Salinity

(ppm)

Product

salinity

(ppm)

Taba Golden Coast RO 750 40,000 350

Maleh Company RO 4000 35,000 400

Nuweiba Helnan RO 240 44,000 400

Hilton RO 300 44,000 400

Shar

m E

l-

Shei

kh

Pyramiza RO 2000 44,000 400

Ramo RO 1000 44,000 400

Metito RO 500 44,000 400

Raga RO 2000 44,000 400

16

Southern Water

Co. RO 7000 44,000 400

Montazah RO 2500 44,000 500

Residence RO 500 44,000 600

Euro Palace RO 500 44,000 400

Meridien RO 500 44,000 400

Aqua Marina RO 2000 44,000 400

Moevenpick RO 1000 44,000 400

Marriott RO 500 44,000 350

Dah

ab

Sheiha Zayed RO 2500 44,000 400

Bacha Coast RO 500 44,000 400

Ghazala RO 500 44,000 400

Helnan RO 800 44,000 400

Pullman RO 500 44,000 400

2.3. Desalination and Future Development in Egypt

With the increasing population and limited renewable water resources, the

government of Egypt has to develop a solution to overcome the scarcity of

water supply and the desalination is one of the promising solutions. However,

the current practice in Egypt indicates that industrial and tourist sectors

undertook the lead in the desalination in Egypt, resulting that few installations

were operated by public water agencies. The government policy will be

directed to develop remote area where natural resources are present and the

desalination of brackish water may be used as the only source of providing the

basic needs for living, mainly water.

Therefore, desalination is being studied and resulted in ranking the most

prominent remote areas to be selected for research and development according

to priorities based upon water scarcity. A summary of the most prominent

remote areas as reported by El-Sadek, 2010, is as follows.

1. Along the Red Sea coast where tourist potential is present with a

brackish water supply of 1000 ppm;

2. Along the northwest coast where new communities and tourist potential

as well with as a brackish to saline water of salinity ranging between

1000 and 10,000 ppm;

17

3. Sinai coastal zone and wadis where there are nourishing tourism,

agricultural, industrial and new communities are built with a brackish

water of 1000 ppm;

4. The northern desert along the delta fringes (Nubaria and vicinity) as the

over-exploiting of some wells caused salinization of groundwater

As abovementioned, the desalination of brackish water is very promising

water resources in these remote areas. An initiative for using such resources

was implemented in Central Sinai by USAID/Egypt and the North Sinai

governorate, under the LIFE Sinai program. The initiative was aimed at

constructing three desalination RO units in Central Sinai to assist Bedouin

communities in the sub-governorate areas of El-Hasna and Nekhl in developing

and improving their livelihoods by supporting them with their required share of

potable water. The three plants located in Al-Meswateyya, Al-Monbateh and

Bir Beda. The construction of the three plants was finished on September 2011

with a total production of 600 m3 of potable water per day (200m

3/day/plant)

benefiting 6000 inhabitants of Central Sinai. Currently, only the Al-Monbateh

desalination plant is running whereas the other two plants are expected to start

operating with full capacity by the end of 2012.

In the following section a review of the available brackish water in Egypt

and its desalination requirements are presented

2.4. Overview of Brackish Water In Egypt

2.4.1. Brackish Water Versus Saline Water

Saline water is defined as the water that contains a significant amount of

total dissolved solids (TDS) and usually expressed in parts per million (ppm) or

milligram per litter (mg/l). The concentration level of the salts classifies saline

water into three main categories. Freshwater mainly covers water with a TDS

up to 1000 mg/l, while brackish water from 1000 to 10,000 and seawater above

35,000 mg/l. Sometimes there exits special cases of brackish water that might

contain 10,000 to 35,000 mg/l TDS and in this case it can be referred as

18

“difficult” brackish water (Buros, 1980). It is worth mentioning that,

sometimes, saline water contains small amounts of organic matter and

dissolved gasses, however, the majority of dissolved materials are inorganic

salts.

Table 2.5 shows the typical salinity limits of waters, some guide limits for

livestock to salinity in drinking water, and ranges of salinity in some of the

popular seas (modified after Salinity Management Handbook, Second Edition).

Table 2.5. The Typical Salinity Limits of Waters

TDS (ppm)

Distilled Water 0.67

Freshwater 0-1000

Brackish Water 1000-10000

Tolerance of livestock to

salinity in drinking water (at

these values, animals may

have an initial reluctance to

drink, but stock should adapt

without loss of production)

Beef Cattle 4000-5000

Dairy cattle 2500-4000

Sheep 5000-10000

Horses 4000-6000

Pigs 4000-6000

Poultry 2000-3000

Salt water swimming pool 4000-6000

Sea Water 10000-35000

Dead Sea 73700

Mediterranean Sea 38000

Red Sea 40000

Brackish water has a salinity between freshwater and seawater. The typical

salinity of brackish water is between 1000 mg/l to 10,000 mg/l of total

dissolved solids (Buros, 1980). If an appropriate desalination scheme is

adopted, brackish water can present an economic and reliable fresh water

supply for many remote areas lacking conventional water supplies.

2.4.2. Sources of Brackish Water

A brackish aquifer is a main brackish water source. It is a geologic deposit

of water-bearing permeable rock or unconsolidated materials from which

brackish groundwater can be usefully extracted using a well. The processes that

generate brackish groundwater depend on the site-specific hydrogeology and

geochemistry. In some cases, high levels of dissolved solids are derived from

19

the presence of connate water (i.e., seawater trapped at the time of original

deposition), but in most inland brackish water systems these original solutes

have long since been flushed away.

Coastal aquifers form another class of natural brackish water created from

mixing of groundwater that is discharging to the ocean. Under natural

conditions most groundwater in coastal areas discharges directly to the ocean.

Brackish water from irrigation return flows can also be utilized as

desalination source water, although the quantity and quality typically vary by

season and region.

2.4.3. Pretreatment of Brackish Water Prior to Desalination

Unlike seawater desalination, the treatments of brackish water require

minimal pretreatment to remove particulates. This may attributed to the fact

that brackish water typically contains very low concentrations of suspended

solids and organic matter. However, in case of using RO desalination

technology, brackish groundwater may require pretreatment to remove

constituents such as manganese, sulfides and dissolved iron which, if oxidized,

can cause fouling of the RO membranes.

2.4.4. Desalination of Brackish Groundwater in Egypt

Brackish groundwater desalination in Egypt has a great potential with

respect to the availability of the resource. All major aquifer systems in Egypt

contain considerable volumes of brackish groundwater (Allam et al., 2002).

Figure 2.6 shows the distribution of the main aquifer systems in Egypt (Attia).

And Table 2.6 shows location, area, salinity and the exploitable volumes of the

main brackish aquifers (Allam et al., 2002).

20

Figure 2.6. Distribution of the main aquifer systems in Egypt (Attia)

Table 2.6. Exploitable Volumes of Brackish Groundwater (Allam et al., 2002)

Aquifer Location Area

(km2)

Salinity

(mg/l TDS)

Exploitable

volume

(billion m3)

Coastal aquifers Coastal dunes

Fluviatil of wadis

Calcarenites

Shallow marine sands

20,000 >2,000 <2

Nile Aquifers Fringes

North coast

>1,500 4

El Moghra aquifer West of Nile Delta 10,000 >3,000

Nubian Sandstone Eastern Desert

Sinai

100,000 1,500-3,500 >100

Fissured carbonate

aquifer

Western Desert

Eastern Desert

500,000 5

Sinai is rich in brackish groundwater through deep seated aquifers. The

thickness of the aquifers varies between 30 to 500 m with a salinity varying

from 2,000 ppm up to 9,000 ppm. The current total deep groundwater

21

extraction from the different aquifer systems in Sinai Peninsula is 3.199 million

m3/year where 1.89 million m

3/year is used in agriculture and the rest is used

for domestic and industrial uses. The investigations in south Sinai have

identified several shallow and deep reservoirs but of limited potential for

development (Allam et al., 2002). Table 2.7 shows the brackish water resources

in Sinai and the numbers of wells constructed as well as the range of salinity of

the obtained water. Table 2.8 gives details about the desalination units in Sinai.

Table 2.7. Brackish Water Resources in Sinai (Abou Rayan et al., 2001)

City Number of

wells

Approximate

depth (m)

Capacity,

(m3/d)

Salinity

(mg/l)

El-Arish 50 40-60 52,000 3000-5500

El-Hasana 12 12-1000 6,250 1800-5000

Nakhl 7 17-1200 3,600 1800-3000

El-Quseima Spring - 1,440 1200

Sheikh Zuwayid 25 30-38 5,000 1200-4000

Rafah 35 35-90 10,000 2700-3000

Table 2.8. Desalinated Brackish Water (modified after Abou Rayan et al., 2001)

City No. of

units

Capacity

(m3/d)

Process

El-Arish 7 2800 ED

El-Hasna 1 300 ED

Nakhl 2 200 RO

El-Kuntilla 1 150 RO

Abu Aweigila 1 100 RO

El-Monbateh 1 200 RO

El-Meswateyya 1 200 RO

Bir-Beda 1 200 RO

2.4.5. Energy Consumption of Brackish Water Desalination (BWD) and

Typical Recovery Rates

The energy consumption is a major concern that must be taken into

consideration when planning and designing of a brackish water desalination

plant. Table 2.9 shows an estimated range of energy consumption for different

desalination technologies. On the other side, system recovery should be

optimized to balance productivity, energy consumption, membrane life, fouling

and cost. Table 2.10 gives the typical values of recovery rates for different

brackish water desalination schemes.

22

Table 2.9. Typical Electicity Consumption for BWD Schemes (Talaat et al., 2002)

Desalination technology Energy consumption

KWh/m3

RO

Low salinity (<1000 mg/l)

Medium salinity (1000-3000mg/l)

High salinity (3000-5000mg/l)

0.5-0.6

1.0-1.5

2.2-2.5

EDR

Low salinity BW

Medium salinity BW

High salinity BW

0.4-0.6

0.8-2

2.2-3.3

IE 0.3-0.4

VC 10.0-12.0

Table 2.10. Typical Recovery Values for BWD Schemes (Talaat et al., 2002)

Desalination technology Recovery rate

%

RO

Low salinity (<1000 mg/l)

Medium salinity (1000-3000mg/l)

High salinity (3000-5000mg/l)

80-90

65-75

50-60

EDR

Low salinity BW

Medium salinity BW

High salinity BW

80-90

65-75

50-60

IE 90-95

VC 30-40

2.4.6. Problems Associated with Brackish Water Desalination

Technological problems should be considered as well when planning for

large-scale brackish water desalination in order to maximize the production of

product water and ensure the sustainability of the production. Table 2.11

suggests the common problems encountered in brackish water desalination

(Talaat et al., 2002).

23

Table 2.11. Common Problems Encountered in Brackish Water Desalination

(Talaat et al., 2002) Technology Major problems encountered

Reverse

osmosis (RO)

Membrane fouling due to improper pretreatment. Fouling materials:

organics, iron, manganese, heavy metals, hardness causing salts.

Membrane deterioration by chemical attack (due to improper

pretreatment e.g. attack by chlorine, hydrogen sulphide).

Membrane compaction (improper operation due to frequent

pressurization - depressurization of membrane).

Membrane clogging (improper pretreatment due to hardness causing

salts).

Flux decline with time (loss of productivity).

Produced water quality decline with time (along membrane life-

time).

High pressure pumps failure (improper operation & maintenance).

Electrodailysis

reversal (EDR)

Membrane fouling

Electrode corrosion

Membrane deterioration due to improper operation

Produced water quality decline (improper operation and membrane

deterioration along membrane life-time)

Clogging by hardness causing salts (improper pretreatment)

Ion-exchange

(IE)

Resin fouling (improper pretreatment by foulants e.g. iron &

manganese and heavy metals).

Resin deterioration by chemical attacks (e.g. chlorine, hydrogen

sulphide, oxidizing agents).

Loss of resin activity (along resin life-time & hence decline of resin

capacity).

Vapor

compression

(VC)

Clogging by scales (improper pretreatment, insufficient cleaning).

Loss of productivity due to fouling of heat transfer surfaces (mainly

due to scale deposition)

Corrosion problems (improper materials selection/improper

pretreatment)

Failure of mechanical parts (e.g. blower in mechanical vapour

compression systems due to improper maintenance).

2.5. Review of Desalination Technologies

2.5.1. Elements for Desalination Process

Desalination technologies and their applications have evolved greatly

during the past 50 years. There are 5 main key elements for the desalination

process for either seawater or brackish desalination, which can be described as

follows: (Figure 2.7)

Intakes: which are the structures that are used to extract water from the

source whether brackish water or seawater and convey it to the process

system;

Pretreatment: the process of removing suspended solids and control

biological growth, for the preparation of water for further processing;

24

Desalination: the process of removing dissolved solids, mainly salts and

other inorganic constituents from water;

Post-treatment: adding chemicals to the product water to prevent any

downstream piping corrosion; and

Reject brine management and disposal: the handling of the concentrate

or reuse of waste residuals from the desalination system.

Figure 2.7 Key Elements of a Desalination System (shown in figure:

membrane-based system) modified after (Buros et al. 1980).

2.5.2. Types of Desalination Plants

Desalination plants can be categorized into two main types. The first type

involves a phase-change during the process of separation of salts from water;

while the second type of desalination plants do not involve phase change. In

such plants extraction of salts takes place while the solution remains in liquid

phase. The first type is known as thermal desalination or thermal distillation

and the second is known as membrane desalination.

Ion exchange is also another desalination technology which is mainly used

for water softening and demineralization. It is known as chemical approach

technology and usually used as a polishing step following another desalting

process of the two mentioned earlier. Therefore, ion exchange is considered

impractical for desalination of water with high level of dissolved solids.

25

Figure 2.8 shows the global evolution of the two main desalination

technologies (i.e. membrane and thermal technologies) over the past six

decades. A detailed review of the two technologies is presented in the

following sections

Figure 2.8. Cumulative global capacity of installed desalination plants for

thermal and membrane technology. Thermal technology includes MED, MSF

and MVC. Membrane technology includes RO, ED and EDR. Points reflect

current online (or presumed online) capacity of both technologies. (GWI,

2006b)

2.5.3. Membrane Desalination Processes

Membrane technologies can be used not only for desalting brackish water

and seawater sources but also for treating wastewater in reuse and recycling

applications, because of their ability to provide removal of non-salinity

contaminants (e.g., organic contaminants, bacteria, and viruses). Membrane

processes use either pressure-driven or electrical-driven technologies. Pressure-

driven membrane technologies include Reverse Osmosis (RO) whereas

Electrodialysis (ED) and Electrodialysis Reversal (EDR) are electrical-driven

technologies. In recent years, more new membrane desalination capacity is

added annually than distillation capacity as shown in Figure 2.8. Until 2006,

26

membrane desalination accounted for 56 percent of the online capacity for

desalination worldwide.

2.5.3.a. Reverse Osmosis (RO)

Reverse Osmosis (RO) is a physical process that uses the osmosis

phenomenon, i.e., the osmotic pressure difference between the saltwater and

the pure water to remove salts from water. It is the most commonly used

method of membrane desalination. In this process, a pressure greater than the

osmotic pressure is applied on saltwater (feedwater) to reverse the flow, which

results in pure water (freshwater) passing through the synthetic membrane

pores separated from the salt and a concentrated salt solution is retained for

disposal. Figure 2.9 shows a typical schematic diagram for the reverse osmosis

process.

Figure 2.9. The mechanism of the osmosis and the reverse osmosis (RO)

processes

27

2.5.3.b. Electrodialysis (ED) and Electrodialysis Reversal (EDR)

Electrodialysis (ED) processes use ion-selective membranes and an

electrical potential driving force to separate ionic species from water. Ionic

species are driven through cation- and anion-specific membranes in response to

the electrical potential gradient while the ion-depleted water passes between the

membranes. The EDR process is similar to the ED process, except that it also

uses periodic reversal of polarity to effectively reduce and minimize scaling

and fouling, thus allowing the system to operate at comparatively higher

recoveries. By reversing the electrical current and exchanging the fresh

(product) water and the concentrate (brine) streams within the membrane stack

several times per hour, fouling and scaling constituents that build up on the ED

membranes in one cycle are washed out in the next cycle. EDR has a higher

recovery rate (up to 94%) because of the feedwater circulation within the

system and alternating polarity (Younos and Tulou, 2005).

Figure 2.10. Typical arrangement of an electrodialysis membranes

2.5.3.c. Recovery Rates of Membrane Desalination Processes

Typically, 35 to 60 percent of the seawater fed into a membrane process is

recovered as product water. For brackish water desalination, water recovery

can range from 50 to 90 percent, depending on initial salinity and the presence

of sparingly soluble salts and silica, although recovery is typically between 60

and 85 percent (Sethi et al., 2006a).

28

2.5.4. Thermal Desalination Processes

The basic concept of the thermal desalination is to heat the saline solution until

water vapor is generated. As the vapor is allowed to condense on cool surface

and liquid water containing very little of the original salt is produced. The three

major thermal processes are Multi-stage Flash (MSF) distillation, Multiple

Effect Distillation (MED), and Vapor Compression (VC).

2.5.4.a. Multi-stage Flash (MSF) distillation

MSF uses a series of chambers, or stages, each with successively lower

temperature and pressure, to rapidly vaporize (flash) water from the bulk

liquid. The vapor is then condensed by tubes of the inflowing feedwater,

thereby recovering energy from the heat of condensation. MSF units are widely

used in Middle East and they account for over 40% of the world's desalination

capacity (El-Sadek, 2010) It is worth noting that one of the largest thermal

MSF plants in Egypt is located on the northern coast in Marsa Matrouh with a

capacity of 2,000m3/d (Khalil, 2004)

Over the past 50 years, the per unit cost of desalination using multi-stage

flash (MSF), the desalination technology that has been used for centuries and

economically suitable for capacities of more than 3,000m3/d/unit (Khalil, 2004)

and has decreased by an average of 44 percent per decade (El-Sadek, 2010).

2.5.4.b. Multiple Effect Distillation (MED)

MED is a thin-film evaporation approach, where the vapor produced by

one chamber subsequently condenses in the next chamber, which exists at a

lower temperature and pressure, providing additional heat for vaporization.

2.5.4.c. Vapor Compression (VC)

VC is an evaporative process where vapor from the evaporator is

compressed and its heat used for subsequent evaporation of feedwater in the

29

same tank of water that produced it thus allows heat recycling in a single-effect

distillation process.

In Thermal Vapour Compression, the compressor is driven by steam, and

such systems are popular for medium-scale desalination because they are

simple, in comparison to MSF.

In Mechanical Vapour Compression, the compressor is driven by a diesel

engine or electric motor.

2.5.5. Ion Exchange Desalination Processes

The ion-exchange system can be described as the interchange of ions

between a solid phase and a liquid phase surrounding the solid. Chemical resins

(solid phase) are designed to exchange their ions with feedwater (liquid phase)

ions, which purify the water. Resins can be made using naturally-occurring

inorganic materials (such as zeolites) or synthetic materials (Younos and

Tulou, 2005). Ion exchange is mainly used for water softening and

demineralization, and applications of ion exchange at the municipal level are

limited. Compared to other desalination technologies, this process makes

economic sense only where there is a small amount of salt to be removed from

the water. Thus, the main application of ion exchange is the production of

ultrapure water as the removal of 1 pound of salt takes about 1.5 pounds of acid

and 1.5 pounds of base to regenerate the exchangers (Xu, 2005)

2.5.6. Hybrid Desalination Processes

Hybrid desalination configurations include combinations of processes

designed to improve process efficiency or reduce energy costs. Hybrid thermal-

membrane facilities incorporate both thermal and membrane desalting

processes that are typically co-located with a power plant to improve overall

process economics.

30

Hybrid desalination facilities may also integrate multiple processes in

series to increase the separation or concentration capabilities of the facility.

These series hybrids are typically smaller in capacity.

Figure 2.11 gives a range of saline water concentration and the appropriate

desalination method to be used.

Figure 2.11. Range of applicability of different desalination technologies in

terms of salt concentration in water (www.lenntech.com accessed: March,

2012)

2.5.7. Energy Consumption of Different Desalination Technologies

Energy consumption in membrane process (i.e. RO and ED) for brackish

and low salinity water is much lower than in thermal distillation processes.

Recent innovations in RO have reduced the energy consumption further.

However, without detailed information on site conditions and the specific

application, one cannot make a clear general statement that a specific

desalination technology whether membrane or thermal is better than the other.

In general, thermal systems are robust and have high tolerance for variable

feedwater quality, while membrane systems have lower capital and energy

costs but are sensitive to fouling (Kennedy M. D. et al.). Figure 2.12 shows a

comparison of energy consumption for brackish water desalination using RO or

EDR.

31

Figure 2.12 Comparison of Energy Consumption by Process for the

Desalination of Brackish Feedwater across a Range of TDS Concentration

(USBR, 2003).

2.5.8. Renewable Energy for Desalination

The integration of renewable energy and desalination systems holds great

promise for increasing water supplies in water scarce regions. Renewable

energies can power desalination plants through solar or wind energy (Tzen,

2005). Figure 2.13 shows the distribution of renewable energy sources

desalination units.

Figure 2.13. Distribution of renewable-powered desalination technologies

(Tzen, 2005)

An effective integration of these technologies is the combination of

photovoltaics with reverse osmosis (PV-RO). As evidenced in Figure 2.13.

32

photovoltaic-powered RO units make up approximately 32% of total renewable

energy sources desalination facilities. PV is highly reliable and is often chosen

because it offers the lowest life-cycle cost, especially for applications requiring

less than 10 kW (Thomson, 2003).

RO desalination has several advantages in using solar energy (i.e. PV-RO) over

solar MSF desalination (El-Kady and EL-Shibini, 2001):

1. The RO process requires one source of energy (electricity) while MSF

needs two sources of energy — electricity for pumping system and

thermal.

2. RO is a one-phase desalination process, while the MSF process has two

phases.

3. RO requires less energy than MSF.

4. An RO plant is made up of modules, which is easy to install, maintain,

operate, and requires little space.

5. The PV cells can be installed on the roof of the RO building, i.e., no

additional area is required for PV panels.

6. PV cells are modular, easy to install, with low maintenance costs.

7. PV cells operate well in arid areas, produce direct current to drive DC

motors, and are independent of the main electricity power supply.

The disadvantages of the PV–RO system are replacement of the RO

membrane every 3 years and replacement of batteries in the PV storage system

every 7 years of operation (El-Kady and EL-Shibini, 2001). Battery lifetime in

PV systems in central Europe is typically 3 to 8 years, but in hot countries, this

reduces to typically 2 to 6 years, since high ambient temperature dramatically

increases the rate of internal corrosion (Thomson, 2003). However, Thomson

(2003) presented the design and testing of the batteryless approach of PV-RO

desalination plants and provided estimates of performance and capital costs of

such systems.

33

Assimacopoulos et al. (2001) and, in broad agreement with the

comprehensive studies, states:

“PV-RO (Photovoltaic powered-Reverse Osmosis) is clearly the favoured

desalination combination for small stand-alone systems”.

2.6. Reject Brine Disposal Methods

Development of desalination has brought to concern the suitable

environmental disposal method of the byproduct of the desalting process. The

disposal of the reject brine is dependable on the location of the desalination

plant, whether the desalination plant is near to the coastal shores or established

in an inland area remote from the coastal areas. Regardless of the desalination

technology used, saline water is separated into two streams after the

desalination process: a freshwater stream with a low salt concentration and

brine or concentrate stream with a high salt concentration that needs to be

disposed.

Desalination plants near the coastal shores usually dispose the concentrate

in the seas or oceans. Accordingly, the effect of the ocean disposal is negligible

because of the minute volume of concentrate compared to the receiving water

bodies. However, the promulgation of more and more stringent environmental

protection regulations will increasingly reduce this opportunity. The negative

influences of the discharged brine may not only damage the environment or

reduce public acceptance, but can also result in financial penalties if toxicity

standards are not met. Macedonio et al. (2011) stated some possible measures

to mitigate the environmental impacts on ocean outfalls:

Lower recovery rates and/or dilution of the brine with seawater prior to

the discharge to reduce its salinity;

Discharge devices, such as multiple port diffusers, spreading the brine

across a larger area and increasing dispersion velocity;

Discharge devices, such as multiple port diffusers, spreading the brine

across a larger area and increasing dispersion velocity;

34

Dilution of the brine with water from other processes, e.g. with cooling

water from power;

Discharge in an area with strong currents and at depths that minimize

impact on benthic life forms.

In cases where the desalination units are installed away from coastal zones

(i.e., inland desalination plants), the design has to take in consideration a safe

option for disposal without harming the environment. Nowadays, the scope is

not only considering the safe disposal of the concentrate but also taking into

account the environmental sustainability of the disposal option and achieving

an economical benefit of the concentrate.

The cost plays an important role in selecting the method of brine disposal.

It could range from 5 to 33 % of the total cost of the desalination plant. The

cost of land disposal is much higher if compared to that discharging brine into

shores (Khordagui, 1997).

There are many options that were identified for the disposal of inland

desalination plants starting with pumping into designed lined evaporation

ponds; disposal into surface water bodies; disposal into any existing municipal

sewerage system; concentration into solid salts; irrigation of plants tolerant to

high salinity levels; and injecting the brine back into deep saline aquifers

(Khordagui, 1997).

The factors that influence the choice of the suitable disposal method were

identified. These factors include the amount of the concentrate (reject brine);

the quality or constituents of concentrate; the geographical and physical

location of the discharge point of the concentrate; the availability of the site,

public acceptance; option permissibility as well as capital and operating cost of

the disposal method (Mickley et al., 1993).

It is worth noting that the chemical characteristics of the reject brine are

function of the feed water quality, desalination technology used, the chemicals

used for pre- and post treatment, and percent recovery (Mickley, 1995).

35

Usually in RO plants, filters need to be backwashed every few days to

clear the accumulation of solids. This filter backwash is not permitted to be

directly discharged to the environment, because it can cause both considerable

discoloration in the water at the discharge site and contamination. However, the

practice may occur in other locations. In addition, anti-scaling substances,

antifoaming additives, oxygen scavengers, and anticorrosion chemicals may be

present in the discharge of the concentrate (Rachid and Abdelwahab, 2005).

2.6.1. Evaporation Ponds

The use of evaporation ponds is considered one of the most widely used

disposal methods. Evaporation ponds comprise the largest portion of disposal

method in countries known for their arid or semi-arid climate conditions. Of

the attractions to use evaporation ponds, presence of high evaporation rates,

ease of construction, low land cost, low maintenance requirements, and the

absence of mechanical equipments except for the pump that conveys the

concentrate to the pond (Mickley et al., 1993).

However, a survey of drinking water desalination plants (membrane plants

of capacity 98 m3/d or more) in the continental US that included 137 plants

showed that 48% of the brine is disposed to surface water, 23% dispose to the

head-works of wastewater treatment plants, 12% utilize a land application

process, 10% dispose through deep well injection, and only 6% use evaporation

ponds (Mickley et al., 1993).

Mickley et al. (1993) also stated that in parallel to the advantages and

attractions to the use of evaporation ponds, there are a lot of disadvantages that

sometimes cause barriers to the utilization of evaporation ponds. For example,

the need of large areas in case of high disposal rate and/or low evaporation

rates, the need of impervious liners of clay or synthetic membrane such as PVS

or Hypalon to avoid any potential of contaminating underlying potable water

aquifers through seepage, and the requirement of level terrain and low land

costs

36

The proper sizing of evaporation pond depends on accurate estimation of

evaporation rate. Pond sizing include two outputs: the surface area of the pond

and the depth. Surface area is determined by the evaporation rate while the

calculation of the depth is based on water storage, storage capacity for salts,

surge capacity, and freeboard for rainfall and wave action (Mickley et al.,

1993).

The pond must be large enough to satisfy needs of the land area being

drained, the volume of subsurface drain water collected, and the rate of

evaporation for the served region. Ponds must have a minimum embankment

top-width of five meters; freeboard of 0.5 m or equal to maximum wave run-

up; an inside slope of 6:1 (h:v) and outside slope of 2:1(h:v); and a foundation

stripped of all vegetation. Internal dikes may be constructed to create cells

within the pond and to allow transfer from cell to cell and disposition of salts in

a progressive evaporation sequence (Tanji et al., 1985).

As the salinity of water affects the evaporation rate, it has been suggested

the use of an evaporation factor of 0.7 for multiplying by the calculated solar

evaporation rate to account for the effect of salinity (Mickley et al., 1993).

The availability of water, salt, solar radiation and flat land open the doors

for the use of solar ponds as an attractive source of renewable energy (Burston

and Akbarzadeh, 1995). The effectiveness for generating electricity for solar

ponds requires: all-year solar exposure, large volumes of brine, as well as an

adequate source of “fresher” water, low cost flat land of low permeability,

distant from shallow aquifers, relatively low winds, and a consistent electricity

demand (Ahmed et al. 2000). The deserts of Egypt are suitable for using solar

ponds as most of the locations meet the requirements for an effective utilization

of solar ponds.

When designing evaporation ponds it is preferred to use small ponds

connected by pipelines than to design large ponds. Smaller ponds are easy to

manage especially in windy weathers where the generated waves can damage

37

the banks of pond requiring costly maintenance (Ahmed et al., 2000). Also, the

use of smaller ponds allow ease of operation during periods where there is a

decrease in disposal rate, and thus less operator’s attention is required for the

smaller pond than a large pond.

2.6.1.a. Enhancements of Evaporation Ponds

Evaporation ponds are also taken as a great opportunity to develop

resource recovery measures such as aquaculture, brine shrimp, beta-carotene

production, harvesting of salts, and as a solar ponds for electricity generation

(Ahmed et al., 2000). The use of evaporation bonds as solar ponds will allow

the integrity of the system as using the generated electricity as a supplementary

source of power, thus relieving the pressure on the current distressed power

sources.

To overcome the drawback of the low evaporation rates and increase the

efficiency of evaporation, wet surfaces (capillaries or clothes) exposed to wind

actions can be used where surface density can be high enough to generate a

reasonable evaporation flow. Thus, the surface would be wetted by capillarity

effect and the water evaporates leaving solids of the brine crystallize on the

surfaces. The final solid waste could then be properly managed by an

authorized company or even could be reused (Arnal et al., 2005)

Wind-Aided Intensified eVaporation (WAIV) is an enhanced evaporation

technology particularly used for reverse osmosis (RO) concentrate to increase

natural evaporation. The technology exploits wind energy to evaporate wetted

surfaces which are packed in high density per footprint. These surfaces are

deployed in arrays with large lateral dimensions and significant heights of

about three or four meters (Gilron et al., 2003). Wetting of the surface is

occurred by a pump that brings the rejected brine from a pond or from a storage

tank to a distribution network on top of the vertical surfaces from which the

vertical surfaces are fed by gravity. The driving power of wind drives away

excess humidity from the vertical installed surfaces raising the magnitude of

38

evaporation of a factor of 15-20 times compared to the conventional

evaporation ponds, then excess brine is allowed to return to pond by gravity

through an impervious surfaces (concrete) sloping back to the pond.

There are many advantages for the WAIV technology including

minimizing the land use of evaporation ponds, increasing the efficiency of

evaporation, simple and yet reliable, ideal for inland desalination plants,

reduction of up to 50% in treatment costs of RO rejected waste and it can be

used in minerals harvesting where product is commercially valuable (Lesico

CleanTech).

Mahmoud (2011) introduced the concept of a two-cell evaporation pond

tends to store water in the buffering cell in order to use it in maintaining

minimum water level instead of using supplementary fresh water to feed the

pond. The pond is divided into a buffering cell and an evaporation cell. The

concept was adopted in order to avoid that tight corner of choosing between the

pond area large enough to avoid overtopping and not too large to avoid draught

during hot summers.

He wrote a computer model in BASIC language to simulate the pond

performance. The model calculates the pond water depth, salinity and

efficiency. The output is introduced in both forms of tabulated data and

graphics. The model is run with real data of the closed basin of “Toshka

Project” in Egypt as a case study. Sensitivity analysis shows that the pond

efficiency is very sensitive to its size and less sensitive to drainage water

salinity, while pond depth has very small effect. The results indicate that there

is significant effect of the salinity increase due to salt accumulation on

evaporation rates and pond size. The pond size is selected with the maximum

efficiency resulting from one hundred year routing. Different scenarios show

that the use of the proposed design of a two-cell pond tends to increase the

overall efficiency from 68-80% for one-cell pond to 97-100% for two-cell

pond.

39

2.6.2. Disposal in Municipal Sewerage Systems

The disposal of brine in municipal sewerage systems is used in many small

RO plants. This process has the advantage of lowering the BOD of the

domestic sewage. Nonetheless, TDS increases and may have some effects on

the microorganisms of the system and may make the treated effluent unsuitable

for irrigation purposes. In addition, the design capacity of the existing sewerage

system may not be able to accommodate the increase in discharge (Ahmed et

al., 2000).

2.6.3. Salt Production

The approach to extract all the salts from the reject brine has been taken

seriously for the advantages of being environmentally friendly and producing

commercial products. Studies have yielded the SAL-PROC technology

introduced by Geo-Processors Pty Limited. SAL-PROC is an integrated

process for sequential extraction of dissolved elements from inorganic saline

waters in the form of valuable chemical products in crystalline, slurry and

liquid forms. The mechanism of the process involves multiple evaporation

and/or cooling, supplemented by mineral and chemical processing adding to

this that no hazardous chemical is used in the process. The technology is based

on simple closed processing and fluids flow circuits which enables utilization

of inorganic saline waters to extract a group of valuable chemicals (Ahmed et

al., 2002).

40

Figure 2.14. A Typical SAL-PROC process

There is a wide range of different product streams that can be produced

from SAL-PROC depending on the chemical composition of the saline waters.

As shown in Figure 2.14, sodium chloride salt is only one of a range of the

chemical products of commercial value which can be processed from using the

technology. The recovered chemical products are of high quality and in

demand by different industries (Ahmed et al., 2002). However, economic

success depends not only on the technology to bring about selective salt

recovery but also on the local salt market and the challenges associated with

marketing each salt (Mickley, 2010).

Application areas for the SAL-PROC products are identified by the market

as follows (Ahmed et al., 2002):

Feedstock, fillers, reagents, coating material and supplements for:

o Animal dietary needs

o Fire retardants

o Manufacture of magnesium metal

o Manufacture of light-weighted and fire-proof plaster boards and

other building products

o Manufacture of salt-tolerant building footing, wall panels and

other construction products

o Application in tanneries

o Production of quality paper products

o Manufacture of plastics, paint, ink, and sealant products

o Soil conditioners for remediation of sodic and acidic soils

o Sealants for irrigation channels and earthen ponds

41

o Stabilizers for road base construction

o Dust suppressant

o Flocculating agents for water/wastewater treatment

Ahmed et al. (2002) performed a desktop pre-feasibility study on reject

brine from Petroleum Development Oman (PDO) desalination plants indicated

that various types of salts including gypsum, sodium chloride, magnesium

hydroxide, calcium carbonate, sodium sulphate and calcium chloride can be

produced. The study showed that by processing 405 ML of reject brine per year

from the PDO desalination plants, it is possible to produce commercial salts

worth US $895,000 annually.

Depending on the chemical composition of the saline feed water the

process route may involve one or more steps of reaction and evaporation and/or

cooling supplemented by conventional mineral and chemical processing steps.

From the studies on the PDO-RO desalination plants, there were three proposed

process routes for the treatment of the rejected brine. Figure 2.15 schematically

shows a comprehensive overview for the three process routes (Ahmed et al.,

2002).

The use of waste effluent as a resource is the main attractiveness of the

SAL-PROC technology as well as the low-value chemicals as the reagents for

the recovery of saleable chemical products which offer higher return from their

sale, surpassing the cost of the operation. Mostly the treatment facility is of low

electricity, where the main usage of electricity would be for the operation of

pumps and agitators in the chemical reactors and fluid transfer circuits (Ahmed

et al., 2002)

42

Figure 2.15. Proposed Process Routes for the Treatment of Reject Brines Generated by PDO-Operated RO Desalination Plants

(Ahmed et al., 2002)

43

2.6.4. Deep Well Injection

Deep well injection is presently applied worldwide for disposal of

industrial, municipal and liquid hazardous wastes. In recent years this

technology has been given serious consideration as an option for brine disposal

from land based desalination plants. For example, for the United States, it’s

considered the third most used method for brine management comprising 17

percent of the disposal methods. It comes after disposal into surface water

which covers 41 percent and disposal into sewers with 31 percent (Mickley,

2006). Of the extremely important aspects in the design of injection wells, is

the site selection, which is dependent upon geological and hydrogeological

conditions. It is worth noting that injection wells should not be located in areas

vulnerable to earthquakes or regions with mineral resources (Ahmed et al.,

2000).

A typical waste injection well injects the waste liquids at depths ranging

from a few hundred to a few thousand feet. Low cost when compared to the

alternatives of landfilling and chemical treatment (often costing 80% less), and

relatively high success rate are among the reasons for the earlier growth of

deep-well injection as a waste disposal option (Lehr, 1986).

Deep-well injection is typically employed for larger desalination plants

(e.g., > 3,800 m3/day) because the costs for developing deep-injection wells

are relatively high and are not largely reduced for smaller flows. For example,

the typical capital cost of a 3,000-m-deep well is reported at $8.1 million for a

concentrate flow of 3,800 m3/day, which decreases to only about $5.1 million

for a concentrate flow of either 380 or 38 m3/day (Malmrose et al., 2004).

These costs exclude any pretreatment or standby disposal system. While capital

costs for well injection are about average of typical inland concentrate

management methods, the annual operating costs are relatively low as a

percentage of total operating costs (Mickley, 2006).

44

Stephen (1986) made some modifications to the existing techniques for

impact assessment calculations and presented several management tools that

can be useful in assessing the environmental impacts resulting from saltwater

disposal injection wells. He stated the most useful of these tools, which

include: (a) calculations of radius of endangering influence using Cooper-Jacob

method and Theis non-equilibrium equation; (b) potentiometric head contour

maps; (c) formation hydraulic transmitting properties; and (d) water quality.

Whereas his modifications included: (a) estimates of radius of endangering

influence that require observed initial hydrostatic heads and aquifer hydraulic

transmitting properties for the injection interval; and (b) the geochemical

characterization of nearby suspected ground water contamination using all

major ion concentrations in a trilinear diagram of water quality analysis,

instead of using only chloride as a brine tracer.

“Once an aquifer is contaminated, these chloride-rich brines are not easily or

inexpensively removed.” –Stephen (1986)

Underground migration of injection fluids possible pathways have been

discussed by Canter, the Environmental Protection Agency, and Fryberger and

Tinlin. These include: (a) corroded or improperly plugged injection wells

where the intended receiving interval or adjacent saline aquifers are

hydraulically connected to freshwater geological horizons; (b) abandoned

exploration wells located within the radius of endangering influence created

from nearby active injection wells; (c) fracturing of geologic units resulting in

the hydraulic interconnection of the injection horizon, adjacent saline aquifers,

and/or freshwater aquifers; or (d) different combinations of the above.

Saripalli et al. (2000) gave a variety of processes that contribute to the

reduction in permeability of the host formation or the perforations or screens

that are placed in the well’s injection interval. These processes include

particle/colloid migration into the formation, bacterial growth, emulsification

of fluids, and precipitation of dissolved matter, flow of unconsolidated sands

into well bores, scale formation and entrapment of gases.

45

He also introduced a measure of the effects of plugging and damage to

subterranean formations on injection well performance. It was expressed by

injectivity (I) which is defined as the ratio of injection rate (q) to the difference

between well flowing pressure (Pwf) and the average formation pressure (Pr)

given by the following equation.

rwf PP

qI

(2.1)

Several factors affect the injectivity, which include the physical and

chemical quality of the injected fluid, injection rate and pressure, as well as the

nature and physical properties of subterranean strata. It’s worth noting that one

of the most important constraints on stable injectivity is the presence of

suspended solids in the injection fluid. High TSS, low injection rate, low

injection pressure, and low porosity and permeability of the well strata all lead

to rapid well plugging and diminished injectivity (Saripalli et al., 2000).

Saripalli et al. (2000) also introduced the half-life concept of an injection

well which is defined as the time required for its injectivity to decline to half its

initial value. This is a good indicator of the well performance.

Two relevant questions that engineers must answer while designing and

maintaining a deep-well injection facility are: (1) what is the water quality

(TSS) criterion to be imposed on the influent waste streams to ensure a given

injector half-life, and, conversely (2) given a certain influent waste stream

quality, what is the expected half-life of the injection well?

He stated that, as the plugging of an injection well occurs and the

formation gets worse, the need of larger injection pressures becomes crucial to

maintain a given flow rate, which can lead to well failure, causing the spread of

contamination and compromising safety. In addition, the build-up of high sub-

surface pressures can cause the fracturing of confining strata and create

pathways for the vertical migration of injected fluid.

46

The tubing-and-packer design considerations were introduced by Shekan

and Kwiatkowski (2000). The tubing and packer assembly is installed inside

the final cemented casing of the injection well. Other design considerations

include compatibility of the concentrate with the tubing material (corrosion

potential), anticipated permeate and concentrate flows, tubing diameter

selection in the retrofit of existing deep injection wells, and annular monitoring

systems for leak detection.

Shahatto (2003) investigated the effect of abstracting brackish water for

desalination and subsequent brine injection into the same coastal aquifer. A 2D

vertical model was built for a cross section perpendicular to the shore line. For

all simulations, the code ROCKFLOW has been used. The results of the

simulation scenarios showed that the injected brine sinks faster than it can be

transported horizontally by the groundwater flow and could form a salty lens at

the aquifer bottom around the injection well or above a low permeability layer.

Also, when the desalination plant is shutdown and in the absence of pumping

and injection, the formed salty plume around the injection well during the

operation of desalination plant is spread to the sea with time. The remediation

could happen by extracting of this brine followed by mixing with less saline

water (dilution) and re-injection in the aquifer or disposal into the sea at an

appropriate distance from the reef.

Williams and Feeney (2003) suggested a hypothetical system of discharge

and recharge wells. The main problem of case study was that disposing reject

brine to the Pacific Ocean was difficult to permit. The marine sanctuary

restricts any discharges into the ocean that may injure a sanctuary resource.

Because a change in seawater salinity may injure plants and animals in the

marine sanctuary, brine disposal is effectively prohibited. So, they took two

actions. First, the desalination plant is operating such that the brine is the same

chemistry as seawater. This was achieved by pumping a mixture of freshwater

from onshore and seawater from offshore as feedwater to the desalination unit.

Second, they designed a system with two production wells and a horizontal

47

brine injection well such that some of the brine is recirculated back into the two

production wells. This system was designed using SEAWAT (Gue and

Langevin, 2002).

Nassar (2007) developed a methodology in assessing the environmental

impacts of desalination plants discharging brine through injection into

underground. He used laboratory and computational methods to simulate the

phenomena of subsurface brine disposal by injection. He prepared a setup for a

seepage tank of dimensions 1.42m long, 0.1m wide, and 0.6m height with two

well configurations to represent to a 2D flow in the vertical plane

experimentally. SEAWAT was used for building the computational model. The

results were used for calibrating the computational model simulated.

Preliminary design charts for the management of the production and injections

well fields for the desalination plants were created in terms of four design

parameters, which are relative salt concentration (RSC); production and

injection rates (Qd, Qr), well spacing (S) and simulation period (T). Also the

study showed that on the long run, the injection well will affect the salinity of

the production well and it was shown through the developed design charts that

assess the time-varying effect for the different design scenarios.

Navarro and Carbonell (2007) have studied the contamination of

groundwater in a specific aquifer in Spain that was caused by the disposal of

some byproducts. These byproducts are resulting from the extraction of dry raw

materials in quarries. The migration of pollutant is resulting from the

fluctuation in the potentiometric head. The hydrogeochemical processes

associated with uncontrolled waste disposal in these landfill areas were studied

along a flow path that crosses the contaminated area. A transport model was

developed to study te reactions associated with the different mineral phases

through inverse modeling. This transport model was used also to simulate the

dilution phenomenon associated with the pollution after the potential removal

of the sources of contamination.

48

Ali (2009) studied the environmental impacts of the disposal of gas

production by-products waste water in deep aquifers. He studied the

sustainability of the proposed option of Mansoura Petroleum Company, a gas

and oil production company with a large concession area in the Nile Delta

around Mansoura city, of using a flooded gas production well for the disposal

of water resulting from the gas extraction process for a timeframe of fifty years.

He used groundwater flow modeling to obtain the flow pattern for the area and

developed a contaminant transport model to predict the extent of contamination

resulting from the disposal of produced water into El-Wastani formation. Ali

(2009) assumed a single injection well with a rate of 220m3/day for 50 years,

and he showed that the 10% relative contamination line will migrate vertically

upward into the main water bearing formation (Meet Ghamr) for a distance of

about 150m while the lateral extent is about 1800 m. He also introduced

multiple future scenarios for a conservative approach for the assessment; these

scenarios indicated negligible impacts on the top 300-400 m of groundwater

aquifer around the disposal well.

He also recommended that such disposal should be accompanied with a

system of monitoring wells that should be placed down gradient and regularly

monitored.

2.6.5. Aquaculture

Aquaculture is a growing industry. Moderately saline effluent can be used

to culture fish though rigorous monitoring is required. Species reported to grow

well in high salinity water in Australia are brine shrimp (Artemia salina),

Barramundi (Lates calcarifer), Black Bream (Acanthopagrux butcheri), Red

napper (Pagrus auratus), Milk Fish (Chanos chanos), Mullet (Mugil cephulux)

and Tilapia (Oreochromis mossambicus). Having brine shrimp production

downstream from finfish culture has advantages in that the brine shrimp utilize

nutrients generated from fish culture, while providing food for fish fry (Ahmed

et al.). It is worth noting that brine shrimp have the ability to live in water of

49

very high salinity levels and can tolerate varying levels of salinity from 5 g/L

up to 250 g/L (Daintith, 1996).

Evaporation basins can also provide a foundation for algae production.

Dunaliella salina grows and produces commercial grades of beta-carotene at

salinities greater than 200 g/L. Other species of salt tolerant ‘algae’ (more

correctly, a blue green bacteria) may also have commercial application (Ahmed

et al. 2000).

51

CHAPTER THREE

ASSESSMENT OF EVAPORATION POND DISPOSAL

ALTERNATIVE

3.1. Introduction

As abovementioned, two main disposal alternatives will be evaluated. The

first alternative is the evaporation pond whereas; the second alternative is

injection in the deep aquifers. In this chapter, the evaporation pond alternaive is

assessed and evaluated.

In order to obtain a faire comparison between the two disposal alternatives,

an existing evaporation pond in Sinai has been selected as a case study. The

data of this evaporation pond were collected and analyzed to be used in the

evaluation of the evaporation pond alternative. The description of the existing

evaporation pond and the details of the evaluation of the pond disposal

alternative are presented below.

3.2. Case Study Description

Three desalination plants were constructed in Central Sinai under the

umbrella of the LIFE Sinai program. The three plants, located in Al-

Meswateya, Al-Monbateh and Bir Beda are now operating and they produce a

total of 600m3 of potable water per day (i.e. 200m

3/plant/day) benefiting more

than 6,000 inhabitants of Central Sinai. The locations of the three desalination

plants are shown in Figure 3.1. The plants are named based on their location.

Al-Monbateh was taken as a case study for the simulation of desalination and

the disposal of the byproduct (reject brine) of the desalination process. Figure

3.2 shows a layout for Al-Monbateh desalination plant and the disposal area.

52

Figure 3.1. The Sinai Peninsula with the Location of the Case Study

Figure 3.2. Layout For The Project Area of Al-Monbateh Desalination Plant

53

The desalination plant receives water from Al-Monbateh deep well which

taps the Lower Cretaceous aquifer of Sinai. The plant runs on two identical

reverse osmosis (RO) units with a recovery rate of 70 percent (recovery rate is

described as the percentage of the permeate water produced to the feed water).

Each unit has a production capacity of 10m3/h of fresh water over an operation

period of 10 hours per day to produce a total production of 200 m3/day. Product

water is collected in a product tank with capacity of 200 m3 from where a

product pumps transfer the water to tankers. The produced water quality is

according to the WHO drinking water standards. The reject brine produced

undergoes a further treatment in order to reduce the final brine volume for final

disposal. Two reject RO modules are provided to receive and treat the by-

product of the two main RO units. The design parameters for the main and the

reject RO units are listed in Table 3.1 and Table 3.2

Table 3.1. The Design Parameters for Each of the Main RO Units

Description Unit

Operation period 10 Hours/day

Feed water design temperature 20-30 oC

Feed water quality ~2200 mg/L TDS

Reverse Osmosis recovery 70 %

Feed water flow 14.29 m3/h

Brine flow 4.29 m3/h

RO permeate flow 10 m3/h

Treated water quality ≤500 mg/L TDS

Table 3.2. The Design Parameters for Each of the Reject RO Units

Description Unit

Operation period 10 Hours/day

Feed water design temperature 20-30 oC

Feed water quality 11176 mg/L TDS

Reverse Osmosis recovery 70 %

Feed water flow 4.29 m3/h

Brine flow 1.29 m3/h

RO permeate flow 3 m3/h

Treated water quality ~173 mg/L TDS

54

The raw water from Al-Monbateh well is branched by pipes connection to a

300m3 raw water tank which collects and stores the water before feeding the

two reverse osmosis (RO) units. The raw water is chlorinated before storage in

the raw water tank as a pretreatment step. Two filter-feed pumps (one per each

RO module) take water from raw water tank and feed it to pretreatment system.

Suspended solids are removed by the manual multimedia filter. Each RO

module has one filter. This filter is operated manually with backwashing

required once per day. One backwash assistance pump is used for

backwashing. A de-chlorination dosing set for dosing sodium meta-bisulphate

is used to remove chlorine before RO membranes.

After filtration, the filtered water complies with the guidelines for feeding

RO membranes, mainly that there shall be no chlorine present, dissolved iron

shall be less than 0.01 mg/l, and the silt density index (SDI) is less than 3. In

addition, antiscalant chemicals are added to prevent scaling of calcium

sulphates where chemical mixing takes place in-line.

Chlorination and pH adjustments dosing sets are used as a post treatment

step for the permeate water produced from each RO unit. The treated product

water is collected in the product tank with a capacity of 200m3 from where a

product pumps transfer the water to tankers to be delivered to the nearby

habitants.

Each main RO module produces 4.29 m3/h of brine (total of 8.58m3/h) at

TDS concentration of about 11,176 mg/L. The reject brine from each main RO

module is collected in a buffer tank from which is treated by a second RO unit

called Reject RO unit. Each reject RO unit produces 3 m3/h of permeate water

at TDS of 173 mg/L and a concentrate with flow of 1.29m3/h at about 35,950

mg/L. The permeate water from both reject RO units (i.e. 6 m3/h) is mixed

with approximately 1 m3/h of the final reject from both reject RO units

resulting a final solution of 7 m3/h at concentration of less than 5,000 mg/L.

This final mixture is pumped to the elevated irrigation tank to be used for

irrigation purposes. The remaining final reject (i.e. 1.6m3/h at 35,950 mg/L) is

55

disposed as final brine and is managed by evaporation using evaporation ponds.

The disposal area is divided into two identical connected lined evaporation

ponds with bottom dimensions of 44.0 m length, 33.0 m wide and 1.25 m depth

and 2:1 horizontal to vertical side slopes.

Figure 3.3 shows a schematic diagram for the operation of the Al-Monbateh

desalination plant.

56

Figure 3.3. Schematic Operation Diagram for the Al-Monbateh Desalination Plant

57

The Al-Monbateh well penetrates two of the most potential aquifers in

Sinai: the Upper Cretaceous and the Lower Cretaceous aquifers. The Lower

Cretaceous, which is the producing aquifer for the Al-Monbateh well, is

considered to be the aquifer with the greatest development potential among the

other aquifer systems in Sinai. The two aquifers are saline aquifers, and

regardless of the spatial salinity variation in both aquifers the Lower

Cretaceous has less groundwater salinity than the Upper Cretaceous in most of

the aquifers extent. Figure 3.4 shows a schematic representation of the

stratigraphic units in Central Sinai.

Figure 3.4. Typical Stratigraphic Succession in Central Sinai (Ghoubachi,

2010)

58

3.3. Description of Al-Monbateh Disposal System

The current disposal system in Al-Monbateh desalination plant consists of

two adjacent connected evaporation ponds with dimensions of 44 m length, 33

m wide and 1.25 m height each and the sides of the ponds are constructed with

a horizontal to vertical slope of 2:1 (Figure 3.5). The ponds are lined to assure

that the concentrate treatment takes place through evaporation only and to

avoid water seepage thus prevent salts from reaching to the underlying

groundwater that might cause contamination to the nearby shallow dug-wells.

The design of the ponds was based on a constant inflow rate of 1.6 m3/hr of

reject brine resulting from the desalination process of the RO units of Al-

Monbateh plant with total daily working hours of 10 hours (i.e. total daily brine

discharge is 16 m3). A constant evaporation rate of 6 mm/day was assumed and

no seepage takes place as the ponds are lined. The provided area should

provide a safe disposal for the current system under the given design criteria

and operation rules. It is worth noting that one of the obvious drawbacks of the

current disposal system is that the system has a limited operation period (e.g.,

no more than 10 hours per day – assuming the availability of electric power).

Figure 3.5. Plan and Cross-Section of the Evaporation Ponds

59

3.4. Overtopping Problem in Al-Monbateh Disposal System

During a field visit to the evaporation ponds, the operator of the plant stated

that after five months from the operation of the desalination plant, continuous

overtopping of the water above the safe allowable level was recorded. As a

quick solution, the operator of the plant has to use additional pumping system

to pump out some of the reject from the pond and unsafely dispose it few

meters away from the evaporation pond leaving the brine to spread on the

ground surface to be treated by both evaporation and seepage. This brings the

critic of constructing lined evaporation ponds to avoid contamination of the

underlying groundwater. It also refutes the design environmental requirements

of a safe disposal of the reject brine. In addition to a running cost of the

pumping unit(s) attached to the system.

Accordingly, the data of Al-Monbateh Disposal System with the current

configuration will not provide a fair comparison to the two disposal

alternatives, as the dimension of the evaporation pond was underestimated. To

overcome this issue, an additional task was taken in this research, where we

had to identify the reasons for the overtopping problem in Al-Monbateh and

obtain the proper dimension of the pond to be used in the comparison.

The main cause of overtopping problem could be one or a combination of

the following possible reasons:

1. Assumption of constant evaporation rates along the 12 months of the

year which was taken equal to 6 mm/day.

2. Ignoring the rainfall on the evaporation pond area as an additional

inflow to the ponds.

3. Neglecting the effect of salinity variation of the reject brine inside the

evaporation ponds on the evaporation rates. Further explanation for the

effect of salinity on the evaporation rates is discussed later in section

3.7.

60

4. Increasing the desalination plant operation period per day more the

designed period (i.e. 10 hours per day) resulting in an increased inflow

to the ponds.

For the comparison and assessment of the current disposal system (i.e.,

evaporation ponds) with the other disposal options to be fair and feasible, the

evaporation pond option has to safely dispose the brine alone without any

supplementary units being attached. Since the current pond cannot act solely as

a safe disposal option, there was a need to find the appropriate dimensions of

the pond that comprise a safe option for disposing the reject brine. This can be

done by identifying the main cause(s) of overtopping and taking into account

these causes then hypothetically modify the dimension of the evaporation pond

to assure safe disposal of the reject brine.

For reject brine to be safely disposed, the disposal option has to be

developed and assessed in terms of its technical viability, its effect on the

environment, and its economic feasibility.

An assessment for the current evaporation ponds was performed using a

MATLAB code which comprises both water and salt balance for the pond. The

flowchart of the code is shown Figure 3.6.

61

Figure 3.6. Flow Chart for the Assessment Process of the Evaporation Pond

3.5. Simulation Model

A MATLAB code is written to simulate pond routing that utilizes water and

salt balance. The input data are the simulation routing period in days, the

desalination plant outflow rates as the inflow rates for the ponds, the rainfall

and evaporation rates on a monthly basis, the inflow water salinity, the pond

dimensions, and the pumping characteristic if needed (i.e., if the depth of water

inside the pond increases above the maximum allowed depth inside the pond

that was taken as 1.0 m which is about 80% of the total depth of the pond). The

model outputs are the cumulative water storage, water depth, water salinity at

every time step, number of overtopping occurrence, the maximum cumulative

water storage, maximum water depth, maximum water salinity at different

pumping rates. Also, the pond efficiency which is defined as the ratio of the

summation of the periods of non-overtopping to the overall period of the

62

routing can then be calculated. The time step is chosen to be one day and the

inflow to the ponds is divided equally on both ponds.

The simulation model is used to assess the current pond dimensions taking

into consideration the reasons causing the overtopping as described in section

3.4.

Due to lack of data at Al-Monbateh desalination plant location, the actual

evaporation rates and rainfall on a monthly basis were obtained from the

nearest available five meteorological stations. The Inverse Distance method

was utilized to calculate the estimated rainfall and the evaporation rates at Al-

Monbateh location. The Inverse Distance method is a weighted average

method, where weights for each meteorological stations are inversely

proportionate to its distance from the point being estimated which is Al-

Monbateh location. The formula of the inverse distance may be written as

(Lam, 1983)

N

i i

N

i

i

i

x

d

pd

P

12

12

1

1

(3.1)

Where,

Px: estimate of rainfall/evaporation rate for the ungauged station

Pi: rainfall/evaporation rate values of rain gauges used for estimation

di: distance from each location the point being estimated

N: numbers of surrounding stations

The available data at the five stations are the monthly rainfall and the

reference evapotranspiration rates ET0. An empirical relation between the

evapotranspiration rates ET0 and the pan evaporation rates Ep is introduced as

follows:

pp EKET 0 (3.2)

where, Kp is the pan coefficient.

63

The most widely used table of Kp values to estimate ET0 from Ep is the one

provided by Doorenbos and Pruitt (1977). The table gives the Kp values for

National Weather Service (NWS) class ‘‘A’’ evaporation pans located over

grass surfaces having a range of upwind grass fetch distances as described in

Doorenbos and Pruitt (1977). More recently, Allen and Pruitt (1991) published

the original Kp values (Table 3.3) used to develop the Doorenbos and Pruitt

(1977) table. The Kp values in the table vary depending on the fetch, wind

speed, and relative humidity (Snyder et al., 2005), a constant value of 0.7 is

assumed for calculation in the current simulation model.

Table 3.3 Kp Values from Allen and Pruitt (1991) Corresponding

to Mean Relative Humidity (%) and Wind Run (km/day) Data

Relative

Humidity

(%)

Wind

Run(Km/day)

Fetch (m)

100 1 10 1,000

30 84 0.74 0.55 0.66 0.77

30 260 0.66 0.5 0.6 0.7

30 465 0.58 0.45 0.52 0.62

30 700 0.5 0.4 0.45 0.55

57 84 0.81 0.64 0.75 0.83

57 260 0.73 0.58 0.68 0.77

57 465 0.66 0.52 0.6 0.7

57 700 0.59 0.45 0.53 0.63

84 84 0.85 0.73 0.82 0.87

84 260 0.78 0.65 0.75 0.81

84 465 0.71 0.59 0.67 0.75

84 700 0.65 0.53 0.61 0.68

The values of pan evaporation rates are multiplied by a pan factor to

convert them to the equivalent lake evaporation values. Usually this

dimensionless factor is equal to 0.7 (Chow et al., 1988), but this value varies by

season and location. It will be assumed equal to 0.75 in this case.

Available Meteorological Stations

As above-mentioned, five meteorological stations are available in the

vicinity of Al-Monbateh plant. The location of these stations are shown in

Table 3.4 and Figure 3.7.

64

Figure 3.7. The Location of the Five Meteorological Stations and Al-

Monbateh Desalination Plant

Table 3.4. Location of the Five Meteorological Stations

Station Aqaba

Airport

Aqaba

Port Port Said

Beer-

Sheva Ismailia

Location Latitude (N) 29.63 29.48 31.28 31.23 30.6

Longitude (E) 35.01 34.98 32.23 34.78 32.25

Distance to Al-Monbateh

plant (km)

137 148 201.5 83 190

Table 3.5 summarizes the available evapotranspiration rates and the

estimated evaporation rates for the five meteorological stations. In addition,

Table 3.6 depicts the recorded rainfall data and estimated rainfall at Al-

Monbateh obtained from the five meteorological stations

Table 3.5. Available Evapotranspiration Rates and Estimated Evaporation

Rates for the Five Meteorological Stations

ET0 (mm/day) Estimated

ET0 at Al-

Monbateh

(mm/day)

Estimated

Evaporatio

n Rate

(mm/day) Month

Aqaba

airport

Aqaba

port

Port

Said

Beer

Sheva Ismailia

January 3.08 2.83 2.06 2.06 2.59 2.41 2.58

February 3.74 3.31 2.83 2.46 3.41 2.94 3.15

March 5.28 4.3 3.29 3.19 4.57 3.87 4.15

April 6.48 5.59 4.1 4.49 6.13 5.14 5.50

65

May 8.93 7.27 4.5 5.91 6.95 6.64 7.12

June 10.46 8.41 5.23 6.07 7.73 7.30 7.83

July 10.12 8.09 5.77 6.06 7.61 7.22 7.74

August 9.96 7.98 5.51 5.53 6.94 6.83 7.32

September 7.96 7.02 5.29 4.59 5.77 5.74 6.15

October 6.32 5.46 3.95 3.72 4.45 4.54 4.87

November 4.88 3.97 2.85 2.89 2.8 3.40 3.65

December 3.38 2.92 2.31 2.12 2.2 2.49 2.67

Table 3.6. Available Rainfall Data and Estimated Rainfall for the Five

Meteorological Stations

Rainfall (mm) Estimated

rainfall at Al-

Monbateh (mm) Month

Aqaba

airport

Aqaba

port

Port

Said

Beer

Sheva Ismailia

January 5 6 18 45 7 26.01

February 6 11 12 40 6 23.92

March 4 8 10 36 7 21.07

April 3 5 5 10 2 6.81

May 1 3 4 1 2 1.65

June 0 0 0 0 0 0.00

July 0 0 0 0 0 0.00

August 0 0 0 0 0 0.00

September 0 0 3 0 0 0.25

October 2 0 8 3 2 2.68

November 4 6 7 20 6 12.58

December 8 2 16 39 5 22.64

3.6. Water and Salt Balance

The water balance for both inflow to the evaporation pond and its outflow can

be expressed as:

Water inflow = Reject brine discharge + Rainfall (3.3)

Water outflow = Evaporation + Pumped-out water

Accordingly, the storage volume and water characteristics inside the

pond can be obtained from the simple mass balance calculations as:

(3.4)

Water storage = Cumulative inflow water – Cumulative outflow (3.5)

66

water

Water depth = Storage water / Area of pond (3.6)

On the other hand, salt balance can be expressed as:

Salt inflow = Salt concentration of reject brine discharge

(3.7)

Salt outflow = Salts dissolved in Pumped-out water (3.8)

Salt storage = Cumulative inflow salts – Cumulative outflow salts (3.9)

Salinity = Mass of dissolved salts in storage / Pond water mass (3.10)

Salt depth = Un-dissolved salt volume / Area of pond (3.11)

Salt storage is divided into dissolved and un-dissolved salts according to

the solubility of the salt mixture. Solubility is the concentration limit of

dissolved slats after which salts are transformed from a solution state to crystal

state. These crystals deposit on the bed of the pond. It is assumed that the

deposited salts never transfer into solution again.

Figure 3.8 illustrates the solubility of the most common salts at different

temperatures (Volland, 2005). In the case study considered here, the average

temperature is around 20 ºC. As there are no available data of salt texture

formation, it is assumed that Sodium Chloride (NaCl) is dominant. Hence, a

solubility value of 34 grams of dissolved salts in 100 grams of water is

considered. This value is equivalent to 340,000 part per million (ppm), which is

about ten times the typical sea water salinity

67

Figure 3.8. Solubility of Some Common Salts at Different Temperatures

(Volland, 2005).

3.7. Effect of Salinity on Evaporation Rates

Different studies have been made to find the relation between evaporation

rate and water salinity. According to the water report number 13 of the Food

and Agriculture Organization (FAO) of the United Nations (Johnston et al.,

1997) a correction factor Y is used to reduce evaporation rate due to the

increase in water electrical conductivity as follows:

ECY 0066.03234.1 for EC up to 60 dS/m (3.12)

Where, Y is a correction factor according FAO's formula [dimensionless]

and EC is the electrical conductivity of water [dS/m]. From the above equation

(3.12), it is clear that the evaporation rate decreases linearly with the increase

of water electrical conductivity, which is a direct measure of water salinity.

Substituting water salinity in parts per million (ppm) instead of electrical

conductivity in dS/m, which is a direct measure to water salinity assuming that

1 dS/m ≈ 640 ppm, and using a dimensionless parameter α = Y/1.3234 to keep

its value equal to unity with fresh water of zero salinity. Equation (3.12)

becomes:

68

0.1108 6 S for water salinity up to 38,400 ppm (3.13)

Where, α is a correction factor [dimensionless], and S = salinity of water

[ppm]. Another study was made to estimate this relation at Lake Qaroun in

Egypt (Ali, 1998), field measurements led to the following formula:


Where, α is s correction factor [dimensionless], and S = salinity of water

[ppm].

Equation (3.13) shows the linear relation of evaporation rate reduction with

the increasing of water salinity. As the salt concentration in evaporation ponds

may go beyond these limitations, equations (3.13) and (3.14) do not satisfy

higher salinity levels. Another study was conducted at the Dead Sea in Jordan

Valley (Niemi et al., 1997). This study depended on field data where

evaporation rates corresponding to 39 different water salinities were measured.

These measured data could be put in the following formula:


Where, α is a correction factor [dimensionless], and S is the salinity of water

[ppm].

The advantage of equation (3.15) is its validity for higher water

concentration values than equations (3.13) and (3.14). That makes it more

suitable to be used to simulate the performance of evaporation ponds,

especially for long durations. Figure 3.9 summarizes the three equations and

shows a comparison among them.

69

Figure 3.9. Salinity effect on evaporation rate (Mahmoud, 2011)

3.8. Results and Discussion

The simulation model is run with the plant inflow and the expected rainfall

as the total inflow to the pond while the outflow is the evaporation and the

pumped out volumes, when needed, with a simulation routing period of 25

years (i.e. 300 months) while the area of the evaporation ponds is the actual

constructed area and no seepage is considered due to the lining of the bottom

and sides of the ponds. The pond started operation in late in 2011 in September,

so that its evaporation rates and rainfall were put first in their arrays in the

model.

The pond performance under the native design criteria the outflow rates is

more than the inflow rates and thus continuous drought occurs no flooding

occurs during the age of the pond as both the inflow and outflow rates are

assumed to be constant during the life time of the pond. During operation the

water salinity concentration increases until it reaches its maximum saturation,

which is taken as 340,000 ppm. Once the water solution reaches its maximum

concentration, the dissolved salts is transformed into un-dissolved salts and

deposits on the bed of the pond. This concentration goes very high when water

depth inside the pond becomes very small or becomes zero. At zero level of

water inside the pond, all accumulated dissolved salts become un-dissolved and

70

create a layer of salt at the bottom of the pond. The thickness of this layer

increases as the sequence of zero water levels continues, however, the

simulation results of the exiting pond shows that the water depth is

accumulating and no zero water levels will be witnessed.

To investigate the causes of the problem, different combinations of the four

mentioned causes in section 3.4 are tested and the pond performance is

checked. Each possible cause is identified alone at first and the depth is then

plotted against the simulation period to check the pond performance. Figure

3.10 shows the effect of each cause alone, where Figure 3.10 (a) shows the

effect of rainfall, Figure 3.10 (b) shows the effect of actual evaporation rates,

Figure 3.10 (c) shows the effect of the salinity, and Figure 3.5d shows the

effect of increasing the working hours of the plant to double the designated

working hours. It can be shown that the effect of rainfall only or the actual

evaporation rates only doesn’t cause the overtopping of the pond after five

months of operation as reported by the plant operator. However from Figure

3.10 (c) and 3.10 (d) overtopping occurs after 1 year and 3 years respectively

which is also not the claim of the working staff at the Al-Monbateh

desalination plant.

72

Figure 3.10. Pond Performance with the Effect of the Four Possible Causes of

Overtopping (a) Effect of Rainfall Only, (b) Effect of Actual Evaporation

Rates, (c) Effect of Salinity, and (d) Effect of Increasing Working Hours

Different combinations of the four possible causes are studied to reach the

most possible reason to the overtopping problem that lead the operator to attach

a pumping unit after the operation of the plant by only five months. Figure 3.11

summarizes three of the considered combinations. From Figure 3.11 (a), it can

be concluded that the combination of both rainfall and the actual evaporation

rates doesn’t cause the overtopping of the brine from the pond, however from

Figure 3.11 (b) the combination of rainfall, actual evaporation rates, and

considering the effect of salinity on evaporation rates resulted in overtopping

occurrence after 16 months and 15 days prior to the pond operation date.

Finally, Figure 3.11 (c) depicts the combination of all the four possible causes

of the overtopping problem. In Figure 3.11 (c) in addition to considering the

actual evaporation rates, actual rainfall, and effect of salinity, the desalination

plant is assumed to run for three different operating hours: 16, 18, and 20 hours

per day instead of 10 hours per day. The combination of these four possible

reasons proved a good agreement between the simulated model and the actual

case which properly investigates the cause of that early overtopping.

73

The pond performance curve shows that the water level exceeds the safe

overtopping level after four months and 12 days, 5 months, and 6 months for

16, 18, and 20 operating hours respectively. Also, the pond will flood after the

first five months and 21 days, 7 months, and 12 months and 9 days for 16, 18,

and 20 plant operating hours respectively.

Thus, it is suggested that the problem may occurred because of a

combination of the four possible reasons.

It is worth to mention that Figure 3.11 reflects the importance of

considering the effect of salinity of the reject brine on the evaporation ponds,

however for more investigation on the salinity effect, sensitivity analysis of the

salinity effect and the reduction factor is discussed in section 3.7.

74

Figure 3.11. Pond Performance Curves after Combination of Different

Possible Causes. (a) Combination of Both Rainfall and Actual Evaporation

Ponds, (b) Combination of Rainfall, Actual Evaporation Ponds and Salinity

Effect on Evaporation Rates, and (c) Combination of Rainfall, Actual

Evaporation Ponds, Salinity Effect on Evaporation Rates, and Increasing the

Plant Working Hours.

75

3.9. Sensitivity Analysis for the Effect of Salinity on the Pond

Performance

The correction factor α which is calculated using equations 3.13, 3.14 and

3.15 derived based on empirical relationships depends on the coefficient of

salinity term in the equations. Thus a general form for the equation can be

formulated replacing the numerical coefficients of each equation to a constant

‘C’. The equation of the salinity effect on the evaporation rates can then be

expressed as:

0.110 6 SC (3.16)

Where, α is evaporation rates reduction factor [dimensionless], C is the

coefficient of salinity [dimensionless], and S is the water salinity [ppm].

Several attempts were reported in literature to estimate the values of the

coefficient C. A detailed review of these attempts is presented in Mahmoud,

2011. To investigate the importance of the salinity effect on reducing the

evaporation rates and accordingly on the design of evaporation ponds, 4

simulations were conducted with different values for the coefficient C. A

summary of these values are presented in the table below.

Table 3.7. Summary of C Values and Comments

Simulation

No.

C Value Comments

1 0 No effect for salinity for comparison

2 2 As estimated from Dead Sea in Jordan Valley

(Niemi et al., 1997).

3 3 As estimated from lake Qaroun study in Egypt

(Ali, 1998)

4 8 As suggested by FAO's formula

Figure 3.12 shows the sensitivity analysis and the evaluation of the effect

of changing the coefficient ‘C’ on the pond performance. Figure 3.7 shows the

pond performance in case of ignoring the salinity effect on the evaporation

76

rates in solid line for comparison. It is evidence from Figure 3.12 that the

salinity has a major influence on the evaporation pond performance. The

comparison indicates that with ignoring the effect of salinity (i.e., C = 0) the

maximum water depth in the pond did not exceed 0.5m, whereas the maximum

depth could reach a value of 1.8 to 3.4m when the value of C changes between

8 to 2. This not only proves the importance of considering the effect of salinity

on the evaporation rates, but also emphasis the importance of accurately

determine the appropriate value of C.

Figure 3.12. The Effect of the Salinity Coefficient Variation on the Pond

Performance

3.10. New Dimensions for the Evaporation Ponds for Safe Disposal

A large number of trials of changing the bottom dimensions, the depth, and

the side slopes of the pond are carried out to obtain the pond dimensions

needed for the evaporation pond to act independently as a safe disposal option

for the reject brine. Eleven alternatives are identified out of the trials as safe

dimensions. Figure 3.13 presented the top areas, the bottom areas and the

depths of the 11 alternatives. The figure shows that the top required area is a

value near 6000 square meters, while the bottom area varies between 2000 to

3000 square meters and the depths ranges from two to three meters.

77

Figure 3.13. Top and Bottom Areas and Depths of the 11 Safe Alternative

Dimensions for the Evaporation Ponds

Table 3.8 lists the dimensions of the 11 alternatives, the depth, bottom

area, side slopes, top area, volume and the relative lining cost compared to the

existing evaporation pond as well as the relative volume.

Since, the 11 alternatives assure that the evaporation pond will act solely

without the need of any supplementary units attached to it (i.e., pumping units),

then the decision on which alternative is the better will be based upon the

economical side through calculating the expected cost of the pond. The cost of

the evaporation ponds are discussed in details in section 3.11. The total cost is

plotted against the 11 alternative in Figure 3.14, it can be seen that the lowest

cost comes from alternative number six with 60m length and 50m wide with

side slopes 6:1 and two meters deep.

The pond performance of the new chosen dimensions is shown through

Figure 3.16 (a) and (b). Figure (a) shows the variation of the water depth along

25-years simulation period. The water inside the pond increases with times to

reach 1.60 m and doesn’t flood out of the pond, while Figure (b) illustrates the

salinity variation over time inside the pond. The results of the simulation model

0

0.5

1

1.5

2

2.5

3

3.5

0

1000

2000

3000

4000

5000

6000

7000

1 2 3 4 5 6 7 8 9 10 11

Dep

th (m

)

Are

a (m

2)

Alternative

Top Area

Bottom Area

Depth

78

show that the solubility is reached 14 times where the salts are separated to

form a salt layer in bottom of the pond.

Figure 3.14. The Total Cost of the 11 Safe Alternative Dimensions of the

Evaporation Pond

1.78

1.84

1.9

1.96

2.02

2.08

1 2 3 4 5 6 7 8 9 10 11

Tota

l C

ost

(E

GP

)

Mil

lion

s

Alternative

79

Table 3.8. The 11 Safe Disposal Dimensions Alternatives of the Evaporation Pond

Pond

Number L (m) W (m) d (m) S (S:1)

Bottom

Area

(m2)

Relative

Lining

Cost

Ltop

(m)

Wtop

(m)

Top

Area

(m2)

Volume

(m3)

Relative

volume

BaseCase 44 33 1.25 2 1452 1.00 49 38 1862.0 2063.44 1.00

1 60 54 2.5 4 3240 3.17 80 74 5920.0 11200.00 5.43

2 55 55 2.75 4 3025 3.18 77 77 5929.0 11979.00 5.81

3 55 50 3 4 2750 3.15 79 74 5846.0 12462.00 6.04

4 55 50 2.5 5 2750 3.22 80 75 6000.0 10546.88 5.11

5 50 50 2.75 5 2500 3.23 77.5 77.5 6006.3 11176.17 5.42

6 60 50 2 6 3000 3.32 84 74 6216.0 8928.00 4.33

7 55 55 2 6 3025 3.34 79 79 6241.0 8978.00 4.35

8 55 50 2.25 6 2750 3.39 82 77 6314.0 9786.94 4.74

9 50 50 2.5 6 2500 3.55 80 80 6400.0 10562.50 5.12

10 50 50 2.5 6 2500 3.44 80 80 6400.0 10562.50 5.12

11 45 45 2.75 6 2025 3.29 78 78 6084.0 10401.19 5.04

Figure 3.15. Typical Cross-Section of the Evaporation Pond

80

Figure 3.16. Pond Performance For The New Dimensions of The Pond For 25-

Years Routing Period. (a) Pond Water Depth and (b) Salinity Variation

Further researches are recommended for a better estimate for the effect of

salinity on the evaporation rates as the given relationships assume a linear

change with salinity. The lower vapor pressure and lower evaporation rate of

saline water result in a lower energy loss and, thus, a higher equilibrium

temperature than that of freshwater under the same exposure conditions. The

81

increase in temperature of the saline water would tend to increase evaporation,

but the water is less efficient in converting radiant energy into latent heat due to

the exchange of sensible heat and long-wave radiation with the atmosphere.

The net result is that, with the same input of energy, the evaporation rate of

saline water is lower than that of freshwater. However, there is no simple

relationship between salinity and evaporation except those presented earlier in

the section 3.7 which are site dependent and changed from a certain location to

another, and there are always complex interactions among site-specific

variables such as air temperature, wind velocity, relative humidity, barometric

pressure, water surface temperature, heat exchange rate with the atmosphere,

incident solar absorption and reflection, thermal currents in the pond, and depth

of the pond. As a result, there is a need to study in depth the effect of salinity

on evaporation rates and the dependency on the geographical location and In

case of location dependency, field experiments are needed formulate

relationships for the different regions of Egypt.

3.11. Cost of the Evaporation Ponds

Although sizing of an evaporation pond is a relatively straightforward

procedure once appropriate evaporation data and inflows are available, the

costs associated with pond construction are highly site specific and quite

variable. Therefore, generic cost estimating of evaporation ponds from typical

handbook-type and previous studies data is very difficult and subject to a wide

range of accuracy. However, by gathering site-specific data, a reasonably

accurate cost estimate can be made.

In general, it is anticipated that evaporation ponds most likely will be

competitive for relatively small plants in remote, inland locations with high

evaporation rates. The major factors contributing to the cost of an evaporation

pond are: land costs, earthwork, lining, operation and maintenance. The cost of

land can vary greatly from site to site, it can easily vary by a factor of 10 or

more, depending on the exact location near the city. In our case, the cost of a

feddan is assumed EGP 10,000 as an average for rural areas. It’s worth noting

82

that in large evaporation ponds, there is a distinction between evaporative area

and total area which is important in determining the land requirements, thus an

area correction factor shall be provided to multiply times the evaporative area

to calculate the total area. Like the cost of land itself, the cost of earthwork is

very site specific, depending on whether the terrain is flat or hilly, rocky or

sandy, forested or clear, etc. In selecting a site for an evaporation pond, such

factors must be considered. The earthwork cost is taken as EGP 20 per cubic

meters. Once it has been constructed, the pond operates essentially

maintenance free. Periodic maintenance is required only for the repair of the

dike or liner, pipe, flow control devices, etc. Operating costs also include

security and damage inspection. A total capital operating costs is assumed to be

5 percent of the total costs.

The costs of installing liners include those for material and construction.

Figure 3.17 illustrates the lining elements of the existing evaporation pond

which is used in calculating the cost of the new evaporation pond. Figure 3.18

shows a group of charts created for estimation of the liners cost based on the

given lining elements and their up-to-date unit costs.

Figure 3.17. Typical Lining Cross-Section for the Existing Evaporation Pond

83

50

75

100

125

150

175

200

225

250

0 2000 4000 6000 8000 10000

Lin

ing C

ost

(E

GP

/m2)

Pond Area (m2)

Side Slope = 2:1 Depth = 1m

Depth = 2m

Depth = 3m

50

100

150

200

250

300

350

400

0 2000 4000 6000 8000 10000

Lin

ing C

ost

(E

GP

/m2)

Pond Area (m2)


Depth = 2m

Depth = 3m

84

Figure 3.18. Cost of Liners per Bottom Area Square Meter for Different Pond

Depths and Side Slopes

For comparing between the current provided disposal option and the other

alternative (i.e., deep well injection) the cost of the option is calculated using

the mentioned above cost elements. For the current evaporation pond, two

other elements are added which are: the cost of the pumping unit and its

operation cost, and the cost of the environmental penalty resulting from the

unmanaged disposal. Figure 3.19 shows the cost of existing evaporation ponds

and the new proposed ponds with the safe dimensions.

Since the frequency of operating the pump, the operating duration and the

operating discharge and head are not known then the cost of the pumping unit

and its associated operation cannot be estimated with the given information.

Soil salinization, loss of crop yield, and contamination of the underlying

groundwater are examples of the environmental penalties resulting from the

unmanaged disposal. The penalties can be expressed by an additional cost

added to the cost of the existing evaporation pond. The two costs along the

lifetime of the ponds are expressed in Figure 3.19 by dashed bars stacked over

50

150

250

350

450

550

650

0 2000 4000 6000 8000 10000

Lin

ing C

ost

(E

GP

/m2)

Pond Area (m2)


Depth = 2m

Depth = 3m

85

the cost of the native cost of the pond. These costs are uncertain and can be

more of less than the stacked values on the graph.

The figure shows that the cost of the safe ponds is about 1.885 million

Egyptian pounds while 0.541 for the existing one. An increase of 250% in the

cost proves the significant importance of considering the effect of salinity on

the evaporation ponds design as discussed in the previous section of this

chapter.

Figure 3.19. The Total Cost of the Existing and the New Evaporation Ponds

0.541

1.885

0

0.2

0.4

0.6

0.8

1

1.2

1.4

1.6

1.8

2

Existing Pond New Pond

Tota

l C

ost

(E

GP

x 1

,000)

Mil

lion

s

Uncertain

Uncertain

87

CHAPTER FOUR

MATHEMATICAL FORMULATION

In order to assess the option of the injection of the reject brine into deep

aquifers and to simulate the transient state flow conditions in this study, the

MODFLOW code is utilized. MODFLOW is the U.S. Geological Survey’s

three-dimensional finite-difference groundwater model (McDonald and

Harbaugh 1988; Harbaugh and McDonald 1996). Originally conceived solely

as a groundwater-flow simulation code, MODFLOW’s modular structure has

provided a robust framework for integration of additional simulation

capabilities that build on and enhance its original scope. The family of

MODFLOW-related programs now includes capabilities to simulate coupled

groundwater/surface-water systems, solute transport, variable-density and

unsaturated-zone flow, aquifer system compaction and land subsidence,

parameter estimation, and groundwater management. Integrated to this code is

another popular transport code, MT3DMS (Zheng et al., 1990, 1999) that uses

the flow output of MODFLOW and runs solute transport simulation in an

efficient manner. Coupled with these two models is the variable density flow

SEAWAT code. It is a generic MODFLOW/MT3DMS-based computer

program designed to simulate three-dimensional variable-density groundwater

flow coupled with multi-species solute and heat transport. The program has

been used for a wide variety of groundwater studies including those focused on

brine migration in continental aquifers as well as those focused on saltwater

intrusion in coastal aquifers. These three codes are considered reasonable

modeling tools for the current simulation.

88

It is worth noting that MODFLOW was modified to solve the variable-

density flow equation by reformulating the matrix equations in terms of fluid

mass rather than fluid volume and by including the appropriate density terms.

Fluid density is assumed to be solely a function of the concentration of

dissolved constituents; the effects of temperature on fluid density are not

considered. Temporally and spatially varying salt concentrations are simulated

in SEAWAT using routines from the MT3DMS program. SEAWAT uses

either an explicit or implicit procedure to couple the groundwater flow equation

with the solute transport equation.

The flow modeling in this study is conducted using MODFLOW within the

framework of the graphical user interface GMS (Groundwater Modeling

System). GMS is a comprehensive graphical user environment for performing

groundwater simulations. The entire GMS system consists of a graphical user

interface (the GMS program) and a number of analysis codes (MODFLOW,

MT3DMS, SEEP2D, SEAWAT, etc…).

This chapter presents the mathematical formulation of MODFLOW code

(used for simulating the flow), MT3D code to simulate the brine transport and

SEAWAT for the variable-density flow simulation of brine.

4.1. Mathematical Model of MODFLOW Code

The following partial differential equation represents the three dimensional

movement of groundwater through the porous medium:

szzyyxxs q

z

hK

yy

hK

yx

hK

xt

hS

)()()( (4.1)

where zzyyxx KandKK ,, are values of hydraulic conductivity along the x, y, and z

coordinate axes, which are assumed to be parallel to the major axes of

hydraulic conductivity [LT-1

]; h is the potentiometeric head [L]; qs is a

volumetric flux per unit volume and represents sources and/or sinks of water

[T-1

]; Ss is the specific storage of the porous material [L-1

]; and t is time [t]

89

Equation (4.1) describes groundwater flow in heterogeneous and

anisotropic medium, under the condition that the principal axes of hydraulic

conductivity are aligned with the coordinate directions. It also represents the

unsteady state conditions.

In case of having a homogeneous medium, then equation (4.1) can be written

as:

ss q

z

h

y

h

x

hK

t

hS

)(

2

2

2

2

2

2

(4.2)

In case of having a steady state condition, then equation (4.2) becomes:

szzyyxx q

z

hK

yy

hK

yx

hK

x

)()()(0 (4.3)

In case of having a homogeneous medium in addition to a steady state

condition, then equation 4.3 reduces to:

sq

z

h

y

h

x

hK

)(0

2

2

2

2

2

2

(4.4)

The analytical solution of the flow equation is possible only for very simple

systems. In real field applications, however, the aquifer conditions of

heterogeneity and anisotropy and the irregularity and complexity of the

geologic structures and boundary conditions preclude the possibility of using

such analytical solution. Therefore, a numerical method must be developed to

get the approximate solution. One such approach is the finite-difference

method, in which the partial derivatives are replaced by terms calculated from

the differences in head values at these points. The process leads to systems of

simultaneous linear algebraic difference equations; their solution yields values

of head at specific points and times.

4.2. Mathematical Model of MT3DMS Code

Once reject brine produced from the desalination plant is injected in the

disposal well, the solutes associated with this water such as the dissolved

solids, heavy metals, and antiscalant (which are added to the water prior to

90

desalination as well as the backwash water of the membranes of the

desalination plants) start to mitigate in the groundwater system. Two main

mechanisms affect the solute migration: advection and dispersion. The mass

balance equation for a solute species is written as a partial differential equation

in three dimensions and has the form (e.g., Javandel et al., 1984):

N

k

kss

i

ij

ij

i

RCq

CVxx

CD

xt

C

1

)()(

(4.5)

where C is the concentration of solutes dissolved in groundwater [ML-3

]; t is

time [T]; x is the distance along the respective Cartesian coordinate axis [L]; Dij

is the hydrodynamic dispersion coefficient [L2T

-1]; V is the seepage or linear

pore water velocity [LT-1

]; qs is the volumetric flux of water per unit volume of

the aquifer representing sources (positive) and sinks (negative) [T-1

]; Cs is the

concentration of the source or sink [ML-3

]; θ is the porosity of the porous

medium [dimensionless]; and

N

k

kR1

is the chemical reaction term [ML-3

T-1

].

The first term on the right hand side of equation (4.5) accounts for solute

dispersion (both mechanical dispersion and molecular diffusion) while the

second term accounts for advective transport. The third term gives the effect of

sources or sinks in the system and the last term deals with the chemical

reactions that may be encountered for some solutes. The components of the

dispersion tensor, Dij, are given as (Bear, 1972):

ij

ji

TLTijij DV

VVVD *).( (4.6)

where δij is the Kroneker delta (δij = 1 for i = j and δij = 0 for i ≠ j), αL and αT are

the longitudinal and transverse local dispersivities, respectively, |V| is the

magnitude of the velocity, and D* is the effective coefficient of molecular

diffusion.

Equation (4.5) is the governing equation underlying the transport model,

and equation (4.6) is an auxiliary equation that relates the dispersion

91

coefficients needed in (4.5) to flow velocity and aquifer dispersivity. The

transport equation is linked to the flow equation through the following

relationship:

i

ii

ix

hKV

(4.7)

where Kii is a principal component of the hydraulic conductivity tensor [LT-1

],

θ is the effective porosity, and h is the hydraulic head [L]. The hydraulic head

is obtained from the solution of the three dimensional groundwater flow

equation:

t

hSq

x

hK

xss

j

ij

i

)(

(4.8)

where Ss is the specific storage of the porous materials [L-1

]. It should be noted

that the hydraulic conductivity tensor (K) actually has nine components.

However, it is generally assumed that the principal components of the

hydraulic conductivity tensor (Kii, or Kxx, Kyy, Kzz) are aligned with the x, y and

z coordinate axes so that non-principal components become zero. This

assumption is incorporated in most commonly used flow models, including

MODFLOW.

Several numerical approaches can be used to solve the transport equation;

for example, finite differences, finite elements, method of characteristics, and

random walk particle-tracking methods. In this study, MT3DMS (Zheng and

Wang, 1999; Zheng, 2006) is used for solving the transport equations.

MT3DMS is a model for simulation of advection, dispersion and chemical

reactions of solutes in groundwater flow systems in either two or three

dimension. The model uses a mixed Eulerian-Lagrangian approach to the

solution of the advection-disperive- reactive equation, based on combination of

the method of characteristics and the modified method of characteristics. The

model program uses a modular structure similar to that implemented in

MODFLOW that is being used here for flow solution. The modular structure of

the transport model makes it possible to simulate the advection, dispersion,

92

source/sink mixing, or chemical reactions independently without reserving

computer memory space for unused options.

The MT3DMS transport model was developed for use with any block-

centered finite difference flow model such as MODFLOW and is based on the

assumption that changes in concentration field will not affect the flow field

significantly. After a flow model is developed and calibrated, the information

needed by the transport model can be saved in disk files which are then

retrieved by the transport model.

The transport model can be used to simulate changes in concentration of

single-species miscible salinity in groundwater considering advection,

dispersion and some simple chemical reactions, with various types of boundary

conditions and external sources or sinks. MT3DMS accommodates the

following spatial discretization capabilities and transport boundary conditions:

(1) confined, unconfined or variably confined/unconfined aquifer layers; (2)

inclined model layers and variable cell thickness within the same layer as

present in both modeled aquifers as discussed in Chapter five; (3) specified

concentration or mass flux boundaries; and (4) the solute transport effects of

external sources and sinks such as wells, drains, rivers, areal recharge and

evapotranspiration.

4.3. Mathematical Model of SEAWAT Code

The SEAWAT (Guo and Langevin, 2002) program was developed to

simulate three-dimensional, variable-density, transient groundwater flow in

porous media. The source code for SEAWAT was developed by combining

MODFLOW and MT3DMS into a single program that solves the coupled flow

and solute-transport equations. Since both aquifers being modeled (the

producing aquifer and the injection aquifer) are saline aquifers, use of a code

that solves the variable-density groundwater flow is important and thus it is the

code implemented in this study. The SEAWAT code was tested by simulating

five benchmark problems involving variable-density groundwater flow. These

93

problems include two box problems, the Henry problem, Elder problem, and

HYDROCOIN problem, and it was found that SEAWAT results compare well

with those of SUTRA (a computer model for simulation of variable-density

saturated-unsaturated flow with solute or energy transport developed by Voss,

1984). Also, the SEAWAT code was used by Nassar (2004) to simulate the

unsteady two-dimensional phenomena of subsurface brine disposal and to

verify using the code in solving the variable density flow through an

experimental seepage tank with known extraction and injection rates as well as

known initial and injection salinities. The results of the simulated model using

SEAWAT showed a good agreement with that of the experimental setup, thus

proving the reliability of using the code in simulating the current density-

dependent groundwater flow and the solute migration in this study.

SEAWAT is based on the concept of freshwater head, or equivalent

freshwater head, in a saline groundwater environment as discussed later in

Section 5.5 (Model Calibration). The governing equation for the variable-

density groundwater flow is as follows:

,'. 00,

0

00

0sss q

t

C

Ch

hSzh

0K (4.7)

where ρ0 is the fluid density [ML-3

] at the reference concentration and reference

temperature; µ is dynamic viscosity [ML-1

T-1

]; K0 is the hydraulic conductivity

tensor of material saturated with the reference fluid [LT-1

]; h0 is the hydraulic

head [L] measured in terms of the reference fluid of a specified concentration

and temperature (as the reference fluid is commonly freshwater). Ss,0 is the

specific storage [L-1

], defined as the volume of water released from storage per

unit volume per unit decline of h0; t is time [T]; θ is porosity [-]; C is salt

concentration [ML-3

]; and q's is a source or sink [T-1

] of fluid with density ρs.

The associated solute transport equation is defined as:

94

,').()..()(

1 k

ss

kkkk

db CqCCt

CK

qD

(4.8)

where, ρb is the bulk density (mass of the solids divided by the total volume)

[ML-3

]; Kdk is the distribution coefficient of species k [L

3M

-1]; C

k is the

concentration of species k [ML-3]; D is the hydrodynamic dispersion

coefficient tensor [L2T

-1]; q is the specific discharge [LT

-1]; and Cs

k is the

source or sink concentration [ML-3

] of species k.

The viscosity effects were neglected in this study so the term

0 is taken

equal to one, and fluid density was treated as a simple linear function of only

one solute species which is the salt concentration of the reject brine dealt with

in this study.

It is worth noting that under the SEAWAT approach, the two separate

computer programs, MODFLOW and MT3DMS, are modified and combined

into one program. Among these modifications are the conversion of volumetric

fluxes to mass fluxes and the addition of relative density-difference terms and

solute-mass accumulation terms to the basic finite-difference equation solved

by MODFLOW. Additionally, modifications are made to each of the stress

packages of MODFLOW because mass fluxes and freshwater heads are used in

SEAWAT. Modifications of MT3DMS are relatively minor and mainly affect

internal data transfer and manipulation (Guo and Langevin, 2002).

95

CHAPTER FIVE

GROUNDWATER FLOW AND TRANSPORT MODELS

This chapter describes the development of a three-dimensional groundwater

model to simulate groundwater flow in both the Lower and the Upper

Cretaceous aquifers of the study area to simulate the case of the injection of the

reject brine. The model is developed using the following steps:

1. Defining the model domain (i.e., areal and vertical extents of the

model);

2. Defining the boundary conditions;

3. Defining sources and sinks in the studied domain (i.e., wells, recharge

zones, rivers or streams if any, etc…); and

4. Calibrating the model by adjusting model parameters (e.g., hydraulic

conductivity) until model performance matches observed field data

A three dimensional domain is selected for flow and transport modeling

around Al-Monbateh desalination plant within which a hypothetical disposal

well is located. To simulate the actual case of the Al-Monbateh production

well, which taps the Lower Cretaceous aquifer in Sinai, and the proposed

injection well with their relatively small production and injection rates, a local

model has to be developed as the salinity plume is not expected to migrate for

long distances. However, the available data and potentiometric maps do not

provide sufficient information to develop a local simulation model. Thus,

creating a regional model and cutting it around the production and injection

wells for locally studying the flow and transport (cake-cutting process) is

adopted for the sake of a better modeling through a finer grid as discussed later

in this chapter.

96

Figure 5.1 shows the location of Al-Monbateh well with respect to the

water contours of the Lower Cretaceous aquifers. It can be seen that the nearest

available contour map is the 50 m above mean sea level (amsl) contour at a

distance of 8 km south of the well while no available data in the other three

directions for a local simulation model.

Figure 5.1. The Potentiometric Map for the Lower Cretaceous Aquifer with the

Location of Al-Monbateh well.

The geometric characteristics of the regional model domain are determined

based on the hydrogeologic framework of the study area and the available

potentiometric maps. The model is composed of four main layers: (1) the upper

conductive aquifer of the Upper Cretaceous formation, (2) the lower

conductive aquifer of the Lower Cretaceous formation, (3) a thin clay layer

separating the two previous aquifers, and (4) an impervious rock layer for the

Jurassic sedimentary rocks underlying the Lower Cretaceous formation. Figure

5.2 shows a typical stratigraphic succession in the studied area. The regional

model dimensions are not uniformly set as discussed in Section 5.3.1. It

97

extends an average of 50 km in the north-south direction and an average of 70

km in the east-west direction.

Figure 5.2. Typical Stratigraphic Succession in Central Sinai (Ghoubachi,

2010)

5.1. Description of Study Region

The study area is located in the central eastern portion of Sinai and is

bounded by longitudes 33° 46′ – 34° 32′ E and latitudes 30° 25′ – 30° 58′ N. It

occupies an area of about 3,000 km2 which encompasses 4.9% of Sinai

Peninsula’s total area. Three aquifer systems are bounded by the chosen model

domain: the Eocene, the Upper Cretaceous aquifer and the Lower Cretaceous

aquifer. The domain encompasses about 20% of the total area of the Lower

Cretaceous sandstone aquifer, known as Malha formation, which is the

producing aquifer of Al-Monbateh well. The Eocene forms the mountainous

areas of the study region so it is assumed that the Upper Cretaceous extends in

these areas as no data is available for the Eocene aquifer.

98

The area is an arid area with scarce and irregular seasonal rainfall. The

surroundings are not very well developed and the number of habitants is small

because of the limited availability of fresh water resources, except for some

spots where deep wells penetrate either the Upper Cretaceous or the Lower

Cretaceous aquifer and produce brackish groundwater mainly used for

irrigation and planting of some salt-tolerant plants such as olives. The water

quality of these aquifers is not promising for a domestic use for those habitants,

however, desalination offers a great alternative for freshwater by desalting the

produced brackish water of the aquifers underlying the study area.

Through a field visit to Al-Monbateh well and its present desalination plant,

it was found that the life of people there is very primitive. During surveying the

habitants of the visited site, it was also found that before the construction of the

desalination plant, they depended on buying freshwater for their domestic uses

from Al-Arish city which is about 70 km away, or sometimes from rainfall

harvesting during the rain seasons. It is also important to mention that the area

began to act as an attraction point for other nearby habitants who are to some

extent still depending on buying freshwater from Al-Arish. The area is now

witnessing small local development which has put more pressure on Al-

Monbateh desalination plant, resulting in more demand and probably resulted

in the operation of the plant more than the designated daily working hours (i.e.

10 hours). This has some important implications for the disposal system as was

discussed in Chapter 3.

5.2. Conceptual Flow Model

The modeling domain and the boundary conditions dictate how

groundwater is perceived to flow in the system. As indicated by the head

contours, groundwater of the Upper Cretaceous aquifer moves from the

southeast to the northwest direction while it flows from the south to the

northeast direction in the Lower Cretaceous aquifer. The two aquifers are

assumed to be hydraulically not connected since both the head contours and the

iso-salinity contours do not comply with each other at any point along the

99

domain. A clay layer with very low hydraulic conductivity is then assumed to

separate the two aquifers.

The salinity of the Upper Cretaceous is found to be higher than that of the

Lower Cretaceous aquifer, thus the injection is assumed to be screened in the

upper formation and due to the presence of the clay layer, and the proposed

injection would be isolated from the Lower Cretaceous aquifer, the producing

aquifer for Al-Monbateh well. However, the maps of the aquifers, whether the

heads or the altitudes of the top or the bottom of the layers, are usually

developed based on data from sparse wells which casts some uncertainty as to

whether the two aquifers are totally separated or not. This is evaluated using

different scenarios in the flow and solute transport simulations.

The Upper Cretaceous aquifer is believed to be unconfined aquifer with

variable thickness ranging from 400 m to 800 m, whereas the Lower

Cretaceous aquifer is considered in many studies (Ghoubashi, 2011 and Dames

and Moore, 1985) to be confined except for a narrow outcropping strip along

the southern terminal of the aquifer, which is outreached in the studied area.

Thus, the aquifer is taken as confined in the studied domain. Its thickness

varies from few meters in a small portion of the study domain to about 600 m

towards the east direction.

5.3. Numerical Flow Model Development

5.3.1. Areal and Vertical Extent

The model consists of two main water-bearing formations, the Upper

Cretaceous Carbonate aquifer and the Lower Cretaceous Sandstone aquifer.

The deposition of Upper Cretaceous is assumed to extend from the land surface

downward to the clay inter-beds of the Lower Cretaceous aquifer. The Eocene

formation which overlies the Upper Cretaceous aquifer only appears in the

mountainous areas of the study region.

The top of the Lower Cretaceous aquifer is defined by the lower surface of

the Upper aquifer which is composed of limestone and marl with shale

100

interbeds. The base of the aquifer is also defined by the top of the Upper

Jurassic and no connectivity is expected between the two aquifers, and

therefore, the model is truncated at the base of the Lower Cretaceous aquifer.

Figures 5.3, 5.4 and 5.5 presents the vertical extends of the two aquifers

through the contour maps of the top and the bottom levels of the aquifer.

5.3.2. Groundwater Levels and Movement

The groundwater levels for both the Upper Cretaceous carbonate and the

Lower Cretaceous sandstone (Malha Formation) are obtained from the

Groundwater Sector in the Ministry of Water Resources and Irrigation

(MWRI). The potentiometric maps are based on the water levels data collected

from numerous deep wells tapping both aquifers dated back to 2002. The

potentiometric map of the Upper Cretaceous (Figure 5.6) shows that

groundwater flows from the southeast to the northwest with an average

hydraulic gradient of 0.005 till the mid of the modeled area and then moves

with a relatively mild hydraulic gradient of 0.0017 for the rest of the domain.

The map indicates that a highest potentiometric level, within the study area, of

300 m.a.m.s.l. (meters above mean sea level) is observed at the southeastern

part of the model where it starts fluxing into the model domain and terminates

the study area at a potentiometric level of 50 m.a.m.s.l. The recharge to the

Upper Cretaceous aquifer occurs through direct infiltration of rainfall or from

surface flow on its exposed areas where the estimated rate of recharge to the

Upper Cretaceous aquifer is about 190,000 m3/day (Dames and Moore, 1985)

The potentiometric surface map of the Lower Cretaceous aquifer (Figure

5.7) indicates that water moves from the southern part of the model with an

average hydraulic gradient of 0.006 and diverts eastward to exit the eastern

boundary to Wadi Arava Rift in Palestine with a water level of 50 m.a.m.s.l..

The Malha aquifer system is believed to function as an unconfined aquifer only

at a limited zone (1 to 2 km) near its southern outcrops at which recharge to the

aquifer takes place, and is confined elsewhere, being capped by the overlying

Upper Cretaceous complex.

101

The potentiometric maps prove that the two aquifers are not hydraulically

connected where the Lower Cretaceous sandstone aquifer is believed to be

confined over the studied domain while the Upper Cretaceous carbonate

aquifer is considered unconfined.

5.3.3. Groundwater Quality

Figure 5.8 represents the iso-salinity contour map of the Upper Cretaceous

aquifer system. It indicates a general trend of salinity increase towards the

north. The concentration ranges from 3,000 mg/L near the southeastern

boundary to 10,000 mg/L at the northwestern boundary with a mild salinity

gradient in the southeastern – northwestern direction till the middle of the study

domain where the slope steeps sharply until the end of studied area. Figure 5.9

represents the iso-salinity contour map of the Lower Cretaceous aquifer. It

shows a general increase of groundwater salinity towards the northwestern

direction. The Lower Cretaceous aquifer within the studied area has a salinity

range of 2,000 to 10,000 mg/L.

It can be shown from the iso-salinity figures that the groundwater quality

of the Lower Cretaceous aquifer is generally better than that of the Upper

Cretaceous aquifer. This supports the presence of more wells tapping the

Lower Cretaceous than the Upper Cretaceous in the study region. For the same

point in a horizontal plane, the average salinity difference between the two

aquifers is about 1,000 mg/L. Thus, the injection of the reject brine will take

place in the Upper Cretaceous aquifer.

102

Figure 5.3. Contours of Top Level of the Upper Cretaceous Aquifer

103

Figure 5.4. Contours of Base Level of the Upper Cretaceous Aquifer

104

Figure 5.5. Contours of Base Level of the Lower Cretaceous Aquifer

105

Figure 5.6. The Potentiometric Map of the Upper Cretaceous Aquifer

106

Figure 5.7. The Potentiometric Map of the Lower Cretaceous Aquifer

107

Figure 5.8. The Iso-salinity Contour Map of the Upper Cretaceous Aquifer in

Sinai (EPIQ Water Policy Team, 1998)

108

Figure 5.9. The Iso-salinity Contour Map of the Lower Cretaceous Aquifer in

Sinai (EPIQ Water Policy Team, 1998)

109

5.3.4. Aquifer Hydraulic Properties

Potentiometeric maps of both the Upper and Lower Cretaceous aquifers

indicate that the two aquifers are not hydraulically connected and are separated

by a clay layer extending between the two aquifers. Thus, three hydrogeologic

units are included in the model and they are defined based on the available data

as follows:

The Upper Cretaceous Carbonate water-bearing formation having a high

hydraulic conductivity

The Lower Cretaceous Sandstone water-bearing formation with a higher

hydraulic conductivity than that of the Upper.

Fine grained thin layer of clay with a very low hydraulic conductivity

which extends in the entire model domain separating the two previous

hydrogeologic units.

The potentiometric maps of both the Upper and the Lower Cretaceous

aquifers are presented in Figures 5.6 and 5.7. It can be seen that the water

levels in the Upper Cretaceous is higher than that of the Lower Cretaceous by

almost 100 meters.

For the Upper Cretaceous carbonate aquifer, a value of 1x10-4

m-1

is

assumed for the specific storage and a specific yield of 0.05. This assumption is

based on the average values for unconfined and carbonates formations. For the

Lower Cretaceous sandstone aquifer, the specific storage is assumed to be

1x10-5

m-1

and a value of 0.15 for the specific yield.

110

5.3.5. Domain Spatial Discretization for the Regional Model

The regional model is divided into a total of nine vertical layers (numerical

layers) of which hydraulic parameters are required to be identified. Six layers

comprise the Upper Cretaceous aquifer and one transition clay layer is assumed

between the Upper Cretaceous and the Lower Cretaceous aquifers with a

thickness of 10 meters. The eighth and the ninth layers comprise the Lower

aquifer. At any location, the thickness of each of these two layers is taken as

half the total thickness of the Lower Cretaceous aquifer at that location. The

first six layers that form the Upper Cretaceous are divided as follows. The

bottom five layers have thickness of 50 meters each while the upper most layer

has a thickness equal to the remaining portion of the aquifer till the top of the

Upper Cretaceous aquifer. In the horizontal direction each of these layers is

divided into grid cells with size of 250m in both X and Y direction. This results

in a total number of model cells of 514,560. However, not all of these cells

contribute in the calculations as some of these cells are inactive cells due to the

irregular geometry of the study region in all directions. These inactive cells are

handled in the MODFLOW environment as an array with certain flags. The

flags are used to indicate whether the cell is active or inactive within each layer

so that it is taken into consideration in the model simulation or not.

A model grid of 268 rows, 320 columns and 9 layers is used to represent the

study area. The study focuses on the salt migration in the host formation of the

injection process which is the Upper Cretaceous aquifer. Therefore, the

discretization is increased in the injection aquifer and the layer thickness is

limited to 50 meters whereas a coarser grid is used in the Lower Cretaceous

aquifer. The nine layers of the model are conceptualized as shown in Figures

5.11 and 5.12. To summarize the model conceptualization:

Layers two to six have a uniform thickness of 50 m each while layer one

is the remaining thickness of the aquifer from the top of layer two to the

top of the Upper Cretaceous aquifer;

111

Layer 7 is a transition clay layer between the two water-bearing

formations which acts as an impervious layer between the Upper and the

Lower Cretaceous aquifers;

Layers 8 and 9 have non-uniform thickness varying from one grid cell to

another with a value equal half the total thickness of the Lower

Cretaceous aquifer.

5.3.6. Domain Spatial Discretization for the Local Model

The local model is extracted from the regional model and is developed to

allow for a finer gird in order to better simulate the salts migration. The salinity

plume migration distance is expected to be small due to the low injection rates

of the Al-Monbateh desalination plant (i.e., 16 m3/day). The local model is

taken 5 km x 5 km (Figure 5.10) where the boundaries are set such that the

southern boundary is 2 km away from the injection well and the western

boundary is 3 km away from the well.

Figure 5.10. The Location of the Local Model With Respect to the Regional

Model

The local model domain is divided into 12 layers, the upper most six layers

comprise the Upper Cretaceous aquifer with the last five layers having a

thickness of 50 meters each while the top layer has a thickness equals to the

112

remaining thickness of the aquifer. The seventh layer is the transition clay layer

separating the two aquifers, and the lower five layers comprise the Lower

Cretaceous aquifer with each layer thickness equal to one fifth the total

thickness of the aquifer. In the horizontal direction each of these layers is

divided into grid cells with size Δx = 50 m and Δy = 50 m resulting in a total

number of cells of 10,000. Figure 5.13 shows a plan view and typical cross-

sections (East-West and South-North) for the local model and the suggested

location of the injection well is also shown in the figure.

113

Figure 5.11. Typical Cross-Sections (East-West) in the Conceptual Regional Model (row 169 and row 230)

Figure 5.12. Typical Cross-Sections (South-North) in the Conceptual Regional Model (column 80 and column 186)

114

Figure 5.13. Plan View and Typical Cross-Sections (East-West and South-North) for the Local Model (The location of the injection

well is shown as a black dot)

115

5.4. Boundary and Initial Conditions and Implementation of

MODFLOW

The dependence of head on fluid density has important implications for

assigning boundary heads, particularly when the density of the boundary head

changes during the simulation. In this study, it is initially assumed that there is

no change in salinity concentration at the boundaries as the injection wells are

far from the domain boundaries and thus the plume migration of the injected

reject brine will not reach the boundaries of the model.

Initial conditions represent starting values for the dependent variable, such

as freshwater head for groundwater flow and concentration for solute transport,

at some starting time. Initial conditions for both flow and transport must be

specified for transient simulations. For this study region, the point-water heads

(saline water heads) are converted to freshwater heads and assigned to the

model boundaries.

In the chosen study region, the model boundaries are chosen based on the

hydrologeologic conditions of the Lower Cretaceous aquifer. The upper

(northern) boundary is taken as no-flow boundary as it comprises the terminal

of the Lower Cretaceous, where the aquifer changes into deeply faulted

limestone region towards the north side, therefore, the northern boundary can

be considered as no-flux. The other three boundaries, the south, the west and

the east are considered specified head boundaries and the assigned freshwater

head values are determined at the points where the boundaries intersect with

the saline water contours.

The boundaries of the Upper Cretaceous aquifer have the same extent as

that of the Lower Cretaceous. However, the four boundaries are specified head

boundaries obtained from the intersection of the model boundaries with the

saline water contours of the Upper Cretaceous aquifer which are converted into

equivalent freshwater heads.

116

5.4.1. Injection Well

First a base case scenario representing the current condition is established.

In the base case the calibration parameters are estimated such that model

predicted heads match measured head to a certain degree of tolerance. After

establishing the base case the flow field is updated to account for the presence

of the proposed injection well at a distance of 200 m north Al-Monbateh. This

location is selected downgradient from the production well as the flow moves

in the south-north direction. The well package in MODFLOW is used to

simulate the production and injection wells in the study region. For each cell at

the well location, a negative flow value (m3/day) is assigned to indicate a

volumetric extraction whereas a positive value is assigned for the injection

rates. Based on the data provided from Al-Monbateh desalination plant, the

reject brine volume is estimated to be 16 m3/day.

The regional study area encompasses nine deep production wells, two of

which tap the Upper Cretaceous aquifer while the other seven tap the Lower

Cretaceous aquifer. Figure 5.14 shows the location of the wells with respect to

the model boundaries while Table 5.1 gives the technical and hydrogeologic

data of the wells

Figure 5.14. Satellite Images Showing the Location of the Study Area and the

Location of the Nine Wells within the Study Region

117

Table 5.1. Technical and Hydrogeologic Data of Wells in Study Region

Well

Location Producing

aquifer

Yield

(m3/hr)

Well Type

and Use

Depth to

Water Lat.

(N)

Long.

(E)

Gebel Libni 30°44' 33°53' U. Cr. 11 Production /

Agriculture 220

Talaat El

Badan 30°29' 34° 3' U. Cr. -

Test

Productive /

Abandoned

163

Halal – 1 30°41' 34°10' L. Cr. 10 Production /

Agriculture 160

Halal – 2 30°41' 34° 9' L. Cr. 35 Production /

Agriculture 140

Monbateh 30°39' 34°13 L. Cr. 30

Production /

Agriculture

and

desalination

167

Sabha 30°43' 34°25' L. Cr. - - -

Hodeibiya 30°35' 34°13' L. Cr. - - -

Gaifi 30°35' 34°22' L. Cr. - - -

Garour 30°29' 34°20' L. Cr. - - -

5.4.2. Hydraulic Conductivity

For better simulation for the aquifers in the study area, the modeled layers

are taken heterogeneous. Thus the values of the hydraulic conductivity for the

Upper Cretaceous aquifer and the Lower Cretaceous aquifer differ spatially,

while a single value (i.e., homogenous conditions) is assigned for the clay layer

separating the two aquifer.

5.5. Model Calibration

Calibration is the process of modifying the input parameters to a

groundwater model (e.g., hydraulic conductivity) until the output resulting

from the numerical simulation model matches an observed set of data (e.g.,

water levels). One of the tools provided in GMS for model calibration is

automated parameter estimation. Automated parameter estimation is supported

in GMS for the MODFLOW simulations using PEST, a general purpose

parameter estimation utility (Doherty, 1994). With automated parameter

estimation, inverse modeling is used to iteratively adjust a set of parameters

118

and repeatedly launch the model until the computed output matches field-

observed values. In this study, the parameter estimation program PEST is used

to calibrate the flow model using the hydraulic conductivity as the calibration

parameter and the water heads as the calibration target. It implements a

nonlinear least squares regression method to estimate model parameters by

minimizing the sum of squared weighted residuals.

Due to variation in the salinity of the groundwater over the area of the two

simulated aquifers, the density varies spatially. Thus, the measured heads,

known as point-water heads which are the heads in terms of the native aquifer

waters, are not the freshwater heads which MODFLOW reads and writes. Since

PEST is supported for the MODFLOW simulations, then the calibration targets

which are the water heads must be read in terms of freshwater heads. To

evaluate the freshwater heads from the point-water heads, the relationship

between salt concentration and fluid density is required as well as the

relationship between the freshwater and point-water heads. For isothermal

conditions, fluid density is predominantly affected by the salt concentration. An

empirical relation between the density of saltwater and concentration was

developed by Baxter and Wallace (1916):

,ECf (5.1)

where is the water density at any concentration level [ML-3

], f is the

freshwater density [ML-3

], E is a dimensionless constant having an

approximate value of 0.7143 for salt concentrations ranging from zero to that

of seawater, and C is the salt concentration [ML-3

].

For two piezometers open to a given point, N, in an aquifer containing

saline water, with piezometer A containing freshwater and piezometer B

containing water identical to that present in the saline aquifer at point N, the

freshwater head at point N is the elevation of the water level in piezometer A

above datum as shown in Figure 5.15, and is given by:

119

N

f

Nf Z

g

Ph

(5.2)

where fh is the equivalent freshwater head [L],

NP is the pressure at point N within

the saline water [ML-1

T-2

], f is the density of freshwater [ML-3

], g is the

gravitational acceleration [LT-2

], and NZ

is the elevation of point N above an arbitrary

datum [L].

Figure 5.15. Two Piezometers, One Filled with Freshwater and the Other with

Saline Aquifer Water, Open to the Same Point in the Aquifer.

For the Lower Cretaceous aquifer, a set of six points is selected to

represent the observation data. Two of them are the available most recent

measured heads of Al-Monbateh and El-Halal-2 wells obtained from the North

Sinai’s General Directorate of Groundwater dating back to 2009, while the

other four observation points are chosen at the locations of known head

contours. For the Upper Cretaceous aquifer, a set of eight points is selected to

represent the observed heads. These eight points are chosen at the location of

known head contours. The locations of points for the Upper and the Lower

Cretaceous aquifers are shown in Figures 5.16 and 5.17, respectively.

The calibration process is performed as follows: a set of observed water

heads is provided, the flow model is executed several times, and the model

solution is imported to GMS each time. GMS automatically compares the

120

computed solution to the observation points, and the residual errors are

calculated. The sum of squared weighted residuals, with the weights assigned

based on the reliability and quality of each observation point, is then calculated

and compared to previous iterations. The process is repeated until the minimum

sum of squared weighted residuals is obtained. A plot showing the value of the

objective function (sum of squared weighted residuals) with the number of

model runs (iterations) is prepared and updated each time. This allows the

modeler to observe the calibration process and judge whether the model is

converging or diverging.

After PEST converges to an optimum solution, the solution is imported to

GMS, and a calibration goodness of fit (target bar) which represents the

magnitude of the residual error is displayed next to each observation point as

shown in Figures 5.16 and 5.17. The size of the target bar is based on the

standard deviation of the measurement error which is determined automatically

by assigning the limit intervals of the observation point. A standard deviation

of 0.75 m is considered for both aquifers.

Two parameterization schemes within PEST can be used; either to estimate

a single value for the hydraulic conductivity assuming a homogeneous aquifer

domain or to use different values at scattered points, known as Pilot Points, to

account for the heterogeneity of the aquifer until the objective function is

minimized. A common strategy is adopted to improve the first scheme by

subdividing the model domain into zones of assumed uniform parameter values

based on geological or other information. Unfortunately, such information is

often absent or unreliable. Furthermore, there can be a considerable degree of

variation of hydraulic conductivity within each geologic unit. As a result, the

pilot points’ methodology is very attractive; where instead of creating a zone

and having the inverse model estimate one value for the entire zone, the value

of the parameter within the zone is interpolated from the pilot points. Using this

technique, PEST is asked to assign hydraulic conductivities to discrete points

within the model domain. The hydraulic conductivity at each cell or node of the

121

numerical groundwater model is then calculated from the hydraulic

conductivities assigned to these pilot points using a spatial interpolation

algorithm such as Kriging or Inverse Distance Weighted (IDW) interpolation.

In this study the pilot points’ parameterization scheme is adopted with six

scatter points for each of the Lower Cretaceous and the Upper Cretaceous

aquifers and the IDW interpolation is utilized. Many trials were performed for

obtaining the most appropriate locations for the pilot points within the model

domain to obtain a minimum sum of squared residuals. Figure 5.16 shows the

locations of the best six pilot points in the Upper Cretaceous aquifer whereas

Figure 5.17 shows the locations of the best six pilot points in the Lower

Cretaceous aquifer.

Figure 5.16. Location of the Six Pilot Points Shown by the Triangular Symbol

on the Model Domain with the Potentiometric Map in the Background. Also

Shown are the Box Plots of the Errors Associated with each Observation Point

as Estimated from the Calibration Process for the Upper Cretaceous Aquifer.

122

Figure 5.17. Location of the Six Pilot Points Shown by the Rhombus Symbol

on the Model Domain with the Potentiometric Map in the Background. Also

Shown are the Box Plots of the Errors Associated with each Observation Point

as Estimated from the Calibration Process for the Lower Cretaceous Aquifer.

It is worth noting that during extracting the local model from the developed

regional model, the same pilot points sets of the Upper and the Lower

Cretaceous aquifer are used for interpolation for creating the heterogeneity

fields of the aquifers. Calibration targets are created in both aquifers for

validating the extraction process of the local model.

5.5.1. Estimated Parameters

The estimated parameters from the inverse modeling approach are the

horizontal hydraulic conductivities of the Lower Cretaceous and Upper

Cretaceous aquifers. An anisotropy value of 1:3 is assumed for obtaining the

values of the vertical hydraulic conductivities of the modeled aquifers. The

parameter estimation program, PEST, requires specifying an acceptable

interval for the estimated parameter. The lower and upper limits defining this

interval are given in Table 5.2. Also, the estimated hydraulic conductivity

values for the six pilot points for each aquifer obtained through the calibration

process are listed in the table.

123

Table 5.2. Parameter Estimation (PEST) Calibration Parameters and Estimated

Values and the Acceptable Intervals

Parameter Estimated

Value (m/day)

Acceptable Interval

Lower limit Upper limit

Hydraulic

conductivity for

the Upper

Cretaceous aquifer

Point 1 0.136568

1 1000

Point 2 17.86299

Point 3 0.00029

Point 4 1.135464

Point 5 1.131695

Point 6 0.186727

Hydraulic

conductivity for

the confining clay

layer

One

value for

the layer

3.26x10-6

0.0001 1x10-7

Hydraulic

conductivity for

the Lower

Cretaceous aquifer

Point 1 0.390443

1 1000

Point 2 0.28771

Point 3 269.2224

Point 4 0.06921

Point 5 12.67441

Point 6 21.58198

5.5.2. Assessment of Calibration

The results of the calibration are assessed by comparing the simulated and

measured heads as shown in Figures 5.18 and 5.19. The computed water heads

after the calibration process are plotted against the observed heads at the

available observation points for both the Upper Cretaceous and Lower

Cretaceous aquifers. The 45 degree line is also shown which represents the

perfect match between the modeled heads and the observed heads. The figure

shows that the simulated heads are very close to the observed heads. It should

be emphasized that the heads for both aquifers are either obtained from the

observed heads from the wells tapping the aquifers or the available

potentiometric maps as discussed earlier.

Tables 5.3 and 5.4 summarize the comparison between the observed and

the computed heads at the eight points of the Upper Cretaceous aquifer and the

six points of the Lower Cretaceous aquifer.

124

Figure 5.18. Comparison between the Computed Heads after Calibration and

the Observed Heads for the Eight Observation Points of the Upper Cretaceous

Aquifer.

Figure 5.19. Comparison between the Computed Heads after Calibration and

the Observed Heads for the Six Observation Points of the Lower Cretaceous

Aquifer.

Table 5.3. Comparison between the Observed Heads and the Model Computed Heads

for the Upper Cretaceous Aquifer.

Point Observed

Head (m)

Computed

Head (m)

Residual

(m)

% of

Residual

1 50 51.13 -1.13 -2.27

2 100 99.70 0.30 0.30

3 100 100.70 -0.70 -0.70

4 150 149.41 0.59 0.39

5 150 147.84 2.16 1.44

6 200 199.54 0.46 0.23

7 200 200.33 -0.33 -0.16

8 250 248.98 1.02 0.41

125

Table 5.4. Comparison between the Observed Heads and the Model Computed

Heads for the Lower Cretaceous Aquifer.

Point Observed

Head (m)

Computed

Head (m)

Residual

(m)

% of

Residual

1 (Halal-2) 30 32.63 -2.63 -8.76

2 (Monbateh) 34 33.58 0.42 1.23

3 100 100.10 -0.10 -0.10

4 50 48.70 1.30 2.59

5 50 50.68 -0.68 -1.35

6 100 100.19 -0.19 -0.19

To ensure that there is no trend in the errors, the percentage error (residuals

divided by the observed heads) are plotted for observation points of the Upper

and Lower Cretaceous aquifers in Figure 5.20. No correlation appears to exist

between the residuals. Also it can be seen from the figure that the maximum

error does not exceed 9% at one point only whereas all the other points have

errors below 3%.

Figure 5.20. Percentage Error at each of the Eight and Six Observations used

in the Model Calibration for the Upper and the Lower Cretaceous aquifer.

-10.00

-8.00

-6.00

-4.00

-2.00

0.00

2.00

4.00

0 1 2 3 4 5 6 7 8

% R

esid

ual

Lower Cretaceous

Upper Cretaceous

126

5.6. Transport Model

The transport model utilizes the flow results from MODFLOW and

incorporates advection and dispersive transport processes. Geochemical

reactions that may take place, especially for heavy metals, are not considered as

they are not present in the case of reject brine injection.

The purpose of the transport simulations is to forecast the release and

migration of the salts associated with the injected reject brine from the

proposed injection well located 200 m north Al-Monbateh production well. The

aim is to use these transport simulations to analyze different injection scenarios

and to examine the effects of different geologic interpretations and structures

on the distance traveled by the salt plume and the area impacted by the

contamination. This allows calculating the average salinity change for the

affected domain of the injection zone and thus an evaluation of the

environmental penalty resulting from the injection process in the Upper

Cretaceous aquifer.

The results of the flow model are used as the input for the transport model

along with relevant transport parameters. The used SEAWAT code was

developed by combining MODFLOW and MT3DMS into a single program that

solves the coupled flow and solute-transport through the governing equations

of the three-dimensional, variable-density, transient groundwater flow and

solute transport in porous media.

5.7. Base Case Scenario

The calibrated parameters are used to perform the base case simulations

which represent the current situation. The production rates of the wells in the

study region are taken as listed in Table 5.1. The resulting heads of the Upper

and Lower Cretaceous aquifers in the model domain are presented in Chapter

six.

In the base case, the clay layer is assumed to separate the Upper and the

Lower Cretaceous aquifers. This assumption is driven from the potentiometric

127

maps of the two aquifers. It might not be a conservative assumption because

there is uncertainty about the extent of the clay layer and whether it covers the

modeled region. The injection will take place in the Upper aquifer while the

production aquifer is the bottom aquifer which raises the risk of leakage of the

concentrate to the producing aquifer if that clay layer has a limited extent.

However, the effect of this assumption and the uncertainty associated with the

clay extension will be examined in Chapter six.

The specific storage and the specific yield are also required for the

modeling of the study region. Specific storage values are assumed 0.0001 m-1

and 0.00001 m-1

for the Upper and Lower Cretaceous aquifers, respectively,

whereas the specific yield is taken 0.05 and 0.15 for the Upper and Lower

aquifers, respectively. The assumption is based on the typical ranges for

specific yield for various aquifer materials described by American Society of

Civil Engineers (1996) in the Hydrology Handbook. The values for carbonate

(limestone) units usually range between 0.5 and 5% and for sandstone units;

they range from 5 to 15%.

For the transport simulations, the parameters used are the porosity and the

porous medium dispersivity for the two water-bearing formations. Porosity is a

critical parameter that determines how fast groundwater is moving and thus

how fast any solute dissolved in water will be moving. Usually the values of

the effective porosity for carbonate (limestone) ranges between 0.07 and 0.56

and for sandstone they range from 0.14 to 0.49 (McWorter and Sunada, 1977).

Values of porosity of both the Upper and Lower Cretaceous aquifers are very

limited and not reported in most reports, however, they are taken as 0.35 and

0.30, whereas a value of 0.4 is assumed for the porosity of the clay layer.

Many flow and transport simulation trials are performed for obtaining

reasonable values for both the longitudinal and transverse dispersivities. A

constant value for the ratio between the transverse to the longitudinal

dispersivity is set equal to 0.2 for both the horizontal and vertical transverse

dispersivities. Trials for longitudinal dispersivity of 100, 200, and 500 m for

128

the Upper Cretaceous aquifer and 50, 100, and 200 m for the Lower Cretaceous

aquifer are carried and the chosen longitudinal dispersivities are 200 m and 50

m for the Upper and Lower Cretaceous aquifers, respectively. The initial

salinity of the water of both aquifer are assigned to the model grid based on the

iso-salinity contour maps presented earlier in this chapter.

After establishing the base case, the flow field is updated to account for the

proposed injection well. The same model is used to obtain the transient state

flow field after adding the injection rate (16 m3/day). Different injection

scenarios are studied and the results are presented and discussed in the next

chapter.

129

CHAPTER SIX

INJECTION SCENARIOS RESULTS AND DISCUSSIONS

This chapter presents the results of the groundwater flow and transport

simulations associated with different injection scenarios. The simulation model

is run for a simulation time frame of 25 years. The base-case calibrated flow

model is first used for transport modeling followed by the modeling of the

different proposed injection scenarios. Some cases are also considered

addressing sensitivity and uncertainty issues as stated in chapter five.

The developed model is then used to simulate the case of extraction of Al-

Monbateh well from the Lower Cretaceous aquifer and the injection of the

reject brine in the Upper Cretaceous aquifer. Although the injection rate might

seem small (i.e., 16 m3/day) but it can result in an environmental deterioration

and economic penalty on a long-term basis. Injection of brine in deeply seated

layers of the Upper Cretaceous aquifer will result in increased concentration of

the salts in the water stored in the aquifer, which will result in an increased cost

of a later desalination of the stored water. This issue is addressed in this chapter

where the affected volume, the corresponding increased water salinity resulting

from the injection, and the economic penalty are calculated from the model

results.

The first case is the simulation of an extraction of 300 m3/day from the

Lower Cretaceous Sandstone aquifer through Al-Monbateh well and an

injection of 16 m3/day of reject brine in the Upper aquifer. The injection well

location is chosen such that the flow in the Upper Cretaceous aquifer moves the

reject brine to the northwest direction away from the domain of extraction,

regardless of the fact that the two aquifers are hydraulically not connected that

was based on the available potentiometric maps. This arrangement ensures that

130

no brine will migrate downwards to the extraction area in case the clay layer is

not present in some areas.

Simulation is run for 25 years to study the extent of the increased salt

concentration plume in both the horizontal and vertical directions around the

injection well. Three injection scenarios are studied:

Scenario one: Injection through a 100 m screen located at the bottom of

the Upper Cretaceous aquifer

Scenario two: Injection through a 50 m screen located at the bottom of

the Upper Cretaceous aquifer

Scenario three: Injection through a 50 m screen the end of which is at a

distance of 50 m from the bottom of the Upper Cretaceous aquifer

6.1. Groundwater Flow Results of the Base-Case Calibrated Model

The calibrated parameters are used to perform the base case simulations

which represent the current situation. The production rates of the wells in the

study region are taken as listed in Table 5.1. The resulting head distribution of

the Upper Cretaceous aquifer in the model domain is shown in Figure 6.1 while

the resulting heads of the Lower Cretaceous aquifer are shown in Figure 6.2.

The same model is used to obtain the transient state flow field after

assigning the injection rate (16 m3/day) for 25-years period. Continuous

injection commonly leads to a buildup of head around the injection well which

can create a strong vertical gradient leading to a vertical movement of the reject

through the clay layer to migrate to the Lower Cretaceous aquifer. However,

due to the small injection rate, the simulation results show an expected small

increase in the heads around the injection well which does not exceed one

meter.

131

Figure 6.1. The Head Distribution of the Upper Cretaceous Aquifer for the

Base-Case Calibrated Model

Figure 6.2. The Head Distribution of the Lower Cretaceous Aquifer for the

Base-Case Calibrated Model

6.2. Results of the First Injection Scenario

The injection takes place in the lowest 100 m of the Upper Cretaceous

aquifer and just above the clay layer separating the two aquifers. Figure 6.3

shows a plan view for the study area and a zoom-in view around the disposal

well to clearly show the concentrated salt plume. It is shown that after 25 years

of continuous injection of the reject brine, the concentrated salt plume with a

132

concentration contour 4.5 kg/m3 (4,500 ppm) migrates a distance of about 225

m west and 150 m north of the injection location while about 100 m east and

south of the injection point. The extent of the plume is more stretched in the

northwest direction which is aligned with the direction of groundwater flow in

this aquifer.

To depict the full three dimensional view of the concentrated salt plume,

two vertical cross sections passing through the well and oriented south-north

and west-east are shown Figure 6.4. It is shown that the plume did not migrate

downward and did not reach the Lower Cretaceous aquifer (i.e., the source of

feed water) because of the presence of the clay layer with the very low

hydraulic conductivity.

Figure 6.5 exhibits the time evolution of the relative concentration at

different distances along the northwest direction from the point of injection. It

can be observed that as we go farther from the injection point, the relative

concentration decreases. At 250 meters away from the injection well, the curve

is almost horizontal and no change in the concentration is witnessed.

Figure 6.3. Injection Results of the First Injection Scenario showing the Salt

Plume Distribution after 25 years of Continuous Injection in Plan View.

133

Figure 6.4. Injection Results of the First Injection Scenario showing the Salt

Plume Distribution after 25 years of Continuous Injection. The Top Part is a

East-West Cross-Sectional View and the Lower Part is a South-North Cross-

Sectional View.

Figure 6.5. Relative Concentration Curves at Different Distances From the

Injection Well for Scenario One.

0

1

2

3

4

5

6

7

8

0 5 10 15 20 25

Rel

ati

ve

Con

cen

trati

on

Time (years)

250 m

200 m

150 m

100 m

50 m

134

6.3. Results of the Second Injection Scenario

The injection in this scenario is assumed to occur through a 50-m screen

located at the bottom of the Upper Cretaceous aquifer. The resulting plume is

shown in Figures 6.6 and 6.7. After 25 years of continuous injection, the

concentrated salt plume with the concentration contour 4.5 kg/m3 migrates a

distance of about 225 m west and 150 m north of the injection location while

about 100 m east and south of the injection point. The extent of the plume is

more stretched in the northwest direction where the flow of the groundwater

occurs.

The two vertical cross sections passing through the well show that the

plume did not migrate downward and did not reach the Lower Cretaceous

aquifer because of the presence of the clay layer. However an upward

migration of about 75 m above the injection location is observed. The presence

of the clay layer beneath the injection zone bounds the salt migration

downwards in the vertical direction and helps the salt to spread more laterally.

The area impacted by the increased salinity is relatively smaller than that of the

first scenario and so is the volume. This is discussed in details in Section 6.6

6.4. Results of the Third Injection Scenario

Similar to the previous two injection scenarios, the injection takes place in

50 m of the Upper Cretaceous aquifer at a distance of 50 m above the clay layer

separating the two aquifers. Figure 6.8 shows a plan view for the resulting

plume and a zoom-in view around the disposal well to clearly show the plume.

It is seen that after 25 years of continuous injection at a rate of 16 m3/day, the

concentrated salt plume migrates a distance of about 175 m west and 125 m

north of the injection location while about 125 m east and south of the injection

point.

Figure 6.9 depict the full three dimensional view of the concentrated salt

plume. It shows two vertical cross sections where the salt plume migrates

downwards a distance of 75 m to reach the clay layer and migrates 100 m

135

above the injection point. Compared to the second injection scenario, the

absence of the clay layer just below the injection zone gives the plume the

freedom to spread downwards in the vertical direction.

The horizontal extent of the plume is relatively smaller than both the first

and second scenarios. However, the vertical migration is slightly larger which

results in a larger impacted volume. This result and the economic implications

are discussed in the environmental penalty of injection in Section 6.6

Figure 6.6. Injection Results of the Second Injection Scenario showing the Salt

plume Distribution after 25 years of Continuous Injection in Plan View.

136

Figure 6.7. Injection Results of Second Scenario showing the Salt plume

Distribution after 25 years of Continuous Injection. The Top Part is a East-

West Cross-Sectional View and the Lower Part is a South-North Cross-

Sectional View.

Figure 6.8. Injection Results of the Third Injection Scenario showing the Salt

plume Distribution after 25 years of Continuous Injection in Plan View.

137

Figure 6.9. Injection Results of the Third Injection Scenario showing the Salt

plume Distribution after 25 years of Continuous Injection. The Top Part is a

East-West Cross-Sectional View and the Lower Part is a South-North Cross-

Sectional View.

6.5. Uncertainty in the Clay Layer Extent

The clay layer separating the Lower and Upper Cretaceous aquifers is

assumed to extend entirely between the two formations based on the available

potentiometric and iso-salinity maps of both aquifers. However, there is no

guarantee that clay is laterally continuous and forms a complete confining layer

to the Lower Cretaceous aquifer along the full extent of the model domain. If

the clay layer is fully separating the Lower Cretaceous aquifer from the Upper

Cretaceous aquifer, then little or no migration downwards from the injection

zone to the lower aquifer would be expected.

The uncertainty in the lateral extent of the clay layer is addressed by

changing the hydraulic properties (i.e., the hydraulic conductivity) of the clay

layer beneath the injection zone for a distance of 250 m in both x and y

directions to the that of either the Upper or the Lower Cretaceous aquifers. The

transport simulation results of the two cases differ slightly. The case of

138

assigning the hydraulic conductivity values of the Upper Cretaceous aquifer is

discussed with the understanding that the other case yields more of less similar

results.

As the injection takes place in the lowest 50 m of the Upper Cretaceous

aquifer, Figure 6.10 displays the results of addressing the impact of uncertainty

in lateral extent of the clay layer beneath the injection zone located in lowest 50

m of the Upper aquifer. The figure shows the salt plume in the Upper aquifer.

The plume in this case spreads laterally equally in all direction comparing to

the previous three injection scenarios and with almost a distance of 200 m

around the injection point. Figure 6.11 shows the salt concentration plume in

plan view at successive 50 meters downwards from the start of the Lower

Cretaceous aquifer till the bottom of the aquifer. The plume in the Lower

aquifer takes a stretched shape like previous scenarios with elongation in the

direction of groundwater flow to the northwest. The cross-sections shown in

Figure 6.12 indicate that the vertical migration is larger than the previous

injection scenarios and reaches 175 m upwards from the point of injection and

downwards to the Lower Cretaceous aquifer.

Figure 6.10. Injection Results of Uncertainty in Clay Layer extension showing

the Concentrated Salt plume Distribution after 25 years of Continuous Injection

in Plan View.

139

1 2

3 4

Figure 6.11. Injection Results of Uncertainty in Clay Layer Extension showing the

Salts Concentration Plume Distribution in Plan at Successive 50 meters downwards

from the Injection Zone after 25 years of Continuous Injection.

140

Figure 6.12. Injection Results of Uncertainty in Clay Layer Extension showing

the Salt plume Distribution after 25 years of Continuous Injection. The Top

Part is a East-West Cross-Sectional View and the Lower Part is a South-North

Cross-Sectional View.

6.6. The Environmental Penalty of Injection

The injection of reject brine into groundwater aquifers has a number of

environmental implications and undesirable effects. These include changing the

hydraulic properties of the receiving aquifer such as reduction in permeability

of the host formation or the perforations or screens that are placed in the well’s

injection interval. This can be caused by particle/colloid migration into the

formation, bacterial growth, emulsification of fluids, and precipitation of

dissolved matter. Another impact is the fracture of the host geologic units

resulting in the hydraulic interconnection of the injection horizon and adjacent

aquifers. Also the build-up of high subsurface pressures can cause the

fracturing of confining strata and create pathways for the vertical migration of

injected fluids. Corroding or plugging of the injection wells are also

environmental penalties which add costs due to required maintenance. If a

plugging of an injection well occurs and the formation gets worse, the need of

141

larger injection pressures becomes crucial to maintain a given flow rate, which

can lead to well failure, causing the spread of concentrated reject brine and

compromising safety.

In this study we only focus on one of these environmental impacts

resulting from the injection of the reject brine in the deep aquifer and translate

it into an equivalent cost. The environmental penalty used here is defined as the

increased cost of desalination of the aquifer water volume that experiences an

adverse increase in salinity after the end of the simulation compared to the

background salinity of the aquifer.

The addressed environmental impact as defined above provides the

convenience of a relative comparison between the different scenarios using this

measure. Another reason behind using it is the possibility of expressing the

impact on the desalination costs – where desalination holds a great potential as

a future water resource.

In this environmental penalty approach, a benchmarked change in salinity

of 1,000 ppm is taken as the reference to account for the total volume

estimation. In other words, if a salinity of a certain volume of water has

increased by 1,000 ppm or more in the grid cells then an active flag is assigned

to the cell and this volume is taken in consideration in estimating the total

volume. The water volume is calculated by multiplying the volume of the grid

cells times the porosity of the layer. This volume of the water is then multiplied

by the salinity of each grid cell after the 25-year simulation period to estimate

the mass of the salts of the affected volume. By summation of the mass salts of

the total affected grid cells and dividing by their total volume, an average

salinity can then be estimated.

Using the same active flag array of the affected grid cells, and with the

same approach used above, the average salinity is calculated for the native

water of the aquifer. It should be noted that the background and the increased

142

salt concentrations vary from one cell to another in the model grid cells. In

summary, the equations used to obtain the average salinity are written as:

gw nVV (6.1)

gi

k

i

gii

k

i

avgV

VSS

1

1

(6.2)

where Vw is the water volume [m3], n is the aquifer effective porosity

[dimensionless], Vg is the grid cell volume [m3], Savg is the average salinity

[kg/m3], S is the salinity [kg/m

3], and K is the number of active flagged cells

(the cells with a salinity increase of 1 kg/m3 or more)

Without regard to the injection taking place in the Upper Cretaceous

aquifer and the production is from the Lower Cretaceous aquifer, the increase

in salinity of the Upper Cretaceous will reflect in an increased cost of

desalination in case of a future utilization of the aquifer. The environmental

penalty is the increased cost of desalination due to the salinity change that the

native aquifer water has experienced.

Excel spreadsheet is used to perform the calculation and MATLAB is used

for preparing and handling the output data of the GMS software to be input for

the spreadsheet. Figure 6.13 shows a screenshot of the Excel spreadsheet

developed.

Figure 6.13. Screenshot of the Excel Spreadsheet Used for Calculation of the

Environmental Penalty Affected Volume.

143

6.6.1. The Environmental Penalty of the Three Injection Scenarios

The injection of the reject brine in the Upper Cretaceous aquifer will

negatively impact a volume of water and change its average salinity. The

simulation of a 25-year period of continuous injection using the first injection

scenario has affected a water volume of about 2.625 million cubic meters of the

Upper Cretaceous. The average salinity of the affected volume has increased by

164 % reaching a value of 9,680 ppm whereas the native average salinity of the

affected volume is 3,660 ppm.

For the second injection scenario, the affected volume of water of the

Upper Cretaceous aquifer reaches 2,143,750 cubic meters where its average

salinity increases to reach 9,790 ppm with a percentage increase of 167% after

25-years of continuous injection.

The third injection scenario involves the injection takes place at distance

50 meters above the bottom of the Upper Cretaceous aquifer and the screen

length is 50 meters. For this scenario, the average salinity of a total affected

water volume of 2,712,500 m3 changes from 3,660 ppm to 8,800 ppm. In other

words the average salinity has increased by 140% after continuous injection of

the reject brine for 25-years.

For the three addressed scenarios, there is no effect on the Lower

Cretaceous aquifer due to the presence of the confining clay layer between the

two aquifers which separates the injection host formation from the extraction

aquifer.

6.6.2. The Environmental Penalty for the Uncertainty in the Clay Layer

Case

Figures 6.11 through 6.12 give a three dimensional perspective for the

injection transport results over 25-years of continuous injection. It is shown

that the salt plume migration in the Lower Cretaceous aquifer is significant

comparing to the first three injection scenarios. Also the vertical migration of

the salts in the Upper aquifer is significantly larger than that of the three

144

injection scenarios. Thus it is expected that a high environmental penalty will

result.

For the Upper Cretaceous, the affected water volume is about 5.95 million

cubic meters and its salinity increases by 75 % reaching a value of 6,410 ppm.

For the Lower Cretaceous, the water volume affected is in the order of 16.9

million cubic meters and its average salinity has increased by 130 % to reach a

value of 4,750 ppm instead of a native average salinity of 2,070 mg/L.

6.7. The Cost of Injection for the Three Injection Scenarios

The costs associated with injection wells are highly site specific and quite

variable. Therefore, generic cost estimates of injection wells from typical

handbook-type data is very difficult and subject to a wide range of uncertainty.

However, by gathering site-specific data, a reasonably accurate cost estimate

can be made. The major factors contributing to the cost of an injection well are:

pretreatment, pumps, site tests (i.e., logging, surveying, and testing), injection

well components, drilling, monitoring, maintenance and operating costs.

For the pretreatment, the rejected water may require pretreatment in an

above-surface facility to prevent plugging in the receiving formation. When

significant suspended solids are present, such as when concentrate is mixed

with membrane pre-filter backwash and periodic cleaning waste, typical

pretreatment consists of total suspended solid removal is required. Also pH

adjustment may be necessary. Pumps are used in above-surface facilities to

inject the concentrate. The flow and pressure requirements are site specific.

The discharge head will vary depending upon the geologic conditions and

depth of the injection zone. In this study, the injected volume is very small,

thus the required head from the simulation did not exceed one meters and the

injection can take place by free-fall of the concentrate.

Deep injection wells are normally multi-cased. Usually the use of more

than one casing to provide transition zones and isolate contaminated aquifers

from water contained in shallower or deeper aquifers is adopted. It is not

145

common to inject water in an aquifer where the source of water is seated in

deeper than the injection aquifer so the contaminated water does not seep

downwards to the source aquifer. However, in this study since the iso-salinity

maps shows that the Upper Cretaceous aquifer has higher salinity values than

that of the Lower Cretaceous aquifer and Al-Monbateh is extracting from the

Lower Cretaceous aquifer, the injection takes place in the Upper Cretaceous

aquifer.

The monitoring is required to ensure compliance with environmental

regulations. Also periodic samples can be taken and analyzed to determine if

there has been any leakage of the concentrate to the feed water aquifer. In

general, the most critical areas are that around the injection location in the

Lower Cretaceous aquifer. The operating costs for disposal wells are generally

low. Well maintenance consists of periodically checking the casing and

repairing it if required. In this study, an inclusive cost for the injection well of

EGP 4,000 per meter are assumed based on surveys of recent deep injection

well costs.

The total cost of the injection wells considered here is the cost of well

installation, operation, and maintenance in addition to the cost of the

environmental penalty which is the difference between the desalination cost of

the total affected water volume before and after the injection. The depths of the

injection well are 611 m, 611 m, and 561 m for the three scenarios,

respectively. Thus the inclusive cost of the injection well is calculated.

Future extraction and utilization of the saline water of the Upper

Cretaceous aquifer in desalination is expected and of a possible choice since

the great achievements in desalination technology have now moved the costs

for desalting in many applications from the realm of "expensive" to

"competitive". The desalination cost is a function of the feed water salinity,

where an increase in the salts concentration of the feed water will reflect on an

increased desalination costs. The costs of unit cubic meter desalination using

RO processes as a function of feed water salinity and the plant capacity are

146

shown in Figure 6.14. Table 6.1 summarizes the total cost estimates for the

three injection scenarios.

Figure 6.14. Unit Cost of Brackish Water RO Desalination with Plant

Capacity (Khidr, 2012).

Table 6.1. Summary of the Costs of the Injection Wells of the Three Scenarios Scenario Depth

(m)

Well Cost

(EGP)

Affected

Volume

(m3)

Salinity

(ppm)

EGP

per

m3

Environmental

Penalty Cost

(EGP)

Total Cost of

Injection

(EGP)

1 611 2,444,000 2,625,000 9,700 1.77 4,646,250 7,090,250.00

2 611 2,444,000 2,143,750 9,800 1.82 3,901,625 6,345,625.00

3 561 2,244,000 2,712,500 8,800 1.44 3,906,000 6,150,000.00

Figure 6.15 shows the cost estimated for the three scenarios. The black

portion gives the cost of the well drilling and ooperating, whereas the grey

component pertains to the environmental penalty. It can be seen from the figure

that the third disposal scenario yields the lowest well installation cost and the

lowest environmental penalty. Therefore, the third scenario gives the lowest

total cost among the three scenarios.

0

0.2

0.4

0.6

0.8

1

1.2

0 1000 2000 3000 4000 5000 6000 7000 8000 9000 10000

Co

st (

US

D/m

3)

Plant Capacity

3000-5000 ppm

5000-10000 ppm

147

Figure 6.15. The Cost of Disposal of the Proposed Three Injection Scenarios.

Figure 6.16 recalls back the costs of the current disposal system of the Al-

Monbateh desalination plant (i.e., the lined evaporation ponds) and sets the

costs of the two different alternatives together in one graph for a relative

comparison. The figure shows that the deep well injection disposal option

yields an exaggerated cost compared to that of the evaporation ponds, whether

compared to the current ponds or the proposed ponds which takes into account

the actual evaporation rates, rainfall, and salinity effects. The assessment

results stand in a favor of the evaporation ponds choice regardless of the fact

that the sizing was not properly done. This supports the use of evaporation

ponds for membrane concentrate disposal as it is most appropriate for smaller

volume flows as well as for regions having a relatively warm, dry climate with

high evaporation rates, level terrain, and low land costs.

2.4

44

2.4

44

2.2

44

7.0

9

6.3

46

6.1

5

0

1

2

3

4

5

6

7

8

Scenario 1 Scenario 2 Scenario 3

Tota

l C

ost

(E

GP

)

Mil

lion

s

148

Figure 6.16. The Total Costs of the Different Disposal Options.

6.8. The Proposed Potential Areas for Future Extraction of Feed Water

within the Study Area

Sinai could be self sufficient in satisfying its domestic water demand under

the proper water management. Extensive development in the socio-economic,

industrial and agricultural sectors is expected to be stressed in the future. The

Lower Cretaceous aquifer is the most prospective aquifer in Sinai as mentioned

earlier by many researches. The aquifer is not yet effectively utilized nor

precisely evaluated, although it represents a strategic reserve for future

economic development. Spatially distributed areas of good quality groundwater

suitable for various types of development are identified and the suitable areas

for domestic and irrigation purposes are delineated in this study based on four

design criteria. Three of the criteria pertain to the Lower Cretaceous aquifer

and these are the depth to water (which can also be referred as depth to

aquifer), the thickness of the aquifer and the quality of the water (e.g., salinity).

The forth criterion is the topography of the area.

The results can be considered to be useful for preparing preliminary

groundwater development plans for the studied domain. The desalination

potential is addressed through a preliminary estimate for the water needed for a

0.541

1.8

85

2.2

44

Uncertain

Costs

6.1

5

0

1

2

3

4

5

6

7

Exiting Pond New Pond Deep Well

Injection

Tota

l C

ost

(E

GP

) M

illi

on

s

149

domestic usage in the area and a projection for the two common brine disposal

options is applied for the proposed exploitation scheme. The costs of the two

disposal options are studied and presented as a function of the demand. The

desalination technology proposed is the RO process with a recovery rate of

70% and a reject brine concentration of 12,000 mg/L as given in Al-Monbateh

desalination plant (i.e., the first stage of desalination only is taken). The cost

elements of the evaporation ponds are taken as discussed earlier in Chapter

three whereas the injection wells as presented in this chapter.

The evaluation of groundwater development in the study area relies on four

criteria as mentioned above. The study area is divided into polygons with high,

fair, and low development priority and values of 3, 2, or 1 is assigned for each

polygon. The four criteria are included in the evaluation by summation of the

polygonal values of each criterion and averaging the results to get a value

between 1 and 3 expressing the priority of the groundwater development with 3

as highest and 1 as lowest. It is worth noting that the water quality criterion was

given a higher weight in the averaging process. Figure 6.17 shows the polygons

of each criterion and their priority with the lighter color giving the highest

while the darker giving the lowest groundwater development priority. Figure

6.18 presents the results of the averaging (i.e., overlying the four criteria) and

the spatial potential areas are expressed in lightest colors.

150

Land Levels

Depth to the Lower Cretaceous

Aquifer

Thickness of the Lower Cretaceous

Aquifer

Salinity of the Lower Cretaceous

Aquifer

Figure 6.17. The Four Criteria for Potential Groundwater Development.

The most suitable areas are chosen as shown in Figure 6.19 and the

expected demand is calculated based on the population density of north Sinai.

The World Bank (2012) (http://maps.worldbank.org/mena/egypt-arab-republic,

accessed on: May, 2012) estimated the current population density as

approximately 44. However, a value of 100 capita per square kilometer and 200

cubic liters as a daily consumption are assumed in estimating the total demand.

Three zones are identified and the estimated future demand is 7600 m3/d, 7100

m3/d, and 4000 m

3/d for the three areas shown in the figure. It is worth noting

the Al-Monbateh well lies in the highest priority zone.

http://maps.worldbank.org/mena/egypt-arab-republic

151

Figure 6.18. The Overly of the Four Potential Criteria (the light color

expresses the high potential while the dark color expresses low

potential).

Figure 6.19. The Proposed Suitable Areas for Groundwater

Development.

152

The developed regional model is utilized to simulate the proposed

extraction and injection of the brine. The groundwater heads after the 25-year

simulation for the Upper and Lower Cretaceous aquifers are shown in Figures

6.20 and 6.21, respectively. The change in regional water heads is minute if

compared to that of the base-case scenario. The resulting flow fields are used in

the solute transport simulation.

Figure 6.22 shows the transport simulation results where the simulation is

conducted under the total extraction and injection rates. The figure shows the

lateral migration of each injection well field on the regional model grid where

each cell has dimensions of 250 x 250 meters. Five injection wells are assumed

for the studied domain as a preliminary step for injection. For the first injection

well field, the maximum lateral migration is about 700 m, while it is 600 m for

the second zone and only 375 m for the third injection area. The same approach

for evaluation of the environmental penalty used in the local disposal scenarios

is applied in the regional case and the affected volumes are estimated as well as

the increased desalination costs in case of future extraction. The calculation of

the environmental penalty is based on the present value of money and does not

consider the expected achievements in the desalination technologies which will

reflect on reduced costs of desalination.

A detailed study regarding the allowable injection rate and required

pressure is required before any step of groundwater development, desalination,

and injection of brine. However, the assumed hypothetical case can be

beneficial in a future detailed study with a proper management scheme for the

extraction and injection wells.

153

Figure 6.20. The Head Contours of the Upper Cretaceous Aquifer for the

Proposed Potential Extraction and Injection Model (The proposed wells and the

existing wells are presented as black dots).

Figure 6.21. The Head Contours of the Lower Cretaceous Aquifer for the

Proposed Potential Extraction and Injection Model (The proposed wells and the

existing wells are presented as black dots).

154

Figure 6.22. Injection Results of the Upper Cretaceous Aquifer for the Proposed Potential Extraction and Injection Model (The

proposed wells and the existing wells are presented as black dots).

155

6.9. The Cost of the Different Disposal Options for the Proposed

Groundwater Development Potential

The cost of disposal for the two proposed alternatives (evaporation pond

and deep well injection) can be estimated based on the cost elements discussed

earlier in this chapter and in Chapter three during the evaluation of the current

disposal system of Al-Monabateh plant. The cost of the environmental penalty

is added also to the cost of implementation of the injection wells.

For calculating the cost of the lined evaporation ponds, the model

discussed in Chapter three is used for the proper sizing of the ponds using the

same criteria of the actual evaporation rates, rainfalls and effect of salinity on

reducing the evaporation rates. Different production rates are assumed and the

corresponding expected reject brine is used as the input water to the pond. The

size is obtained then the cost is roughly estimated based on this size. Table 6.2

shows the estimated costs with different extraction discharges.

Table 6.2. The Estimated Costs of the Lined Evaporation Ponds with Different

Extraction Rates

Extraction (m3/d) Product (m

3/d) Brine (m

3/d) Pond Cost (EGP)

3200 2240 960 53,992,185

6400 4480 1920 105,785,145

9600 6720 2880 156,461,681

16000 11200 4800 261,733,144

24000 16800 7200 392,857,283

The cost of the deep well injection disposal option is estimated by

assuming a number of injection wells for the three proposed groundwater

development zones. Both the well implementation costs and the environmental

penalty costs are added to express the total cost of the wells. Table 6.3 shows

the estimated costs of the deep injection disposal option.

156

Table 6.3. The Estimated Cost of the Deep Well Injection Disposal Option

Extraction

(m3/d)

Production

(m3/d)

Reject Brine

(m3/d)

Number of

Injection Wells

Implementation

Cost

(EGP)

Volume

Affected

(m3)

Penalty Cost

(EGP)

Total Cost

(EGP)

4650 3255 1395 1 2,592,000 43,750,000 39,943,750 42,535,750

9300 6510 2790 2 5,168,000 86,406,250 78,888,906 84,056,906

15200 10640 4560 3 7,640,000 146,562,500 133,811,563 141,451,563

21000 14700 6300 4 10,108,000 190,312,500 173,755,313 183,863,313

23860 16702 7158 5 13,016,000 217,656,250 198,720,156 211,736,156

26720 18704 8016 5 13,016,000 234,062,500 213,699,063 226,715,063

157

Figure 6.23 presents the costs of the two brine disposal options as a

function of the product water quantity. Shown are the total cost of the

evaporation ponds and the total cost of the deep well injection. The two cost

components of the deep well injection are presented with the dashed lines. The

figure shows that the cost of the deep well injection is generally less than that

of the lined evaporation ponds and the cost difference increases with increasing

the product water (i.e., increasing the reject brine) which can give a preliminary

claim that deep well injection is a favorable option in case of large brine

volumes. However evaporation ponds for membrane concentrate disposal are

most appropriate for smaller volume flows, which was the case in Al-Monbateh

desalination plant.

Figure 6.23. The Cost of Disposal of Different Alternatives with the Product

Water Quantity.

Although the deep well injection proves an economic feasibility, but there

are many advantages associated with the use of the evaporation ponds: (1) they

are relatively easy and straightforward to construct; (2) properly constructed

evaporation ponds are low maintenance and require little operator attention

0

50

100

150

200

250

300

350

400

450

0 5000 10000 15000 20000

Tota

l C

ost

s (E

GP

)

Mil

lion

s

Product Water (m3/day)

Total Cost of Evaporation Ponds

Total Cost of Injection

Injection Wells Capital Cost

Penalty of Injection

158

compared to mechanical equipment; (3) except for pumps to convey the

concentrate water to the pond, no mechanical equipment is required; and (4) for

smaller volume flows, evaporation ponds are frequently the least costly means

of disposal, especially in areas with high evaporation rates and low land costs.

Taking into consideration that injecting back the reject brine is a conservative

water mass attitude which means that the reject water will be locally stored

again in the aquifer domain and thus decrease the quantity loss of water.

Achievements in the desalination technologies are expected in the future which

may overcome the higher costs associated with higher salinity waters.

Despite the inherent advantages of evaporation ponds, they are not without

disadvantages that can limit their application. First, they can require large tracts

of land if they are located where the evaporation rate is low or the disposal rate

is high. Second, they mostly require impervious liners of clay or synthetic

membranes such as polyvinylchloride (PVC) or Hypalon, and this requirement

substantially increases the costs of evaporation ponds. Third, seepage from

poorly constructed evaporation ponds can contaminate underlying potable

water aquifers and cause an increased environmental penalty. Lastly, there is

little economy of scale for this land-intensive disposal option. Consequently,

disposal costs can be large for all but small-sized membrane plants.

It is worth noting, however, that regardless of the high cost of the lined

evaporation ponds, they can be utilized as solar ponds and thus they provide a

renewable energy source for the desalination plant where the energy is to be

harnessed for operating high compression pumps needed for reverse osmosis

modular systems – the promising desalination technology. And the reasons are

obvious, since Egypt has great potential of brackish water wells, immense

amounts of solar radiation in remote areas and future integrated development

projects are located at a distance from the Nile water.

159

CHAPTER SEVEN

SUMMARY, CONCLUSIONS AND

RECOMMENDATIONS FOR FUTRE WORK

7.1. Summary

Awareness of increasing water scarcity has driven efforts to seek for non-

conventional water resources. Atop of these resources is the saline water of

both the sea and the groundwater aquifers. Desalination of brackish water holds

a great promise as a freshwater resource which helps ameliorate the stress on

the Nile River as the only renewable source of water in Egypt. Brackish

groundwater is usually present in vast quantities, where inland desalination can

be utilized. In parallel with the implementation of an inland desalination plant,

a disposal system for the produced reject brine has to be developed. This

disposal should have the mildest effects on the environment and be cost

effective. The two main disposal alternatives, the evaporation ponds and the

deep injection into saline aquifers, are addressed through a field case study, Al-

Monbateh desalination plant that already exists in Central Sinai.

For the assessment of the current disposal system of the plant, the

evaporation ponds, a MATLAB simulation code utilizing the water and salt

balance was developed for the evaluation process considering the effect of

salinity on the evaporation rates and its projection on the area of the pond. The

current disposal system malfunctioned after a couple of months of operation

and concentrate started to flood, and thus pumping units were attached and an

unmanaged disposal took place on land surface few meters away from the

constructed lined evaporation ponds. For a fair comparison between the current

disposal and other alternatives, the code is used for investigating the cause of

160

the problem. It is found that the main issue is the reduced evaporation rates due

to the high salinity of the reject. New dimensions are calculated for an

independent evaporation pond disposal alternative and the associated

approximate cost of the pond is calculated. Sensitivity analysis for the effect of

salinity is carried out for assessing its importance in the design of the ponds.

Deep injection of the reject brine in saline aquifers is studied as an

alternative for the evaporation pond. The injection host is the Upper Cretaceous

aquifer while the extraction of Al-Monbateh well is from the Lower Cretaceous

aquifer. The fact that the two aquifers are not hydraulically connected is

strengthened by the different water head and salinity values for the same spatial

point at the two aquifers. For the proposed injection scheme, the groundwater

flow direction around the disposal well is simulated and the possibility of such

disposal and its short- and long-term sustainability are evaluated in this study.

An environmental penalty is defined for the study as the volume of the

groundwater flow that is adversely affected by the increased salinity of more

than 1,000 ppm and its projection on the cost of desalination, in case of future

utilization of the resource.

In order to investigate the abovementioned issues, a regional three-

dimensional numerical groundwater flow and solute transport model is applied

and used to evaluate the impact of the proposed disposal and address the

uncertainty associated with the subsurface characteristics, processes and

injection location of the reject brine. The different versions of the groundwater

model are developed and run using well established groundwater packages.

The USGS groundwater flow model, MODFLOW, and the associated solute

transport model, MT3DMS, in conjunction with SEAWAT for variable density

groundwater flow simulation are used in this study. The simulation timeframe

is taken as 25 years of continuous injection of the reject brine from Al-

Monbateh desalination plant. For the small injection rate associated with the

plant, a local model is extracted from the regional model and a finer grid

161

discretization is used for a more accurate simulation of the groundwater flow

and salt transport.

Two water-bearing formations are modeled, the Upper and the Lower

Cretaceous aquifers. Confined aquifer conditions are assumed for the Lower

Cretaceous while unconfined conditions are assumed for the Upper aquifer.

Specified head boundaries are assigned to the lateral model boundaries expect

for the northern boundary of the Lower Cretaceous aquifer that is assigned as a

no-flow boundary. This is because it represents a deeply faulted limestone unit

terminating the Lower Cretaceous aquifer. No recharge is assumed for both

modeled aquifers.

The groundwater flow model is calibrated using the head values extracted

from potentiometric maps of both aquifers and from two wells tapping the

Lower Cretaceous aquifer. Hydraulic conductivity is considered the calibration

parameter, and the calibration is performed using the Pilot Points method

integrated with the PEST (parameter estimated) code. The calibration process

is assessed using the sum of squared errors (the difference between the

observed and the simulated heads). Heterogeneity fields for both aquifers are

created as a result of calibration. After the calibration is done, the base case

scenario for the flow and solute transport is developed based on the current

production from the two aquifers from the present nine wells in the regional

study domain. The reject brine injection is at a rate of 16 m3/day, and the local

model is utilized in the simulation of the injection scenarios.

Preliminary groundwater development potential areas are identified in the

study area based on four criteria; the topography of the area, the depth to the

producing aquifer which is the Lower Cretaceous aquifer, the thickness of the

aquifer and the salinity of the water. A projection of the two disposal

alternatives is carried on the potential areas for extraction in order to assess the

environmental and economic feasibility of the different alternatives for brine

disposal in case desalination is utilized in future development expansion. The

162

regional model is used to assess the environmental impacts and penalty

resulting from the deep well injection option.

7.2. Conclusions

After analyzing the results presented throughout this study and the case

study on the two disposal alternatives, the following conclusions can be drawn

from the analysis:

1. Salinity effect on evaporation rates is a critical factor in the design of

evaporation ponds.

2. For small disposal volumes the evaporation pond design is very

sensitive to the side slopes while not in case of large volumes.

3. Deep well injection is an attractive disposal alternative for large disposal

volumes but probably not so for small volumes compared to lined

evaporation ponds, however several aspects involve the decision on

which is the better alternative and the most appropriate alternative is

highly site specific.

4. It is anticipated that evaporation ponds most likely is a competitive

option for relatively small plants in remote, inland locations with high

evaporation rates.

5. Although deep well injection causes unpleasant impacts which are

reflected on an increased cost for desalination, but the option preserves

the amount of reject brine produced in the host formation instead of

losing these volumes through evaporation. These volumes might be a

great resource of water in the future especially with the rapid evolution

of the desalination technologies and the ever decreasing costs.

6. For the case study, the injection at a distance of 50 meters above the

clay lens is found to be the most economic among the deep injection

options.

7. For the case study, the lack of a clay lens beneath the injection zone has

high environmental penalty for the Lower Cretaceous, thus accurate

163

aquifer characterization is important before choosing the injection well

location.

7.3. Recommendations for Future Work

Based on the results of this study and through the literature for the

previous studies of the disposal options, the following is recommended for

future studies:

1. Study in depth the effect of salinity on evaporation rates and the

dependency on the geographical location.

2. In case of location dependency, field experiments are needed to

formulate relationships for the different regions of Egypt.

3. Consider the chemical characterization of the injected reject brine in the

injection simulation as well as the temperature of the injected fluid to

account for the effects of viscosity variations on groundwater flow.

4. Survey the potential aquifers for brackish groundwater for desalination

with an estimate of the stored volumes and the allowable exploitable

volumes for the main brackish aquifers in Egypt.

5. Study the possible development schemes for Central Sinai and other

areas rich with brackish water through the appropriate management and

decision-making tools.

6. Investigate the possible enhancements for the addressed disposal

alternatives and especially for the evaporation ponds, like the use of the

evaporation pond as a source of renewable energy (i.e., solar pond) for

operating the desalination plant thus achieving the integrity of an

independent system of producing fresh water,

7. Study the use of unlined evaporation ponds, with the environmental

impacts and penalties associated with the expected seepage

8. Study the feasibility of using mechanical evaporation in the disposal of

the concentrate.

164

REFERENCES

1. Abd El Rahman, H. (2001). Evaluation of Groundwater Resources in

Lower Cretaceous Aquifer System in Sinai, Water Resources

Management 15, pp. 187–202.

2. Abou Rayan, M., Djebedjian, B., & Khaled, I. (2001). Water supply and

demand and a desalination option from Sinai, Egypt. Desalination,

136, pp. 73-81.

3. Ahmed, M., Shayya, W. H., Hoey, D., Mahendran, A., Morris, R., & Al-

Handaly, J. (2000). Use of evaporation ponds for brine disposal in

desalination plants. Desalination, 130(2), pp. 155-168.

4. Ahmed, M., Shayyab, W. H., Hoey, D., & Al-handaly, J. (2001). Brine

disposal from reverse osmosis desalination plants in Oman and the

United Arab Emirates, Desalination, 133, pp. 135-147.

5. Ahmed, M., Hoey, D., Thumarukudyd, M. R., Goosen, M. F. A., Al-

haddabi, M., & Al-belushi, A. (2003). Feasibility of salt production

Corn inland RO desalination plant reject brine: a case study.

Desalination, 158, pp. 109-117.

6. Ali, A. A. M. (2009). Assessing Environmental Impacts of Disposal of

Gas Production Byproducts in Deep Aquifers: Application to El

Wastani Formation. A master’s thesis. Department of Irrigation and

Hydraulics, Cairo University, Egypt.

7. Allam, A. R., Saaf, E., & Dawoud, M. A. (2002). Desalination of brackish

groundwater in Egypt, Desalination, 152, pp. 19-26.

8. Arnal, J. M., Sancho, M., Iborra, I., Gozálvez, J. M., Santafé, a., & Lora,

J. (2005). Concentration of brines from RO desalination plants by

natural evaporation. Desalination, 182(1-3), pp. 435-439.

9. Attia, F. A. Groundwater resources management in Egypt in the concept

of IWRM, Mediterranean Document on Groundwater.

10. Baxter, G.P., and Wallace, C.C., 1916, Changes in volume upon solution

in water of halogen salts of alkali metals: IX, American Chemical

Society Journal, no. 38, p. 70-104.

11. Buros, O. K., Cox, R. B., Nusbaum, I., El-Nashar, A.M., and Bakish R.

165

(1980). The USAID Desalination Manual. Englewood Cliffs, NJ:

IDEA Publications.

12. Burston, I. and A Akbarzadeh. (1995). Integration of Solar Ponds into Salt

Affected Areas. Conference Proceedings, International Symposium

on Energy, Environment and Economics, University of Melbourne.

13. Chow, V.T., Maidment, D.R. and Mays, L.W. (1988) Applied Hydrology,

McGraw-Hill, New York, NY.

14. Committee on Advancing Desalination Technology, National Research

Council. (2008). Desalination: A National Perspective. Washington,

DC: The National Academies Press.

15. Cosgrove, W.J. and Rijsberman, F.R. (2000). World Water Vision:

Making Water Everybody’s Business. Earthscan Publications,

London.

16. Daintith, M. (1996). Rotifers and Artemia for Marine Aquaculture: A

Training Guide. University of Tasmania.

17. Doherty, J. (1994). PEST. Watermark Computing, Corinda, Australia.

18. El-Kady, M., El-Shibini, F. (2001). Desalination in Egypt and the future

application in supplementary irrigation, Desalination, 136, pp. 63–72.

19. El-Sadek, A. (2010). Water desalination: An imperative measure for water

security in Egypt. Desalination, 250, pp. 876-884.

20. EPIQ Water Policy Team. (1998). Hydrogeology of Deep Aquifers in the

Western Desert and Sinai, report no. 10. Ministry of Public Works

and Water Resources in Egypt and the US Agency for International

Development.

21. Gilron, J., Folkman, Y., Savliev, R., & Waisman, M. (2003). WAIV -

wind aided intensified evaporation for reduction of desalination brine

volume, Desalination, 158, pp. 205-214.

22. Guo, W., and Langevin, C.D., 2002, User’s guide to SEAWAT: A

computer program for simulation of three-dimensional variable-

density ground-water flow: U.S. Geological Survey Techniques of

Water-Resources Investigations, book 6, chap. A7.

23. Harbaugh, A.W., and McDonald, M.G. (1996), User's documentation for

MODFLOW-96 an update to the U.S. Geological Survey modular

finite-difference ground-water flow model: U.S. Geological Survey

Open-File Report 96-485.

24. Javandel, I., Doughty, C., and Tsang, C.F. (1984) Groundwater Transport:

166

Handbook of Mathematical Models, American Geophysical. Union,

Water Resources Monograph.

25. Khalil, E. E. (2004). Water strategies and technological development in

Egyptian coastal areas, Desalination, 165, pp. 23–30.

26. Khidr, S. (2012). Economic of Using Desalinated Seawater and Brackish

Water in Supplementary Irrigation. A Master’s thesis. Department of

Irrigation and Hydraulics, Cairo University, Egypt.

27. Khordagui H. (1997). Environmental aspects of brine reject from

desalination industry in the ESCW region, ESCWA, Beirut,

Lebanon.

28. Lam, N. S. (1983). Spatial interpolation methods review. The American

Cartographer, 10, pp. 129-149.

29. Macedonio, F., Katzir, L., Geisma, N., Simone, S., Drioli, E., & Gilron, J.

(2011). Wind-Aided Intensified eVaporation (WAIV) and Membrane

Crystallizer (MCr) integrated brackish water desalination process:

Advantages and drawbacks. Desalination, 273(1), pp. 127-135.

30. Mahmoud A. M. H. (2011). Evaporation Pond Modeling for Closed Basin

Drainage. A doctoral thesis. Irrigation and Hydraulic Department,

Cairo University, Egypt

31. Malmrose, P., J. Lozier, M. Mickley, R. Reiss, J. Russell, J. Schaefer, S.

Sethi, J. Manuszak, R. Bergman, and K. Atasi. (2004). Current

perspectives on residuals management for desalting membranes.

AWWA, 96 (12), pp. 73-87.

32. McDonald, M.G., and Harbaugh, A.W. (1988). A modular three-

dimensional finite-difference ground-water flow model: Techniques

of Water-Resources Investigations of the United States Geological

Survey, Book 6, Chapter A1.

33. McWorter, D.B., and D.K. Sunada, 1977, Groundwater Hydrology and

Hydraulics, Water Resources Publications, Fort Collins, Colo.

34. Mickley, M., Hamilton, R., Gallegos, L., & Truesdall. J. (1993).

Membrane concentration disposal. American Water Works

Association Research Foundation. Denver, Colorado.

35. Mickley, M. (1995). Environmental considerations for the disposal of

desalination concentrate. International Desalination Association

Conference. Abu Dhabi. United Arab Emirates.

36. Mickley, M. (2010). Brackish Groundwater Concentrate Management.

Mickley & Associates Boulder, Colorado.

167

New Mexico State University.

37. Midwater. (2010). Design basis and plant description of Al-Monbateh

desalination plant - Life Sinai Program. Cairo, Egypt.

38. Moawad, A. (2008). Future of Desalination in Egypt. The First Arab

Water Forum. Riyadh, Saudi Arabia

39. Mohamed, A., Maraqa, M., & Al Handhaly, J. (2005). Impact of land

disposal of reject brine from desalination plants on soil and

groundwater. Desalination, 182(1-3), pp. 411-433.

40. Nassar, M. K. K. (2004). Design and management of discharge and

recharge well fields for coastal desalination plants. A master’s thesis.

Department of Irrigation and Hydraulics, Cairo University, Egypt.

41. Navarro A. and Carbonell M. (2007). Assessment of groundwater

contaminant caused by uncontrolled dumping in old gravel quarries

in the Beso’s aquifer (Barcelona, Spain). Environmental

Geochemistry Health, 30. pp. 273-289

42. Nile Consultants. (2003). Water Resources Supply for Bedouins

Settlement in Sinai. Final Report, Vol. 1. In conjunction with the

Ministry of Water Resources and Irrigation, Cairo, Egypt.

43. Rachid, M., and C. Abdelwahab. (2005). Ecotoxicological marine impacts

from seawater desalination plants. Desalination 182. pp. 403-410.

44. Saripalli, K. P., M. M. Sharma and S. L. Bryant, (2000). Modeling

injection well performance during deep-well injection of liquid

wastes, Hydrology, 227. pp. 41-55.

45. Sethi, S., S. Walker, P. Xu, and J. Drewes. (2006a). Existing and

emerging concentrate minimization and disposal practices for

membrane systems, Florida Water Resources Journal, pp. 38-48.

46. Skehan, S., & Kwiatkowski, Peter J. (2000). Concentrate disposal via

injection wells permitting and design considerations, Florida Water

Resources Journal, pp.19-21.

47. Snyder, R. L., Orang, M., Matyac, S., Grismer, M. E., & Asce, M. (2005).

Simplified Estimation of Reference Evapotranspiration from Pan

Evaporation Data in California. Journal of Irrigation and Drainage

Engineering, (June), 249-253.

48. Talaat, H.A., Sorour, M.H., Abulnour, A.G., Shaalan, H. F. (2002). The

potential role of brackish water desalination within the Egyptian

water supply matrix. Desalination, 152, pp. 375-382.

168

49. Thomson, M. (2003). Reverse-osmosis desalination of seawater powered

by photovoltaics without batteries. A doctoral thesis. Loughborough

University.

50. Tzen, E. (2005). "Successful RES Desalination Applications," Proceeding

of the International Seminar, Hammamet, Tunisia.

51. USBR (U.S. Bureau of Reclamation). (2003). Desalting Handbook for

Planners, 3rd Edition. Washington, D.C., USA.

52. Wangnick Consulting. (2000). IDA worldwide desalting plants inventory.

Report no. 16. Gnarrenburg, Germany,

53. Xu, T. (2005). "Ion exchange membranes: State of their development and

perspective," Journal of Membrane Science 263(1-2), pp. 1-29

54. Younos T., & Tulou, K. E. (2005). Overview of desalination techniques,

Virginia Polytechnic Institute and State University, Universities

Council On Water Resources, Journal Of Contemporary Water

Research & Education, 132, pp. 3-10.

55. Yousef, R. M., Sakr M.L., & Shakweer, A. F. (2007). Desalination

Technology Roadmap 2030, The Cabinet Information and Decision

Support Center Center for Future Studies, Cairo, Egypt.

56. Zheng, C. (1990). MT3D: A Modular Three-Dimensional Transport

Model for Simulation of Advection, Dispersion and Chemical

Reactions of Contaminants in Groundwater Systems. Report to the

USEPA, 170 pp.

57. Zheng, C. and Wang, P.P. (1999), MT3DMS: A Modular Three-

Dimensional Multispecies Transport Model for Simulation of

Advection, Dispersion and Chemical Reactions of Contaminants in

Groundwater Systems; Documentation and User’s Guide, Contract

Report SERDP-99-1, U.S. Army Engineer Research and

Development Center, Vicksburg, MS.

�� !��"��#�$�%&��'��(��)�%�#�� *��+��,��-��.��/� ��0�� 1��2��0��3�0��4��-��456��7��(��89�:��/�0��;�<��)��8<��-�"=��>�-��;�<��)� ��%&��?�� @A� B�*�C��D)�1��%&��#�$� ��8<��456��( ��:��+E��7�F=�)��*��56��7�G)�;��3��:��.��HI�)�� B�J��=��-��3��0��8��K��<�L8��LM��7��65K��!��3��B��C��*��3-��3�1-��'�-L=��HI)�N�5O� ��*��-L�+��,��.��%&��?�� P�B�*�C�� Q��K$��R��$�:�-&��?$��S�B�J��=��:6$�%-T�� ()�4��-��-��L��1��G�� ?�&3��)��B�<I��L��2��-��6��U"��8 ��D��-��V9��W�D��C�K��3*��I�� X�B�0*-��+��.��Y��Z��8��K��3�-I��7��B�,�[��7��(�\]��-��*)�4��-��-O��L�7�*��#�$� �%)�Q��1,�V��1��(�� *��\�3�1 ��-��V9��[��*��<�[��8<��*)��,8�3�� _�V��]&��0*-��B�>�_��?7 ��R��$�;�<��(,$�U*F $�=�� a <J��bO=��%"�� 6�DJ��-��9<LPc�DJ��

�� !�� "�� #��$��%��&��'��(��)�*�+ ,-��(.'� /��0��(�� 1+�(. � ��"�2 3��!�/��0��(��4�� 5�67��8��9��:��8+ ; ��<� ��"��"=��>�?��"��@?�+A� �BC��?� ��4��!�D �E�F��8��&$G�� G��4�� ?��H=��!��< �BC�� G��I��!�JK��"L��H��$��'�� G��(�*�� !�F��5�67 9��&��L��M8��'��(��)�*�+ NO��(.'� �H� ��(��<� ��P�Q��9��: G��!��!�@6G:��(��6<�Q�8+RSTU�'V�W�G��2 3�X �&�+��7 ��&�G��6<��P��4��F�� Y�Z>��[��<� �/RT\] _\�V�@�7�� P�%��)�+ �RTab]�V�X�8�:��@�7��c�6��RUde ea�V� ��"�� BC��@�7�� 7 ��P��P�0*��fZ�>��'��7F�[��&��6<�"��P��g�(��I?��2 3'��)�*�+ ��(��h��(.'� i��(��1:j��(�7��+��F�k��l+ ��XZ>��(��6<��P�)�+ ��[��:�m�8PL��?��[��&��= �� n�g��F�� 8$&��P�@?�E��P�%��o��&�� 7 "��G;W��p��5�67� �m�&�� I��P��6<�6>=�F�+��"�@?�E��8G$��*�� 4��K ��[��&��!��&!��G�!�'�F� ��P��G��c��L��2 3�F�� )�+ ��@�7��Q��9��: G��l+ ��q+=��F��8��F��5�67� �@ +�� @ +��'�[��&��(��2 1��q+=��(��H��E�+5�67� �)�+ ��L��KL�� !�F� [��&��6<�9��:�'��[��&��2 3�F��79�&��L��[��&��K�'��(��64H��q+=�? r�� G8��s�G�� BC��H�tG&��Q��"��!��X�8�:��[��&��(��2 3��2Z�L��X�8�:I��L��[��&��l�.��!�F��q+=��(��6<�Y�Z>'��)�*�+ ��(��N,��(.'� �?��(�� &�+�8��I�"� ��(��[��&��9��:��(��6<�X��[��&��Q��H��?��L��/�"�� *��c��" ��= �� 8PL�@?�+rH�@ BC��%��&��(�*��!� �BC�� "�� !�D=@?�E��G8��8��8!�F�� 8��u�$:��P��"�qPL�.��F4&�H��:��

�� !�"��#�$%��&'��(�")�%��*+��,��*��-&.�'/��)' �/�� 0�-�1.�2&��3� ��4��5� �67�� +��8��9+��:/��;')�<�+��=��>? �@��A*1&�"�� +��&�8��B �CDEFD��G�/)��4HIE��BJ$��$�?�'6�*��@�J��&�( �� +��;2'��?'��2� J+��,��C2�/��&�6�K�89C��J)��*LMFEEE�@NO&�(O@��=P��QL�R.2�� )' �/�� S��&2�*+��T��L�(�*��'J,��>�1�UR*%+�� +��;� V�� +$,��(�� (*W��C�<9��( �� +�� 8�$��+�%+��K�89C7��?�� +,��C2�/��X�1�Y5*,�R*%+�� +��?�=�+� ,� ��3 6�� 6�%�� 0��:�+�� Z��*C��[�9�� \��/)�]��4�A*>*��BW)W�-?' � !��?�� /��?��_ ��W��8��9+��' ��a /��4��\�2'��b'��a!��?�U�)&��cd��-?' ��Z��*C��[�9��R*%+�� +�� =��@�'9�\��+L�?�� % 2�� )',��4��\��K�J��K�*6. ��: +��9 ��/ +J�� (*W��C�<9��(��/��?�=� ��\�2'��e+��K�*6.�@�'9�\�,�Z��*��[�9��a�)�3 6�f�� \�g\?��/:�+,�h: �+��:��K�*6d��i2�1�� +��ej�(.�3 6��: +��9 ��k�+��+��l.��%+1�'J,�=�;�V��(�� ;')�<�+�� +��6*��#.�K�*6d��?��9 ��7'J��L� !��% ��B %��L��;&�)m��4�-&n� K*��6�%��,*�:+�� %�,FoE�=p�A�/�C��iq�+C�?�� (*W�� +��#�)�\�iq�+C�Ar1�� +� , �/��#.��+�� +��+�� +��+�%+��A'J+,FH�@* ��BJ$��l� l� :��#�<9��( �;'+��-*�J��Ds�-&n)�@��\�2'��(��6��/+��/��7�6�-'6��(��;'�<��6*�+��&�'��4��4 �+5��/�� +��*6��4DsE�� !��?��/(d��W�7��(��D=H��BJ$��#* �� 4��6*��')<��#�<9�� tIEE��;&�)m�� %��#* �+��(�f<5FHo�p��4�-&n��L�? �� %�,�� 8�$��;&�)mHE�p��+��c�]��q��:�+�FEEE�@* ��BJ$��=� ��*�(��(��$��[9��)��+ (�?�A*�(��J \��\��#*$��T� A?d��8��'/��8��QL�A?��)��u /�� 8 V�?��\�2'��QL��;*5�+��b�'Ld��?�3� ��v*j*��3� ��QL�A*�(��(��$,�&2?��+��[9��A�4��(�jw�,�b�'Ld��QL=��8��#*$�)?I� ��8>=��C�x��8��K��a��8��QL��(�?�� +�� /��(��J\�?�� 8�1�K��"��(�� +�� 6 �� _ �� !��?�-?'W��?��c�:��yr��\��?�� 5q*+��&�&��\7��7'J�=A?��)��z).��8��

�� !"#$% &'�(�)*��+(��,(�� &-./#01"&-2#341"&5678619:;<&&=>?9@1"&-AB9C<&DEFE & &-2#341"&-.GH&I1%&-B#@B&-1928<&=>?9@1"&-AB9C&>.JKC901"&-C8!&LG$&M6NO1"&P9QGRJB&SB&TUVH& &LW&-2#3?9:.1X8#.41"&X&Y>1"& && &S.3OJ001"&-3V1&SB&#0JA7Z [[[[[[[[[[[[[[[[[[[[[[[[[[[[[\\\\\\\ 86JH#1"&]9J2"_&��a�b��c��d�e�f��d [[[[[[[[[[[[[[[[[[[[[[[[[[[[[\\\\\\\& 86JH#1"&]9J2"&_*g(�c�h��+(�i�jk��d \\\\\\\[[[[[[[[[[[[[[[[[[[[[[[[[[[[[& 86JH#1"&]9J2"_�*g(�c��(��d & &&& & &-.GH&-2#341"<&=>?9@1"&-AB9C& &=U.V1"<&-.;>A1"&>NB&-78640C& DEFD& &

�� !"#$% &'�(�)*��+(��,(�� &-./#01"&-2#341"&5678619:;<&&=>?9@1"&-AB9C<&DEFE & &-2#341"&-.GH&I1%&-B#@B&-1928<&=>?9@1"&-AB9C&>.JKC901"&-C8!&LG$&M6NO1"&P9QGRJB&SB&TUVH& &LW&-2#3?9:.1X8#.41"&X&Y>1"& Z">[%&\O]& )��_��a��_� &&&b9J2c&9:.1X>.41"&X&Y>1"&dKe& -2#341"&-.GH& &-AB9C=>?9@1"& )�f��f��gh� &&&b9J2c&#$9KB&9:.1X>.41"&X&Y>1"&dKe& -2#341"&-.GH&=>?9@1"&-AB9C& &-2#341"&-.GH<&=>?9@1"&-AB9C& &=U.V1"<&-.;>A1"&>NB&-78640C& DEFD& i

�� !"#$% &'�(�)*��+(��,(�� &-./#01"&-2#341"&5678619:;<&&=>?9@1"&-AB9C<&DEFE & &-2#341"&-.GH&I1%&-B#@B&-1928<&=>?9@1"&-AB9C& &-C8!&JG$&K6LM1"&N9OGPQB&RB&STUH>.QVC901"& &JW&-2#3?9:.1X8#.41"&X&Y>1"& &&-2#341"&-.GH<&=>?9@1"&-AB9C& &=T.U1"<&-.;>A1"&>LB&-78640C& DEFD& Z

Assessing Alternatives for Disposal of Reject Brine from Inland Desalination Plants

Documents

Transcript of Assessing Alternatives for Disposal of Reject Brine from Inland Desalination Plants